CC BY-NC-ND
0Arabian Journal of Chemistry (2015) xxx, xxx-xxx
King Saud University Arabian Journal of Chemistry
www.ksu.edu.sa www.sciencedirect.com
ORIGINAL ARTICLE
Efficient treatment of lead-containing wastewater by hydroxyapatite/chitosan nanostructures
Ahmad M. Mohammad a,b *, Taher A. Salah Eldin c *, Mohamed A. Hassan c, Bahgat E. El-Anadouli a
a Chemistry Department, Faculty of Science, Cairo University, PO 12613 Giza, Egypt
b Department of Chemical Engineering, Faculty of Engineering, The British University in Egypt, PO 11837 Cairo, Egypt c Nanotechnology and Advanced Materials Central Lab, Agriculture Research Center, Giza, Egypt
Received 12 June 2014; accepted 14 December 2014
KEYWORDS
Adsorption isotherm; Wastewater treatment; Lead removal; Hydroxyapatite; Chitosan
Abstract The development of hydroxyapatite nanorods (nHAp) and hydroxyapatite/chitosan nanocomposite (nHApCs) was sought as potential sorbents for the removal of lead ions from aqueous lead-containing solutions in a batch adsorption experiment. The high resolution transmission electron microscopy, energy dispersive X-ray analysis, X-ray diffraction, Fourier transform infrared spectrophotometry and Zeta potential measurements were all combined to reveal the morphology, composition, crystal structure, functionality and stability of the prepared sorbents. The equilibrium concentration of Pb2+ ions was identified by the atomic absorption spectrophotometry. The kinetics of the sorption process was investigated together with the influence of initial lead ions concentration, sorbent dosage and solution pH on the sorption capacity. The sorption process followed pseudo-second-order kinetics, where 20 min was quite enough to attain equilibrium. Two models of adsorption isotherms (Freundlich and Langmuir) were employed to correlate the data in order to understand the adsorption mechanism. Interestingly, in one of the experiments, for a 200 mL solution (pH = 5.6) containing 100 ppm lead ions, a sorbent dosage of 0.4 g nHAp could achieve a complete removal for lead ions. However, typically, the sorption capacities of nHAp and nHApCs to lead ions were 180 and 190 mg/g respectively, which appear excellent for lead removal. © 2015 Production and hosting by Elsevier B.V. on behalf of King Saud University. This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/).
Corresponding authors at: Chemistry Department, Faculty of Science, Cairo University, PO 12613 Giza, Egypt (A.M. Mohammad). Tel.: +20 100 1030534; fax: +20 2 35713250.
E-mail addresses: ammohammad@cu.edu.eg (A.M. Mohammad), t1salah@hotmail.com (T.A. Salah Eldin), mohamedali656@gmail. com (M.A. Hassan), bahgat30@yahoo.com (B.E. El-Anadouli). Peer review under responsibility of King Saud University.
1. Introduction
The violent spreading of kidney and liver failure, cancer as well as the common waterborne diseases such as the bacterial diarrhea, hepatitis, and typhoid fever has called for a sincere urgent treatment for wastewater emerged from sewages, industrial effluents and stormwater before draining into lakes, groundwater and/or rivers. Unfortunately, the existence of large dosages of heavy metals in these effluents represents
http://dx.doi.org/10.1016/j.arabjc.2014.12.016
1878-5352 © 2015 Production and hosting by Elsevier B.V. on behalf of King Saud University.
This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/).
the greatest challenge in water purification, where exceeding the allowed concentration limit of these metals in the human body can end up with acute or chronic death. Of these heavy metals, lead is recognized as longstanding environment contaminant. It is released into the atmosphere from mining industries and fossil fuel burning, and has long been involved in the production of batteries, ammunition, metal products, and devices to shield X-rays (Mudipalli, 2007). Like other heavy metals, lead may contaminate surface and drinking water from industrial effluents, and the corrosion of household plumbing systems, respectively, which ultimately makes drinking risky (Hammer and Mark, 1996). Lead affects the central nervous system and inhibits the ability to synthesize red blood cells. Blood concentrations of lead above 40 ig/dL can produce observable clinical symptoms in domestic animals. It can also cause miscarriages and subtle abortions. According to the US environmental protection agency (EPA), a diet of 2-8 mg of lead per kilogram of bodyweight per day over a period of time will cause death (Mudipalli, 2007). Therefore, the development of proper treatments reducing the lead content in wastewater is essential in order to avoid lead poisoning.
Several approaches have been employed to remove heavy metal ions from wastewater such as the chemical precipitation, adsorption, cations-exchange, reverse osmosis, electrodialysis, and electrochemical reduction (Fu and Wang, 2011). However, generally, the adsorption process has been proved convenient in terms of cost, simplicity and flexibility (Rengaraj et al., 2002). In this regard, several sorbents have been recommended for the removal of Pb2 + ions, such as the activated carbon, and mesoporous and nanoporous materials as clays, zeolites, chitosan and apatite (Ziagova et al., 2007). Of these sorbents, calcium hydroxyapatite (Ca10(PO4)6(OH)2, abbreviated as HAp), has shown a remarkable sorption efficiency for long-term containments (Sun et al.,
2013). Recently, biological molecules such as chitin, chitosan, and lignin have also been recommended for the removal of toxic metals (Shen et al., 2013). Chitosan biopolymers (Cs), in particular, with their non-toxicity, hydrophilicity, biocom-patibility, and biodegradability, exhibited effective sorption efficiency for heavy metals (Shen et al., 2013). In this investigation, artificially-made calcium hydroxyapatite nanorods (nHAp) and hydroxyapatite/chitosan nanocomposite (nHApCs) have been employed in the removal of lead ions from lead-containing aqueous samples using the adsorption technique. Several parameters including the adsorption time, initial Pb2+ concentration, adsorbent dosage, and pH have been investigated to evaluate their influence on the sorption efficiency.
2. Materials and methods
2.1. Sorbent preparation
All chemicals used in this investigation were of analytical grade and used without further purification. Calcium hydroxyapatite nanorods (nHAP) were synthesized by the sol gel method according to the procedure described earlier (Salah et al.,
2014). Typically, 100 mL 0.479 M H3PO4 (Sigma-Aldrich No. 04107) was dropwisely added (3 mL/min) to a 100 mL 0.8 M solution of Ca(NO3)2-4H2O (Sigma-Aldrich No. 31218) in order to achieve a Ca/P stoichiometric ratio of 1.67. The pH was adjusted at 10 ± 0.05 using ammonia
solution (33%). The following equation represents the reaction involved therein at room temperature.
10Ca(NO3)2 + 6H3PO4 + 20NH4OH
! Ca!e(PO4)6(OH)2 # +20NH4NO3 + 18H2O (1)
The suspension was left under moderate stirring for 16 h, and after 24 h of aging at room temperature, the precipitate was rinsed with deionized water and dried at 80 0C under vacuum.
Similarly, the hydroxyapatite/chitosan nanocomposite (nHApCs) was prepared by adding 0.479 M H3PO4 dropwisely to 0.8 M solution of Ca(NO3)2-4H2O containing 0.25 g chitosan (Cs, Sigma, No. 48165) in order to keep the stoichiometric ratio of Ca/P at 1.67, and similar procedure was followed up but the precipitate drying was achieved at 60 0C. By the way, Cs was initially dissolved in acetic acid (HAc) before mixing with calcium nitrate solution.
In acid medium Cs - NH2 + HAc ^ Cs - NHj" + Ac- (2) In Basic medium Cs - NH+ + OH-! Cs - NH2 #+H2O (3)
2.2. Materials characterization
The morphology, crystal structure, functionality and stability of the prepared sorbents were revealed using a set of advanced techniques. The high resolution transmission electron microscope (HR-TEM, Tecnai G20, FEI, Netherland) supported with an energy dispersive X-ray (EDX) unit was used for the sake of imaging and crystal structure revelation. Two different modes of imaging were employed; the bright field at 200 kV using LaB6 electron source gun and the diffraction pattern imaging. Before imaging, the HAp particles were deposited from a dilute aqueous suspension onto Cu grids with the support of a carbon film. The crystal structure of the prepared sorbents was identified using X-ray diffraction (XRD) (X'Pert PRO PANalytical, Netherland), which operated at 45 kV and 30 mA using filtered Cu Ka radiation (k = 1.5406 A) in the 2h range from 50 to 800 and high score plus software. In order to investigate the sorbent' functionality, Fourier transform infrared spectroscopy (FTIR, Jasco 6100 - Japan) was conducted in KBr pellet at room temperature over the range from 400 to 4000 cm-1. Zeta potentials (Zetasizer Nano S, Malvern Instruments, UK) of nHAp and nHApCs aqueous suspensions of different pH (2-11) were measured to evaluate the point of zero charge (PZC) of these colloidal suspensions. Before measurements, nHAp and nHApCs were dispersed in deionized water (Milli-Q Millipore, Billerica, MA, USA) and the pH was adjusted using NaOH and HCl, and the suspension was left under sonication (Ultrasonicator, SB-120DTN, Taiwan) for 10 min.
2.3. Sorption experiment
2.3.1. Determination of the sorption capacity The batch equilibrium technique was employed at room temperature. Lead ions solutions with different concentrations were prepared by dissolving Pb(NO3)2 (Eastern fine chemicals, Italy) in deionized water. Typically, a certain amount of nHAp or nHApCs is added to 200 mL of the Pb2+ solutions after adjusting the pH. Next, the mixture was shaken (Shaker, Lab-line 4625-1CE, USA) at a speed of 300 rpm for a given
time. The sorbent (nHAp, nHApCs) was next filtered from the solution by a 0.2 pm syringe filter. After that, the Pb2+ ion concentration was measured by atomic absorption spectropho-tometry (Analytik-jina Zeenit 700p, Germany).
The equilibrium sorption capacity, qe, of nHAp and nHApCs, which is defined by the equilibrium amount (mg) of Pb2 + ions adsorbed per unit mass (gram) of sorbent (simply the unit will be noted as mg/g), was calculated using the general equation:
(C0 - Ce)V
qe =■
where C0 and Ce represent, respectively, the Pb2+ concentrations before and after (equilibrium) adsorption (mg/L), V is the volume of the Pb2 + solution (L) and M is the amount (g) of the sorbent used in the reaction mixture.
2.3.2. Effect of contact time
The influence of contact time on the sorption capacity was studied by adding 0.1 g of the sorbent (nHAp, nHApCs) to 200 mL of a Pb2+ solution (100 mg/L) and the mixture was shaken at room temperature for a certain time (5-90 min). The regular procedure of filtration and analysis was next employed to calculate the sorption capacity.
2.3.3. Effect of initial Pb2+ ion concentration
The influence of the initial Pb2+ ions concentration on the adsorption kinetics was investigated by mixing 0.1 g of nHAp with aqueous Pb2+ ions solutions of various concentrations (100, 200, 300, 400 and 500 mg/L) and the regular procedure of shaking, filtration and analysis was followed. The data were fitted according to the Freundlich and Langmuir adsorption isotherms, which are often used to describe the adsorption of solutes from aqueous solutions.
2.3.4. Effect of pH and sorbent dosage
Finally, the effect of pH in the range of 3-11 and the sorbent dosage from 0.025 to 0.4 g on the sorption capacity was investigated in a regular batch equilibrium adsorption experiment.
3. Results and discussion
3.1. Sorbent characterization
We have previously reported a detailed description for the analysis of nHAp and nHApCs. Shortly, the TEM imaging (Fig. 1A) of nHAp particles prepared as described earlier in Section 2.1 indicated the crystallization of nHAp in thin (ca. 5 nm in diameter) and short (20-50 nm in length) nanorods with minor aggregations. The electron diffraction pattern (inset of Fig. 1A) of nHAp showed polycrystalline diffraction rings corresponding to the crystallographic planes (002), (1 02), (1 3 0), (202) and (211) of nHAp crystals. On the other hand, the EDX analysis (Fig. 1B) was consistent with the starting relative ratio of calcium to phosphorus (1.67) in nHAp. The XRD pattern (Fig. 1C) of nHAp agreed well with the typical one of HAp (JCPDS no. 01-073-8417) with lattice parameters a « b « 0.9410 nm, c « 0.6879 nm and a space group P63/m. This indicates the deposition of nHAp in a hexagonal structure, which is believed stable even after the composite's
formation. The d-spacing between adjacent lattice planes of nHAp and nHApCs is displayed in Table 1. Alternatively, two well-defined signals have been appeared for Cs at 2h = 20.88 and 10.58. The signals appeared at nHApCs were very much similar to those of nHAp except in peak height and width. The disappearance of signals characterizing Cs in nHApCs was attributed to the low doping level of Cs in the composite, which agreed with previous expectation for certain HAp/Cs ratios (Danilchenko et al., 2009; Nikpour et al., 2012; Salah et al., 2014). Fig. 1D comparing the FTIR spectra of nHAp, Cs and nHApCs, interestingly, reveals a direct evidence confirming the formation of nHApCs composite (Salah et al., 2014). Generally, all signals of nHApCs coincided to a great extent with the major signals of nHAp (see Table 2). However, the signals at 2900 and 1388 cm-1 corresponding to stretching and bending (-CH) group appeared only in Cs and nHApCs but disappeared from nHAp, which verify the nHApCs complex formation. Furthermore, the Zeta-potential measurements (Fig. 2) pointed out a PZC at pH of 7 and 8, respectively, for nHAp and nHApCs, which further infers the nHApCs composite formation.
3.2. Sorption study 3.2.1. Effect of contact time
The pseudo-second-order reaction kinetics was used to determine the sorption rate constant.
qt kqe qe
where qt represented the sorption capacity (mg/g) at any given time (t, min), qe was the equilibrium sorption capacity (mg/g) and k was the second order reaction rate constant of adsorption (g/mg min). The following expression denoted the sorption rate h (mg/g min) (Azizian, 2004):
h — kq]
Fig. 3A, displaying the adsorption isotherm of Pb2+ ions (pH = 5.6) by nHAp and nHApCs, indicates a quick removal of Pb2+ ions within the first 20 min of adsorption. Next, the kinetics of adsorption was slower until attaining a state of equilibrium. Generally, we noticed an obvious increase in the sorption capacity of nHApCs over nHAp, which may be influenced by the nature of bonding between Pb2+ ions and the different sorbents. The PZC of nHAp and nHApCs suggests carrying positive charges at pH lower than 7, which support the predomination of the cation exchange mechanism. The difference here is the existence of numerous amino groups in nHApCs (embedded in Cs), which are reactive sites for metal ions. These amino groups may further launch another route of interaction (via chelation) with Pb2+ ions, which can increase the Pb2+ ions uptake by nHApCs. Fig. 3A also indicates that a contact time of ~65 min was sufficient to ensure saturation of nHAp and nHApCs with Pb2+ ions.
In order to understand the kinetics of Pb2+ ions sorption by nHAp and nHApCs, not only the sorption process but the diffusion steps involved should be considered. The diffusion steps involve the mass transfer of Pb2+ ions from the bulk of the Pb2+ ions-containing solution to the active surface sites of nHAp and nHApCs, which are capable to interact with Pb2+ ions. The intra-particle diffusion of Pb2+ ions within
400 350 300 250 200 150 100 50 0
Chitosan
nHApCs
45 2 Theta
.O Energy (keV)
400 900 1400 1900 2400 2900 S400 3900
Wave number (im4)
Figure 1 (A) TEM image of nHAp prepared as described earlier in Section 2.1. The inset of (A) shows the electron diffraction pattern of nHAp, (B) the EDX analysis of nHAp, (C) X-ray diffraction pattern and (D) FTIR of nHAp, Cs, and nHApCs respectively.
Table 2 Characteristic FTIR nHAp and nHApCs. wavenumbers of Chitosan,
Characteristic group Chitosan nHAp nHApCs
(cm-1) (cm-1) (cm-1)
-NH (stretching) 3421 - -
-CH (stretching) 2900 - 2900
-CH (bending) 1388 - 1388
C-N 1643 - -
C-O-C (stretching) 1039 - 1039
-OH (stretching) - 3450 3450
-OH (bending) - 1643 1643
-PO4 - 1036 1036
- 865 865
- 602 602
- 566 566
Table 1 XRD lattice parameters of nHAp and nHApCs
sorbents.
2 theta (deg.) d-spacing (A) h k l
25.88 3.439 0 0 2
31.88 2.811 2 1 1
32.28 2.777 1 1 2
34.08 2.628 2 0 2
39.78 2.260 1 3 0
49.58 1.839 2 1 3
Figure 2 Zeta-potential measurements for nHAp and nHApCs suspensions at different pHs.
the adsorbent might limit the rate of the adsorption process. Several parameters including the physical nature, composition, crystallinity of the adsorbent as well as temperature and pH can be critical in identifying the sorption mechanism. In this investigation, the linear form of pseudo-second-order kinetic model was employed to test the batch experimental data of Pb2+ ions adsorption by nHAp and nHApCs, (see Fig. 3B). The second-order constants (k, qe, and h) for nHAp and nHApCs were evaluated from Eqs. (5) and (6) and listed in Table 3. Interestingly, the pseudo-second-order kinetic model was perfect in fitting the experimental results, where a close
t (min) t (min)
Figure 3 (A) Effect of contact time on the sorption capacity of Pb2+ ions (pH = 5.6) onto nHAp and nHApCs and (B) the linear fitting of the experimental data using the pseudo-second-order kinetic equation (Eq. (5)).
Table 3 The parameters obtained from fitting the sorption data of Pb2+ ions onto nHAp and nHApCs using the pseudo-second-order equation (Eq. (5)). h = kq] and R2 is the regression coefficient for the linear plot.
Sorbents k (g/mg min) qe (mg/g) h (mg/g min) R2
nHAp 0.0013 192.30 46.30 0.9977
nHApCs 0.0020 196.10 75.80 0.9989
Figure 4 Effect of initial Pb2+ ions concentration on sorption capacity of nHAp sorbent to Pb2+ ions (pH = 5.8, adsorbent dosage 0.1 g).
matching was attained between the experimental and calculated results. This indicates that the rate of Pb2 + ions adsorption onto nHAp and nHApCs has a second-order mechanism. The sorption capacities of Pb2 + ions by nHAp and nHApCs were compared to other sorbent materials (see Table 4) and the comparison showed the excellence of nHAp and nHApCs for Pb2 + ions removal (Bereket et al., 1997; Gunay et al., 2007; Gharaibeh et al., 1998; Goel et al., 2005a,b; Gupta et al., 2001; Li et al., 2005; McLellan and Rock, 1988; Rivera-Utrilla et al., 2003; Wang et al., 2007).
3.2.2. Effect of initial Pb2+ concentration
We have so far calculated the Pb2 + ions sorption capacity of nHAp but at a single initial concentration (100 mg/L) of Pb2 + ions. As we know, the initial concentration of Pb2 + ions may be different in wastewater samples and this may influence the sorption kinetics. To investigate this, a series of several
solution containing different concentrations of Pb2+ ions was prepared and the sorption capacity of 0.1 g nHAp for Pb2 + ions was evaluated. Fig. 4 indicated a significant increase in the sorption capacity of nHAp for Pb2+ ions with the initial Pb2 + ions concentration. When the initial Pb2 + ions concentration increased from 100 to 500 mg/L, the uptake capacity of nHAp increased from 182.3 to 251.2 mg/g. A reasonable explanation for this phenomenon may originate from the
Table 4 A comparison of adsorption capacity (qe) for Pb2+ ions by several sorbents at room temperature.
Sorbents Adsorption capacity, qe (mg/g) Reference
Olive mill product 21.56 Gharaibeh et al. (1998)
GAC saturated with bacteria 26.40 Rivera-Utrilla et al. (2003)
Carbon aerogel 34.72 Goel et al. (2005a)
Carbon nanotubes 12.41 Li et al. (2005)
Sphagnum moss peat 19.90 McLellan and Rock (1988)
Red mud 64.79 Gupta et al. (2001)
Bentonite 15.38 Bereket et al. (1997)
Modified activated carbon 29.44 Goel et al. (2005b)
Natural clinoptilolite 80.93 Gunay et al. (2007)
Pretreated clinoptilolite 122.40 Gunay et al. (2007)
Manganese oxide-coated carbon nanotubes 78.74 Wang et al. (2007)
nHAp 192.30 Present work
nHApCs 196.10 Present work
diffusion constrains of the Pb2+ sorption process. At low initial Pb2+ ion concentration, the ratio of Pb2+ to the number of available adsorption sites or ions exchange site is small, which leads to mass transfer resistance between the aqueous and solid phases thus decreased the uptake. A higher initial concentration provided an important driving force to overcome the mass transfer resistances of the pollutant thus increased the uptake (Aksu and Tezer, 2005).
The Freundlich isotherm proposes an empirical model for adsorption on heterogeneous surfaces with the form:
ln qe = ln kf + (1 ) ln Ce
where qe is the equilibrium sorption capacity (mg/g), Ce is the equilibrium concentration of metal ion (mg/L), and (kf and n) are the Freundlich isotherm constants (Bhattacharyya and Gupta, 2007).
On the other hand, the Langmuir isotherm assumes a monolayer adsorption of the adsorbate onto a finite number of identical active sites of the adsorbent surface (Langmuir, 1916). Mathematically, the model is expressed by the following equation:
where qe is the equilibrium sorption capacity (mg/g), Ce is the equilibrium concentration of metal ion (mg/L), qmax (mg/g) is the maximum sorption capacity, and b (L/mg) is the Langmuir constant, which correlates to the energy of adsorption (Langmuir, 1916).
Kinetically, the Freundlich isotherm (Eq. (7)) was used to address the dependence of the Pb2+ ions uptake on the initial Pb2+ concentration (see Fig. 5A). The Freundlich isotherm did not perfectly fit the experimental data, and the parameters, kf and 1/n, were 159.20 and 0.09, respectively. The small numerical value of 1/n (<1) reveals a physical nature for the bonding between adsorbent and adsorbate (Lin et al., 2009). On the other hand, the Langmuir isotherm, considering a monolayer adsorption process of molecules on solid surfaces, offered a better fitting for the experimental data of Pb2+ ions adsorption on nHAp. The model basically assumes the existence of a certain number of active sites on the adsorbent to which the same number of particles bind (1:1 ratio). In addition, the model ignores the dependency of occupying a specific site on the status of the adjacent site. Furthermore, it describes adsorption
processes where no interaction between the sorbate species occurs on sites having the same sorption energies regardless the surface coverage. According to this model (see Eq. (8) and Fig. 5B), the maximum sorption capacity, qm, was estimated as 250.0 mg/g. The Langmuir constant, b, was 0.120 L/mg, which indicates a high sorption energy between Pb2 + and nHAp. One of the reasons for the great sorption affinity of Pb2 + ions on nHAp may originate from the electrostatic nature of bonding between Pb2 + and nHAp. Actually, the surface of nHAp is rich with negative entities as OH- and PO3- which act as hard Lewis bases. On the other hand Pb2 + ions act as a borderline Lewis acid. Therefore a strong interaction derives the intensive uptake.
3.2.3. Effect of pH
The influence of pH on the sorption process of Pb2 + ions on nHAp was studied in the pH range from 3 to 11 (see Fig. 6). The pH can influence the adsorption kinetics and the sorption capacity of Pb2 + ions by changing the surface charges of nHAp, the Pb2 + complex type (Faur-Brasquet et al., 2002; Machida et al., 2005), and the number of active sites available for the adsorption process (Krivosheeva et al., 2012). In this investigation, the highest Pb2+ ions uptake by nHAp was achieved at high pH values. This can likely be attributed to a competition between H+ and Pb2 + ions to be adsorbed on nHAp, where at lower pH values, H + ions are able to exclude a significant number of adsorption sites at nHAp from the Pb2 + adsorption process (Doyurum and Celik, 2006; Gohari
Figure 6 Effect of pH on the adsorption of Pb2+ ions (initial concentration = 100 mg/L) by 0.1 g of nHAp.
0 —.—.—.—.—r—.—.—.—.—I—.—.—.—.—I—.—I—.—.—I—.—.—.—.—F 0
0 0.1 0.2 0.3 0.4 0.5
nHAP(g)
Figure 7 Effect of nHAp dosage on the removal of Pb2 + ions (initial concentration 100 mg/L, pH = 5.8).
et al., 2013). At high pH values, the H + ions competition disappears and the positively charged Pb2+ and Pb(OH) + ions can easily attach to the free binding sites, increasing the Pb2+ ions uptake (Gohari et al., 2013). It is known that precipitation plays a major role in removing of Pb2 + ions in alkaline media (Srivastava et al., 2006). The dominant Pb2 + species at pH higher than 8.0 is Pb(OH)2. Therefore, it can be assumed that Pb2+ removal by nHAp was dominantly controlled by adsorption at pH values between 6.0 and 8.0, but it could be slightly enhanced by lead hydroxide precipitation at pH higher than 8.0 (Mobasherpour et al., 2011).
3.2.4. Effect of adsorbent dosage
The effect of nHAp dosage on the sorption capacity was next investigated and plotted in Fig. 7. As expected, the Pb2 + sorption increased rapidly with the nHAp dosage, as a consequence of increasing the adsorption sites available for the Pb2+ ions removal. Interestingly, starting with 0.4 g of nHAp could achieve a Pb2 + removal efficiency of ~99.7%. Although the Pb2+ removal efficiency increases with the nHAp dosage, the sorption capacity, which normalizes the Pb2 + uptake to the nHAp dosage, decreases (Bhattacharyya and Guptaa, 2007). Therefore, one should customize the process to obtain the highest possible removal efficiency for Pb2+ ions with the least amount of the nHAp dosage.
4. Conclusion
This investigation highlights the effectiveness of nHAp and nHApCs sorbents for the removal of Pb2+ ions from aqueous lead-containing solutions to be used in wastewater treatments. The adsorption process was investigated as a function of the contact time of Pb2+ ions with the sorbents, initial Pb2+ ions concentration, pH and sorbent dosage. The kinetics of the sorption process could be fitted to a pseudo-second-order reaction model, and the sorption capacity of Pb2+ by nHAp increased with the initial Pb2+ ions concentration. The Freundlich and Langmuir adsorption isotherms have been employed to evaluate the adsorption behavior, but the Langmuir model was much better, and a maximum adsorption capacity for nHAp of 250.0 mg/g was obtained. Interestingly, the Pb2+ ions sorption by nHAp increased with pH and nHAp dosage while the sorption capacity decreased with the nHAp dosage.
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