Environ Sci Pollut Res DOI 10.1007/s11356-017-0508-y
RESEARCH ARTICLE
Adsorption/reduction of Hg(II) and Pb(II) from aqueous solutions by using bone ash/nZVI composite: effects of aging time, Fe loading quantity and co-existing ions Antonio Gil 3 & Mohammad Javad Amiri 1 & Jahangir Abedi-Koupai 2 & Saeid Eslamian 2
Received: 28 February 2017 / Accepted: 18 October 2017 # Springer-Verlag GmbH Germany 2017
Abstract In this research, a versatile and highly efficient method for the stabilization of nanoscale zerovalent iron particles (nZVI) on the surface of ostrich bone ash (OBA) was presented as a novel inorganic adsorbent (OBA/nZVI) for the removal of Hg(II) and Pb(II) ions from aqueous solutions, even after 1 year of storage under room conditions. The removal behavior of the OBA/nZVI was assessed as a function of the initial pH, contact time, initial pollutants concentration, temperature, amount of adsorbent, effect of competitive metal ions, and ionic strength. The synthesized adsorbent was characterized by several techniques including N2 adsorption at − 196 °C, FT-IR spectroscopy, scanning electron microscopy, X-ray diffraction, and zeta potential. The results confirmed that the OBA is a good candidate as support of nZVI. The maxima adsorption capacity for Hg(II) and Pb(II) ions found from experimental results were 170 and 160 mg g−1, when the loading quantities of Fe were 20%. The equilibrium sorption data obeyed a Langmuir–Freundlich isotherm type model.
Responsible editor: Philippe Garrigues * Mohammad Javad Amiri
[email protected] Jahangir Abedi-Koupai
[email protected] Saeid Eslamian
[email protected] 1
Department of Water Engineering, College of Agriculture, Fasa University, Fasa 74617-81189, Iran
2
Department of Water Engineering, College of Agriculture, Isfahan University of Technology, Isfahan 84156-83111, Iran
3
Department of Applied Chemistry, Public University of Navarra, Campus of Arrosadia, 31006 Pamplona, Spain
The kinetic data of the adsorption followed the mechanism of the pseudo-second-order model. The thermodynamic experiments indicated that the removal of metal ions were feasible, endothermic, and spontaneous. It can be found that fresh and aged OBA/nZVI maintained its usability even after five cycles in the order: fresh (OBA/nZVI)-Hg(II) > fresh (OBA/ nZVI)-Pb(II) > aged (OBA/nZVI)-Hg(II) > aged (OBA/ nZVI)-Pb(II), which indicate that OBA/nZVI can be regenerated as adsorbent. The existence of Fe in the OBA/nZVI was proved by SEM-EDX results and X-ray diffraction analysis also confirmed adsorption/reduction of some of the Hg(II) to Hg0 and Pb(II) to Pb0. Keywords Zerovalent iron . Ostrich bone ash . Composite adsorbent . Lead adsorption . Mercury adsorption
Introduction The release of various toxic and hazardous heavy metal (HM) ions into the water resources is the most environmental concern due to bioaccumulating tendency and nonbiodegradable nature (Hansen et al. 2010). Industrialization and urbanization constitute the major sources of metal pollution in water bodies. Thus, removal of hazardous HM ions from wastewater has attracted great consideration in recent years. According to the Agency for Toxic Substances and Disease Registry, lead and mercury due to their highly toxic, cumulative poison, and environmental effects in living organisms, are considered as priority pollutants (www.atsdr.cdc.gov). According to the Environmental Protection Agency (EPA), the recommended concentration level of mercury and lead in drinking water are 0.001 and 0.015 mg L−1, and for wastewater are 0.005 and 0.05 mg L −1 (The Council of the European Communities 1976).
Environ Sci Pollut Res
Previous researches showed that reverse osmosis, membrane filtration, chemical precipitation, flotation, ion exchange, phytoextraction, coagulation, ultra filtration, and adsorption using commercial adsorbents can be applied for removal of Hg(II) and Pb(II) ions from the contaminated waters (Wang and Chen 2009). However, most of these methods have drawbacks and restrictions. Therefore, researchers turn their interest in the biosorption method particularly using bio-composites with low cost and high adsorption capacity (Tran et al. 2010). In recent years, several types of animal bone wastes have been used as versatile adsorbents for a broad variety of contaminants from aqueous solutions due to the eco-friendly and low-cost, high adsorption capacity, and easily available (Hassan et al. 2008; Kizilkaya et al. 2010; Amiri et al. 2013, 2016; Arshadi et al. 2015). Bone wastes contains organic compounds (30%), including organic tissue, fat and collagen, and inorganic phase (70%), comprising hydroxyapatite (HAP), which can be used as adsorbent to remove inorganic cations by ion exchange mechanism (Kizilkaya et al. 2010; Amiri et al. 2013). Ostrich bone waste which, generated in very high amounts in Iran, can be applied rather than ion exchange resins or activated carbon for wastewater treatment (www.worldrecordacademy.com). However, the development of animal bone waste in removing HMs is limited due to the difficulty in separation and recycling (Feng et al. 2010). To resolve this drawback, Fe3O4 incorporated to HAP (Dong et al. 2010; Feng et al. 2010) and nZVI particles used as a support substance (Zhang et al. 2010, 2011; Arshadi et al. 2014; Soleymanzadeh et al. 2015) have been developed. Over the last few years, nZVI as a powerful ecofriendly reducing agent has been widely applied for elimination of broad types of organic and inorganic pollutants due to the high surface to volume ratio and also high reaction activity. nZVI has the benefit for the application as support material of composite adsorbent since it can be simply separated from the aqueous solution through a magnetic field. Therefore, employing support substances for stabilization of nZVI is a practical method to inhibit oxidation and aggregation of uncoated nZVI for the remediation of contaminated groundwater (Zhang et al. 2010; Liu et al. 2010; Shi et al. 2011). The goal of this work is to assess the effectiveness and performance of OBA/nZVI as a recyclable nanobiomaterial for the removal of Hg(II) and Pb(II) ions in batch reactors even after 1 year of storage under ambient conditions. At first, the synthesis and characterization of the OBA/nZVI composite were studied. Then, the effect of several operational parameters including initial Hg(II) and Pb(II) ions concentrations, pH, contact time, temperature, adsorbent dosage, competitive metal ions, ionic strength, and Fe loading quantity was also investigated. In addition, the uptake of Hg(II) and Pb(II) ions in terms of kinetic, isotherm, and thermodynamic
models was studied to comprehend information about the proposed mechanism.
Experimental Materials All of the chemicals including Hg(NO3)2·2H2O, Pb(NO3)2, FeCl2·4H2O, NaBH4, HCl solution, and NaOH were provided from Aldrich and Merck and were used without further purification. Solvents were purified and distilled before their use according to standard procedures (Perrin et al. 1983). Ostrich bone ash (OBA) was provided under laboratory conditions. At first, fresh ostrich bone wastes obtained from local butcher’s store was rinsed several times with water and removed from fate and residual protein pieces by boiling in water for 2 h. Then, the bones were dried in the oven at 70 °C overnight. The burning of dried bones was performed using an air furnace at 550 °C for 24 h. The white powder was transferred to the vacuum desiccator and dried in the oven at 65 °C. OBA was pulverized with a 45– 80 range mesh (180–355 μm) using standard of American Society for Testing and Materials (ASTM) sieves with geometric average size of 300 μm. OBA supported nZVI was obtained by employing Fe(II) a reduction procedure reported previously (Arshadi et al. 2014; Soleymanzadeh et al. 2015). Briefly, Fe(II) solution was obtained by dissolving 5.4 g of FeCl2·4H2O into a mixture of ethanol and water in a 4/1 (v/v) (72 mL ethanol +18 mL deionized water). Afterwards, an amount of 3.7 g of OBA was weighted and added to a 100 mL of this solution. The content was then placed in an ultrasonic shaker for 30 min in order to disperse the OBA grains. Meanwhile, a certain amount of borohydride solution (3.9 g/100 mL) was transferred drop wisely into the stirred mixture while the Fe(II)-OBA were mixed continuously under nitrogen using an magnetic stir bar. After the addition of the first drop of borohydride solution, the black nZVI particles observed promptly. The suspension is subjected to mechanical stirring for 30 min on heater following the entire addition of the borohydride solution. Then, the content was filtered using a Büchner vacuum filtration funnel and denoted as OBA/nZVI. Characterization techniques Various techniques including SEM–EDX (Seron-AIS2100), XRD (Philips X’PERT MPD diffractometer), FTIR (Jasco FT/IR-680), N2 adsorption at − 196 °C (Quantachrome, USA), and ICP-AES (Shimadzu ARL 34000) were applied for the characterization of OBA/nZVI. The pH of the solutions was measured using a pH/mV meter (Metrohm, 827 pH Lab).
Environ Sci Pollut Res
The point of zero charge (pHZPC) of samples was calculated by the solid addition procedure (Cason and Lester 1997). Adsorption measurements Adsorption experiments of Pb(II) and Hg(II) ions were conducted by adding 0.1 g of adsorbents to flasks with 30 mL of the metal pollutant, at a concentration of between 5 and 1000 mg L−1, at a controlled temperature of 25 °C, under batch conditions. The flasks were shaken in a shaker (Edmund Buhler, SM 30 control) at 200 rpm at 25 °C. After shaking, the solution at a given time, determined by preliminary kinetic measurements, was filtered through the centrifuge at 3000 rpm for 5 min. Pb(II) and Hg(II) ions concentration in filtrate were determined by atomic absorption spectrophotometer. The adsorption capacities of the samples (mg g−1) were calculated according to Eq. (1):
qe ¼
ðC o −C e Þ V m
ð1Þ
where Co and Ce are initial and equilibrium concentrations (mg L−1), V is volume of the solution (L), and m is the weight of the adsorbent (g). The sorption percentage (%) of metal ions was calculated according to Eq. (2): E ð% Þ ¼
C o −C e 100 Co
ð2Þ
Pseudo-first-order, pseudo-second-order, and intra-particle diffusion models were applied to fit the experimental data of Hg(II) and Pb(II) ions removal using OBA and OBA/nZVI. The pseudo-first-order, pseudo-second order, and intraparticle diffusion equations suppose that the adsorption process includes physisorption, chemisorption and intrapore diffusion mechanism. All the studied mathematical kinetic models are summarized in Table 1, where qe and qt (mg·g−1) are the adsorption capacities at equilibrium and at time t. The parameters kf (min−1), ks (g·mg−1·min−1), and kint (mg·g−1·min0.5 ) are the rate coefficients for pseudo-first-order, pseudosecond-order and intra-particle diffusion models (Kizilkaya et al. 2010). Freundlich, Langmuir, Redlich–Peterson (R-P), and Langmuir–Freundlich (L-F) models were also applied to the experimental results of adsorption of Hg(II) and Pb(II) ions under equilibrium conditions on OBA and OBA/nZVI. The Langmuir model supposes that the sorption surface is homogenous, that monolayer adsorption is possible. In the Freundlich model, distribution of active sites and their energies is heterogeneous, correspond to the multiple adsorption. The Redlich–Peterson and Langmuir–Freundlich isotherm models have the hybrid features of the Freundlich and the Langmuir which can be used in both homogeneous and heterogeneous systems. The mathematical isotherm models used
are summarized in Table 1, where qe is the adsorbed amount and Ce is the equilibrium concentration of metal ions. In the Langmuir model, qmax is the maximum solute adsorbed at the equilibrium state for the completion of a layer (mg g−1) and kL is a constant that depends on the energy of adsorption which shows the enthalpy of adsorption. It is also an index to describe the binding energy of surface adsorption. In the Freundlich model, kF and n are the coefficients attributed to the adsorption capacity and the adsorption intensity of adsorbent. In the Langmuir–Freundlich (L-F) model, kL (L·mg−1) and n are the metal ions attributed to the L-F equilibrium constant and the exponent of the L-F equation. kRP, aRP, and β in the Redlich–Peterson (R-P) isotherm are the model constants. kRP is the solute adsorptivity (L g−1), aRP is relevant to the adsorption energy (L mg−1), and β is the heterogeneity constant (0 < β < 1). In this study, the thermodynamic parameters including standard free energy (ΔG), standard enthalpy (ΔH), and standard entropy (ΔS) were also investigated. The mathematical thermodynamic equations considered are included in Table 1, where R is the universal gas constant, T is the solution temperature (K), and kd (g−1) is the ratio of adsorbed material amount by the adsorbent (qe, mg·g−1) to the remaining amount in the solution (Ce, mg·L−1). ΔH and ΔS were obtained from the slope and intercept of the plots of lnKd versus 1/T (Arshadi et al. 2015).
Results and discussion Characterization of OBA/nZVI OBA and OBA/nZVI were characterized and their properties are presented in Table 2. The results of samples components indicated that calcium and phosphorus are the principal elements in OBA and OBA/nZVI composite. Weight percentage of HAP in the OBA/nZVI and OBA nearly obtained as 38.3 and 46.1, respectively. The Ca/P mole ratio of the composite was 1.71, much greater than the observed for OBA (1.58) or even the ideal stoichiometric apatite composition (1.67). Therefore, the framework of OBA/nZVI become more efficient than OBA for elimination of Hg(II) and P(II) ions from contaminated waters. The iron values of OBA were increased from 1.45 to 18.9 wt.% when nZVI stabilizes on it. The specific surface area, average pore size and pore volume of the OBA/nZVI composite were 109 m2 g−1, 110 Å, and 0.31 cm3· g−1, compared to 67 m2 g−1, 83 Å, and 0.24 cm3·g−1 for the OBA alone. The increment in the specific surface area, average pore size and pore volume of OBA/nZVI is due to the non-aggregation of the nZVI particles resulting mass transfer between the surface of the composite and Hg(II) and Pb(II) ions in solution and the consequent interaction between the active sites of OBA/nZVI and HMs (Pan et al. 2009). Therefore, nZVI can be supported onto larger
Environ Sci Pollut Res Table 1
Mathematical equations of the used kinetic and thermodynamic models
Model
Equation
Parameter and dimension
qt = qe(1 − exp(−kft))
kf (min−1) qe, qt (mg·g−1) kS (mg·g−1·min−1) t (min)
Kinetic models Pseudo first-order Pseudo second-order
k q2 t
qt ¼ 1þqs ek s t e
kint (mg·g−1·min-0.5)
qt = kint t0.5
Intra-particle diffusion Isotherm models Langmuir
qmax (mg·g−1) kL (L·mg−1) kF (mg·g−1) n (−) kLF (L·mg−1) n(−) (0 < n < 1)
L qmax C qe ¼ k1þk LC
Freundlich
1
qe ¼ k F C ne
Langmuir–Freundlich (L-F)
n
max ðk LF C e Þ qe ¼ q1þ ðk LF C e Þn
Redlich–Peterson (R-P)
k RP C e qe ¼ 1þa Cβ
kRP (L·g−1) aRP (L·mg−1) β (−) (0 < β < 1)
ΔH lnk d ¼ Δs R − RT
kd (g−1) T (k)
RP
e
Thermodynamic equations ΔS and ΔH kd
qe k d ¼ Ce
ΔG Ea
ΔG = − RT · ln kd Ea = ΔH + RT(Zarrouk et al. 2011)
porous substance such as OBA to improve its mobility in environmental remediation and inhibit nZVI aggregation and deposition in the subsurface (Zhang et al. 2010; Arshadi et al. 2014; Soleymanzadeh et al. 2015). SEM micrographs for the OBA, OBA/nZVI, and OBA/ nZVI after Pb(II) and Hg(II) adsorption are presented in Fig. 1. OBA is a white powder prepared with the combustion of ostrich by-products in an air condition (see Fig. 1(A)).
Table 2
Chemical and chemical properties of OBA and OBA/nZVI
Chemical properties Ca P Ca/P Fe Pb Hg Physical properties Specific surface area (m2 g−1) Average pore size (Å) Pore volume (cm3 g−1)
OBA/nZVI wt.%
OBA wt.%
24.2 14.1 1.71 18.9 < 0.1 < 0.1
28.27 17.85 1.58 1.45 < 0.1 < 0.1
109 110 0.31
67 83 0.24
Investigation of the images from SEM (see Fig. 1(B)) indicates the surface of OBA is sponge-like porous morphology with various pore sizes which can be used as a good candidate to inhibit oxidation and aggregation of uncoated nZVI in the aqueous solutions. In fact, the high specific surface area of OBA (67 m2 g−1) confirms that it can play an important role as the carrier of nZVI. The black color of OBA/nZVI demonstrated that nZVI successfully supported on the OBA surface (see Fig. 1(C)). The surface of OBA/nZVI depicted roughness which was aggregated with numerous nanoparticles compared to the surface of OBA (see Fig. 1(D)). The color change of OBA/nZVI was observed (black to reddish-brown), because of the quick reduction of Hg(II)-Hg(0) and Pb(II)–Pb(0) on the surface of composite simultaneously with surface oxidation of Fe(0)–Fe(II) (see Fig. 1(E)). SEM results indicates alters in the surface morphology of OBA/nZVI before and after adsorption of Pb(II) and Hg(II) ions (see Fig. 1(F, G)). It can be observed that new precipitates were formed on the surface of OBA/nZVI composite after the uptake of Pb2+ (see Fig. 1(G)) and Hg2+ (see Fig. 1(F)) ions. Typical EDX spectra for the OBA, OBA/nZVI, and OBA/ nZVI after Pb(II) and Hg(II) adsorption are presented in Fig. 2. The EDX results indicated that OBA mainly includes calcium and phosphorus which are the basic components of HAP (Ca10(PO4)6(OH)2) (see Fig. 2a). The presence of iron and the major constituents of HAP in the structure of OBA/
Environ Sci Pollut Res Fig. 1 The photography and SEM images of samples: (A) OBA powder, (B) OBA, (C) OBA/nZVI powder, (D) OBA/ nZVI, (E) OBA/nZVI powder after metal ions adsorption, (F) OBA/nZVI after Hg2+ adsorption (G)OBA/nZVI after Pb2+ adsorption
A
B
C
D
E
F
nZVI which is come from nZVI and OBA, presenting successful stabilization of nZVI (see Fig. 2b). The strong lead and mercury peaks appeared when OBA/nZVI was impregnated by Pb(II) and Hg(II) solutions (see Fig. 2c, d). The obtained results demonstrated that the adsorption of Hg(II) and Pb(II) ions on the surface of OBA/nZVI composite was successful. Moreover, after OBA/nZVI loaded with Pb(II) and Hg(II) ions, an obvious decrease of Ca could be detected which may be attributed to the participation of particular amounts of interchange in the lead and mercury adsorption (Feng et al. 2010). The X-ray diffraction of OBA, freshly synthesized OBA/ nZVI and aged OBA/nZVI after 1year is presented in Fig. 3. XRD analysis of OBA indicated the presence of peaks related to the HAP phase: 2θ = 25.87, 29.83, 31.88, 32.99, 34.58, 39.66, 46.51 and 49.47 °. Apparent peak at 2θ = 45 ° in the XRD pattern of freshly and aged OBA/nZVI composite is attributed to Fe0. Various types of iron oxide such as magnetite, maghemite and hematite which are characterized by
G
a 2θ = 36 ° (Gotic and Music 2007), were not found in the XRD pattern of freshly and aged OBA/nZVI composite presenting successful stabilization of nZVI on the structure of OBA even after 1 year. However, some weak peaks of iron oxides were appeared in the aged OBA/nZVI which may be attributed to the surface oxidation of nZVI. The FT-IR spectra of OBA and OBA/nZVI are included in Fig. 4. The FT-IR analysis of OBA can give an indication of the elimination of organic components after the combustion bone. The spectra of bones indicates the presence of the organic components including fat tissues, collagen and amide groups from the protein constituents and also main inorganic species, carbonate groups and phosphate (Bahrololoom et al. 2009). The bands at 1650, 1560, and 1250 cm−1 associated to amide functional groups do not appeared in the OBA structure (see Fig. 4(A)). Moreover, there are no absorption bands attributed to C-H bonds in the OBA structure which indicated total elimination of the organic compound of bone. Thus, all bands detected in the OBA (see Fig. 4(A)) are related to the
Environ Sci Pollut Res Ca
A P
Na
Ca
Al
Mn 10
5 Energy (Ke V) Ca
B P
Fe Na
Ca Al Mn Fe 5 Energy (Ke V)
C
10
Ca P Pb
Fe Ca
Na
Al
Pb
Mn
Pb
Fe
Pb
Pb 10
5 Energy (Ke V)
D
Ca P Hg
Na
Al
Adsorption experiments
Fe Ca
Hg
Fe
5 Energy (Ke V)
2009). The bands at 1456, 1417, and 873 are assigned to – CO32− groups. There is also a broad band at 3436 cm−1 that it is associated to the hydroxyl group. An increase in the band intensities at 1158, 1096 and 1033 cm−1 (vibration bands of HAP), 603 cm−1 (v4 frequency which is due to P-O stretch), and 561 cm−1 (v4 frequency which is generally related to P-O bending and P–O stretch) was also appeared as compared to ostrich bone waste (Amiri et al. 2016) due to removing of organic compounds by calcination. The immobilization of nZVI changes the peak shape of the FT-IR spectrum (see Fig. 4(B)). A strong broad absorption peak in hydroxyl (3361 cm−1) and carboxyl groups (1744 cm−1) exhibited clear alters. Moreover, the IR bands at 823 and 705 cm−1 are assigned to iron oxides and also one -OH bending frequency observed at 871 cm−1 that associated to Fe-OH-Fe groups. The stretching and bending modes of Fe–O were related to bans at 634 and 621 cm−1 (Wang and Ro 2007). However, the spectrum of OBA/nZVI shows a remarkable shift from 1237 to 1220 cm−1 and 1096 to 997 cm−1 due to the molecular vibration of HAP. The point of zero charge (pHZPC) of the OBA and OBA/ nZVI were obtained in NaCl aqueous media at various pH. The pHZPC of OBA and OBA/nZVI were measured to be 6.82 and 5.87. The shift of pHZPC from 6.82 to 5.87 shows the changes on the surface of OBA after supported by nZVI. This also displays that the entrapment of nZVI onto OBA was successful. pHZPC of the pure iron oxides typically was reported between 7 and 9 from the literature (Sun et al. 2006; Uzuma et al. 2008). As a consequence, the lower pHZPC of the OBA/nZVI versus OBA yield acidic surfaces which was related to the introduction of multiple oxygen-containing functional groups.
Hg
Hg
Hg
10
Fig. 2 EDX spectra of samples. a OBA. b OBA/nZVI. c OBA/nZVI after Pb2+ adsorption. d OBA/nZVI after Hg2+ adsorption
inorganic materials of bone. The spectrum of OBA and OBA/ nZVI shown peaks at 1165 cm−1 (out-of-plane and in-plane deformation modes of P–O–H), 1028, and 1100 cm−1 (v3 frequency which is associated to the asymmetric stretching of PO bonds), 960 cm−1 (v1 frequency which is assigned to the symmetric stretching of P-O bonds), 603 cm−1 (v4 frequency which is due to P-O stretch), 560 cm−1 (v4 frequency which is generally related to P-O bending and P–O stretch) representing the characteristic vibration bands of O–P–O bonds of calcium phosphate (Peters et al. 2000; Sun et al.
Initial studies displayed that OBA has a high potential to entrap nZVI particles. OBA was HAP and did not contain any organic components of the bone. Indeed, the specific surface area (67 m2 g−1) of OBA shows that this bio-material is a privileged candidate as carrier of nZVI. In fact, oxidation, agglomeration, and great mobility in the aqueous solutions are the main limitations of nZVI particles in environmental remediation (Eslamian et al. 2013) which could be solved by a stable immobilized nZVI on the surface of a biological origin support such as OBA. For these reasons, kinetic and thermodynamic investigations of Pb(II) and Hg(II) ions adsorption on OBA and OBA/nZVI was reported. The removal efficiency of metal ions of OBA and OBA/ nZVI at various pH between 1 and 9 was presented in Fig. 5. The lowest removal efficiency of metal ions was observed at an initial solution pH of 1. At acidic medium, the uptake capacity of Hg(II) and Pb(II) ions may be related to the competition between hydrogen ion with metal ions for binding to
Environ Sci Pollut Res
Ca10(PO4)6(OH)2 Fe2O3/Fe3O4
Fig. 3 X-ray diffraction patterns of OBA, OBA/nZVI, and aged OBA/nZVI after 1 year
Fe
o
OBA/nZVI – after one year o
Intensity (c.p.s)
Fe
OBA/nZVI
Ca10(PO4)6(OH)2
OBA
2 theta (º) vacant adsorbent site. Dissolution of calcium, phosphorus, and iron, which are the fundamental components of OBA/ nZVI composite, may happen at low pH. For this regards, the Fe leaching amount of OBA/nZVI into the solution under various pH conditions was also investigated (data not presented). The results showed that the leaching of Fe is negligible when the initial pH is above 3.0, which could demonstrate its application for the remediation of contaminated groundwater with a pH between 5 and 9. However, highest amount of Fe leaching (68 mg L−1) was found at the lowest initial pH solution (pH = 1), which can considerable impact on its application in heavily acidic conditions. A sharp linear decrease of Fe leaching through OBA/nZVI into the solution was observed between pH 1 to 3. Accordingly, the removal efficiency of Pb(II) and Hg(II) ions by OBA/nZVI decreased substantially
when the pH of solution was lower than 3. In this condition, the high amount of iron ions in the solution is attributed to the corrosion of metallic iron and production of hydrogen. As it is observed in Fig. 5, the OBA/nZVI presents higher removal efficiency of metal ions than OBA in the pH range between 1 and 9. In the case of OBA/nZVI, as the pH raised to 3 for Hg(II) and Pb(II) ions, the adsorption efficiency of metal ions rapidly raised to more than 96% and then keeps almost constant at the interval pH of 3 and 9. This behavior can be attributed to quick reduction of Hg(II)-Hg(0) and Pb(II)– Pb(0) on the surface of composite and also the adsorption of Hg(II) and Pb(II) ions using the external hydroxyl species on the surface of nZVI (Sun et al. 2006). In the case of OBA, as the pH raised to 5 and 6 for Hg(II) and Pb(II) ions, the adsorption efficiency of metal ions gradually raised to more than
T (%)
Fig. 4 FT-IR spectra of samples: (A) OBA and (B) OBA/nZVI
υ1, P-O stretching Fe-O stretching CO3-2 group υ4, P-O stretching and P-O bending
P-O-H Hydroxyl group υ3, P-O stretching
4000
3000
Iron oxide υ1, P-O stretching
2000
1000 -1
Wavenumbers (cm )
400
Environ Sci Pollut Res 100
80 60
OBA OBA/nZVI
40
Pb(II) Removal (%)
Pb(II) Removal (%)
100
OBA OBA/nZVI
60
40
20
20
0
0
100
80 60 40
OBA OBA/nZVI
20
Hg(II) Removal (%)
100 Hg(II) Removal (%)
80
80 60
OBA OBA/nZVI
40 20 0 0.00
0 2
4
6
8
0.05
pH Fig. 5 Effect of the pH on the adsorption of Pb (II) and Hg(II) ions by OBA and OBA/nZVI
70%. This trend can be attributed to the OH− consumption via deprotonation of –COOH, -P-OH, and ≡Ca-O-H2+ dominant active sites. So, negatively charged sites of the OBA (–COOH and -P-O−) and the ≡Ca-O-H functional groups dominate in alkaline ranges, causing that the number of negatively charged sites on the surface of OBA increases. Consequently, the electrostatic attraction of OBA to metal ions enhanced with rising pH. However, the point of zero charge of OBA/nZVI is lower than the OBA which indicated that the surface of OBA/nZVI become more negatively charged than OBA in the alkaline range. Above these points, pH equal to 5 and 6 for Hg(II) and Pb(II), adsorption efficiency of metal ions by OBA tended slightly to decrease which is related to the reduction of the mobility of metal ions due to the decline in the exchangeable form (Amiri et al. 2013). Therefore, OBA/nZVI composite has been showed as a versatile material for the removal of Hg(II) and Pb(II) ions from natural water resources with a pH value between 5 and 9 (Bundschuh et al. 2004). An increase in the adsorbent dosage caused an increase in the adsorption efficiency of the metal ions by OBA and OBA/ nZVI due to the availability of a greater number of adsorption sites (see Fig. 6). Between 0.02 to 0.2 g, OBA/nZVI shows higher removal efficiency of metal ions than OBA. It was also found that the adsorption efficiency of Pb(II) and Hg(II) ions by OBA increased from 46.6 to 81.7% and from 38.2 to
0.10
0.15
0.20
Adsorbent dosage (g)
10
Fig. 6 Effect of the amount of dosage on the adsorption of Pb (II) and Hg(II) ions by OBA and OBA/nZVI
69.5%. In the same way, the adsorption efficiency of Pb (II) and Hg(II) ions by OBA/nZVI increased from 53 to 96% and from 45 to 97%. However, equilibrium adsorption capacity was achieved at higher adsorption dosages. In order to consider the adsorption efficiency, 0.1 g of adsorbent dosage was selected for the rest of the study. The removal efficiency of Hg(II) and Pb(II) ions with initial concentration of 10 mg L−1 by OBA and OBA/nZVI as a function of shaking time (0–360 min) is illustrated in Fig. 7. It
100 Removal Efficiency (%)
0
80 60 OBA-Pb OBA-Hg OBA/nZVI-Pb OBA/nZVI-Hg
40 20 0 0
100
200 Time (min)
300
400
Fig. 7 Adsorption kinetics of Pb (II) and Hg(II) ions on OBA and OBA/nZVI
Environ Sci Pollut Res
indicates that more than 96% of Hg(II) and 91% of Pb(II) by OBA/nZVI were removed within the first 45 and 15 min. The black zero-valent iron nanoparticles entrapped in OBA were quickly solubilized, the system changes reddish-brown when there is oxygen in it, which demonstrated that the OBA/nZVI reacted immediately with oxidative matters in solution. So, OBA/nZVI reacted with metal ions much faster than the OBA under the same situation. The adsorption efficiency of Hg(II) and Pb(II) ions by OBA/nZVI was 99.2 and 94.1% at 360 min. To establish the adsorption equilibrium of Hg(II) and Pb(II) ions by OBA/nZVI, a contact time of about 45 min was needed. In the case of OBA, the adsorption efficiency of Hg(II) and Pb(II) ions was 71.4 and 88.8% at 120 min. Kinetics studies indicated that a fast adsorption efficiency happened within the first time and the metal uptake slowly approaches equilibrium. The kinetic data from Hg(II) and Pb(II) ions adsorption experiments were fitted using the pseudo-first-order, pseudo-second-order, and intra-particle diffusion models according to equations of Table 1 and the results are summarized in Table 3. The kinetic data indicated a good accord with the pseudo-second-order model and the correlation coefficients (R2) were higher than 0.99 for OBA and OBA/nZVI. The compliance of the kinetic data with the Kinetic parameters for the adsorption of Pb(II) and Hg(II) ions
System First-order model qe (mg·g−1) kf (min−1) OBA Pb(II) Hg(II) OBA/nZVI Pb(II) Hg(II)
10.5 9.6
R2
9 7.2
0.97 0.97
0.19 9.3 1.62 9.8 Second-order model
0.47 0.76
ks (g·mg−1·min−1) qe (mg·g−1) ks (mg·g−1·min−1) OBA Pb(II) Hg(II) OBA/nZVI Pb(II) Hg(II)
OBA Pb(II) Hg(II) OBA/nZVI Pb(II) Hg(II)
R2 100
0.009 0.013 0.087 0.043
9.2 7.1
0.83 0.70
9.4 7.8 10 4.3 Inter-particle model kint (mg·g−1·min-0.5)
0.999 0.999 1 1 R2
Removal Efficiency (%)
Table 3
pseudo-second-order kinetic equation demonstrates that the rate-limiting step for the adsorption of metal ions on OBA/ nZVI is chemisorption, which involves valence forces through sharing or exchange of electrons between OBA/nZVI and metal ions. Thus, the uptake of the metal ions on OBA/nZVI may be suggested to comprise of two processes with initial sorption rate of 7.8 and 4.3 mg g−1 min−1 for Pb(II) and Hg(II). In the case of OBA/nZVI, the experimental (qe,exp) and calculated (qe,cal) uptake capacities for Pb(II) ion were obtained to be 9.4 mg/g. Similarly, qe,exp and qe,cal for Hg(II) ion were found to be 9.9 and 10 mg/g. These results show that qe,cal are very close to the qe,exp. To evaluate the participation of diffusion mechanism in the sorption of metal ions by OBA/ nZVI, the Weber and Morris plot, qt versus t0.5, was applied to study the intra-particle diffusion mechanism (see Table 3). It is seen that the plots of Pb(II) and Hg(II) ions were not linear over the entire time denoting that several mechanisms are involved in the adsorption process (Arshadi et al. 2015). These results confirmed by the pseudo-first-order model which indicated the relation between ln(qe − q) versus t over the whole time range was not linear, implying that more than one process affected the adsorption (Pan et al. 2009). The effect of the initial concentration of the metal ions on the adsorption efficiency by OBA and OBA/nZVI in a solution with optimum pH after 360 min is seen in Fig. 8. The results included in this figure indicates that the adsorption efficiency of metal ions decreased with increasing the initial Hg(II) and Pb(II) ions concentration, which can be related to the adsorption sites and the specific surface area. The high adsorption efficiency was found at the 5 mg L−1 of metal solution by OBA and OBA/ nZVI. Four isotherms were applied to explain the experimental data for the adsorption process of Hg(II) and Pb(II) ions on OBA and OBA/nZVI, namely Freundlich, Langmuir, Langmuir–Freundlich and Redlich–Peterson adsorption models (see Table 1).
OBA-Pb OBA-Pb OBA-nZVI-Hg OBA-nZVI-Pb
80 60 40 20
0.32 0.25
0.68 0.68
0.026 0.12
0.88 0.57
0 0
200
400
600
800
1000
Initial concentration of metal ions (mg L-1)
Fig. 8 Effect of the initial concentration of Pb (II) and Hg(II) ions on the adsorption by OBA and OBA/nZVI
Environ Sci Pollut Res Table 4
Isotherm constants for Pb(II) and Hg(II) ions adsorption by OBA and OBA/nZVI qmax
kL
kF
n
aRP
β
R2
Sorption model
113.7
0.044
–
–
–
–
0.89
Langmuir
– 115.2
– 0.01
7.19 –
1.64 0.727
– –
– –
0.994 0.907
Freundlich Langmuir–Freundlich
– 94.68
– 0.031
– –
0.39 –
3219 –
446 –
0.999 0.87
Redlich−Peterson Langmuir
–
–
5.16
1.49
–
–
0.993
Freundlich
97.36 –
0.015 –
– –
0.86 0.78
– 1846
– 315
0.89 0.991
Langmuir–Freundlich Redlich−Peterson
OBA
OBA/nZVI Pb(II)
Hg(II)
175.9
0.025
–
–
–
–
0.974
Langmuir
–
–
19.39
2.97
–
–
0.881
Freundlich
161.4 –
0.0016 –
– –
0.542 1.2
– 3.2
– 0.005
0.985 0.985
Langmuir–Freundlich Redlich−Peterson
181.6 – 170.7
0.031 – 0.0092
– 21.57 –
– 3.03 0.695
– – –
– – –
0.988 0.889 0.994
Langmuir Freundlich Langmuir–Freundlich
–
–
–
1.13
4.4
0.01
0.994
Redlich−Peterson
Adsorption isotherms of metal ions were carried out by mixing 0.1 g of OBA and OBA/nZVI with 100 mL of various metal ion concentrations, between 5 and 1000 mg L−1. Correlation coefficient (R2) of the nonlinear regression analysis method was used for applicability of the isotherm models according to Calisir et al. (2009) (see Table 4). The isotherm data of OBA indicated that Freundlich and Redlich–Peterson models show better fitting with higher R2 values compared to the other models. In the case of OBA/nZVI, Langmuir– Freundlich and Redlich–Peterson models fitted the experimental data better (R2 > 0.98) compared to the other two models. In the all cases, the Redlich–Peterson isotherm model indicates high R2 values which it provides a considerably better fit compared to the two-parameter isotherm equations. The results indicated that the Freundlich isotherm equation of OBA is changed to Langmuir–Freundlich isotherm model after entrapment of nZVI onto OBA. In fact, the heterogeneous adsorption surface of OBA is tended to homogeneous after the iron nanoparticles immobilization. It is believed that the nanoscale iron supported on OBA may play an important role in the elimination of metal ions (see Fig. 9). The maximum adsorption capacity calculated from experimental data of OBAPb, OBA-Hg, OBA/nZVI-Pb, and OBA/nZVI-Hg was found to be 88, 66, 160, and 170 mg g−1. Adsorption capacity of OBA/nZVI and other adsorbents previously applied for the removal of Hg(II) and Pb(II) ions is listed in Table 5. A comparison of this research with those of literature displays that the used adsorbents behaves in a comparable way, or even better, in most cases.
A
200
150 qe (mg g-1)
Hg(II)
100 Freundlich Langmuir Redlich-Peterson Langmuir-Freundlich Experimental data
50
0 200
B
150 qe (mg g-1)
Pb(II)
100 Freundlich Langmuir Redlich-Peterson Langmuir-Freundlich Experimental data
50
0 0
200
400 600 Ce (mg L-1)
800
1000
Fig. 9 Effect of the equilibrium concentration on the adsorption of a Hg(II) and b Pb(II) ions on OBA/nZVI
Environ Sci Pollut Res Table 5 Adsorption capacities for various adsorbents used for Hg(II) and Pb(II) removal from aqueous solutions
Hg (mg g−1)
Adsorbent
–
598.8
(Dong et al. 2010)
Kaolinite-supported nZVI
–
90
(Zhang et al. 2010)
Sineguelas waste supported nZVI
–
225
(Arshadi et al. 2014)
Kaolin supported nZVI
–
48
(Zhang et al., 2011)
Zeolite-nZVI
–
806
(Kim et al., 2013)
Chitosan/magnetite composite
–
63
(Tran et al. 2010)
Camel bone charcoal
28.2
–
(Hassan et al. 2008) (Kadirvelu et al. 2004)
Activated carbon from sago waste
55.6
–
Nano-TiO2
166.6
–
(Ghasemi et al. 2012)
Pumice-nanoscale zero-valent iron
107.1
–
(Liu et al. 2015)
OBA OBA/nZVI
66 170
88 160
This work This work
120
Removal efficiency (%)
100 80 60
OBA-Pb OBA-Hg OBA-nZVI-Hg OBA-nZVI-Pb
20 0
0
20
Reference
HAp/Fe3O4
The effect of the temperature on the adsorption efficiency of metal ions (80 mg L−1) by OBA and OBA/nZVI in a solution with optimum pH after 360 min is seen in Fig. 10. The results indicate that as the temperature raised from 25 to 80 °C, the removal efficiency of metal ions by OBA/nZVI enhanced significantly, which it is related to the more adsorption sites and mobility of metal ions at higher temperatures. In the case of OBA, when the temperature was raised from 50 to 80 °C, the removal efficiency of metal ions decreased dramatically, which is related to the reduction of attractive forces between OBA surface and metal ions at high temperature. The desorption tendency of the metal ions from the OBA surface was increased at high temperature. Various thermodynamic parameters, such as the activation energy (Ea), enthalpy change (ΔH), the Gibbs free energy of adsorption (ΔG), and the entropy change (ΔS) have been calculated for better assessment of increasing temperature effect on the removal of metal ions by OBA/nZVI (see Table 6). The ΔG values of Pb(II) ions adsorption on OBA/nZVI for the temperatures of 25, 50,
40
Pb (mg g−1)
40 60 o Temperature ( C)
80
100
Fig. 10 Effect of the temperature on the adsorption of Pb (II) and Hg(II) ions on OBA and OBA/nZVI
and 80 °C were found as − 2.4, − 5.5, and − 9.3 kJ mol−1. In the same way, the ΔG values of Hg(II) ions were obtained as − 1.8, − 8.5, and − 16.5 kJ mol−1. The negative amounts of ΔG demonstrate that the adsorption of metal ions on the OBA/nZVI is spontaneous. The decrease in ΔG with rising temperature displays that the adsorption process is more desirable at higher temperatures. In the case of OBA/nZVI, the values of ΔH, Ea and ΔS for Pb(II) ions were obtained to be 35, 37.7 kJ mol−1 and 125.4 J mol−1 K−1. In the same way, the amounts of ΔH, Ea and ΔS for Hg(II) ions were determined to be 78.2, 80.9 kJ mol−1 and 268.4 J mol−1 K−1. The positive amounts of ΔH displays that the removal of Hg(II) and Pb(II) ions by OBA/nZVI are endothermic. Authors reported that the values of ΔH for the physical adsorption are in the range of 2.1 < ΔH < 20.9, and for the chemical adsorption are between 20.9 < ΔH < 418.4 kJ mol−1 (Zhang et al. 2015). The values of ΔH for Hg(II) and Pb(II) ions adsorption on OBA/nZVI displays chemical adsorption because they are higher than 20.9 k mol−1. The positive values of ΔS demonstrated the raised randomness and an enhancement in the degrees of freedom at OBA/nZVI-solution interface within the immobilization of the metal ions on the available active site of the OBA/ nZVI. The Ea values found in this research for the adsorption of both metal ions on OBA/nZVI are in the range of chemisorption (Şeker et al., 2008). Thus, the values of Ea, ΔG and ΔH found in this work for the adsorption of Hg(II) and Pb(II) on OBA/nZVI composite can be mainly described by a chemisorption process. The effect of loading amounts of nZVI on OBA for removal of Hg(II) and Pb(II) ions from wastewater is presented in Fig. 11. As the immobilization of nZVI on OBA increased from 0 to 20%, the adsorption efficiency of metal ions increased drastically. This can be attributed to the fact that the rising nZVI loadings could enhance the specific surface area of OBA/nZVI. A rising in immobilization of nZVI on OBA from 0, 5, 10, 15, to 20% resulted in an increase of the specific
Environ Sci Pollut Res Table 6 Thermodynamic parameters for the adsorption of Pb(II) and Hg(II) ions on OBA and OBA/nZVI
ΔH (kJ mol−1)
System
ΔS (J mol−1 K−1)
T (K)
ΔG (kJ mol−1)
Ea (kJ mol−1)
OBA Pb(II)
27.26
Hg(II)
75.1
13.64
41.2
298 323
−2.1 −4.0
353
−2.9
298
−0.4
323 353
−3.5 −1.6
298 323
−2.4 −5.5
353 298
−9.3 −1.7
22.8
39.7
OBA/nZVI 35
78.2
268.4
surface areas from 67, 75, 81, 91, to 109 m2 g−1. However, a further increase in the immobilization of nZVI on OBA from 20 to 25% resulted in a slight decrease in the adsorption efficiency of metal ions. When the immobilization of nZVI on OBA increased from 20 to 25%, the specific surface areas decreased slightly from 109 to 107 m2 g−1. Therefore, a reduction of the specific surface area and adsorption efficiency of the metal ions was occurred which it is related to the aggregation of nZVI particles. Similar results have been reported by Zhang et al. (2010, 2006). An increase in the amount of nZVI leads to an increase in particle size, i.e., a decrease in the dispersion of nZVI. This behavior can be observed from the relation between the loaded nZVI content (wt.%) and the adsorption value based on the specific surface area (mg/m2) (Ads/S). In this way, the relationship between the loaded nZVI content (wt.%) and the adsorption capacity of Pb (II) and Hg (II) ions based on the specific surface area (mg·m2) was investigated (see Fig. 12). According to the Fig. 12a, the Ads/S value for Pb(II) slightly depends on the amount of nZVI 120 110
140
100 120 90 100 80
Pb(II) Hg(II) Surface area
80
Total surface area (m2 g-1)
Pb(II) and Hg(II) adsorbed (mg L-1)
160
70
60
60 0
5
10 15 Loaded nZVI content (wt.%)
20
25
Fig. 11 Effect of the nZVI loadings on the adsorption of Hg(II) and Pb(II) ions
323
−8.5
353
−16.5
37.7
80.9
up to 20 wt.%. If the amount of nZVI increases, this further leads to a decrease in the ratio of Ads/S. This point may indicate the change in the number of active sites on the OBA/ 1.40
Pb(II) adsorbed based on specific surface area (mg m2)
Hg(II)
125.4
A
1.38 1.36 1.34 1.32 1.30 1.28 1.26
Data points
1.24 1.22 1.20
0
5
10
15
20
25
Total surface area (m2 g-1)
B
1.4
Hg(II) adsorbed based on specific surface area (mg m2)
Pb(II)
1.2 1.0 0.8
Data points
0.6 0.4 0.2 0.0 0
5
10
15
20
25
Total surface area (m2 g-1) Fig. 12 a Correlation between the loaded nZVI content (wt.%) and the adsorption capacity of Pb(II) based on the specific surface area (mg m−2) and b correlation between the loaded nZVI content (wt.%) and the adsorption capacity of Hg(II) based on the specific surface area (mg m−2)
Environ Sci Pollut Res
ϕ0 Fe2þ = ¼ −0; 44 V
ð3Þ
Fe
So, each metal ion which has a more positive standard reduction potential than − 0.440 V can be undergone reduction by Fe. The selectivity order for adsorption of various metal ions by OBA/nZVI composite was Hg2+ > Pb2+ > Ni2+ > Cd2+. However, the results included in Fig. 13 follow the standard reduction potential order of the metal ions, that is, the standard reduction potential of Hg(II) (ϕ0 Hg2þ = ¼ þ0:79 V ) being more positive than Hg
Ni(II),
and
Cd(II)
(
ϕ
Pb2þ
100 80 60 40 20 0 Cd vy ea H
Ni Pb Cu
s al et m
=Pb ¼ −0:126 V, 0 0 ϕ Ni2þ = ¼ −0:257 V, ϕ Cd2þ = ¼ −0:352 V ). OBA/ Ni Cd nZVI could remove Ni2+ and Cd2+ ions with standard reduction potential slightly more positive than Fe. The reaction of OBA/nZVI composite for these metals comprises two possible mechanisms. The first mechanism is the direct reduction of Fe0. The adsorption of Hg(II) ions Pb(II),
0
A oval (%)
Fe→Fe2þ þ 2e− ;
by OBA/nZVI composite with standard reduction potential greatly more positive than Fe is mostly as the first mechanism. The second mechanism is the initial sorption of metals on the core-shell of nZVI followed by the reduction of adsorbed metals. The adsorption of Pb(II), Ni(II), and Cd(II) by OBA/nZVI composite with standard reduction potential slightly more positive than Fe is likely as the second mechanism. In order to investigate the effect of co-existing metal ions on Hg(II) and Pb(II) adsorption to OBA/nZVI, Hg(II) and Pb(II) concentration was retained at a constant amount of 200 mg L−1, while the concentrations of co-existing metal ions changed from 250 to 500 mg L−1. Cd(II), Ni(II), Pb(II), and Cu(II) ions had no significant effect on the Hg(II) uptake capacity by OBA/nZVI (see Fig. 14), which confirmed the selectively of OBA/nZVI to remove Hg(II) ions from aqueous solutions in the presence of the above competing metal ions. However, Hg(II) removal efficiency decreased from 58 to 27% with the presence of Cr(VI). This was probably the re duction potential of total Cr ϕ0 CrðVIÞ=Cr ðIIIÞ ¼ þ1:23 V being
Hg(II) Rem
nZVI surface. It is possible to propose that the adsorption on the surface of OBA/nZVI is due to the presence of Fe0. The difference in the trends of Hg(II) and Pb(II) ions can be explained by the size of the ions, which are 1.0 Å for Hg(H2O)62+ and 1.2 Å for Pb(H2O)62+ (Persson 2010). The adsorption efficiency of Hg(II), Pb(II), Ni(II), and Cd(II) ions with the initial concentration of 10 mg L−1 by OBA/nZVI as a function of shaking time is presented in Fig. 13. OBA/nZVI was more efficient to the removal of Hg(II) ions from aqueous media than Pb(II), Ni(II), and Cd(II). In fact, OBA/nZVI composite has a considerably low standard reduction potential which can act as an efficient and eco-friendly electron donor to Hg(II) and Pb(II) ions, transforming metal ions to their reduced form, whereas iron is converted from Fe to Fe2+ accordance with the following reaction (Li and Zhang 2006):
500 Cr
(m ions
t
ntra
nce
co etal
-1 )
gL
250
M
B oval (%)
100
60
Cd Ni Pb Hg
40
80
Pb(II) Rem
80
60 40 20 0 Cd Ni Cu Hg Cr
ls eta ym av He
Removal efficiency (%)
100
20
500 250
0 0
100
200 Time (min)
300
400
Fig. 13 Adsorption kinetics of Hg(II), Pb(II), Ni(II), and Cd(II) ions on OBA/nZVI
-1 gL )
ns (m
ratio ncent
l Co
Meta
Fig. 14 Effect of the co-existence heavy metal ions on the adsorption of Hg(II) (a) and Pb(II) (b) ions
Environ Sci Pollut Res
Hg
total Cd(II), Hg(II), and Cu (II) ions had significant effects on the Pb(II) uptake capacity by OBA/nZVI. OBA/nZVI has the ability of selectively and effectively elimination of Pb(II) ions from aqueous solutions in the presence of the Ni(II) and Cd(II) ions. The effect of the ionic strength on the removal of Hg(II) and Pb(II) ions by OBA/nZVI is shown in Fig. 15. The concentration of NaCl and KCl have low effect on the Hg(II) and Pb(II) removal by OBA/nZVI. It has been proved that the removal of Hg(II) and Pb(II) ions is not only an adsorption process, but also a reduction process. The removal of Hg(II) and Pb(II) ions by OBA/nZVI was efficient in the presence of NaCl and KCl proposing that OBA/nZVI would be beneficial in the elimination of Hg(II) and Pb(II) ions from electroplating wastewater. As mentioned above, the electrochemical reduction is the main sequestration mechanism by OBA/nZVI which can be expressed as follows (Ponder et al. 2000):
Removal efficiency (%)
100
Cl
80 60 40 20
0 0.001M co 0.01M nc en tra 0.1M tio ns (M )
Fresh OBA/nZVI-Hg(II) Fresh OBA/nZVI-Pb(II) Aged OBA/nZVI-Hg(II) Aged OBA/nZVI- Pb(II)
80
60
40
20
0 1
2
3 Cycle
4
5
Fig. 16 Removal efficiency of Hg(II) and Pb(II) ions using the fresh and aged OBA/nZVI composite after five cycles
2Fe þ 3Pb2þ þ 4H2 O 3Pb þ 2FeOOH þ 2 Hþ
ð4Þ
2Fe þ 3Hg2þ þ 4H2 O 3Hg þ 2FeOOH þ 2 Hþ
ð5Þ
The reusability of fresh and aged OBA/nZVI for the sequestration of Hg(II) and Pb(II) ions from wastewater was also investigated (see Fig. 16). The experimental results displayed that OBA/nZVI (10 g L−1) maintained its usability towards Hg(II) and Pb(II) ions for initial concentrations of 45 mg L−1 even after five cycles in the order: fresh (OBA/nZVI)-Hg(II) > fresh (OBA/nZVI)-Pb(II) > aged (OBA/nZVI)-Hg(II) > aged (OBA/nZVI)-Pb(II), which indicate that OBA/nZVI can be used as a regenerated adsorbent. This trend may be related to the oxidation and aggregation of Fe on the surface of the aged OBA/nZVI.
A
Na
100 Removal of Hg(II) and Pb(II) (%)
more positive than Hg(II) (ϕ0 Hg2þ = ¼ þ0:79 V ). Similarly,
Hg Pb
tals y me Heav
B Removal efficiency (%)
100
KC
80 60 40 20
0 0.001M on ce 0.01M nt 0.1M ra tio n (M )
lc
Hg Pb
tals y me Heav
Fig. 15 Effects of the ionic strength (a NaCl and b KCl) on the adsorption of Hg(II) and Pb(II) ions by OBA/nZVI
Fig. 17 X-ray diffraction patterns of Pb(II) and Hg(II) loaded on OBA/ nZVI
Environ Sci Pollut Res
In order to confirm the proposed mechanism, XRD patterns of the OBA/nZVI composite were recorded after impregnation in Pb(II) and Hg(II) solutions (see Fig. 17). Comparison of the experimental XRD data to reference patterns showed that Hg(II) and Pb(II) are reduced to Hg and Pb, and also the presence of other insoluble phases. There are three characteristic peaks for amorphous Pb(OH)2 at 2θ = 22.45, 26.93, and 47.1° (see Fig. 17a) and one for PbO·xH2O at 2θ = 29.74 ° (see Fig. 17b). The peak at 2θ = 56.3 is likely related to iron oxide (see Fig. 17c), while that at 2θ = 35.42, 52.61, 62.42, and 64.51 associated to the zerovalent lead (see Fig. 17d). The peaks at 2θ = 27.2, 33.1, 54.86, 78.2, and 87.2°, appearing in OBA/nZVI after impregnation with Hg(II) solutions, can be assigned to the formation of reduced Hg that represent an important role of nZVI core in the reduction process (see Fig. 17e). Furthermore, some weak signals of oxidized nZVI were also observed between 2θ of 34.96 and 39.98 (see Fig. 17f). Therefore, adsorption/reduction is the predominant mechanisms for the sequestration of metal ions by OBA/nZVI composites.
Summary and conclusions The performance and effectiveness of OBA/nZVI biocomposite as an efficient and eco-friendly adsorbent for the removal of Hg(II) and Pb(II) ions from wastewater were investigated. The effect of initial Hg(II) and Pb(II) ions concentrations, pH, contact time, temperature, adsorbent dosage, effect of competitive metal ions, and ionic strength were studied. The results showed as follows: 1. The kinetic data indicated a good accord with the pseudosecond-order model and the correlation coefficients (R2) were higher than 0.99. The equilibrium sorption data show that the Langmuir–Freundlich isotherm model satisfactorily explained the experimental results. The thermodynamic findings indicate that the sorption process was an endothermic and spontaneous in nature. 2. The efficient adsorption of Hg(II) and Pb(II) ions can be approached in a broad pH range. 3. The results confirmed the selectively of OBA/nZVI to remove Hg(II) ions from aqueous solutions in the presence of the Cd(II), Ni(II), Pb(II), and Cu (II) ions. Similarly, OBA/nZVI has the ability of selectively and effectively separation of Pb(II) ions from aqueous media in the presence of the Ni(II) and Cd(II) ions. Furthermore, the removal of Hg(II) and Pb(II) ions by OBA/nZVI was efficient in the presence of NaCl and KCl. 4. It can be found that fresh and aged OBA/nZVI maintained its usability even after five cycles in the order: fresh (OBA/nZVI)-Hg(II) > fresh (OBA/nZVI)-Pb(II) > aged (OBA/nZVI)-Hg(II) > aged (OBA/nZVI)-Pb(II), which
indicate that OBA/nZVI can be used as a regenerated adsorbent. 5. Low durability, aggregation, oxidation, and poor stability of the uncoated nZVI are the major problems of these particles for applications to in situ groundwater remediation which could be easily solved by stabilization of nZVI on the OBA surface. 6. The dual adsorption/reduction are the predominant mechanisms for the removal of metal ions by OBA/nZVI composite. 7. In view of the practical application of adsorbent, OBA/ nZVI composite can be applied in permeable reactive barriers. Funding information The authors would like to thank Iranian Nanotechnology Initiative for supporting of this work.
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