Verónica Villacorta*a,
César Augusto Barreroa,
María-Belén Turriónb,
Francisco Lafuenteb,
Jean-Marc Grenechec and
Karen Edilma Garcíaa
aSolid State Group, Faculty of Exact and Natural Sciences, University of Antioquia – UdeA, Street 67 N°53-108, Medellín, Colombia. E-mail: vvillacortal@unal.edu.co
bInstitute on Sustainable Forest Management, Department of Agroforestry Sciences, Area of Soil Science and Agricultural Chemistry, University of Valladolid, Palencia, Spain
cInstitut des Molécules et Matériaux du Mans – IMMM UMR CNRS 6283, Université du Maine, Le Mans, Cedex 9, 72085, France
First published on 24th November 2020
Adsorption kinetics models have been used to evaluate the adsorption behaviour of pollutants on different materials but there are no reports for the adsorption of As5+, As3+, Sb3+ and Hg2+ on co-precipitated akaganeite nanoparticles which were previously formed in the presence of these ions. In this research, the performance of pure and co-precipitated akaganeite nanoparticles as adsorbents of As3+, As5+, Sb3+ and Hg2+ in aqueous solutions was evaluated using the nonlinear kinetics models of Langmuir, Lagergren, Ho–McKay, Bangham, Elovich and simplified Elovich. In addition, transmission 57Fe Mössbauer spectrometry was used for the first time to compare the physico-chemical properties of akaganeite before and after the adsorption processes. The results showed that co-precipitated akaganeites had much better adsorption capacities than pure akaganeites. On the other hand, the Sb3+ and Hg2+ were the fastest and slowest pollutants respectively adsorbed on all akaganeites. The kinetics models that best described the experimental data for As3+, As5+ and Sb3+ were those of Elovich and simplified Elovich. For Hg2+, the kinetic model that best described the experimental data was that of Bangham. The 300 K and 77 K Mössbauer spectrometry showed only slight variations in some of the hyperfine parameters for the akaganeites after adsorption.
On the other hand, surface-modified akaganeites have shown great potential for environmental applications.13 In this respect, Deliyanni et al.14 developed an adsorbent based on a nanocrystalline–akaganeite hybrid surfactant, which was synthesized using a cationic surfactant, hexadecyl-trimethylammonium bromide, as a modifier. The samples showed a significantly higher adsorption capacity for arsenate than for pure nanocrystalline akaganeite, as well as fast kinetics, which follows a pseudo second-order rate equation. Moreover, the authors observed that adsorption data could not be fitted to the Langmuir equation due to the heterogeneity of the sorbent surface following the modification. Tufo and co-workers,15 studied the adsorption properties of As5+ on Al for Fe substituted akaganeites. The results indicated that Al-substitution enhanced the reactivity of the surface akaganeites towards As-adsorption. Sun and co-workers,16 reported that Zr4+ substituted akaganeites had better adsorption capacities for As3+ and As5+ than pure akaganeites. Harijan and Chandra,17 developed akaganeite nanorods and akaganeite nanorods decorated with graphene oxide sheets at 5%, 10% and 15% for adsorption of phosphate ions from water at pH 7. The FTIR spectra of β-FeOOH nanorods/graphene oxide showed peaks corresponding to akaganeite and a low intense peak corresponding to graphene oxide. When the authors compared the samples after and before of the phosphate adsorption by this technique, they found additional bands corresponding to the P–O stretching vibration in the HPO42− groups. On the other hand, the adsorption experiments showed the maximum removal capacity for an adsorbent dose of 0.4 g L−1, an initial phosphate concentration of 20 ppm, a time of adsorption of 2 h and a temperature of 30 °C. Under these conditions the maximum removal capacity followed the next order: β-FeOOH nanorods/graphene oxide at 5% > β-FeOOH nanorods/graphene oxide at 10% > β-FeOOH > β-FeOOH nanorods/graphene oxide at 15% > graphene oxide. Deliyanni and co-workers,18 prepared and characterized an akaganeite modified with a cationic surfactant. The material was then used as an adsorbent to remove arsenite ions from aqueous solutions. The surfactant modified akaganeite had a much higher arsenite adsorption capacity than the pure akaganeite, which could be attributed to the presence of the surfactant. The kinetics adsorption follows a second-order rate equation and the isothermal sorption the Langmuir and Freundlich models. The maximum adsorption capacity obtained was found between 98 and 100% at pH 7, 25 °C, an adsorbent dose of 1 g L−1, an adsorbate concentration of 10 mg L−1 and 24 h. The authors concluded that arsenite ions reacted with surface OH− functional groups suggesting that specific adsorption occurred. The same authors observed that the maximum adsorption capacity for pure akaganeite under the same conditions was 135.2 mg g−1. Finally, Deliyanni et al.19 synthetized hybrid surfactant akaganeite nanoparticles which showed an efficiency of removal of As3+ from aqueous solution of 100–120 mg g−1 at pH 7.5, 25 °C and a contact time of 2 hours.
According to the bibliography reviewed, most authors have studied the phenomenon of pollutant adsorption on pure akaganeite but only a few on surface-modified akaganeites. We noted that modifying the surface properties of the akaganeites is a good strategy to improve their adsorption properties. One way to do this is to add surfactants and another is to prepare the samples in the presence of different types of cations. We mentioned above that, in the second case, there are reports of akaganeites prepared in the presence of Al3+ and Zr4+ ions. To the best of our knowledge, there are no reports on the use of akaganeite to adsorb Hg2+. Moreover, there are no studies on the adsorption of Sb3+, As5+, and As3+ on modified akaganeites previously co-precipitated in the presence of these ions. In addition, most of the reported works use the isothermal models to define the sorption capacity of akaganeite and for a small number of them, the kinetics models.
In the present study, pure and co-precipitated akaganeites (in the presence of mercury, arsenic and antimony) were used as adsorbents of Sb3+, As5+, As3+ and Hg2+ in aqueous solutions. To understand the adsorption phenomena, we adjust the experimental data to different nonlinear kinetics models. Moreover, the solid samples before and after adsorption were characterized by 57Fe Mössbauer spectrometry at 300 and 77 K. The results showed that the co-precipitated akaganeites had much better adsorption capacities than pure akaganeites, as well as a faster initial average adsorption velocity. The kinetics models that best described the experimental data for As3+, As5+ and Sb3+ adsorbates onto the adsorbents were Elovich and simplified Elovich. In the case of Hg2+ adsorbate was the Bangham nonlinear kinetic model. Mössbauer spectrometry showed slight variations in some of the hyperfine parameters for the akaganeites after adsorption compared to the parameters for the akaganeites before adsorption.
(1) |
The suspensions were then stirred for 60 minutes at 300 rpm, so that at 1, 3, 6, 9, 12, 20, 40 and 60 minutes, 5 mL aliquots were taken and filtered with a 0.45 μm membrane disc; subsequently, the concentrations of As3+, As5+, Sb3+ and Hg2+ in the aliquots were analyzed by inductively coupled plasma-optical emission spectrometry (ICP-OES) using a Varian 725-ES apparatus.
The temperature of the experiments was controlled at 25 °C, while the pH was about 2. Kolbe and co-workers,8 reported that the best results for the removal of antimonate and arsenate onto akaganeite were achieved under acidic conditions. On the other hand, in the case of acid mine drainage, whose pH can be 2, these pollutants are present, and in nature, mineraloids are the most efficient materials in their decontamination.22
The amount of metal ion adsorbed on the pure and co-precipitated akaganeites at time t, was given by the following equation:
(2) |
Finally, for the adsorption kinetics, the samples were code-named as follows: pAk-Hg, pAk-Sb and pAk-As3+ for pure akaganeites adsorbing Hg2+, Sb3+ and As3+ pollutants, respectively. Ak60Hg–Hg, Ak60Hg–Sb, Ak60Hg–As3+ and Ak60Hg–As5+ for Ak60Hg samples adsorbing Hg2+, Sb3+, As3+ and As5+ pollutants, respectively. Ak60Sb–Hg, Ak60Sb–Sb, Ak60Sb–As3+ and Ak60Sb–As5+ for Ak60Sb samples adsorbing Hg2+, Sb3+, As3+ and As5+ pollutants, respectively. And, finally, Ak60As–Hg, Ak60As–Sb, Ak60As–As3+ and Ak60As–As5+ for Ak60As samples adsorbing Hg2+, Sb3+, As3+ and As5+ pollutants, respectively.
The experimental data were fitted to the nonlinear kinetics models of Langmuir, Lagergren or pseudo order 1, Bangham, Ho and McKay or pseudo order 2, simplified Elovich and Elovich, with the objective of deciding which model (or models) best describe(s) the data. The equations corresponding to these models are showed in Table S1.†
In all adsorption kinetics experiments, the pH values were approximately 2 for adsorbing Hg2+, As5+/As3+ and Sb3+. Using the MINTEQ software,21 and considering the pH values of our experiments, we determined that the predominant chemical forms of Hg2+, As5+, As3+ and Sb3+ were: Hg2+ with a total concentration of 96.5%, H3AsO4 and H2AsO4− with a total concentration of 64.8 and 35.2%, respectively, H3AsO3 with a total concentration of 100%, and finally, Sb(OH)2+ and Sb(OH)3 with a total concentration of 20.6 and 79.4%, respectively. These results are consistent with those reported in the literature.23
In the construction of the adsorption kinetics for Sb3+ and Hg2+, the three points of the same colour corresponded to Q values collected at the same time and for three different wavelengths of the ICP-OES equipment. In the case of antimony, the readings were at 206.834, 217.582, and 231.146 nm, while mercury at 184.887, 194.164, and 253.652 nm. On the other hand, for As3+ and As5+, the Q data at a given time were collected with four different wavelengths corresponding to four readings at 188.980, 193.696, 197.198 and 234.984 nm.
Now, from Fig. 1, it can be seen that the Q values at 40 and 60 minutes are very similar for co-precipitated samples adsorbing Hg2+. This result could indicate that at 60 minutes the experimental maximum adsorption capacity was probably reached. This situation was different for the co-precipitated samples adsorbing As3+, As5+ and Sb3+ whose kinetics showed a slight continuous increase from 40 to 60 minutes. This result probably suggests the existence of a secondary reaction, surface precipitation or ternary adsorption, which can be seen as the formation of a new surface phase or as a multilayer adsorption due to the continuity of the process beyond adsorption.24–27 In the case of arsenate and arsenite adsorption, it is important to stand out that in agreement with Cornell and Schwertmann,28 the kinetics of adsorption on goethite and ferrihydrite seem to involve two stages: initial rapid adsorption followed by slower adsorption, whose data can be adjusted to the Elovich equation, as in this research was also observed.
For comparative purposes the experimental Q value at 60 minutes was correlated with experimental maximum adsorption capacity, Qmax. Table S2† lists the mean experimental Qmax for the different experiments at t = 60 min. Several interesting observations can be noted from these data. The co-precipitated akaganeites had higher adsorption capacities for the pollutants than pure akageneites. In fact, in the case of Hg2+, As3+ and Sb3+, the co-precipitated akaganeites adsorbed between 1.1 to 1.5, 1.4 to 1.8 and 2.2 to 2.6 times more than the pure akaganeite, respectively. This observation was perhaps due to the fact that the co-precipitated samples had smaller mean particle sizes than pure akaganeite (134 ± 19, 20 ± 1, 17 ± 1 and 13 ± 1 nm for pure akaganeite, Ak60Hg, Ak60Sb and Ak60As, respectively).20 Moreover, the higher adsorption capacity can also be due to the fact that the surface of the co-precipitated akaganeites presented a greater heterogeneity motivated by the presence of different active sites. This idea is in agreement with the structural changes observed in the co-precipitated samples compared with the pure sample due to the incorporation of the pollutants to the surface during the synthesis.20
On the other hand, Sb3+ was the best adsorbed pollutant by pure and co-precipitated akaganeites compared to the other contaminants, while Hg2+ was the least adsorbed one. In fact, the experimental maximum adsorption capacities of pollutants by co-precipitated akaganeites varied, from the highest to the lowest values, in the following order: Sb3+ > As5+ > As3+ > Hg2+. A very interesting result is that the highest experimental Qmax values for a given pollutant were found to be independent of the type of co-precipitated akaganeite, i.e. the type of cation used to synthetized it, suggesting that for the adsorption experiments, the morphological properties of the samples are more important than the type of cations used to prepared it. This result is in good agreement with that reported by Tufo and co-workers,15 who found that the morphological characteristics, in comparison with the Al3+ co-precipitated akaganeites, are the most important factors for the adsorption properties.
Now, in order to get an idea of the rate at which the pollutants are adsorbed by the adsorbents, we calculated the average sorption velocity, , in mg g−1 min−1, for all samples from the experimental data, using the following equation:
(3) |
Fig. 6 Kinetics of adsorption of Hg2+ onto pAk, Ak60Hg, Ak60Sb, and Ak60As. The experimental data (blue circles) are modelled with Bangham (orange) nonlinear models. |
In general, on the basis of a simple visual examination of Fig. 3–6, and on the E values reported in Table S3,† it can be seen that the experimental data are best described for Hg2+ by the Bangham nonlinear kinetic model, and for As3+, As5+ and Sb3+ by Elovich and simplified Elovich nonlinear kinetics models.
According to,29 the Bangham model is an internal diffusion model, meaning that the slowest stage of adsorption refers to diffusion of the adsorbate to active sites located on the adsorbent surface. The values of Qmax obtained from the Bangham equation, 4.2, 4.6, 5.4 and 6.1 mg g−1 to pAk–Hg, Ak60Hg–Hg, Ak60Sb–Hg and Ak60As–Hg, respectively, were very close to those obtained from the experimental Hg adsorption data (see Table S2†).
On the other hand, for Sb3+, As5+ and As3+ the Elovich and simplified Elovich models should be highlighted; indeed, in this case both models are adsorption models in which the slowest step of adsorption is the adsorption itself. In addition, chemisorption prevails over physisorption, and an energetic heterogeneity characterize the surface.29 It is important to recall that α and β parameters refer respectively to the initial sorption rate and an energy activation constant for chemisorption. In Table S3,† it can be observed that the values of α (mg g−1 s−1) increased for co-precipitated akaganeites adsorbing Sb3+, As5+ and As3+ compared with the pure sample. These observations are in agreement with the experimental initial mean sorption rates for Sb3+, As5+ and As3+ adsorption calculated by eqn (2) (see Fig. 2), which changed from the high to the low values in the following order Sb3+ > As5+ > As3+; and with the observation that the Ak60As was the adsorbent on which all adsorbates were adsorbed the fastest during the first minutes (Fig. 2). Similarly, the values of β (g mg−1) decreased in all cases compared with those of the pure sample.
The observations found in the evolution of α and β parameters indicated that As3+, As5+ and Sb3+ probably accessed to the surface of co-precipitated akaganeites easier than to the surface of pure akaganeites, due to the increase of the α parameter and to the decrease of the β parameter compared with those of the pure sample. Some factors that probably influenced the values of α and β parameters are the mean particle size, that was lower for co-precipitated samples compared with pure akaganeite, the heterogeneity of the structure and the different surface active sites.
The agreement between the experimental data and the theoretical predictions based on the Elovich and simplified Elovich models confirms the heterogeneous sorption mechanism that could be responsible for the adsorption of As3+, As5+, and Sb3+ ions. This result is in agreement with the idea of the existence of different adsorption sites, which in the case of akaganeite would refer to tunnel sites and surface sites;16 such as hydroxyl groups28 and in our case the adsorbed As5+ and Sb3+ sites.
On the other hand, the incorporation of the adsorbed species depends fundamentally on the physicochemical characteristics of the adsorbent surface, such as the heterogeneity, specific surface area, the coordination sites, density and arrangement of the functional groups and the surface charge, among others.26,28,30 Moreover, the ionic environment and pH of the solution in contact, the size and nature of adsorbate and the duration of the adsorption, also influence the adsorption process.26,28,30 Here, it is important highlight the relation among the surface charge and the ionic species of the pollutants in solution. According to MINTEQ software, at pH 2 the predominant chemical forms of Hg2+, As5+, As3+ and Sb3+ are: Hg2+ with a total concentration of 96.5%, H3AsO4 and H2AsO4− with total concentrations of 64.8 and 35.2%, respectively, H3AsO3 with a total concentration of 100%, and finally, Sb(OH)2+ and Sb(OH)3 with total concentrations of 20.6 and 79.4%, respectively. In agreement with Deliyanni and co-workers,31 the pzc of akaganeite is approximately 7.3, so below this pH the surface of the akaganeite is positively charged. In this way, the adsorption of As5+ as H2AsO4− could be favoured. However, according with the results obtained in the adsorption kinetics models, the chemisorption prevailed over physisorption and in consequence the formation of internal sphere complexes probably changed the point zero charge since it involved the formation of specific chemical reactions between the surface and the adsorbates.
To complement this information, the free energy of adsorption can be defined as the sum between the chemical energy of adsorption and the electrostatic component: ΔGadsorption = ΔGchemical + ΔGcoulombic.28 In this way, when the affinity between the surface of the material and the ions present in the solution is very high, the chemical component of the adsorption energy predominates, and in consequence the generated bonds have a strong covalent character, causing internal sphere complexes. This situation originates that the adsorption of neutral species or with the same charge of the surface can take place.25,28,30
Finally, the experimental data were less well adjusted to the Lagergren model, which is also an adsorption model where the slowest stage of the process is that of adsorption itself.29 In this model, the saturated monolayer is considered as the maximum adsorption.29 From this result, it could be thought that in adsorption of As3+, As5+, and Sb3+ a multilayer adsorption occurred, due to the loss of the plateau and to the increase of the adsorption (see Fig. 1).
Fig. S1–S3† show the 300 K Mössbauer spectra of samples pAk, Ak60Hg, Ak60Sb, and Ak60As after adsorption of Hg2+, As3+ and Sb3+. All spectra were first correctly fitted by introducing two quadrupolar doublets named D1 and D2 and the refined values of hyperfine parameters are listed in Table S4.† It is noted that, within the error bars, the hyperfine parameters (isomer shifts and quadrupole splittings) of the two ferric iron sites for all samples are quite similar. Now, samples pAk and pAk-As3+ showed similar relative spectral areas. On the contrary, the relative spectral areas of the two components for the samples Ak60Hg, Ak60Sb and Ak60As without and with adsorbed pollutants showed some differences. It is important to note that other fitting models can be considered assuming free values of linewidth and/or absorption areas, and/or some constraints: as the hyperfine structure of the quadrupolar spectra is poorly resolved, the main relevant hyperfine parameters are the mean values which are independent of the fitting models. As shown in Table S6,† the mean values of isomer shift and quadrupolar splitting are rather similar.
As shown in Fig. S4–S6,† all spectra obtained at 77 K exhibit rather similar complex hyperfine structures resulting from magnetic sextets composed of asymmetrical lines. The initial fitting procedure consists of four sextets named S1, S2, S3 and S4. These components arise from the existence of four non-equivalent iron sites related to the two monoclinic iron sites (Fe1 and Fe2) in the vicinity of occupied (Cl) or unoccupied (VCl) chloride sites.20 Therefore, S1, S2, S3 and S4 are assigned to Fe1–Cl, Fe2–Cl, Fe1–VCl and Fe2–VCl, respectively. During fitting, the values of the isomer shift, quadrupolar shift, linewidth and absorption area for components S1 and S2 on the one hand, and those for components S3 and S4, on the other hand, can be constrained independently or not, either to be equal, or adjustable. In addition, the values of the hyperfine field were refined without constraint. Finally, many fitting models can be well realized with various sets of hyperfine parameters, leading to similar factors of goodness. At this stage, it remains difficult to establish which of these different models is the best solution. Table S5† shows an example of the values of the hyperfine parameters at 77 K characteristic of the samples pAk, Ak60Hg, Ak60Sb and Ak60As after adsorption of Hg2+, As3+ and Sb3+. Therefore, as in the case of quadrupolar doublets, we need to consider the mean values of each hyperfine parameter, including their respective error bars. They are also given in Table S7;† it is now difficult to establish some significant correlation between such these hyperfine parameters, the nature and content of contaminant, but they show unambiguously the presence of ferric species. The lack of strong evolution can be explained first of all by the low content of contaminant which must preferentially deform the superficial structure, causing thus a small effect.
The results found here suggest that in some cases the adsorption of pollutants by akaganeite slightly modifies certain hyperfine parameters. The isomer shift is the least sensitive parameter. To understand the Mössbauer results, it is worth discussing three aspects: first, how the atomic environment surrounding the Fe3+ ion and its molecular bonding with the ligands affect the hyperfine interactions; second, the relative abundance of iron ions at the surface of akaganeites; and third, the types and distribution of active sites for adsorption. With respect to the first aspect, the p–d electron population of the Fe ion affects the magnitude of the quadrupole shift and the hyperfine magnetic field. The population of this valence shell is affected by the molecular bonding of the iron with its neighbors. In akaganeite, Fe directly bonds to oxygen in the octahedral site. Oxygen share bonding with pollutants located at the surface and with hydrogen, which also bonds to chloride into tunnels and to pollutants at the surface. In summary, the slight variation in some of the hyperfine parameters could be partially explained by the fact that the adsorption of the pollutants at the surface and at tunnel sites, indirectly affects the Fe 3d orbital population. Now, with regard to the second aspect, the small variations in the hyperfine parameters can also be understood if we take into consideration that the adsorption process is mainly a superficial phenomenon, i.e. the pollutants are adsorbed at surface sites, and in the case of akaganeite, also tunnels can be also part of the adsorption sites. Here it is interesting to note that,20 we studied the relative abundance of the surface iron ions with respect to the core irons for the different akaganeites. We found that for a shell thickness of x = 0.2 nm, the shell-volume to particle-volume ratio accounts for 1%, 10%, 14% and 22% for pAk, Ak60Hg, Ak60Sb and As60Ak, respectively. Thus, there are more irons located at the core than at the surface of the akaganeites, therefore explaining the very small variations observed in the hyperfine parameters. Finally, and concerning the third point, it is important to take into account that not all the surface sites available in akaganeite are active for adsorption. In this respect, Song and Boily,32 theoretically and experimentally determined the identity and distribution of the adsorption sites at the surfaces of the akaganeite. They reported that reactions occurring on surfaces sites may involve hydroxyl, oxygen and water groups. In pure akaganeites, terminal (001) and (100) planes expose a mixture of the three types of hydroxyl groups (singly, doubly and triply coordinated to iron). On the other hand, terminal (010) plane exposes a mixture of hydroxyl, oxygen and water groups. Not all of these sites are active for adsorption and their relative abundance is different.
Footnote |
† Electronic supplementary information (ESI) available. See DOI: 10.1039/d0ra08075f |
This journal is © The Royal Society of Chemistry 2020 |