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Use of waste ceramics in adsorption technologies
CLAY-03765; No of Pages 8 Applied Clay Science xxx (2016) xxx–xxx
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Use of waste ceramics in adsorption technologies Barbora Dousova a,⁎, David Kolousek a, Martin Keppert b, Vladimir Machovic a, Miloslav Lhotka a, Martina Urbanova c, Jiri Brus c, Lenka Holcova a a b c
University of Chemistry and Technology Prague, Technicka 5, 166 28 Prague 6, Czech Republic Faculty of Civil Engineering, Czech Technical University in Prague, Thakurova 7, 166 29 Prague 6, Czech Republic Institute of Macromolecular Chemistry AS CR, Heyrovskeho nam. 2, 162 06 Prague 6, Czech Republic
a r t i c l e
i n f o
Article history: Received 29 October 2015 Received in revised form 8 January 2016 Accepted 15 February 2016 Available online xxxx Keywords: Waste ceramics Brick dust Toxic cations Toxic anions Adsorption capacity Adsorption efficiency Leaching test
a b s t r a c t Waste brick dust (WBD) was tested as a potential sorbent of cationic and anionic contaminants, including radioactive residues. For adsorption experiments, model water solutions of highly toxic and/or ecologically harmful cations (Cd, Pb, Cs) and anions (As, Sb, Cr, U) were selected. The adsorption of Cd2+ and Pb2+ on WBD was most effective (N 95%) at a very low sorbent dosage (up to 6 g L−1). In terms of anionic contaminants, UVI was adsorbed as cationic complex particles [(UO2)n(OH)2n − 1]+ almost quantitatively (N 95%) at a sorbent dosage of 3 g L−1. The effective adsorption of AsV (N 90%) occurred at around a dosage of 15 g L−1. The adsorption of Cs+, CrVI and AsIII on WBD was almost ineffective. Except for Cs+ and CrVI, all investigated ions were adsorbed according to the Langmuir isotherm model, at the theoretical adsorption capacities Q t ≈ approximately 0.1 mmol g − 1 for Cd 2 + , Pb2 + and UVI, and approximately. 0.04 for AsV and AsIII. The leachability of toxic particles from saturated WBD was very low for selectively adsorbed particles (≈ 0.01–0.08% wt.) and their stability decreased in the order: Pb2+ ≈ Cd2+ N UVI N AsV N AsIII ≫ Cs+ ≫ CrVI. The approximate consumption of WBD per gram of toxic element was found to be about ≈60 g for Pb2+ and UVI, ≈100 g for Cd2+ and N400 g for AsV. © 2016 Elsevier B.V. All rights reserved.
1. Introduction Waste ceramics, including brick dust, represent a feasible and challenging material due to several aspects, such as its chemical stability, availability, environmental safety issues, fineness, and silicate properties. The waste ceramic materials can be generally recycled either as aggregates in concrete production (Silva et al., 2014) or, when the waste particles are fine one can take advantage of ceramic composition (high content of active silica and alumina) and apply it as pozzolanic component of cement based materials and thus reduce the Portland cement consumption (Kulovana et al., 2015). In analogy with natural aluminosilicates, these materials may also be applied as cheap but effective sorbents of ionic species in technological or environmental adsorption processes. Illite is the principal component of many raw materials used for the production of traditional ceramics, including bricks and roof tiles (Dondi et al., 2014); illite improves the plasticity of clayey matter. During the drying and firing process, illite first loses its interlayer water (50– 400 °C); illite dihydroxylation then takes place in two steps between 450 and 800 °C. A glass is then formed above 900 °C (Carroll et al., 2005; Ferrari and Gualtieri, 2006) and works as a flux in ceramics. The kinetics of illite dehydroxylation have been described by Ferrari and Gualtieri (2006). Favorable adsorption properties result from a proper
⁎ Corresponding author. E-mail address: [email protected] (B. Dousova).
chemical composition (Al, Si, Fe content) and surface hydration of the aluminosilicate structure (Lin and Puls, 2000; Zanelli et al., 2015). The mobility of environmental contaminants can be particularly controlled by their chemistry and phase stability related to the pH/Eh of investigated system (Mc Lean and Bledsoe, 1992). Typical cationic (heavy metals) and anionic (metalloids, non-metals) contaminants generally attract differently charged active sites on a sorbent surface, therefore the pH of zero point of charge, pHZPC (Fiol and Villaescusa, 2009) can be considered the first indicator of sorbent selectivity. In natural water systems, solids with a low pH ZPC (aluminosilicates, quartz) represent mostly cation active sorbents (Jiang et al., 2009; Padilla-Ortega et al., 2013), while a high pHZPC is typical for anion active sorbents (Fe/Al oxy(hydroxides), gibbsite) (Cai et al., 2015; Oliviera et al., 2003). The surface of mixed materials (soils, sediments, brick dust) mostly consists of diverse active sites, and the pHZPC represents more or less the average value of particular components of the solid matrix. Therefore, the sorption selectivity of mixed materials is not unique, and under appropriate conditions they can attract both cationic and anionic particles. The aim of this work was to verify the sorption properties of waste brick dust in a raw and water-leached state, without any technological pre-treatment. For this purpose, selected toxic heavy metals (PbII, CdII), anionic contaminants (AsV/AsIII, CrVI) and radioactive wastes (CsI, UVI) were investigated in model aqueous systems. Perspectively, a waste brick dust saturated with toxic ions might be incorporated into cement building material.
http://dx.doi.org/10.1016/j.clay.2016.02.016 0169-1317/© 2016 Elsevier B.V. All rights reserved.
Please cite this article as: Dousova, B., et al., Use of waste ceramics in adsorption technologies, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/ j.clay.2016.02.016
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2. Materials and methods 2.1. Waste brick dust The studied sorbent, waste brick dust (WBD), is generated as a waste (grinding dust) during the production of vertically perforated ceramic blocks intended for thin joint masonry. Currently the grinding dust is partially recycled in the production line – it is dosed as non-plastic component of green ceramic mixture. The amount recyclable by this way is obviously limited and thus the surplus ceramic waste is dumped. The elementary chemical and mineralogical composition of the raw ceramic mixture (“green”) used for the production of red-clay ceramics is given in Table 1. The dominant clay mineral was illite (ca. 30%), while the green body further contained chlorite, quartz (ca. 25%), hematite, albite, orthoclase, muscovite, calcite and dolomite. During firing, illite and chlorite were transformed into amorphous matter, and calcite and dolomite decomposed. Akermanite (Ca2MgSi2O7) and hedenbergite (CaFeSi2O6) are the crystalline products of the reactions between illite and carbonates present in the material. To reduce the high alkalinity and improve the attraction to anions, the raw WBD was washed five times in distilled water at laboratory temperature (20 °C) at a solid-liquid ratio of 1:50. During the procedure, the pH value decreased almost by 3 units, i.e. from 12.5 to 9.8. The washed WBD is referred to as WBDw. 2.2. Model solutions − 2− Model solutions of Pb2 +, Cd2 +, Cs+, H2AsO− and 4 , AsO2 , Cr2O7 2− U2O7 were prepared from inorganic salts (PbCl2, Cd(NO3)2, CsCl, KH2AsO4, NaAsO2, (NH4)2Cr2O7, (NH4)2U2O7) of analytical grade and distilled water, in the concentrations of 0.1 and 0.5 mmol·L−1 and the natural pH (i.e. pH ≈ 3.5 for cationic solutions and pH 5–6 for anionic solutions). The concentration range was selected as appropriate for the simulation of a slightly increased amount of the contaminant in a water system to a heavily contaminated solution. The solution of AsIII as NaAsO2 was prepared under an N2 atmosphere, at pH ≈ 7.4–8.5 (Doušová et al., 2009, 2011).
2.3. Sorption experiments During the adsorption experiments the cationic particles (Cd2 +, Pb and Cs+, respectively) were adsorbed on the raw WBD, whereas the anions (AsIII/V, CrVI, UVI) were adsorbed on the water-washed WBDw. A suspension of the model solution (50 mL) and defined dosage (0.5–15 g L−1) of WBD (WBDw) was shaken using a batch procedure at laboratory temperature (20 °C) for 24 h (Doušová et al., 2006). The product was filtered and the filtrate was analyzed for residual cations/anions, while the solid residue (saturated WBD/WBDw) was kept for leaching tests. WBD saturated with Pb2+ and WBDw saturated with CrVI, AsIII and AsV, respectively, were also tested by the solid-state NMR spectroscopy and SBET measurements to verify the structural and/ or surface changes caused by adsorption. All sorption data were fitted to the Langmuir equation (Jeong et al., 2007, Misak, 1993), which was 2+
verified as a suitable adsorption model for natural oxides, aluminosilicates and soils. The data illustrated the parameters of adsorption (qmax – maximum equilibrium sorption capacity; Qtheor – theoretical sorption capacity; R2 – correlation factor; KL – Langmuir adsorption constant), which were used to compare the adsorption affinities of the investigated ions on brick dust. 2.4. Analytical methods Powder X-ray diffraction (XRD) was performed with a Seifert XRD 3000P diffractometer with CoKα radiation (λ = 0.179026 nm, graphite monochromator, goniometer with Bragg-Brentano geometry) in the 2θ range of 5–60° with a step size of 0.05° 2θ. X-ray fluorescence (XRF) analyses of the solid phase were determined with an ARL 9400 XP+ spectrometer with a voltage of 20–60 kV, probe current of 40–80 mA and effective area of 490.6 mm2. UniQuant software was used for data evaluation. The specific surface area (SBET) was measured on a Micromeritics ASAP 2020 (accelerated surface area and porosimetry) analyzer using the gas sorption technique. The ASAP 2020 model assesses single and multipoint BET surface area, Langmuir surface area, Temkin and Freundlich isotherm analysis, pore volume and pore area distributions in the micro- and macro-pore ranges by the BJH method. The micropore option used the Horvath-Kavazoe method, with N2 as the analysis adsorptive and an analysis bath temperature of −195.8 °C. The samples were degassed at 313 K for 1000 min. The IR spectra were collected on a Nicolet 6700 FTIR (Thermo Nicolet Instruments Co.) with N2 purging system. Spectra were acquired using a single reflection ATR SmartOrbit accessory equipped with a singlebounce diamond crystal (angle of incidence: 45°). A total of 64 scans were averaged for each sample and the resolution was 2 cm− 1. The spectra were rationed against a single-beam spectrum of the clean ATR crystal and converted into absorbance units by ATR correction. Data were collected in the wavenumber range of 4000–400 cm−1. Solid-state NMR spectra were measured at 11.7 T using a Bruker AVANCE III HD 500 WB/US NMR spectrometer. The 27Al MAS NMR spectra were acquired at a spinning frequency of 11 kHz, Larmor frequency of 130.287 MHz and recycle delay of 2 s, and the spectra were referenced to the external standard Al(NO3)3 (0 ppm). The number of scans for the acquisition of a single 27Al MAS NMR spectrum was 512 (total experimental time was ca. 10 min). The 29Si MAS NMR spectra were acquired at a spinning frequency of 11 kHz, Larmor frequency of 99.325 MHz and recycle delay of 10 s. The number of scans for the acquisition of a single 29Si MAS NMR spectrum was 6144 (total experimental time was ca. 17 h). The spectra were referenced to the external standard M8Q8 (− 109.8 ppm). The NMR experiments were performed at a temperature of 24 °C, and temperature calibration was performed to correct for the frictional heating of the samples. The concentration of Pb/Cd/Cs in aqueous solutions was determined by atomic absorption spectrometry (AAS) using a SpectrAA-880 VGA 77 unit (Varian) in flame mode. The concentration of As in aqueous solutions was determined by Hydride Generation Atomic Fluorescence Spectrometry (HG-AFS) using a PSA 10.055 Millennium Excalibur apparatus. The samples were
Table 1 Chemical and mineralogical composition of the green ceramic mixture and waste brick dust (WBD). Chemical composition (% wt)
Green
WBD
Na2O/K2O MgO/CaO Al2O3 SiO2 P2O5 Fe2O3 TiO2
1.1/4.2 3.8/10.6 20.9 51.8 0.9 6.6 0.9
1.3/4.1 4.5/11.5 20.0 51.3 0.9 5.9 0.8
Mineralogical composition
Green
WBD
Quartz (25%) Chlorite Illite (30%) Albite Orthoclase Muscovite Calcite Dolomite
Quartz (22%) Microcline Hematite Albite Orthoclase Muscovite Akermanite Hedenbergite
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pre-treated with a solution of HCl (As-free, 36% v/w) and KI (50%) with ascorbic acid (10%). The instrumental parameters included ppm and ppb modes, HCl (12%) with KI + ascorbic acid solution as the reagent blank and 7% NaBH4 in 0.1 mol·L− 1 NaOH as the reductant. The declared detection limit was 0.05 ppm and the standard deviation was experimentally determined as 2.5%. The concentration of Cr as Cr2 O 27 − in aqueous solutions was measured with a UV/VIS spectrophotometer (UNICAM 5625) at 350 nm following acidification with H2SO4 (10% wt.) (Malat, 1973). Uranium concentrations were measured using an atomic (optical) emission spectrometer with inductively coupled plasma (ICP-OES; Optima 2000-DV, Perkin-Elmer Instruments). The concentration of U is given as mg L−1 of U. The limit of detection for the determination of the concentrations of arsenic was 3 μg L−1. 2.5. Leaching test The stability of toxic ions in the saturated sorbents was verified by a standard procedure EN 12457–2 (CEN, 2002); a suspension of dry sorbent and distilled H2O at a ratio of 1:10 was agitated in a sealed polyethylene reactor at laboratory temperature for 24 h. Then, the suspension was filtered off and the filtrate was analyzed for residual cations/anions. 2.6. Analytical quality control, reproducibility of obtained data An accuracy of AAS analyses was guaranteed by the Laboratory of Atomic Absorption Spectrometry, ICT Prague (CR); the detection limit was 0.5 μg L−1 with a standard deviation of 5–10%). For HG-AFS, the declared detection limit was 0.05 mg L−1 and the standard deviation was experimentally determined as 2.5%. The reproducibility of leaching tests and UV/VIS spectrophotometry was checked in parallel experiments; the standard deviation resulting from 10 experimental sets ranged from 4.5 to 7% for the leaching tests, while it did not exceed 5% for the spectrophotometric analyses. 3. Results and discussion For the adsorption experiments, highly toxic and/or ecologically harmful cations (Cd, Pb, Cs) and anions (As, Sb, Cr, U) were selected according to their current environmental concern on the international scale. The study proceeded in two directions: (i) the adsorption of cationic/anionic particles from aqueous solution on WBD/WBDw for the determination of the optimal sorption parameters, and (ii) the verification of the stability of fixed particles in the saturated sorbent.
decrease in alkali material, particularly CaO, and soluble sulfates such as SO3, which corresponded well with about a ten-fold increase in the SBET (Table 2) and porosity (Fig. 1) of WBDw compared to raw WBD. These results suggest that weakly bound soluble inorganic salts (CaSO4, Na2SO4), causing the negligible porosity of raw WBD, were removed by water-washing. The XRD of WBDw illustrated a slight increase in crystallinity compared to WBD, which corresponded well with the previous data. The structural, chemical and mineralogical changes to the raw (WBD) and water-washed (WBDw) sorbents were studied by XRD, XRF, solid-state NMR and SBET measurements. In the XRD of WBDw quartz intensities showed a slight increase indicating a minor arise of sample crystallinity at the expense of any amorphous phase released by the water-washing of raw WBD. The significant decrease of alkali phases in WBDw was verified by the ATR-FTIR spectroscopy of the solid residue of WBD water leachate, which identified mostly CaCO3 and CaSO4, respectively (Fig. 2). The recorded 27Al and 29Si MAS NMR spectra of raw (WBD) and modified (WBDw) brick dust are shown in Fig. 3. The signal assignment was performed based on literature data (Mackenzie and Smith, 2002) and our previous results (Brus et al., 2012). A broad signal resonating at ca. 59 ppm with a half width of ca. 1600 Hz was detected in the 27Al MAS NMR spectra of raw WBD and modified WBDw sorbents. The resonance frequency reflects the tetrahedral coordination of these aluminum sites, while resonance broadening indicates the amorphous character of these systems. By washing the system by water two lowintensity 27Al MAS NMR signals at ca. 14 and 10 ppm in the modified WBDw show the presence of two different types of hexa-coordinated aluminum sites. With high probability a small amount of sodium ions is removed from the surface area. Consequently, the system is not completely charge-compensated and a limited transformation of AlO4 framework units to free (dissolved) extra-framework AlO6 monomer moieties occurs. As indicated by the intensity of these signals that is b5% we, however, suppose that this process does not affect considerably the overall structure of the synthesized sorbent. The 29Si MAS NMR spectra consisted of a single signal centered around − 91 ppm with a two small signals at − 73 and − 108 ppm. The main signal is usually attributed to fully condensed aluminumsubstituted silica tetrahedra Q4(4Al) and Q4(3Al) accompanied by minor fractions of Q0 and SiO2 units. After water-washing WBD, the transformation of Q0 units occurred, since the amount of Q0 units decreased from 10% to 5%. Simultaneously, the amount of Q4 (4Al) units increased and SiO2 units remained constant at about 23%.
3.1. Characterisation of raw WBD and water-washed WBDw Table 2 illustrates the chemical and surface changes induced by water washing of WBD. The water-washed WBDw showed a significant Table 2 Chemical composition, pH of the zero point of charge and SBET of the raw (WBD) and water-washed (WBDw) sorbent. Composition (% wt)
WBD
WBDw
Na2O/K2O MgO/CaO Al2O3 SiO2 P2O5 SO3 Fe2O3 TiO2 MnO2 SBET (m2 g−1) pHZPC
1.3/4.1 4.5/11.5 21.0 51.3 0.9 1.1 5.9 0.8 0.04 3.31 4.4
0.7/3.8 4.4/ 20.1 50.6 0.7 0.4 6.2 0.9 0.05 32.51 4.8
3
Fig. 1. BHJ pore-volume distribution of WBD/WBDw.
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Fig. 2. WBD-ATR-FTIR spectrum of the solid residue of water leachate.
3.2. Adsorption parameters for cationic and anionic contaminants The adsorption parameters calculated according to the Langmuir model for cations (Table 3) showed maximum values of adsorption capacity for Cd2 + and Pb2 + adsorbed on WBD, which corresponded well to the previous results of Doušová et al. (2006, 2009) and indicated the formation of inner-sphere surface complexes (Sherman and Randall, 2003). The less balanced, non-Langmuir adsorption run towards the larger ionic radius cations (Cd2+ b Pb2+ ≪ Cs+) could be related to the very low SBET and negligible porosity of WBD, resulting in a steric barrier to the surface binding of larger particles. This phenomenon was previously described by Doušová et al. (2009), who compared the intensity of Fe2+/Fe3+ surface modification of kaolins. A negligible effect of Cs+ adsorption resulted primarily from steric factors. According to Cornell (1993), the differences in Cs uptake on different minerals may be related to different cation exchange capacities, structure, composition of the mineral and the surface area/particle size ratio. As mentioned before, illite, which is generally considered the most selective Cs adsorbent among the clay minerals (Cornell, 1993; Staunton and Roubaud, 1997), is transformed during the firing process to an amorphous phase (Vejmelková et al., 2012). The WBD sorbent consisted mostly of quartz,
feldspars and hematite, which are poor Cs adsorbents (Bergaoui et al., 2005). The comparable adsorption parameters for anionic contaminants (Table 4), which were adsorbed on WBDw, indicated a generally lower sorption selectivity of WBDw to anions, which may be mainly related to the pH/pHZPC ratio. The pHZPC b 5 and pH/pHZPC ≫ 1 for both the WBD and WBDw denoted prevailing cation active surface sites and charge density, which inhibited the fixation of anions due to strong repulsion (Fiol and Villaescusa, 2009). Another reason may be the different ability of oxyanions to undergo surface binding. Arsenates (AsV) were specifically adsorbed, forming stable inner-sphere, predominantly bidentate complexes with iron oxides (Fendorf et al., 1997; Grossl et al., 1997). Arsenites (AsIII) and chromates (CrVI) tended to form both inner- and outer-sphere complexes; in the case of innersphere complexes, a weaker monodentate form prevailed (Grossl et al., 1997). The effect of pH on adsorption indicated a different trend. The adsorption of AsV/CrVI on iron oxides decreased with increasing pH, while AsIII in the anionic form (H2AsO− 3 ), which is fundamental for selective adsorption, appeared at pH ≥ 9 (Takeno, 2005). All these findings correspond well with the adsorption data. The adsorption stability and selectivity for anionic particles declined in the order: AsV N AsIII ≫ CrVI. The adsorption on WBD/WBDw was most probably associated with variable arrangements of adsorption mechanism related to a high pH value. According to Stumm (1992), at a high pH the formation of surface precipitates can appear and participate markedly in surface binding. 3− During adsorption on WBD/WBDw anionic particles (AsO3− and 4 , AsO3 2− CrO4 , resp.) were captured at anion active surface sites, but also partially linked by chemical bonds to released alkali cations (Na+, K+, Ca2+),
Table 3 Langmuir parameters for adsorption of cations on WBD.
Fig. 3. 27Al and 29Si MAS NMR spectra of WBD and water-washed WBD.
Cation
Initial concentration (mmol L−1)
qmax (mmol g−1)
Qtheor (mmol g−1)
KL (L mmol−1)
R2
Cd2+ Cd2+ Pb2+ Pb2+ Cs+ Cs+
0.1 0.5 0.1 0.5 0.1 0.5
0.06 0.26 0.11 0.51 0.01 0.04
0.05 0.13 0.12 –⁎ –⁎ –⁎
3063.5 3951.8 1017.3 − − −
0.875 0.895 0.695 − − −
⁎ Non-Langmuir run.
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Table 4 Langmuir parameters for adsorption of anions on WBDw. Anion
Initial concentration (mmol L−1)
qmax (mmol g−1)
Qtheor (mmol g−1)
KL (L mmol−1)
R2
AsIII AsIII AsV AsV CrVI CrVI UVI UVI
0.1 0.5 0.1 0.5 0.1 0.5 0.1 0.5
0.005 0.015 0.01 0.04 0.006 0.01 0.08 0.13
–⁎ 0.04 0.02 0.04 –⁎ –⁎
– 1.3 41.8 281.1 – – 368.0 3232.5
– 0.784 0.695 0.963 – – 0.984 0.785
0.08 0.11
⁎ Non-Langmuir run.
forming less soluble or insoluble surface precipitates. This effect prevailed in the raw WBD with a higher salt content and has been clearly shown in the infrared spectra of AsV adsorption on WBD/WBDw (Fig. 4). The changes in tetraedric symmetry of AsO3− complex, which can be caused 4 by the solution pH, type of surface complexation (inner-sphere vs. outersphere), type of soluble cations and H2O structure resulted in the shift and splitting of spectral bands (Goldberg and Cliff, 2001; Myneni et al., 1998; Roddick-Lanzilotta et al., 2002). Subtracted A and B spectra in Fig. 4 illustrate AsV adsorption on WBD (A) and WBDw (B). The bands at about 858 cm−1 dominate in HAsO3− spectrum acquired at pH ≈ 9.2 4 (Roddick-Lanzilotta et al., 2002), therefore, a surface complexation via HAsO3− can be considered. According to Myneni et al. (1998), typical 4 bands at 798–812 and 855–905 cm−1 can be assigned to stretching vibration of As\\OCa, indicating the formation of different crystalline calcium arsenates as surface precipitation. The unique sorption parameters of UVI (Table 4) resulted from the 2+ variability of uranium phases and stable UO+ forms under 2 and UO2 oxidic conditions and in the pH range of ≈ 3.5–10 (Takeno, 2005). Uranium as UVI was most likely bound as a cationic complex particle [(UO 2) n (OH) 2n − 1 ] +. The optimal adsorption capacities obtained under the same experimental conditions for all the investigated elements are compared in Fig. 5. 3.3. Adsorption efficiency and sorbent consumption To assess the possible practical uses of WBD, not only its adsorption capacities to selected elements but also the actual adsorption efficiencies
Fig. 5. Maximum and theoretical sorption capacities for investigated contaminants.
suggestive of its approximate consumption in technological processes had to be studied. Both Cd and Pb adsorption showed a high sorption efficiency at a low sorbent dosage (see Fig. 6a). Almost quantitative adsorption from less concentrated solutions (≈0.1 mmoL−1) was achieved at a dosage of 1 g L−1 for Pb2+ and 2 g L−1 for Cd2+. The dosage of about 6 g L−1 was sufficient for the quantitative removal of Pb2+/Cd2+ from highly concentrated solutions (≈0.5 mmol L−1). Of the tested anions, UVI and AsV are potential candidates for technological applications (Fig. 6b). Effective adsorption was observed at the sorbent dosage of 3 g L−1 for UVI and 15 g L−1 for AsV. Regardless the value of the maximum adsorption capacities (Table 4), CrVI was adsorbed more effectively at a lower initial concentration, while AsV and AsIII showed the opposite trend. In the case of UVI, the adsorption efficiency was independent of the initial concentration. These results confirmed the adsorption theory, in that non-specific adsorption (electrostatic bonding mechanisms, weak binding energy, outer-sphere complexation) strongly depends on the ionic strength (Stumm, 1992). Inner-sphere surface complexes are specifically bound to active sites, involving covalent bonding and/or combinations of covalent and ionic bonding. Therefore, Pb2+, Cd2+ and UVI adsorption on WBD/WBDw occurred almost specifically with a negligible effect of the surrounding ionic particles and initial concentration. In this
Fig. 4. Subtracted ATR-FTIR spectra of AsV adsorption on WBD/WBDw. A – substracted spectrum of WBD after AsV adsorption and raw WBD, B – subtracted spectrum of WBDw after AsV adsorption and WBDw.
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Fig. 6. Adsorption efficiency as a function of WBD dosage. a) cationic contaminants, b) anionic contaminants, c0 ≈ initial concentration of contaminant.
case, the binding of cations directly to tertahedral (SiO4) or octahedral (AlO6) coordinations of clay structure can be considered. Conversely, the non-specific adsorption of AsIII and CrVI has been markedly affected by the ionic strength of adsorbate and the charge distribution of AsO3− 3 and CrO2− oxyanions. 4 The approximate consumption of sorbent per gram of toxic element (Fig. 7) indicated an important technological specification and allowed us to estimate the applicability of the method under certain conditions. The approximate mass consumption of sorbent per gram of toxic element was found to be about 60 g for Pb and U, about 100 g for Cd and more than 400 g for As as AsV. The other contaminants (As as AsIII, Cr, Cd) likely have no perspective to be removed with WBD due to extremely high sorbent consumption, mainly due to insufficient adsorption efficiency (b30%).
Fig. 7. Hypothetic consumption of WBD per 1 g of contaminated element.
Fig. 8. 27Al MAS NMR spectra of WBD/WBDw after adsorption of Pb2+, CrVI, AsIII and AsV, respectively.
Please cite this article as: Dousova, B., et al., Use of waste ceramics in adsorption technologies, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/ j.clay.2016.02.016
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3.4. Structural and surface changes of WBD/WBDw during adsorption
Acknowledgements
Since adsorption represents in principle a surface process, only small structural changes can be expected. From the comparison of 27Al MAS NMR spectra before and after adsorption of toxic elements (Fig. 8), it was evident that the WBD/WBDw structure was uniform and the adsorption had no influence on the structure of aluminosilicate networks. Similarly, the SBET and porosity of WBD/WBDw saturated with AsO3− 4 did not change markedly (Fig. 1).
This work was the part of the project 13-24155S (Grant Agency of Czech Republic).
3.5. Leaching test The stability of cations/anions in saturated WBD/WBDw was evaluated by the leaching test described in Section 2.6. As evident from Table 5, the stability of the saturated sorbent was closely related to the selectivity of particular adsorption; Pb2+ and Cd2+ adsorption was characterized by high sorption capacities and efficiencies, indicating the formation of inner-sphere surface complexes resulting in very stable binding in both leaching environments (no more than 0.08% of Pb2+/Cd2+ was leached). Favorable stability (about 1% of the initial amount leached) was observed for AsV and UVI (Table 5), which corresponded well with the adsorption data. The non-selective Cs+/CrVI adsorption showed the worst binding stability during the leaching test. Under the applied experimental conditions, WBD/WBDw saturated with Pb2+, Cd2+ and AsV, respectively, could be considered a non-hazardous waste (CSN EN 12457). However, the expected fixation of saturated WBD/WBDw in building materials, such as cement and concrete, might markedly change the stability of toxic particles. 4. Conclusions Waste brick dust is an effective sorbent of selected cationic and anionic toxic particles. The main advantages the low price of this material and the possibility of incorporating the saturated sorbent into a cementitious building material, which may eliminate the hazardous impact of toxic elements due to its stabilization/solidification efficiency. According to the phase composition and pHZPC value, WBD is more effective as a cation active sorbent, but not uniquely. The water-washing of raw WBD resulted in a reduction in alkalinity by ˃2.5 pH units, and in more than a ten-fold increase in SBET. Nevertheless, the adsorption properties of WBDw to anionic particles did not improve markedly, which suggested the participation of surface precipitates in the binding mechanism on raw WBD. Heavy metals such as Pb2+/Cd2+ and uranium as [(UO2)n(OH)2n − 1]+ were adsorbed selectively onto WBD/ WBDw at high adsorption efficiency and low sorbent consumption. Arsenic as AsO34 − was also adsorbed, but with more than four times greater sorbent consumption. The adsorption of cesium as Cs+ and chromium as CrO24 − on WBD/WBDw was totally ineffective. The leaching stability of the adsorbed particles decreased in the order: Pb2+ ≈ Cd2+ ˃ UVI ˃ AsV ˃ AsIII ≫ Cs+ ≫ CrVI. For prospective use, the study of long-term stability of saturated WBD/WBDw incorporated into building material will be essential.
Table 5 Leachability of toxic particles from WBD(1) and WBD(2) w . Toxic particle
Initial amount (mg/g)
leached in KCl (%)
leached in H2O (%)
Class of leachabilitya
AsV (2) AsIII (2) CrVI (2) UVI (1) Pb2+ (1) Cd2+ (1) Cs+ (1)
2.80 0.80 0.01 14.70 5.90 7.00 1.40
1.2 2.3 35.5 0.6 0.02 0.01 18.1
0.8 2.4 41.1 1.2 0.07 0.08 4.3
II – other waste III - hazardous waste III - hazardous waste not limited I – inert waste II – other waste Not limited
(1), (2) a
the type of used sorbent. Regulation No. 294/2005 (CSN EN 12457).
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Use of waste ceramics in adsorption technologies Barbora Dousova a,⁎, David Kolousek a, Martin Keppert b, Vladimir Machovic a, Miloslav Lhotka a, Martina Urbanova c, Jiri Brus c, Lenka Holcova a a b c
University of Chemistry and Technology Prague, Technicka 5, 166 28 Prague 6, Czech Republic Faculty of Civil Engineering, Czech Technical University in Prague, Thakurova 7, 166 29 Prague 6, Czech Republic Institute of Macromolecular Chemistry AS CR, Heyrovskeho nam. 2, 162 06 Prague 6, Czech Republic
a r t i c l e
i n f o
Article history: Received 29 October 2015 Received in revised form 8 January 2016 Accepted 15 February 2016 Available online xxxx Keywords: Waste ceramics Brick dust Toxic cations Toxic anions Adsorption capacity Adsorption efficiency Leaching test
a b s t r a c t Waste brick dust (WBD) was tested as a potential sorbent of cationic and anionic contaminants, including radioactive residues. For adsorption experiments, model water solutions of highly toxic and/or ecologically harmful cations (Cd, Pb, Cs) and anions (As, Sb, Cr, U) were selected. The adsorption of Cd2+ and Pb2+ on WBD was most effective (N 95%) at a very low sorbent dosage (up to 6 g L−1). In terms of anionic contaminants, UVI was adsorbed as cationic complex particles [(UO2)n(OH)2n − 1]+ almost quantitatively (N 95%) at a sorbent dosage of 3 g L−1. The effective adsorption of AsV (N 90%) occurred at around a dosage of 15 g L−1. The adsorption of Cs+, CrVI and AsIII on WBD was almost ineffective. Except for Cs+ and CrVI, all investigated ions were adsorbed according to the Langmuir isotherm model, at the theoretical adsorption capacities Q t ≈ approximately 0.1 mmol g − 1 for Cd 2 + , Pb2 + and UVI, and approximately. 0.04 for AsV and AsIII. The leachability of toxic particles from saturated WBD was very low for selectively adsorbed particles (≈ 0.01–0.08% wt.) and their stability decreased in the order: Pb2+ ≈ Cd2+ N UVI N AsV N AsIII ≫ Cs+ ≫ CrVI. The approximate consumption of WBD per gram of toxic element was found to be about ≈60 g for Pb2+ and UVI, ≈100 g for Cd2+ and N400 g for AsV. © 2016 Elsevier B.V. All rights reserved.
1. Introduction Waste ceramics, including brick dust, represent a feasible and challenging material due to several aspects, such as its chemical stability, availability, environmental safety issues, fineness, and silicate properties. The waste ceramic materials can be generally recycled either as aggregates in concrete production (Silva et al., 2014) or, when the waste particles are fine one can take advantage of ceramic composition (high content of active silica and alumina) and apply it as pozzolanic component of cement based materials and thus reduce the Portland cement consumption (Kulovana et al., 2015). In analogy with natural aluminosilicates, these materials may also be applied as cheap but effective sorbents of ionic species in technological or environmental adsorption processes. Illite is the principal component of many raw materials used for the production of traditional ceramics, including bricks and roof tiles (Dondi et al., 2014); illite improves the plasticity of clayey matter. During the drying and firing process, illite first loses its interlayer water (50– 400 °C); illite dihydroxylation then takes place in two steps between 450 and 800 °C. A glass is then formed above 900 °C (Carroll et al., 2005; Ferrari and Gualtieri, 2006) and works as a flux in ceramics. The kinetics of illite dehydroxylation have been described by Ferrari and Gualtieri (2006). Favorable adsorption properties result from a proper
⁎ Corresponding author. E-mail address: [email protected] (B. Dousova).
chemical composition (Al, Si, Fe content) and surface hydration of the aluminosilicate structure (Lin and Puls, 2000; Zanelli et al., 2015). The mobility of environmental contaminants can be particularly controlled by their chemistry and phase stability related to the pH/Eh of investigated system (Mc Lean and Bledsoe, 1992). Typical cationic (heavy metals) and anionic (metalloids, non-metals) contaminants generally attract differently charged active sites on a sorbent surface, therefore the pH of zero point of charge, pHZPC (Fiol and Villaescusa, 2009) can be considered the first indicator of sorbent selectivity. In natural water systems, solids with a low pH ZPC (aluminosilicates, quartz) represent mostly cation active sorbents (Jiang et al., 2009; Padilla-Ortega et al., 2013), while a high pHZPC is typical for anion active sorbents (Fe/Al oxy(hydroxides), gibbsite) (Cai et al., 2015; Oliviera et al., 2003). The surface of mixed materials (soils, sediments, brick dust) mostly consists of diverse active sites, and the pHZPC represents more or less the average value of particular components of the solid matrix. Therefore, the sorption selectivity of mixed materials is not unique, and under appropriate conditions they can attract both cationic and anionic particles. The aim of this work was to verify the sorption properties of waste brick dust in a raw and water-leached state, without any technological pre-treatment. For this purpose, selected toxic heavy metals (PbII, CdII), anionic contaminants (AsV/AsIII, CrVI) and radioactive wastes (CsI, UVI) were investigated in model aqueous systems. Perspectively, a waste brick dust saturated with toxic ions might be incorporated into cement building material.
http://dx.doi.org/10.1016/j.clay.2016.02.016 0169-1317/© 2016 Elsevier B.V. All rights reserved.
Please cite this article as: Dousova, B., et al., Use of waste ceramics in adsorption technologies, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/ j.clay.2016.02.016
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2. Materials and methods 2.1. Waste brick dust The studied sorbent, waste brick dust (WBD), is generated as a waste (grinding dust) during the production of vertically perforated ceramic blocks intended for thin joint masonry. Currently the grinding dust is partially recycled in the production line – it is dosed as non-plastic component of green ceramic mixture. The amount recyclable by this way is obviously limited and thus the surplus ceramic waste is dumped. The elementary chemical and mineralogical composition of the raw ceramic mixture (“green”) used for the production of red-clay ceramics is given in Table 1. The dominant clay mineral was illite (ca. 30%), while the green body further contained chlorite, quartz (ca. 25%), hematite, albite, orthoclase, muscovite, calcite and dolomite. During firing, illite and chlorite were transformed into amorphous matter, and calcite and dolomite decomposed. Akermanite (Ca2MgSi2O7) and hedenbergite (CaFeSi2O6) are the crystalline products of the reactions between illite and carbonates present in the material. To reduce the high alkalinity and improve the attraction to anions, the raw WBD was washed five times in distilled water at laboratory temperature (20 °C) at a solid-liquid ratio of 1:50. During the procedure, the pH value decreased almost by 3 units, i.e. from 12.5 to 9.8. The washed WBD is referred to as WBDw. 2.2. Model solutions − 2− Model solutions of Pb2 +, Cd2 +, Cs+, H2AsO− and 4 , AsO2 , Cr2O7 2− U2O7 were prepared from inorganic salts (PbCl2, Cd(NO3)2, CsCl, KH2AsO4, NaAsO2, (NH4)2Cr2O7, (NH4)2U2O7) of analytical grade and distilled water, in the concentrations of 0.1 and 0.5 mmol·L−1 and the natural pH (i.e. pH ≈ 3.5 for cationic solutions and pH 5–6 for anionic solutions). The concentration range was selected as appropriate for the simulation of a slightly increased amount of the contaminant in a water system to a heavily contaminated solution. The solution of AsIII as NaAsO2 was prepared under an N2 atmosphere, at pH ≈ 7.4–8.5 (Doušová et al., 2009, 2011).
2.3. Sorption experiments During the adsorption experiments the cationic particles (Cd2 +, Pb and Cs+, respectively) were adsorbed on the raw WBD, whereas the anions (AsIII/V, CrVI, UVI) were adsorbed on the water-washed WBDw. A suspension of the model solution (50 mL) and defined dosage (0.5–15 g L−1) of WBD (WBDw) was shaken using a batch procedure at laboratory temperature (20 °C) for 24 h (Doušová et al., 2006). The product was filtered and the filtrate was analyzed for residual cations/anions, while the solid residue (saturated WBD/WBDw) was kept for leaching tests. WBD saturated with Pb2+ and WBDw saturated with CrVI, AsIII and AsV, respectively, were also tested by the solid-state NMR spectroscopy and SBET measurements to verify the structural and/ or surface changes caused by adsorption. All sorption data were fitted to the Langmuir equation (Jeong et al., 2007, Misak, 1993), which was 2+
verified as a suitable adsorption model for natural oxides, aluminosilicates and soils. The data illustrated the parameters of adsorption (qmax – maximum equilibrium sorption capacity; Qtheor – theoretical sorption capacity; R2 – correlation factor; KL – Langmuir adsorption constant), which were used to compare the adsorption affinities of the investigated ions on brick dust. 2.4. Analytical methods Powder X-ray diffraction (XRD) was performed with a Seifert XRD 3000P diffractometer with CoKα radiation (λ = 0.179026 nm, graphite monochromator, goniometer with Bragg-Brentano geometry) in the 2θ range of 5–60° with a step size of 0.05° 2θ. X-ray fluorescence (XRF) analyses of the solid phase were determined with an ARL 9400 XP+ spectrometer with a voltage of 20–60 kV, probe current of 40–80 mA and effective area of 490.6 mm2. UniQuant software was used for data evaluation. The specific surface area (SBET) was measured on a Micromeritics ASAP 2020 (accelerated surface area and porosimetry) analyzer using the gas sorption technique. The ASAP 2020 model assesses single and multipoint BET surface area, Langmuir surface area, Temkin and Freundlich isotherm analysis, pore volume and pore area distributions in the micro- and macro-pore ranges by the BJH method. The micropore option used the Horvath-Kavazoe method, with N2 as the analysis adsorptive and an analysis bath temperature of −195.8 °C. The samples were degassed at 313 K for 1000 min. The IR spectra were collected on a Nicolet 6700 FTIR (Thermo Nicolet Instruments Co.) with N2 purging system. Spectra were acquired using a single reflection ATR SmartOrbit accessory equipped with a singlebounce diamond crystal (angle of incidence: 45°). A total of 64 scans were averaged for each sample and the resolution was 2 cm− 1. The spectra were rationed against a single-beam spectrum of the clean ATR crystal and converted into absorbance units by ATR correction. Data were collected in the wavenumber range of 4000–400 cm−1. Solid-state NMR spectra were measured at 11.7 T using a Bruker AVANCE III HD 500 WB/US NMR spectrometer. The 27Al MAS NMR spectra were acquired at a spinning frequency of 11 kHz, Larmor frequency of 130.287 MHz and recycle delay of 2 s, and the spectra were referenced to the external standard Al(NO3)3 (0 ppm). The number of scans for the acquisition of a single 27Al MAS NMR spectrum was 512 (total experimental time was ca. 10 min). The 29Si MAS NMR spectra were acquired at a spinning frequency of 11 kHz, Larmor frequency of 99.325 MHz and recycle delay of 10 s. The number of scans for the acquisition of a single 29Si MAS NMR spectrum was 6144 (total experimental time was ca. 17 h). The spectra were referenced to the external standard M8Q8 (− 109.8 ppm). The NMR experiments were performed at a temperature of 24 °C, and temperature calibration was performed to correct for the frictional heating of the samples. The concentration of Pb/Cd/Cs in aqueous solutions was determined by atomic absorption spectrometry (AAS) using a SpectrAA-880 VGA 77 unit (Varian) in flame mode. The concentration of As in aqueous solutions was determined by Hydride Generation Atomic Fluorescence Spectrometry (HG-AFS) using a PSA 10.055 Millennium Excalibur apparatus. The samples were
Table 1 Chemical and mineralogical composition of the green ceramic mixture and waste brick dust (WBD). Chemical composition (% wt)
Green
WBD
Na2O/K2O MgO/CaO Al2O3 SiO2 P2O5 Fe2O3 TiO2
1.1/4.2 3.8/10.6 20.9 51.8 0.9 6.6 0.9
1.3/4.1 4.5/11.5 20.0 51.3 0.9 5.9 0.8
Mineralogical composition
Green
WBD
Quartz (25%) Chlorite Illite (30%) Albite Orthoclase Muscovite Calcite Dolomite
Quartz (22%) Microcline Hematite Albite Orthoclase Muscovite Akermanite Hedenbergite
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pre-treated with a solution of HCl (As-free, 36% v/w) and KI (50%) with ascorbic acid (10%). The instrumental parameters included ppm and ppb modes, HCl (12%) with KI + ascorbic acid solution as the reagent blank and 7% NaBH4 in 0.1 mol·L− 1 NaOH as the reductant. The declared detection limit was 0.05 ppm and the standard deviation was experimentally determined as 2.5%. The concentration of Cr as Cr2 O 27 − in aqueous solutions was measured with a UV/VIS spectrophotometer (UNICAM 5625) at 350 nm following acidification with H2SO4 (10% wt.) (Malat, 1973). Uranium concentrations were measured using an atomic (optical) emission spectrometer with inductively coupled plasma (ICP-OES; Optima 2000-DV, Perkin-Elmer Instruments). The concentration of U is given as mg L−1 of U. The limit of detection for the determination of the concentrations of arsenic was 3 μg L−1. 2.5. Leaching test The stability of toxic ions in the saturated sorbents was verified by a standard procedure EN 12457–2 (CEN, 2002); a suspension of dry sorbent and distilled H2O at a ratio of 1:10 was agitated in a sealed polyethylene reactor at laboratory temperature for 24 h. Then, the suspension was filtered off and the filtrate was analyzed for residual cations/anions. 2.6. Analytical quality control, reproducibility of obtained data An accuracy of AAS analyses was guaranteed by the Laboratory of Atomic Absorption Spectrometry, ICT Prague (CR); the detection limit was 0.5 μg L−1 with a standard deviation of 5–10%). For HG-AFS, the declared detection limit was 0.05 mg L−1 and the standard deviation was experimentally determined as 2.5%. The reproducibility of leaching tests and UV/VIS spectrophotometry was checked in parallel experiments; the standard deviation resulting from 10 experimental sets ranged from 4.5 to 7% for the leaching tests, while it did not exceed 5% for the spectrophotometric analyses. 3. Results and discussion For the adsorption experiments, highly toxic and/or ecologically harmful cations (Cd, Pb, Cs) and anions (As, Sb, Cr, U) were selected according to their current environmental concern on the international scale. The study proceeded in two directions: (i) the adsorption of cationic/anionic particles from aqueous solution on WBD/WBDw for the determination of the optimal sorption parameters, and (ii) the verification of the stability of fixed particles in the saturated sorbent.
decrease in alkali material, particularly CaO, and soluble sulfates such as SO3, which corresponded well with about a ten-fold increase in the SBET (Table 2) and porosity (Fig. 1) of WBDw compared to raw WBD. These results suggest that weakly bound soluble inorganic salts (CaSO4, Na2SO4), causing the negligible porosity of raw WBD, were removed by water-washing. The XRD of WBDw illustrated a slight increase in crystallinity compared to WBD, which corresponded well with the previous data. The structural, chemical and mineralogical changes to the raw (WBD) and water-washed (WBDw) sorbents were studied by XRD, XRF, solid-state NMR and SBET measurements. In the XRD of WBDw quartz intensities showed a slight increase indicating a minor arise of sample crystallinity at the expense of any amorphous phase released by the water-washing of raw WBD. The significant decrease of alkali phases in WBDw was verified by the ATR-FTIR spectroscopy of the solid residue of WBD water leachate, which identified mostly CaCO3 and CaSO4, respectively (Fig. 2). The recorded 27Al and 29Si MAS NMR spectra of raw (WBD) and modified (WBDw) brick dust are shown in Fig. 3. The signal assignment was performed based on literature data (Mackenzie and Smith, 2002) and our previous results (Brus et al., 2012). A broad signal resonating at ca. 59 ppm with a half width of ca. 1600 Hz was detected in the 27Al MAS NMR spectra of raw WBD and modified WBDw sorbents. The resonance frequency reflects the tetrahedral coordination of these aluminum sites, while resonance broadening indicates the amorphous character of these systems. By washing the system by water two lowintensity 27Al MAS NMR signals at ca. 14 and 10 ppm in the modified WBDw show the presence of two different types of hexa-coordinated aluminum sites. With high probability a small amount of sodium ions is removed from the surface area. Consequently, the system is not completely charge-compensated and a limited transformation of AlO4 framework units to free (dissolved) extra-framework AlO6 monomer moieties occurs. As indicated by the intensity of these signals that is b5% we, however, suppose that this process does not affect considerably the overall structure of the synthesized sorbent. The 29Si MAS NMR spectra consisted of a single signal centered around − 91 ppm with a two small signals at − 73 and − 108 ppm. The main signal is usually attributed to fully condensed aluminumsubstituted silica tetrahedra Q4(4Al) and Q4(3Al) accompanied by minor fractions of Q0 and SiO2 units. After water-washing WBD, the transformation of Q0 units occurred, since the amount of Q0 units decreased from 10% to 5%. Simultaneously, the amount of Q4 (4Al) units increased and SiO2 units remained constant at about 23%.
3.1. Characterisation of raw WBD and water-washed WBDw Table 2 illustrates the chemical and surface changes induced by water washing of WBD. The water-washed WBDw showed a significant Table 2 Chemical composition, pH of the zero point of charge and SBET of the raw (WBD) and water-washed (WBDw) sorbent. Composition (% wt)
WBD
WBDw
Na2O/K2O MgO/CaO Al2O3 SiO2 P2O5 SO3 Fe2O3 TiO2 MnO2 SBET (m2 g−1) pHZPC
1.3/4.1 4.5/11.5 21.0 51.3 0.9 1.1 5.9 0.8 0.04 3.31 4.4
0.7/3.8 4.4/ 20.1 50.6 0.7 0.4 6.2 0.9 0.05 32.51 4.8
3
Fig. 1. BHJ pore-volume distribution of WBD/WBDw.
Please cite this article as: Dousova, B., et al., Use of waste ceramics in adsorption technologies, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/ j.clay.2016.02.016
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Fig. 2. WBD-ATR-FTIR spectrum of the solid residue of water leachate.
3.2. Adsorption parameters for cationic and anionic contaminants The adsorption parameters calculated according to the Langmuir model for cations (Table 3) showed maximum values of adsorption capacity for Cd2 + and Pb2 + adsorbed on WBD, which corresponded well to the previous results of Doušová et al. (2006, 2009) and indicated the formation of inner-sphere surface complexes (Sherman and Randall, 2003). The less balanced, non-Langmuir adsorption run towards the larger ionic radius cations (Cd2+ b Pb2+ ≪ Cs+) could be related to the very low SBET and negligible porosity of WBD, resulting in a steric barrier to the surface binding of larger particles. This phenomenon was previously described by Doušová et al. (2009), who compared the intensity of Fe2+/Fe3+ surface modification of kaolins. A negligible effect of Cs+ adsorption resulted primarily from steric factors. According to Cornell (1993), the differences in Cs uptake on different minerals may be related to different cation exchange capacities, structure, composition of the mineral and the surface area/particle size ratio. As mentioned before, illite, which is generally considered the most selective Cs adsorbent among the clay minerals (Cornell, 1993; Staunton and Roubaud, 1997), is transformed during the firing process to an amorphous phase (Vejmelková et al., 2012). The WBD sorbent consisted mostly of quartz,
feldspars and hematite, which are poor Cs adsorbents (Bergaoui et al., 2005). The comparable adsorption parameters for anionic contaminants (Table 4), which were adsorbed on WBDw, indicated a generally lower sorption selectivity of WBDw to anions, which may be mainly related to the pH/pHZPC ratio. The pHZPC b 5 and pH/pHZPC ≫ 1 for both the WBD and WBDw denoted prevailing cation active surface sites and charge density, which inhibited the fixation of anions due to strong repulsion (Fiol and Villaescusa, 2009). Another reason may be the different ability of oxyanions to undergo surface binding. Arsenates (AsV) were specifically adsorbed, forming stable inner-sphere, predominantly bidentate complexes with iron oxides (Fendorf et al., 1997; Grossl et al., 1997). Arsenites (AsIII) and chromates (CrVI) tended to form both inner- and outer-sphere complexes; in the case of innersphere complexes, a weaker monodentate form prevailed (Grossl et al., 1997). The effect of pH on adsorption indicated a different trend. The adsorption of AsV/CrVI on iron oxides decreased with increasing pH, while AsIII in the anionic form (H2AsO− 3 ), which is fundamental for selective adsorption, appeared at pH ≥ 9 (Takeno, 2005). All these findings correspond well with the adsorption data. The adsorption stability and selectivity for anionic particles declined in the order: AsV N AsIII ≫ CrVI. The adsorption on WBD/WBDw was most probably associated with variable arrangements of adsorption mechanism related to a high pH value. According to Stumm (1992), at a high pH the formation of surface precipitates can appear and participate markedly in surface binding. 3− During adsorption on WBD/WBDw anionic particles (AsO3− and 4 , AsO3 2− CrO4 , resp.) were captured at anion active surface sites, but also partially linked by chemical bonds to released alkali cations (Na+, K+, Ca2+),
Table 3 Langmuir parameters for adsorption of cations on WBD.
Fig. 3. 27Al and 29Si MAS NMR spectra of WBD and water-washed WBD.
Cation
Initial concentration (mmol L−1)
qmax (mmol g−1)
Qtheor (mmol g−1)
KL (L mmol−1)
R2
Cd2+ Cd2+ Pb2+ Pb2+ Cs+ Cs+
0.1 0.5 0.1 0.5 0.1 0.5
0.06 0.26 0.11 0.51 0.01 0.04
0.05 0.13 0.12 –⁎ –⁎ –⁎
3063.5 3951.8 1017.3 − − −
0.875 0.895 0.695 − − −
⁎ Non-Langmuir run.
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Table 4 Langmuir parameters for adsorption of anions on WBDw. Anion
Initial concentration (mmol L−1)
qmax (mmol g−1)
Qtheor (mmol g−1)
KL (L mmol−1)
R2
AsIII AsIII AsV AsV CrVI CrVI UVI UVI
0.1 0.5 0.1 0.5 0.1 0.5 0.1 0.5
0.005 0.015 0.01 0.04 0.006 0.01 0.08 0.13
–⁎ 0.04 0.02 0.04 –⁎ –⁎
– 1.3 41.8 281.1 – – 368.0 3232.5
– 0.784 0.695 0.963 – – 0.984 0.785
0.08 0.11
⁎ Non-Langmuir run.
forming less soluble or insoluble surface precipitates. This effect prevailed in the raw WBD with a higher salt content and has been clearly shown in the infrared spectra of AsV adsorption on WBD/WBDw (Fig. 4). The changes in tetraedric symmetry of AsO3− complex, which can be caused 4 by the solution pH, type of surface complexation (inner-sphere vs. outersphere), type of soluble cations and H2O structure resulted in the shift and splitting of spectral bands (Goldberg and Cliff, 2001; Myneni et al., 1998; Roddick-Lanzilotta et al., 2002). Subtracted A and B spectra in Fig. 4 illustrate AsV adsorption on WBD (A) and WBDw (B). The bands at about 858 cm−1 dominate in HAsO3− spectrum acquired at pH ≈ 9.2 4 (Roddick-Lanzilotta et al., 2002), therefore, a surface complexation via HAsO3− can be considered. According to Myneni et al. (1998), typical 4 bands at 798–812 and 855–905 cm−1 can be assigned to stretching vibration of As\\OCa, indicating the formation of different crystalline calcium arsenates as surface precipitation. The unique sorption parameters of UVI (Table 4) resulted from the 2+ variability of uranium phases and stable UO+ forms under 2 and UO2 oxidic conditions and in the pH range of ≈ 3.5–10 (Takeno, 2005). Uranium as UVI was most likely bound as a cationic complex particle [(UO 2) n (OH) 2n − 1 ] +. The optimal adsorption capacities obtained under the same experimental conditions for all the investigated elements are compared in Fig. 5. 3.3. Adsorption efficiency and sorbent consumption To assess the possible practical uses of WBD, not only its adsorption capacities to selected elements but also the actual adsorption efficiencies
Fig. 5. Maximum and theoretical sorption capacities for investigated contaminants.
suggestive of its approximate consumption in technological processes had to be studied. Both Cd and Pb adsorption showed a high sorption efficiency at a low sorbent dosage (see Fig. 6a). Almost quantitative adsorption from less concentrated solutions (≈0.1 mmoL−1) was achieved at a dosage of 1 g L−1 for Pb2+ and 2 g L−1 for Cd2+. The dosage of about 6 g L−1 was sufficient for the quantitative removal of Pb2+/Cd2+ from highly concentrated solutions (≈0.5 mmol L−1). Of the tested anions, UVI and AsV are potential candidates for technological applications (Fig. 6b). Effective adsorption was observed at the sorbent dosage of 3 g L−1 for UVI and 15 g L−1 for AsV. Regardless the value of the maximum adsorption capacities (Table 4), CrVI was adsorbed more effectively at a lower initial concentration, while AsV and AsIII showed the opposite trend. In the case of UVI, the adsorption efficiency was independent of the initial concentration. These results confirmed the adsorption theory, in that non-specific adsorption (electrostatic bonding mechanisms, weak binding energy, outer-sphere complexation) strongly depends on the ionic strength (Stumm, 1992). Inner-sphere surface complexes are specifically bound to active sites, involving covalent bonding and/or combinations of covalent and ionic bonding. Therefore, Pb2+, Cd2+ and UVI adsorption on WBD/WBDw occurred almost specifically with a negligible effect of the surrounding ionic particles and initial concentration. In this
Fig. 4. Subtracted ATR-FTIR spectra of AsV adsorption on WBD/WBDw. A – substracted spectrum of WBD after AsV adsorption and raw WBD, B – subtracted spectrum of WBDw after AsV adsorption and WBDw.
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Fig. 6. Adsorption efficiency as a function of WBD dosage. a) cationic contaminants, b) anionic contaminants, c0 ≈ initial concentration of contaminant.
case, the binding of cations directly to tertahedral (SiO4) or octahedral (AlO6) coordinations of clay structure can be considered. Conversely, the non-specific adsorption of AsIII and CrVI has been markedly affected by the ionic strength of adsorbate and the charge distribution of AsO3− 3 and CrO2− oxyanions. 4 The approximate consumption of sorbent per gram of toxic element (Fig. 7) indicated an important technological specification and allowed us to estimate the applicability of the method under certain conditions. The approximate mass consumption of sorbent per gram of toxic element was found to be about 60 g for Pb and U, about 100 g for Cd and more than 400 g for As as AsV. The other contaminants (As as AsIII, Cr, Cd) likely have no perspective to be removed with WBD due to extremely high sorbent consumption, mainly due to insufficient adsorption efficiency (b30%).
Fig. 7. Hypothetic consumption of WBD per 1 g of contaminated element.
Fig. 8. 27Al MAS NMR spectra of WBD/WBDw after adsorption of Pb2+, CrVI, AsIII and AsV, respectively.
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3.4. Structural and surface changes of WBD/WBDw during adsorption
Acknowledgements
Since adsorption represents in principle a surface process, only small structural changes can be expected. From the comparison of 27Al MAS NMR spectra before and after adsorption of toxic elements (Fig. 8), it was evident that the WBD/WBDw structure was uniform and the adsorption had no influence on the structure of aluminosilicate networks. Similarly, the SBET and porosity of WBD/WBDw saturated with AsO3− 4 did not change markedly (Fig. 1).
This work was the part of the project 13-24155S (Grant Agency of Czech Republic).
3.5. Leaching test The stability of cations/anions in saturated WBD/WBDw was evaluated by the leaching test described in Section 2.6. As evident from Table 5, the stability of the saturated sorbent was closely related to the selectivity of particular adsorption; Pb2+ and Cd2+ adsorption was characterized by high sorption capacities and efficiencies, indicating the formation of inner-sphere surface complexes resulting in very stable binding in both leaching environments (no more than 0.08% of Pb2+/Cd2+ was leached). Favorable stability (about 1% of the initial amount leached) was observed for AsV and UVI (Table 5), which corresponded well with the adsorption data. The non-selective Cs+/CrVI adsorption showed the worst binding stability during the leaching test. Under the applied experimental conditions, WBD/WBDw saturated with Pb2+, Cd2+ and AsV, respectively, could be considered a non-hazardous waste (CSN EN 12457). However, the expected fixation of saturated WBD/WBDw in building materials, such as cement and concrete, might markedly change the stability of toxic particles. 4. Conclusions Waste brick dust is an effective sorbent of selected cationic and anionic toxic particles. The main advantages the low price of this material and the possibility of incorporating the saturated sorbent into a cementitious building material, which may eliminate the hazardous impact of toxic elements due to its stabilization/solidification efficiency. According to the phase composition and pHZPC value, WBD is more effective as a cation active sorbent, but not uniquely. The water-washing of raw WBD resulted in a reduction in alkalinity by ˃2.5 pH units, and in more than a ten-fold increase in SBET. Nevertheless, the adsorption properties of WBDw to anionic particles did not improve markedly, which suggested the participation of surface precipitates in the binding mechanism on raw WBD. Heavy metals such as Pb2+/Cd2+ and uranium as [(UO2)n(OH)2n − 1]+ were adsorbed selectively onto WBD/ WBDw at high adsorption efficiency and low sorbent consumption. Arsenic as AsO34 − was also adsorbed, but with more than four times greater sorbent consumption. The adsorption of cesium as Cs+ and chromium as CrO24 − on WBD/WBDw was totally ineffective. The leaching stability of the adsorbed particles decreased in the order: Pb2+ ≈ Cd2+ ˃ UVI ˃ AsV ˃ AsIII ≫ Cs+ ≫ CrVI. For prospective use, the study of long-term stability of saturated WBD/WBDw incorporated into building material will be essential.
Table 5 Leachability of toxic particles from WBD(1) and WBD(2) w . Toxic particle
Initial amount (mg/g)
leached in KCl (%)
leached in H2O (%)
Class of leachabilitya
AsV (2) AsIII (2) CrVI (2) UVI (1) Pb2+ (1) Cd2+ (1) Cs+ (1)
2.80 0.80 0.01 14.70 5.90 7.00 1.40
1.2 2.3 35.5 0.6 0.02 0.01 18.1
0.8 2.4 41.1 1.2 0.07 0.08 4.3
II – other waste III - hazardous waste III - hazardous waste not limited I – inert waste II – other waste Not limited
(1), (2) a
the type of used sorbent. Regulation No. 294/2005 (CSN EN 12457).
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