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Superior removal of Hg (II) ions from wastewater using hierarchically porous, functionalized carbon
Journal of Hazardous Materials 371 (2019) 33–41
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Superior removal of Hg (II) ions from wastewater using hierarchically porous, functionalized carbon
T
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Yao Lia,1, Mengdan Xiaa,1, Fufei Ana, Nianfang Mab, Xueliang Jiangc, Shenmin Zhua, , ⁎ Dawei Wangd, Jun Mae, a
State Key Laboratory of Metal Matrix Composites, Shanghai Jiao Tong University, Shanghai 200240, PR China Guangdong provincial bioengineering institute(Guangzhou Sugarcane industry research institute), Guangdong Key Laboratory of Sugarcane Improvement and Biorefinery, Guangzhou, 510316, China c School of Materials Science and Engineering, Wuhan Institute of Technology, Wuhan 430205, China d School of Chemical Engineering, UNSW Australia, UNSW Sydney, NSW 2052, Australia e School of Advanced Manufacturing & Mechanical Engineering, University of South Australia, Australia b
A R T I C LE I N FO
A B S T R A C T
Keywords: Heavy metal ions Porous carbon Wastewater Hg (II)
The removal of heavy metal ions from industrial wastewater by adsorption has been central to the environment for decades, where common adsorbent materials are often limited by poor efficiency, complex fabrication and long processing time. Porous carbon derived from biospecies holds promise to address the limitations. In this study we converted bagasse into a carbon composite having hierarchically porous structure; the composite’s dispersion phases – iron oxide and manganese oxide – were synthesized by a simple one-step liquid-phase reaction method. Featuring large specific surface area of 350.8 m2 g−1, the composite demonstrated exceptional Hg (II) removal efficiency of 96.8%, adsorption rate of up to 96.8% within 150 min and adsorption capacity of 9.8 mg g−1. In comparison with other removal materials, our work is outstanding in terms of both removal efficiency and synthesis simplicity. The high efficiency is attributed to the synergy between physical adsorption referring to hierarchically porous structure and chemical adsorption relating to functional complexation processes. It provides a new avenue for the development of high-performance adsorbent materials for heavy metal removal from aqueous media.
1. Introduction The toxicity, carcinogenicity and long-term accumulation effect of heavy metal ions cause enormous risks to both the environment and human health. In particular, mercury (Hg) produced from industrial water is a major toxic element found in the environment [1]. The presence of Hg2+ can lead to a variety of human health problems such as brain and heart diseases, liver and kidney damage, and even cancer [2]. Research into eco-friendly, cost effective and efficient means for Hg2+ removal has gained more interests. A number of approaches have been developed, mainly including chemical precipitation [3], ion exchange [4], solvent extraction [5], evaporation [6], ultrafiltration [7] and adsorption [8]; of these, adsorption is considered as the most promising due to its simplicity, high efficiency, low cost and operation convenience [9]. Thus, a variety of adsorbent materials have been
investigated for Hg2+ removal, including metal oxides [10] and activated carbon [11]. Porous carbon fabricated from biospecies holds promise, because of its interconnected, three-dimensional, porous structures with high specific surface area. As a major type of adsorbents, porous carbon removes Hg (II) from aqueous media mainly through physical adsorption which involves van der Waals or electrostatic interactions taking advantage of large surface area. G. Skodras et al. studied Hg removal through activated carbon which was prepared from agricultural residues and waste tires [12]. M. Zabihi et al. fabricated porous carbons with surface area of 780 m2 g−1 from walnut shells, which exhibited a high monolayer adsorption capacity of 151.5 mg g-1 for Hg (II) removal [13]. Zhang et al. studied various types of activated carbon extracted from organic sludge via chemical activation. ZnCl2-activated carbon exhibited quite high adsorption capacity for Hg (II), and 60–80% of the adsorbed Hg (II) could be recovered via sonication [14].
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Corresponding authors. E-mail addresses: [email protected] (S. Zhu), [email protected] (J. Ma). 1 These authors contributed equally. https://doi.org/10.1016/j.jhazmat.2019.02.099 Received 22 December 2018; Received in revised form 25 February 2019; Accepted 26 February 2019 Available online 27 February 2019 0304-3894/ © 2019 Elsevier B.V. All rights reserved.
Journal of Hazardous Materials 371 (2019) 33–41
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Hierarchical porous materials may be promising candidates for adsorption, because its full accessibility of narrow cavities and efficient mass transport property. C. Fischer et al. developed bioinspired carbidederived carbons with hierarchical pore structure for Hg (II) adsorption [15]. Compared with traditional activated carbons, their micropore size can be adjusted with high accuracy. Bioinspired carbon exhibits high Hg (II) adsorption capacities. From the microcosmic perspective, the 3D hierarchically pore system involves mainly micropores and mesopores [16]. The micropores directly lead to enhanced adsorption capacity; the well-developed mesopores promoted the optimum adsorption kinetics [17]. The mesopore channels also serve as liquid flow pathways and allow fast transport of Hg (II) to active adsorption sites. Hence, it is obvious that hierarchically structured materials should play a critical role in the removal of heavy metal ions [18–21]. Although porous carbon is advantageous of fast adsorption, it displays unideal adsorption efficiency and difficult recovery of adsorbents as its performance is contributed by physical adsorption only [22]. Chemical adsorption refers to a phenomenon where an adsorbent reacts with contaminant(s) through interactions. Lu et al. removed traces of Hg (II) through chemical adsorption where Hg (II) ions were complexed with the hydroxyl groups on the surface of in situ formed manganeseferric hydroxide. It was reported that the removal mechanisms comprised of surface complexation and flocculation-precipitation processes, leading to transformation from liquid to solid phase of Hg (II). Fe-Mn oxide at 40 mg L−1 could effectively remove 80% Hg (II) starting from initial concentration of 30 μg L−1, which represents high adsorption efficiency [23]. Noteworthy is that the chemical adsorption can be further enhanced by modifying the magnetic adsorbent surface with carefully selected functional organic groups. Compared to physical adsorption, chemical adsorption is faster and exhibits higher removal efficiency for Hg (II). To improve adsorption efficiency, physical adsorption has been combined with chemical adsorption. Faulconer et al. impregnated iron oxide with the activated carbon to obtain a new magnetic adsorbent which achieved Hg2+ removal of 96.3 ± 9% and sorbent recovery 92.5 ± 8.3% by magnetic separation. In the adsorption process, iron oxide proved to form strong binding affinity with Hg species and offered plentiful adsorption sites for Hg2+ through coordination with oxygen atoms. This adsorbent could remove Hg (II) to attain a final concentration of 0.2 μg L−1 [24]. Further investigation found that manganese oxides demonstrated similar or even superior chemical adsorption performance. Moghaddam et al. prepared MnO2-coated carbon nanotubes to remove Hg (II) from aqueous solution and obtained adsorption capacity of 58.8 mg g-1, where MnO2 went through oxidationreduction reactions [25]. However, the adsorption value is not satisfactory owing to the low specific surface area of the composite. Inspired by nature, we design and fabricate a carbonaceous material from an industrial waste – bagasse, where graphene-like structure is in situ formed; the carbonaceous material is then combined with metal oxides Fe2O3 and MnxOy by using a one-step liquid-phase reaction method, to prepare porous carbon coupled with Fe-Mn oxide (denoted porous carbon/Fe-Mn). Since the graphene-like structure is introduced to enhance conductivity, the porous matrix behaves like a conductor which not only gathers charge on the pore surface but works as a conductive network linking dielectric metal oxide nanoparticles; this would result in highly effective electrostatic interactions for Hg adsorption. Fe-Mn binary oxides with surface charge and variable valence would have higher surface activity than single oxide, suggesting great potential for water treatment [26]. As a result, the composite – which features a biological structure that cannot be manually synthesized – exhibits outstanding mercury adsorption capacity. Later sections discuss in detail the relationship between composite structure and adsorption properties.
Fig. 1. Schematic fabrication of the porous carbon/Fe-Mn oxide composite and its adsorption mechanism for Hg (II).
2. Experimental 2.1. Chemicals All chemicals and reagents used in this study were of analytical grade or higher, including mercury nitrate monohydrate (Hg (NO3)2·H2O), iron sulfate heptahydrate (FeSO4·7H2O), potassium permanganate (KMnO4), nitric acid (HNO3), sodium hydroxide (NaOH), and potassium hydroxide (KOH). All solutions were prepared with N2purged deionized (DI) water. 2.2. Materials preparation and characterization Fig. 1 describes the preparation of porous carbon with binary oxide nanoparticles in situ formed. In a typical synthesis, a precursor bagasse was employed to fabricate porous carbon via carbonization at 650℃, followed by activation at 700℃ using KOH as an activator [27]. Graphitization was performed in vacuum at 800℃ for 3 h. The powder was washed with 11% dilute hydrochloric acid and dried under vacuum, to obtain porous graphitic carbon. Carboxylation was conducted by mixing 2.0 g of porous carbon with 300 mL of concentrated nitric acid and stirring for 3 h under water-bath heating at 60℃. The sample was first separated from the solution via vacuum filtration. Afterwards, the sample was rinsed to a neutral pH with DI water. After drying the sample at 60℃ for over 24 h, the carboxyl porous carbon was attained. The porous carbon/Fe-Mn composite was prepared following a revised method reported by Tang et al [28]. First, 1.202 g of carboxyl porous carbon was suspended in 30 mL DI water. FeSO4 solution (2.88 × 10−1 M) and KMnO4 solution (9.58 × 10−2 M) were prepared with DI water. Then, 6 mL of FeSO4 and 18 mL of porous carbon suspension were dispersed in 270 mL of DI water under vigorous magnetic stirring. Next, 6 mL of KMnO4 solution was added dropwise into the mixture which pH was adjusted to about 7.5 using HNO3 (1 M and 0.1 M) and NaOH (1 M and 0.1 M). The mixture was continuously stirred 1 h and aged at 25 ± 2℃ for 12 h. The resultant precipitate was collected through vacuum filtration, washed with DI water, and dried at 90℃ for over 48 h, to create porous carbon/Fe-Mn composite (molar ratio of Fe/Mn = 3/1). As a benchmark, a porous carbon/Fe oxide composite (denoted porous carbon/Fe) was synthesized using the same approach but without adding Mn. The surface morphology and chemical composition were observed by scanning electron microscopy (SEM) performed on a JEOL JSM6360LV field emission microscope at 15 kV. Transmission electron microscopy (TEM) was carried out on a JEOL 2010 microscope at 200 kV. The Brunauer-Emmett-Teller (BET) adsorption method was performed 34
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wrinkled carbon matrix, resulting in a rough surface. As for porous carbon/Fe-Mn, a greater number of nanoparticles are present as aggregated flocs formed in situ on the porous carbon surface. The fine morphology of porous carbon and its composites is further confirmed by TEM. In Fig. 3, The interplanar distance of porous graphitic carbon is about 0.33 nm. The 3D porous structure is observed in all three materials. For both composites, abundant nanoparticles are distributed on the surface of porous carbon in the form of aggregated flocs. The 3D hierarchical structure constructed by porous carbon matrix and metal oxide nanoparticles would provide well-developed pore channels, serving as liquid flow pathways for the transport of Hg (II) to active adsorption sites. Fig. 4(a)-(b) contains the nitrogen adsorption/desorption isotherms and pore size distribution curves of synthesized composites. The BET surface area of porous carbon and its composites-carbon/Fe and carbon /Fe-Mn- are 1125.2, 945.3 and 350.8 m2 g−1, respectively, corresponding to pore size 3.5, 3.9 and 6.9 nm. The porous carbon produced from bagasse has a far larger specific surface area than previous efforts, such as 803 m2 g−1 of activated carbon powder from walnut shells [13]. This means great potential for adsorption. Embedding the metal oxides inside porous carbon would slightly reduce the surface area, because iron and manganese oxides may stay on the carbon surface occupying certain micropores. More importantly, the embedding should result in larger pore size, which is desired for Hg (II) adsorption. All three adsorption/desorption isotherms show hysteresis loops in the P/P0 range of 0.45–1.0, with pore widths centered at ˜3.9 nm, indicating mesoporous nature. The porous carbon/Fe-Mn composite still possesses relatively high specific surface area of 350.8 m2 g−1, which would benefit the adsorption of Hg2+. This represents an advance over previous effort where the surface area values of manganese dioxide and iron oxide were reported as 266.9 m2 g−1 and 172.3 m2 g-1 [29]. In Fig. 4 (c), the XRD pattern of porous carbon shows diffraction peaks at 26.6° and 43.5°. The former peak can be indexed to the (003) plane of the graphene-like structure, suggesting some degree of graphitic order. The layered structure of TEM images in Fig. 3(a) corresponds to the (003) plane of the graphene-like structure in XRD analysis. For porous carbon/Fe, the diffraction at 26.6° became unobvious because the iron oxide coated on the surface may interact with porous carbon. The peaks located at 26.8°, 36.2°, 47.3°, 54.9° and 60.0° were all assigned to crystalline phase α-Fe2O3, indicating that amorphous FeOOH was readily transformed into crystalline iron (III) oxides [30]. The manganese oxide component is mainly MnO2, corresponding to the peaks at 34.0°,42.9°, and 62.7°. The diffraction peaks of carbon are weakened implying that crystalline phase of iron oxides transformed into amorphous phase. This indicates that the formation of crystals would be restrained by the co-existence of iron and manganese oxides during the synthesis. For crystalline Fe2O3, the particle size is calculated to be 28.2 nm according to the Scherrer formula. Fe and Mn oxides offer plenty of active complexation sites, which is essential for the adsorption of Hg2+. Electrostatic attraction also contributes to adsorption process. Fig. 4(d) reveals Zeta potentials of porous carbon, porous carbon/Fe and porous carbon/Fe-Mn composites as a function of pH. As it shows, there is an isoelectric point near 6. As the adsorbent surface charge influences the interfacial reaction with its adsorbate, we further explored the interaction between Hg (II) and the adsorbent by utilizing visual MINTEQ (version 3.1) to analyze mercury speciation under different pH. When it is low, Hg (II) exists primarily as Hg2+ cation. At pH 3–5, the adsorbents are positively charged, making electrostatic repulsion unfavorable for Hg (II) adsorption. When pH is over 5, Hg(OH)2 becomes a dominant component. The adsorbents are negatively charged at pH over 6. When pH is over 6, excessive OH− ions compete with the negatively charged ions of the adsorbent to bind with Hg2+, unfavorable for Hg (II) adsorption. Consequently, the isoelectric point at pH 6 appears to be an optimal pH for adsorption.
using an ASAP2020 volumetric adsorption analyzer to measure the surface area, pore size, and pore volume of the adsorbents. The crystalline and amorphous phase compositions were identified using X-ray diffraction (XRD) on a Rigaku D/max 2550 V L/PC system operated at 35 kV and 200 mA with Cu Kα radiation (λ = 1.5406 Å), at a scan rate of 5°min−1 and a step size of 0.02°. Zeta (ζ) potential versus pH was measured using MALVERN Zetasizer at 25℃. Thermal gravimetric analysis (TGA) was conducted on a PE TGA-7 instrument with 10 ℃ min−1 to analyze element content variation. Fourier transform infrared (FTIR) spectroscopy was conducted with a PE Paragon 1000 spectrophotometer on KBr pellets to determine the chemical functional groups. Raman spectroscopy was characterized by a Renishaw in Via Raman Microscope. The adsorbent surface elemental compositions before and after Hg adsorption were analyzed by X-ray photoelectron spectroscopy (XPS) collected on a physical electronics PHI5400 using Mg K radiation as the X ray source. All the spectra were corrected with the C 1 s (285.0 eV) band. The adsorbents after mercury adsorption were separated by centrifugation and freeze-dried under vacuum at −60℃. The residual mercury concentration in the supernatant was determined using a Duo View ICP-OES spectrometer (THERMO iCAP 6000 Radial). 2.3. Mercury batch adsorption tests Hg (II) stock solution (5 mg/L) was prepared by dissolving Hg (NO3)2·H2O into DI water. Initial solutions of varying concentrations were obtained by diluting the stock solution with DI water. The mercury concentration is given as mercury ion concentration. Batch adsorption experiments were performed in sealed 25-mL glass vials. Briefly, the reaction was initiated by mixing a certain dosage of adsorbent into Hg2+solutions. The vials were then sealed with stirring on a thermostated magnetic shaker at 800 rpm at room temperature (25 ± 2℃) with contact time of 24 h. After reaching an equilibrium, suspension was centrifuged at 5000 rpm for 10 min, and the supernatant was collected for the remaining mercury concentration analysis. To investigate the effect of adsorbent dosages on Hg2+ removal, different dosages (10, 30, 50, 70 and 90 mg L−1) were added into the solutions with constant initial Hg2+concentration of 1 mg L−1 at 25℃ without pH adjustment. To investigate the effect of pH, Hg2+ removal experiments were carried out at the pH of 3, 5, 7, 9 and 11 with the optimal adsorbent dosage and the fixed initial Hg2+ concentration of 1 mg L−1 at 25℃. The effect of initial Hg2+concentration was determined by performing Hg2+ removal tests at varying Hg2+ concentrations of 0.5, 1.0, 2.0, 3.0 and 5.0 mg L-1 respectively using the constant dosage and optimum pH. The effect of reaction time on mercury removal was determined by batch experiments conducted in duplicate. To study the kinetic model of Hg2+ adsorption, a batch experiment was performed at 25℃ for 48 h in which the optimum dosages of adsorbent and Hg2+ solutions were mixed at the optimal pH value. At predetermined time intervals (from 30 min to 48 h), the reaction was artificially terminated in duplicate vials to attain a transient state. Immediately, solids and liquids to be sampled were separated by centrifugation. The experimental data from this study was used to determine the adsorption kinetic models and isotherm models. 3. Results and discussion 3.1. Composite analysis for physical adsorption capability Since physical adsorption is mainly supported by large surface area and hierarchical porous 3D morphology, we investigated the microstructure of the porous carbon, the porous carbon/Fe composite and the porous carbon/Fe-Mn composite. In Fig. 2, the pristine porous carbon typically demonstrates tunnel-like structures with hierarchical 3D mesopore units. For porous carbon/Fe, abundant nanoparticles are on the 35
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Fig. 2. SEM images of (a–b) porous carbon and its composites: (c–d) porous carbon/Fe and (e–f) porous carbon/Fe-Mn.
composite. FTIR spectra of the three samples porous carbon, porous carbon/Fe and porous carbon/Fe-Mn are compared as shown in Fig. S1(a). Porous carbon has a broad adsorption band at 3406 cm−1 corresponding to −OH stretching vibration. The band at 1716 cm−1 is characteristic of the C]O stretching vibration which attributes to carboxyl COOH, and the band at 1571 cm−1 relates to asymmetric−COO–stretching. The broad band at ˜1133 cm-1 is assigned to CeO groups stretching vibration [32]. Comparatively, the porous carbon/Fe composite shows a new band at 709 cm−1, attributing to Fe–O–Fe, whilst the C]O band is diminished. In the porous carbon/Fe-Mn composite, the band intensity at 1719 cm−1 (C]O) and 1129cm-1 (CeO) are strengthened, and the band of asymmetric carboxyl −COO– stretching vibration shifts from 1571 to 1581 cm−1. Also, a new band appearing at 604 cm−1 indicates the presence of Mn–O–Mn groups. These changes could be explained by the interaction through oxygen-containing functional groups (C]O, −COO, CeO etc.) which bond the Fe-Mn oxide nanoparticles with porous carbon matrix [33]. Raman spectra are shown in Fig. S1(b). All three samples exhibit a characteristic D band at ˜ 1335 cm−1 and a G band at ˜ 1604 cm−1.The disordered D band arises from sp3 hybridization of CeC bonds, while the ordered G band is suggestive of sp2 C]C bond stretching vibration [34]. Compared with porous carbon, D band at 1354 cm−1 and G band at 1601 cm−1 of porous carbon/Fe-Mn show a slight frequency shift (D: + 9 cm−1, G: –3 cm−1), which confirm the interaction between the carbon matrix and the nanoparticles. The intensity ratio of the D to G
The TGA profiles of porous carbon, porous carbon/Fe and porous carbon/Fe-Mn composites are shown in Fig. 4(f). The weight loss below 120℃ corresponds to the loss of surface water; weight loss percentiles are 0.3 wt%, 1.4 wt%, and 2.7 wt% for porous carbon, porous carbon/ Fe and porous carbon/Fe-Mn, respectively. The largest weight loss occurs at 120–400℃ and above 400℃. The weight loss between 120℃ and 400℃ is induced by the pyrolysis of certain oxygen-containing functional groups, i.e. C]O, −COO, CeO, CeOeC, etc. The proportion of oxygen-containing functional groups is 1.1 wt% for porous carbon, 1.2 wt% for porous carbon/Fe and 6.4 wt% for porous carbon/Fe-Mn. The weight loss above 400℃ arises from the decomposition of carbon skeletons. Carbon content is found to be 98.6 wt%, 71.8 wt% and 37.4 wt% respectively for porous carbon, porous carbon/Fe and porous carbon/Fe-Mn. At the end of the heating process, no porous carbon remains, but porous carbon/Fe remains 25.6 wt% by weight, and porous carbon/Fe-Mn remains 53.5 wt%. Based on these data, we calculated that porous carbon/Fe contained ˜8.96 wt% Fe, and porous carbon/Fe-Mn contained 27.4 wt% Fe and 9 wt% Mn. The loss at 120–240℃ relates to the loss of labile oxygen-containing groups, and the loss at 240–400℃ may correspond to the pyrolysis of more stable oxygen groups such as C]O [31].
3.2. Composite analysis for chemical adsorption capability Chemical adsorption mainly involves complexation interaction and functional groups bonding contributed from the metal oxides in the 36
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Fig. 3. TEM images of (a–b) porous carbon and its composites: (c–d) porous carbon/Fe and (e–f) porous carbon/Fe-Mn.
MnOOH. Furthermore, after adsorbing Hg2+, the composite indicates the presence of Hg (II). Peaks at 100.8 eV and 104.6 eV are characteristic of Hg4f7/2 and Hg4f5/2 consistent with Hg–O bonds [23]. After Hg2+ adsorption, the peak of CeC/C]C increases likely due to π-π electrostatic interactions. The Mn2p peak shift may suggest that MnO2 and MnOOH strongly interact with Hg, as Hg can implant into the internal structure of manganese oxides through ligand exchange [36]. This further confirms that Hg (II) is adsorbed in an oxidized state.
band (ID/IG) was qualified as a measure of the degree of graphitization [35]. Porous carbon has an ID/IG ratio of 0.81. It increases to 0.94 for porous carbon/Fe and decreases to 0.89 for porous carbon/Fe-Mn. These ratios mean that the final composite still contain a graphene-like layer structure with a well-defined degree of order. When the characteristic peaks at 200–450 cm−1 are magnified in Fig. 5(c), new peaks at 220, 284.6 and 398.8 cm−1 are found for porous carbon/Fe, corresponding to Fe–O–Fe bonds which attribute to Fe2O3. For the porous carbon/Fe-Mn composite, these peaks respectively shift to lower wave numbers at 216.5, 279.4, and 369.6 cm−1, implying the coexistence of iron and manganese oxides and their interaction. Another emerging peak at 651.4 cm-1 is assigned to the Mn–O–Mn group and the Mn−OH bonds from MnO2. XPS spectra of porous carbon, porous carbon/Fe and porous carbon/ Fe-Mn composites before and after Hg2+ adsorption was studied to determine their chemical oxidation states and reaction mechanisms. The XPS total survey spectra are shown in Fig. S1(d)-(f). Fig. 5 shows the XPS spectra for C1 s, Fe2p, Mn2p, and Hg4f for the porous carbon/ Fe-Mn composite before and after uptake of Hg2+. The characteristic C1 s peaks at 284.7 eV, 286.3 eV and 288.5 eV can be well-indexed respectively to CeC/C]C, CeOeC and C]O, in agreement with FTIR results. For Fe2p, the binding energy peak at 710.5 eV is representative of Fe2p3/2 attributed to α–Fe2O3. Binding energy peaks at 712.8 eV and 724.5 eV are ascribed to Fe2p3/2 and Fe2p1/2 of FeOOH respectively. The weak peak at 719 eV pertains to Fe2p1 of zero-valent iron. For Mn2p, the peaks at 641.2 eV and 642.9 eV are assigned to Mn2p3/2 of MnO2 and MnOOH respectively. Hence, the metal oxides in the composite primarily exist in the forms of Fe2O3, FeOOH, MnO2 and
3.3. Heavy Hg2+ removal performance The Hg (II) removal performance by porous carbon/Fe-Mn composite was dependent on pH, adsorbent dosage and initial Hg concentration. The initial concentration prior to the treatment is defined as C0. The normalized removal concentration percentage, calculated by (C0–C)/C0, is described as removal efficiency. As shown in Fig. 6 (a), the removal efficiency increases from 82.1% to 98.4% when the carbon/Fe-Mn composite concentration changes from 10 to 90 mg L−1. The larger dosage of adsorbent would provide more adsorption sites and functional groups to combine with, resulting in apparent increments in Hg removal efficiency. However, when the adsorbent dosage exceeds ˜80 mg L−1, the removal efficiency ceases to increase likely because the adsorption reached an equilibrium. To optimize Hg uptake and make full use of the adsorbent, we used an optimal dosage 80 mg L−1 in the following experiments. The efficiency of Hg removal was also strongly affected by pH. In Fig. 6 (b), Hg removal efficiency increases with pH from 3 to 5 and then decreases sharply until pH reaches 11. The graph presents the highest 37
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Fig. 4. (a) Nitrogen adsorption/desorption isotherms, (b) pore size distribution (c) XRD pattern, (d) Zeta potential (e) Mercury speciation (Hg2+ 1 mg L−1, pH 1–9, molar Fe/Mn 3/1). (f) TGA curves of porous carbon and its composites: carbon/Fe and carbon/Fe-Mn.
performance of porous carbon three times than iron oxide. With the adsorbent dosage of 80 mg L−1 and initial Hg2+ concentration of 1.0 mg L−1, the removal efficiency of Hg2+ by porous carbon/Fe-Mn is as high as 96.8%. Under the same experimental conditions, the GO/FeMn composite synthesized by Tang et al. had a Hg removal efficiency of just 89.0% [37].
removal efficiency at pH 5–7. The pH not only affects the surface charge of adsorbent but influences the species of Hg. As demonstrated by the Zeta potential graph (Fig. 3) and the mercury speciation curve over different pH values, overly high or overly low pH values are unfavorable for Hg removal. Taking all factors into account, we determined the optimum pH value to be 6. At this pH value, Hg (II) exists as an oxidized state of Hg(OH)2. Therefore, it is inferred that metal oxides on the composite surface can interact with Hg(OH)2 to form complex species such as Fe2O3=Hg (II) or FeOOH=Hg (II) [37]. In brief, during this process some complexation interactions occurs for the surface functional groups between metal oxides and Hg(OH)2 at pH 6. The effect of initial Hg concentration on removal efficiency is trivial as shown in Fig. 6(c). Over the concentration range, the adsorbent offers abundant available active sites for Hg to anchor to. For convenience, batch experiments were performed at a fixed initial Hg concentration 1 mg L−1. The efficiency of Hg removal was measured as a function of contact time. In Fig. 6(d), the removal rates for all the samples increase sharply until at ˜150 min and then they increase slowly reaching their own equilibriums. The equilibration removal efficiency of Hg was 96.8% for porous carbon/Fe-Mn, compared to 85.0% for porous carbon/Fe and 81.1% for porous carbon. It is worthy to note that the adsorption performance of the porous carbon/Fe-Mn composite was much higher than that of porous carbon. Fe-Mn composite improves the adsorption
3.4. Mechanisms for high Hg2+ removal performance Porous carbon/Fe-Mn exhibits not only a fast adsorption rate but a high adsorption capacity. The Hg (II) amount adsorbed by unit mass of adsorbent at the equilibrium is defined as qe for measurement of the equilibrium adsorption capacity (mg g−1). The equilibration adsorption capacity of the porous carbon/Fe-Mn composite was calculated to be 9.8 mg g−1. In Fig. S2(a), adsorption kinetic tests examine the efficiency of Hg removal by the porous carbon/Fe-Mn composite. The removal rate increases significantly at the initial stage. During the first 30 min, Hg removal efficiency of 91.8% is reached as most accessible adsorption sites are occupied. After 3 h removal efficiency reaches a stable level, and 12 h later the adsorption sites are saturated. Data processing of the graph yields a maximum Hg removal efficiency of 96.8%. Accordingly, the commonly used pseudo-first order model and pseudo-second order model were adopted to simulate the kinetics data. The formulas are 38
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Fig. 5. XPS spectra of carbon/Fe-Mn for (a) C1 s, (b) Fe2p, (c) Mn2p and (d) Hg4f.
In these equations, qe and q correspond to the adsorption capacity respectively at an equilibrium and at time; K1 and K2 are the adsorption rate constants of pseudo-first order and pseudo-second order models; and R2 is the resultant fitting parameter. Apparently, the pseudo-second order model (R2 = 0.999) fits better. Hence, the rate limiting step was the adsorption process instead of the diffusion process [38], and the Hg removal mechanism was mainly ligand exchange and surface complexation between porous carbon/Fe-Mn and Hg. Adsorption isotherms were investigated to further identify the
expressed as Eqs. (1)–(2). Pseudo-first order model: ln(qe-q)= lnqe-K1t
(1) −1
Where qe = 9.8 (mg g ), K1 = 0.005 (g mg Pseudo-second order model:
−1 ⋅
min
−1
2
−1
2
) and R = 0.944
t/q = 1/(K2qe2)+ t/qe Where qe = 9.8 (mg g
(2)
−1
), K2 = 0.083 (g mg
−1 ⋅
min
) and R = 0.999
Fig. 6. (a) Hg removal with various adsorbent concentration (pH 6, Hg2+ 1 mg L−1, 24 h at 25 ± 2℃). (b) Hg removal with various pH (adsorbent 80 mg L-1, Hg2+ 1 mg L−1, 24 h at 25 ± 2℃). (c) Hg removal with various Hg2+ concentration (pH 6, adsorbent 80 mg L−1, 24 h at 25 ± 2℃). (d) Hg removal rate by different adsorbent materials (pH 6, adsorbent 80 mg L−1, Hg2+ 1 mg L−1, at 25 ± 2℃).
39
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pseudo-second order model, and the adsorption isotherms are consistent with the Freundlich model. Furthermore, the equilibration behavior is multilayer adsorption on heterogeneous surface. This research opens new avenues for the development of cost-effective, environmentally-friendly materials for removal of Hg ions and other heavy metal ions from wastewater.
adsorption mechanism and analyze how Hg interacted with porous carbon/Fe-Mn. The adsorption isotherm describes the amount of Hg adsorbed by unit mass of adsorbent at a constant temperature as a function of Hg concentration at equilibration. In Fig. S2 (b)-(e), Langmuir and Freundlich isotherm models are used to simulate the experimental data in table S1. The Langmuir isotherm model is applicable to monolayer adsorption on the homogeneous surface, with all binding sites equal and adsorption energy constant [39]. The Freundlich isotherm model describes multilayer adsorption on the heterogeneous surface, whose binding sites have different adsorption energies [40]. The linear forms of the two models are given as Eqs. (3)–(4). Langmuir isotherm model:
Ce 1 C = + e qe KL⋅qmax qmax
The authors gratefully acknowledge schematic diagram design work provide by Mr. Rong Jiang at School of Design of Shanghai Jiao Tong University and financial support for this research from the National Key R&D Program of China (2016YFA0202900, 2016YFC1402400), National Natural Science Foundation of China (51672173, 51801121), Shanghai Science and Technology committee (17JC1400700, 18ZR1421000, 18520744700), Science and Technology Planning Project of Guangdong Province (2016A010103018), Program of Guangdong Academy of Sciences (2017GDASCX-0503), Shanghai Research institute of criminal science and technology (2016XCWZK15). We also thank the Shanghai Synchrotron Radiation Facility for the measurements.
(3) −1
Where qmax = 51.8 (mg g ), KL = 9.280 (L mg Freundlich isotherm model:
ln qe = ln kF +
Acknowledgments
1 ⋅ln Ce n
−1
2
) and R = 0.502
(4) −1
2
Where 1/n = 0.898, Kf = 304.3 (L g ) and R = 0.989 In these equations, Ce is the equilibration concentration of Hg (mg L−1); qmax is the maximum adsorption capacity; KL and Kf are the Langmuir and Freundlich constants relating to the adsorption energy; and 1/n is the constant related to the adsorption intensity. The Freundlich isotherm model (R2 = 0.989) is more favorable, implying that the adsorption process was multilayer adsorption on the heterogeneous surface. In summary, the process of Hg removal covers both physical and chemical adsorption. On and near the surface of the adsorbent, the dominant mechanism would be chemical adsorption including ligand exchange and surface complexation. Further away from the surface, physical adsorption such as electrostatic attraction in the multilayer adsorption process would take a major role. The comparison of different adsorbents reported with our sample was summarized in Table S2. Activated carbon as adsorbent works through physical adsorption; Mn-Fe oxide and modified Fe oxide work through chemical adsorption; magnetic powered activated carbon (MPAC) and GO/Fe-Mn composite involve both physical and chemical adsorption. The results of our work are outstanding in both Hg(II) removal efficiency and adsorption capacity. The G-RHC/Fe-Mn composite possessed excellent Hg(II) removal efficiency of 96.8%, higher than those of reported activated carbons (82%), in-situ formed Mn-Fe oxides (80%), MPAC (96.3%) and GO/Fe-Mn composite (89%). The Hg(II) removal efficiency of modified Fe oxide is higher but its adsorption capacity is much lower than our sample. Our sample displayed a considerable Hg(II) adsorption capacity of 9.81 mg g−1. This value is much higher than those of reported activated carbons (0.87 mg g−1), Mn-Fe oxides (0.6 mg g−1), modified Fe oxide (0.59 mg g−1) and MPAC (0.096 mg g−1). The adsorption capacity of GO/Fe-Mn is higher but its removal efficiency is lower than our sample. The adsorption capacity of our G-RHC/Fe-Mn composite is far higher than that of the similar counterpart MPAC, owing to the synergetic effect of the active binary Fe-Mn oxides.
Appendix A. Supplementary data Supplementary material related to this article can be found, in the online version, at doi:https://doi.org/10.1016/j.jhazmat.2019.02.099. References [1] M. Yeganeh, M. Afyuni, A.-H. Khoshgoftarmanesh, L. Khodakarami, M. Amini, A.R. Soffyanian, R. Schulin, Mapping of human health risks arising from soilnickel and mercury contamination, J. Hazard. Mater. 244–245 (2013) 225–239. [2] Z. Chen, Z. Geng, Z. Zhang, L. Ren, T. Tao, R. Yang, Z. Guo, Synthesis of magnetic Fe3O4@C nanoparticles modified with -SO3H and -COOH groups for fast removal of Pb2+, Hg2+, and Cd2+ ions, Eur. J. Inorg. Chem. 2014 (20) (2014) 3172–3177. [3] M.M. Matlock, B.S. Howerton, D.A. Atwood, Irreversible precipitation of mercury and lead, J. Hazard. Mater. 84 (1) (2001) 73–82. [4] S. Chiarle, M. Ratto, M. Rovatti, Mercury removal from water by ion exchange resins adsorption, Water Res. 34 (11) (2000) 2971–2978. [5] D. Sevdić, L. Fekete, H. Meider, Macrocyclic polythiaethers as solvent extraction reagents—III, J. Inorg. Nucl. Chem. 42 (6) (1980) 885–889. [6] K. Kim, S.H. Lee, W. Yi, J. Kim, J.W. Choi, Y. Park, J.-I. Jin, Efficient field emission from highly aligned, graphitic nanotubes embedded with gold nanoparticles, Adv. Mater. 15 (2003) 1618–1622. [7] D.S. Han, M. Orillano, A. Khodary, Y. Duan, B. Batchelor, A. Abdel-Wahab, Reactive iron sulfide (FeS)-supported ultrafiltration for removal of mercury (Hg(II)) from water, Water Res. 53 (2014) 310–321. [8] V. Antochshuk, M. Jaroniec, 1-Allyl-3-propylthiourea modified mesoporous silica for mercury removal, Chem. Commun. 3 (3) (2002) 258. [9] O. Hakami, Y. Zhang, C.J. Banks, Thiol-functionalised mesoporous silica-coated magnetite nanoparticles for high efficiency removal and recovery of Hg fromwater, Water Res. 46 (2012) 3913–3922. [10] P. Bonnissel-Gissinger, M. Alnot, J.-P. Lickes, J.-J. Ehrhardt, P. Behra, Modeling the adsorption of mercury(II) on (hydr)oxides II: α-FeOOH (Goethite) and amorphous silica, J. Colloid Interface Sci. 215 (2) (1999) 313–322. [11] F.-S. Zhang, J.O. Nriagu, H. Itoh, Mercury removal from water using activatedcarbons derived from organic sewage sludge, Water Res. 39 (2005) 389–395. [12] G. Skodras, I. Diamantopoulou, A. Zabaniotou, G. Stavropoulos, G.P. Sakellaropoulos, Enhanced mercury adsorption in activated carbons from biomass materials and waste tires, Fuel Process. Technol. 88 (8) (2007) 749–758. [13] M. Zabihi, A. Haghighi Asl, A. Ahmadpour, Studies on adsorption of mercury from aqueous solution on activated carbons prepared from walnut shell, J. Hazard. Mater. 174 (1–3) (2010) 251–256. [14] F.-S. Zhang, J.O. Nriagu, H. Itoh, Mercury removal from water using activated carbons derived from organic sewage sludge, Water Res. 39 (2) (2005) 389–395. [15] C. Fischer, M. Oschatz, W. Nickel, D. Leistenschneider, S. Kaskel, E. Brunner, Bioinspired carbide derived carbons with hierarchical pore structure for the adsorptive removal of mercury from aqueous solution, Chem. Commun. 53 (35) (2017) 4845–4848. [16] S.-W. Cao, Y.-J. Zhu, M.-Y. Ma, L. Li, L. Zhang, Hierarchically nanostructured magnetic hollow spheres of Fe3O4 and g-Fe2O3: preparation and potential application in drug delivery, J. Phys. Chem. C 112 (2008) 1851–1856. [17] M. Rose, Y. Korenblit, E. Kockrick, L. Borchardt, M. Oschatz, S. Kaskel, G. Yushin, Hierarchical micro and mesoporous carbide derived carbon as a high performance electrode material in supercapacitors, Small 7 (8) (2011) 1108–1117. [18] P. Zhang, X. Ma, Y. Guo, Q. Cheng, L. Yang, Size-controlled synthesis of hierarchical NiO hollow microspheres and the adsorption for Congo red in water, Chem. Eng. J.
4. Conclusion A novel hierarchical carbon/Fe-Mn composite readily fabricated from biomass was utilized as an adsorbent for Hg (II) removal. Since the composite contained Fe2O3 and MnO2 as dispersion phase which had oxygen-containing functional groups FeOOH and MnOOH, it was found to readily bond with mercury ions in wastewater through ligand exchange and surface complexation. This combined with the large surface area of 350.8 m2 g−1 creating a synergic effect on Hg ion removal. The composite exhibited fast adsorption rate, high removal efficiency of 96.8%, and considerable adsorption capacity of 9.8 mg g−1, superior to those reported adsorbents. The adsorption kinetics agrees with the 40
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[30] R.M. Cornell, R. Giovanoli, W. Schneider, Review of the hydrolysis of iron (III) and the crystallization of amorphous iron(III) hydroxide hydrate, J. Chem. Technol. Biotechnol. 46 (2) (2010) 115–134. [31] M. Lv, X. Wang, J. Li, X. Yang, Ca. Zhang, J. Yang, H. Hu, Cyclodextrin-reduced graphene oxide hybrid nanosheets for the simultaneous determination of lead(II) and cadmium(II) using square wave anodic stripping voltammetry, Electrochim. Acta 108 (2013) 412–420. [32] J. Wang, Z. Chen, B. Chen, Adsorption of polycyclic aromatic hydrocarbons by graphene and graphene oxide nanosheets, Environ. Sci. Technol. 48 (9) (2014) 4817–4825. [33] V. Chandra, K.S. Kim, Highly selective adsorption of Hg2+by apolypyrrole-reduced graphene oxide composite, Chem. Commun. 47 (2011) 3942–3944. [34] P.N. Diagboya, B.I. Olu-Owolabi, D. Zhou, B.-H. Han, Graphene oxide–tripolyphosphate hybrid used as a potent sorbent for cationic dyes, Carbon 79 (2014) 174–182. [35] R. Droppa, C.T.M. Ribeiro, A.R. Zanatta, M.C. dos Santos, F. Alvarez, Comprehensive spectroscopic study of nitrogenated carbon nanotubes, Phys. Rev. B 69 (4) (2004) 045405. [36] P. Thanabalasingam, W.F. Pickering, Sorption of mercury(II) by manganese (IV) oxide, environmental pollution series B, Chem. Phys. 10 (1985) 115–128. [37] Y.S. Ho, G. McKay, Pseudo-second order model for sorption processes, Process Biochem. 34 (5) (1999) 451–465. [38] K.G. Bhattacharyya, S. Sen Gupta, Adsorption of chromium(VI) from water by clays, Ind. Eng. Chem. Res. 45 (21) (2006) 7232–7240. [39] Z.M. Lei, Q.D. An, Y. Fan, J.L. Lv, C. Gao, S.R. Zhai, Z.Y. Xiao, Monolithic magnetic carbonaceous beads for efficient Cr(VI) removal from water, New J. Chem. 40 (2) (2015) 1195–1204. [40] H. Parham, B. Zargar, R. Shiralipour, Fast and efficient removal of mercury from water samples using magnetic iron oxide nanoparticles modified with 2-mercaptobenzothiazole, J. Hazard. Mater. 205 (2012) 94–100.
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Superior removal of Hg (II) ions from wastewater using hierarchically porous, functionalized carbon
T
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Yao Lia,1, Mengdan Xiaa,1, Fufei Ana, Nianfang Mab, Xueliang Jiangc, Shenmin Zhua, , ⁎ Dawei Wangd, Jun Mae, a
State Key Laboratory of Metal Matrix Composites, Shanghai Jiao Tong University, Shanghai 200240, PR China Guangdong provincial bioengineering institute(Guangzhou Sugarcane industry research institute), Guangdong Key Laboratory of Sugarcane Improvement and Biorefinery, Guangzhou, 510316, China c School of Materials Science and Engineering, Wuhan Institute of Technology, Wuhan 430205, China d School of Chemical Engineering, UNSW Australia, UNSW Sydney, NSW 2052, Australia e School of Advanced Manufacturing & Mechanical Engineering, University of South Australia, Australia b
A R T I C LE I N FO
A B S T R A C T
Keywords: Heavy metal ions Porous carbon Wastewater Hg (II)
The removal of heavy metal ions from industrial wastewater by adsorption has been central to the environment for decades, where common adsorbent materials are often limited by poor efficiency, complex fabrication and long processing time. Porous carbon derived from biospecies holds promise to address the limitations. In this study we converted bagasse into a carbon composite having hierarchically porous structure; the composite’s dispersion phases – iron oxide and manganese oxide – were synthesized by a simple one-step liquid-phase reaction method. Featuring large specific surface area of 350.8 m2 g−1, the composite demonstrated exceptional Hg (II) removal efficiency of 96.8%, adsorption rate of up to 96.8% within 150 min and adsorption capacity of 9.8 mg g−1. In comparison with other removal materials, our work is outstanding in terms of both removal efficiency and synthesis simplicity. The high efficiency is attributed to the synergy between physical adsorption referring to hierarchically porous structure and chemical adsorption relating to functional complexation processes. It provides a new avenue for the development of high-performance adsorbent materials for heavy metal removal from aqueous media.
1. Introduction The toxicity, carcinogenicity and long-term accumulation effect of heavy metal ions cause enormous risks to both the environment and human health. In particular, mercury (Hg) produced from industrial water is a major toxic element found in the environment [1]. The presence of Hg2+ can lead to a variety of human health problems such as brain and heart diseases, liver and kidney damage, and even cancer [2]. Research into eco-friendly, cost effective and efficient means for Hg2+ removal has gained more interests. A number of approaches have been developed, mainly including chemical precipitation [3], ion exchange [4], solvent extraction [5], evaporation [6], ultrafiltration [7] and adsorption [8]; of these, adsorption is considered as the most promising due to its simplicity, high efficiency, low cost and operation convenience [9]. Thus, a variety of adsorbent materials have been
investigated for Hg2+ removal, including metal oxides [10] and activated carbon [11]. Porous carbon fabricated from biospecies holds promise, because of its interconnected, three-dimensional, porous structures with high specific surface area. As a major type of adsorbents, porous carbon removes Hg (II) from aqueous media mainly through physical adsorption which involves van der Waals or electrostatic interactions taking advantage of large surface area. G. Skodras et al. studied Hg removal through activated carbon which was prepared from agricultural residues and waste tires [12]. M. Zabihi et al. fabricated porous carbons with surface area of 780 m2 g−1 from walnut shells, which exhibited a high monolayer adsorption capacity of 151.5 mg g-1 for Hg (II) removal [13]. Zhang et al. studied various types of activated carbon extracted from organic sludge via chemical activation. ZnCl2-activated carbon exhibited quite high adsorption capacity for Hg (II), and 60–80% of the adsorbed Hg (II) could be recovered via sonication [14].
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Corresponding authors. E-mail addresses: [email protected] (S. Zhu), [email protected] (J. Ma). 1 These authors contributed equally. https://doi.org/10.1016/j.jhazmat.2019.02.099 Received 22 December 2018; Received in revised form 25 February 2019; Accepted 26 February 2019 Available online 27 February 2019 0304-3894/ © 2019 Elsevier B.V. All rights reserved.
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Hierarchical porous materials may be promising candidates for adsorption, because its full accessibility of narrow cavities and efficient mass transport property. C. Fischer et al. developed bioinspired carbidederived carbons with hierarchical pore structure for Hg (II) adsorption [15]. Compared with traditional activated carbons, their micropore size can be adjusted with high accuracy. Bioinspired carbon exhibits high Hg (II) adsorption capacities. From the microcosmic perspective, the 3D hierarchically pore system involves mainly micropores and mesopores [16]. The micropores directly lead to enhanced adsorption capacity; the well-developed mesopores promoted the optimum adsorption kinetics [17]. The mesopore channels also serve as liquid flow pathways and allow fast transport of Hg (II) to active adsorption sites. Hence, it is obvious that hierarchically structured materials should play a critical role in the removal of heavy metal ions [18–21]. Although porous carbon is advantageous of fast adsorption, it displays unideal adsorption efficiency and difficult recovery of adsorbents as its performance is contributed by physical adsorption only [22]. Chemical adsorption refers to a phenomenon where an adsorbent reacts with contaminant(s) through interactions. Lu et al. removed traces of Hg (II) through chemical adsorption where Hg (II) ions were complexed with the hydroxyl groups on the surface of in situ formed manganeseferric hydroxide. It was reported that the removal mechanisms comprised of surface complexation and flocculation-precipitation processes, leading to transformation from liquid to solid phase of Hg (II). Fe-Mn oxide at 40 mg L−1 could effectively remove 80% Hg (II) starting from initial concentration of 30 μg L−1, which represents high adsorption efficiency [23]. Noteworthy is that the chemical adsorption can be further enhanced by modifying the magnetic adsorbent surface with carefully selected functional organic groups. Compared to physical adsorption, chemical adsorption is faster and exhibits higher removal efficiency for Hg (II). To improve adsorption efficiency, physical adsorption has been combined with chemical adsorption. Faulconer et al. impregnated iron oxide with the activated carbon to obtain a new magnetic adsorbent which achieved Hg2+ removal of 96.3 ± 9% and sorbent recovery 92.5 ± 8.3% by magnetic separation. In the adsorption process, iron oxide proved to form strong binding affinity with Hg species and offered plentiful adsorption sites for Hg2+ through coordination with oxygen atoms. This adsorbent could remove Hg (II) to attain a final concentration of 0.2 μg L−1 [24]. Further investigation found that manganese oxides demonstrated similar or even superior chemical adsorption performance. Moghaddam et al. prepared MnO2-coated carbon nanotubes to remove Hg (II) from aqueous solution and obtained adsorption capacity of 58.8 mg g-1, where MnO2 went through oxidationreduction reactions [25]. However, the adsorption value is not satisfactory owing to the low specific surface area of the composite. Inspired by nature, we design and fabricate a carbonaceous material from an industrial waste – bagasse, where graphene-like structure is in situ formed; the carbonaceous material is then combined with metal oxides Fe2O3 and MnxOy by using a one-step liquid-phase reaction method, to prepare porous carbon coupled with Fe-Mn oxide (denoted porous carbon/Fe-Mn). Since the graphene-like structure is introduced to enhance conductivity, the porous matrix behaves like a conductor which not only gathers charge on the pore surface but works as a conductive network linking dielectric metal oxide nanoparticles; this would result in highly effective electrostatic interactions for Hg adsorption. Fe-Mn binary oxides with surface charge and variable valence would have higher surface activity than single oxide, suggesting great potential for water treatment [26]. As a result, the composite – which features a biological structure that cannot be manually synthesized – exhibits outstanding mercury adsorption capacity. Later sections discuss in detail the relationship between composite structure and adsorption properties.
Fig. 1. Schematic fabrication of the porous carbon/Fe-Mn oxide composite and its adsorption mechanism for Hg (II).
2. Experimental 2.1. Chemicals All chemicals and reagents used in this study were of analytical grade or higher, including mercury nitrate monohydrate (Hg (NO3)2·H2O), iron sulfate heptahydrate (FeSO4·7H2O), potassium permanganate (KMnO4), nitric acid (HNO3), sodium hydroxide (NaOH), and potassium hydroxide (KOH). All solutions were prepared with N2purged deionized (DI) water. 2.2. Materials preparation and characterization Fig. 1 describes the preparation of porous carbon with binary oxide nanoparticles in situ formed. In a typical synthesis, a precursor bagasse was employed to fabricate porous carbon via carbonization at 650℃, followed by activation at 700℃ using KOH as an activator [27]. Graphitization was performed in vacuum at 800℃ for 3 h. The powder was washed with 11% dilute hydrochloric acid and dried under vacuum, to obtain porous graphitic carbon. Carboxylation was conducted by mixing 2.0 g of porous carbon with 300 mL of concentrated nitric acid and stirring for 3 h under water-bath heating at 60℃. The sample was first separated from the solution via vacuum filtration. Afterwards, the sample was rinsed to a neutral pH with DI water. After drying the sample at 60℃ for over 24 h, the carboxyl porous carbon was attained. The porous carbon/Fe-Mn composite was prepared following a revised method reported by Tang et al [28]. First, 1.202 g of carboxyl porous carbon was suspended in 30 mL DI water. FeSO4 solution (2.88 × 10−1 M) and KMnO4 solution (9.58 × 10−2 M) were prepared with DI water. Then, 6 mL of FeSO4 and 18 mL of porous carbon suspension were dispersed in 270 mL of DI water under vigorous magnetic stirring. Next, 6 mL of KMnO4 solution was added dropwise into the mixture which pH was adjusted to about 7.5 using HNO3 (1 M and 0.1 M) and NaOH (1 M and 0.1 M). The mixture was continuously stirred 1 h and aged at 25 ± 2℃ for 12 h. The resultant precipitate was collected through vacuum filtration, washed with DI water, and dried at 90℃ for over 48 h, to create porous carbon/Fe-Mn composite (molar ratio of Fe/Mn = 3/1). As a benchmark, a porous carbon/Fe oxide composite (denoted porous carbon/Fe) was synthesized using the same approach but without adding Mn. The surface morphology and chemical composition were observed by scanning electron microscopy (SEM) performed on a JEOL JSM6360LV field emission microscope at 15 kV. Transmission electron microscopy (TEM) was carried out on a JEOL 2010 microscope at 200 kV. The Brunauer-Emmett-Teller (BET) adsorption method was performed 34
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wrinkled carbon matrix, resulting in a rough surface. As for porous carbon/Fe-Mn, a greater number of nanoparticles are present as aggregated flocs formed in situ on the porous carbon surface. The fine morphology of porous carbon and its composites is further confirmed by TEM. In Fig. 3, The interplanar distance of porous graphitic carbon is about 0.33 nm. The 3D porous structure is observed in all three materials. For both composites, abundant nanoparticles are distributed on the surface of porous carbon in the form of aggregated flocs. The 3D hierarchical structure constructed by porous carbon matrix and metal oxide nanoparticles would provide well-developed pore channels, serving as liquid flow pathways for the transport of Hg (II) to active adsorption sites. Fig. 4(a)-(b) contains the nitrogen adsorption/desorption isotherms and pore size distribution curves of synthesized composites. The BET surface area of porous carbon and its composites-carbon/Fe and carbon /Fe-Mn- are 1125.2, 945.3 and 350.8 m2 g−1, respectively, corresponding to pore size 3.5, 3.9 and 6.9 nm. The porous carbon produced from bagasse has a far larger specific surface area than previous efforts, such as 803 m2 g−1 of activated carbon powder from walnut shells [13]. This means great potential for adsorption. Embedding the metal oxides inside porous carbon would slightly reduce the surface area, because iron and manganese oxides may stay on the carbon surface occupying certain micropores. More importantly, the embedding should result in larger pore size, which is desired for Hg (II) adsorption. All three adsorption/desorption isotherms show hysteresis loops in the P/P0 range of 0.45–1.0, with pore widths centered at ˜3.9 nm, indicating mesoporous nature. The porous carbon/Fe-Mn composite still possesses relatively high specific surface area of 350.8 m2 g−1, which would benefit the adsorption of Hg2+. This represents an advance over previous effort where the surface area values of manganese dioxide and iron oxide were reported as 266.9 m2 g−1 and 172.3 m2 g-1 [29]. In Fig. 4 (c), the XRD pattern of porous carbon shows diffraction peaks at 26.6° and 43.5°. The former peak can be indexed to the (003) plane of the graphene-like structure, suggesting some degree of graphitic order. The layered structure of TEM images in Fig. 3(a) corresponds to the (003) plane of the graphene-like structure in XRD analysis. For porous carbon/Fe, the diffraction at 26.6° became unobvious because the iron oxide coated on the surface may interact with porous carbon. The peaks located at 26.8°, 36.2°, 47.3°, 54.9° and 60.0° were all assigned to crystalline phase α-Fe2O3, indicating that amorphous FeOOH was readily transformed into crystalline iron (III) oxides [30]. The manganese oxide component is mainly MnO2, corresponding to the peaks at 34.0°,42.9°, and 62.7°. The diffraction peaks of carbon are weakened implying that crystalline phase of iron oxides transformed into amorphous phase. This indicates that the formation of crystals would be restrained by the co-existence of iron and manganese oxides during the synthesis. For crystalline Fe2O3, the particle size is calculated to be 28.2 nm according to the Scherrer formula. Fe and Mn oxides offer plenty of active complexation sites, which is essential for the adsorption of Hg2+. Electrostatic attraction also contributes to adsorption process. Fig. 4(d) reveals Zeta potentials of porous carbon, porous carbon/Fe and porous carbon/Fe-Mn composites as a function of pH. As it shows, there is an isoelectric point near 6. As the adsorbent surface charge influences the interfacial reaction with its adsorbate, we further explored the interaction between Hg (II) and the adsorbent by utilizing visual MINTEQ (version 3.1) to analyze mercury speciation under different pH. When it is low, Hg (II) exists primarily as Hg2+ cation. At pH 3–5, the adsorbents are positively charged, making electrostatic repulsion unfavorable for Hg (II) adsorption. When pH is over 5, Hg(OH)2 becomes a dominant component. The adsorbents are negatively charged at pH over 6. When pH is over 6, excessive OH− ions compete with the negatively charged ions of the adsorbent to bind with Hg2+, unfavorable for Hg (II) adsorption. Consequently, the isoelectric point at pH 6 appears to be an optimal pH for adsorption.
using an ASAP2020 volumetric adsorption analyzer to measure the surface area, pore size, and pore volume of the adsorbents. The crystalline and amorphous phase compositions were identified using X-ray diffraction (XRD) on a Rigaku D/max 2550 V L/PC system operated at 35 kV and 200 mA with Cu Kα radiation (λ = 1.5406 Å), at a scan rate of 5°min−1 and a step size of 0.02°. Zeta (ζ) potential versus pH was measured using MALVERN Zetasizer at 25℃. Thermal gravimetric analysis (TGA) was conducted on a PE TGA-7 instrument with 10 ℃ min−1 to analyze element content variation. Fourier transform infrared (FTIR) spectroscopy was conducted with a PE Paragon 1000 spectrophotometer on KBr pellets to determine the chemical functional groups. Raman spectroscopy was characterized by a Renishaw in Via Raman Microscope. The adsorbent surface elemental compositions before and after Hg adsorption were analyzed by X-ray photoelectron spectroscopy (XPS) collected on a physical electronics PHI5400 using Mg K radiation as the X ray source. All the spectra were corrected with the C 1 s (285.0 eV) band. The adsorbents after mercury adsorption were separated by centrifugation and freeze-dried under vacuum at −60℃. The residual mercury concentration in the supernatant was determined using a Duo View ICP-OES spectrometer (THERMO iCAP 6000 Radial). 2.3. Mercury batch adsorption tests Hg (II) stock solution (5 mg/L) was prepared by dissolving Hg (NO3)2·H2O into DI water. Initial solutions of varying concentrations were obtained by diluting the stock solution with DI water. The mercury concentration is given as mercury ion concentration. Batch adsorption experiments were performed in sealed 25-mL glass vials. Briefly, the reaction was initiated by mixing a certain dosage of adsorbent into Hg2+solutions. The vials were then sealed with stirring on a thermostated magnetic shaker at 800 rpm at room temperature (25 ± 2℃) with contact time of 24 h. After reaching an equilibrium, suspension was centrifuged at 5000 rpm for 10 min, and the supernatant was collected for the remaining mercury concentration analysis. To investigate the effect of adsorbent dosages on Hg2+ removal, different dosages (10, 30, 50, 70 and 90 mg L−1) were added into the solutions with constant initial Hg2+concentration of 1 mg L−1 at 25℃ without pH adjustment. To investigate the effect of pH, Hg2+ removal experiments were carried out at the pH of 3, 5, 7, 9 and 11 with the optimal adsorbent dosage and the fixed initial Hg2+ concentration of 1 mg L−1 at 25℃. The effect of initial Hg2+concentration was determined by performing Hg2+ removal tests at varying Hg2+ concentrations of 0.5, 1.0, 2.0, 3.0 and 5.0 mg L-1 respectively using the constant dosage and optimum pH. The effect of reaction time on mercury removal was determined by batch experiments conducted in duplicate. To study the kinetic model of Hg2+ adsorption, a batch experiment was performed at 25℃ for 48 h in which the optimum dosages of adsorbent and Hg2+ solutions were mixed at the optimal pH value. At predetermined time intervals (from 30 min to 48 h), the reaction was artificially terminated in duplicate vials to attain a transient state. Immediately, solids and liquids to be sampled were separated by centrifugation. The experimental data from this study was used to determine the adsorption kinetic models and isotherm models. 3. Results and discussion 3.1. Composite analysis for physical adsorption capability Since physical adsorption is mainly supported by large surface area and hierarchical porous 3D morphology, we investigated the microstructure of the porous carbon, the porous carbon/Fe composite and the porous carbon/Fe-Mn composite. In Fig. 2, the pristine porous carbon typically demonstrates tunnel-like structures with hierarchical 3D mesopore units. For porous carbon/Fe, abundant nanoparticles are on the 35
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Fig. 2. SEM images of (a–b) porous carbon and its composites: (c–d) porous carbon/Fe and (e–f) porous carbon/Fe-Mn.
composite. FTIR spectra of the three samples porous carbon, porous carbon/Fe and porous carbon/Fe-Mn are compared as shown in Fig. S1(a). Porous carbon has a broad adsorption band at 3406 cm−1 corresponding to −OH stretching vibration. The band at 1716 cm−1 is characteristic of the C]O stretching vibration which attributes to carboxyl COOH, and the band at 1571 cm−1 relates to asymmetric−COO–stretching. The broad band at ˜1133 cm-1 is assigned to CeO groups stretching vibration [32]. Comparatively, the porous carbon/Fe composite shows a new band at 709 cm−1, attributing to Fe–O–Fe, whilst the C]O band is diminished. In the porous carbon/Fe-Mn composite, the band intensity at 1719 cm−1 (C]O) and 1129cm-1 (CeO) are strengthened, and the band of asymmetric carboxyl −COO– stretching vibration shifts from 1571 to 1581 cm−1. Also, a new band appearing at 604 cm−1 indicates the presence of Mn–O–Mn groups. These changes could be explained by the interaction through oxygen-containing functional groups (C]O, −COO, CeO etc.) which bond the Fe-Mn oxide nanoparticles with porous carbon matrix [33]. Raman spectra are shown in Fig. S1(b). All three samples exhibit a characteristic D band at ˜ 1335 cm−1 and a G band at ˜ 1604 cm−1.The disordered D band arises from sp3 hybridization of CeC bonds, while the ordered G band is suggestive of sp2 C]C bond stretching vibration [34]. Compared with porous carbon, D band at 1354 cm−1 and G band at 1601 cm−1 of porous carbon/Fe-Mn show a slight frequency shift (D: + 9 cm−1, G: –3 cm−1), which confirm the interaction between the carbon matrix and the nanoparticles. The intensity ratio of the D to G
The TGA profiles of porous carbon, porous carbon/Fe and porous carbon/Fe-Mn composites are shown in Fig. 4(f). The weight loss below 120℃ corresponds to the loss of surface water; weight loss percentiles are 0.3 wt%, 1.4 wt%, and 2.7 wt% for porous carbon, porous carbon/ Fe and porous carbon/Fe-Mn, respectively. The largest weight loss occurs at 120–400℃ and above 400℃. The weight loss between 120℃ and 400℃ is induced by the pyrolysis of certain oxygen-containing functional groups, i.e. C]O, −COO, CeO, CeOeC, etc. The proportion of oxygen-containing functional groups is 1.1 wt% for porous carbon, 1.2 wt% for porous carbon/Fe and 6.4 wt% for porous carbon/Fe-Mn. The weight loss above 400℃ arises from the decomposition of carbon skeletons. Carbon content is found to be 98.6 wt%, 71.8 wt% and 37.4 wt% respectively for porous carbon, porous carbon/Fe and porous carbon/Fe-Mn. At the end of the heating process, no porous carbon remains, but porous carbon/Fe remains 25.6 wt% by weight, and porous carbon/Fe-Mn remains 53.5 wt%. Based on these data, we calculated that porous carbon/Fe contained ˜8.96 wt% Fe, and porous carbon/Fe-Mn contained 27.4 wt% Fe and 9 wt% Mn. The loss at 120–240℃ relates to the loss of labile oxygen-containing groups, and the loss at 240–400℃ may correspond to the pyrolysis of more stable oxygen groups such as C]O [31].
3.2. Composite analysis for chemical adsorption capability Chemical adsorption mainly involves complexation interaction and functional groups bonding contributed from the metal oxides in the 36
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Fig. 3. TEM images of (a–b) porous carbon and its composites: (c–d) porous carbon/Fe and (e–f) porous carbon/Fe-Mn.
MnOOH. Furthermore, after adsorbing Hg2+, the composite indicates the presence of Hg (II). Peaks at 100.8 eV and 104.6 eV are characteristic of Hg4f7/2 and Hg4f5/2 consistent with Hg–O bonds [23]. After Hg2+ adsorption, the peak of CeC/C]C increases likely due to π-π electrostatic interactions. The Mn2p peak shift may suggest that MnO2 and MnOOH strongly interact with Hg, as Hg can implant into the internal structure of manganese oxides through ligand exchange [36]. This further confirms that Hg (II) is adsorbed in an oxidized state.
band (ID/IG) was qualified as a measure of the degree of graphitization [35]. Porous carbon has an ID/IG ratio of 0.81. It increases to 0.94 for porous carbon/Fe and decreases to 0.89 for porous carbon/Fe-Mn. These ratios mean that the final composite still contain a graphene-like layer structure with a well-defined degree of order. When the characteristic peaks at 200–450 cm−1 are magnified in Fig. 5(c), new peaks at 220, 284.6 and 398.8 cm−1 are found for porous carbon/Fe, corresponding to Fe–O–Fe bonds which attribute to Fe2O3. For the porous carbon/Fe-Mn composite, these peaks respectively shift to lower wave numbers at 216.5, 279.4, and 369.6 cm−1, implying the coexistence of iron and manganese oxides and their interaction. Another emerging peak at 651.4 cm-1 is assigned to the Mn–O–Mn group and the Mn−OH bonds from MnO2. XPS spectra of porous carbon, porous carbon/Fe and porous carbon/ Fe-Mn composites before and after Hg2+ adsorption was studied to determine their chemical oxidation states and reaction mechanisms. The XPS total survey spectra are shown in Fig. S1(d)-(f). Fig. 5 shows the XPS spectra for C1 s, Fe2p, Mn2p, and Hg4f for the porous carbon/ Fe-Mn composite before and after uptake of Hg2+. The characteristic C1 s peaks at 284.7 eV, 286.3 eV and 288.5 eV can be well-indexed respectively to CeC/C]C, CeOeC and C]O, in agreement with FTIR results. For Fe2p, the binding energy peak at 710.5 eV is representative of Fe2p3/2 attributed to α–Fe2O3. Binding energy peaks at 712.8 eV and 724.5 eV are ascribed to Fe2p3/2 and Fe2p1/2 of FeOOH respectively. The weak peak at 719 eV pertains to Fe2p1 of zero-valent iron. For Mn2p, the peaks at 641.2 eV and 642.9 eV are assigned to Mn2p3/2 of MnO2 and MnOOH respectively. Hence, the metal oxides in the composite primarily exist in the forms of Fe2O3, FeOOH, MnO2 and
3.3. Heavy Hg2+ removal performance The Hg (II) removal performance by porous carbon/Fe-Mn composite was dependent on pH, adsorbent dosage and initial Hg concentration. The initial concentration prior to the treatment is defined as C0. The normalized removal concentration percentage, calculated by (C0–C)/C0, is described as removal efficiency. As shown in Fig. 6 (a), the removal efficiency increases from 82.1% to 98.4% when the carbon/Fe-Mn composite concentration changes from 10 to 90 mg L−1. The larger dosage of adsorbent would provide more adsorption sites and functional groups to combine with, resulting in apparent increments in Hg removal efficiency. However, when the adsorbent dosage exceeds ˜80 mg L−1, the removal efficiency ceases to increase likely because the adsorption reached an equilibrium. To optimize Hg uptake and make full use of the adsorbent, we used an optimal dosage 80 mg L−1 in the following experiments. The efficiency of Hg removal was also strongly affected by pH. In Fig. 6 (b), Hg removal efficiency increases with pH from 3 to 5 and then decreases sharply until pH reaches 11. The graph presents the highest 37
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Fig. 4. (a) Nitrogen adsorption/desorption isotherms, (b) pore size distribution (c) XRD pattern, (d) Zeta potential (e) Mercury speciation (Hg2+ 1 mg L−1, pH 1–9, molar Fe/Mn 3/1). (f) TGA curves of porous carbon and its composites: carbon/Fe and carbon/Fe-Mn.
performance of porous carbon three times than iron oxide. With the adsorbent dosage of 80 mg L−1 and initial Hg2+ concentration of 1.0 mg L−1, the removal efficiency of Hg2+ by porous carbon/Fe-Mn is as high as 96.8%. Under the same experimental conditions, the GO/FeMn composite synthesized by Tang et al. had a Hg removal efficiency of just 89.0% [37].
removal efficiency at pH 5–7. The pH not only affects the surface charge of adsorbent but influences the species of Hg. As demonstrated by the Zeta potential graph (Fig. 3) and the mercury speciation curve over different pH values, overly high or overly low pH values are unfavorable for Hg removal. Taking all factors into account, we determined the optimum pH value to be 6. At this pH value, Hg (II) exists as an oxidized state of Hg(OH)2. Therefore, it is inferred that metal oxides on the composite surface can interact with Hg(OH)2 to form complex species such as Fe2O3=Hg (II) or FeOOH=Hg (II) [37]. In brief, during this process some complexation interactions occurs for the surface functional groups between metal oxides and Hg(OH)2 at pH 6. The effect of initial Hg concentration on removal efficiency is trivial as shown in Fig. 6(c). Over the concentration range, the adsorbent offers abundant available active sites for Hg to anchor to. For convenience, batch experiments were performed at a fixed initial Hg concentration 1 mg L−1. The efficiency of Hg removal was measured as a function of contact time. In Fig. 6(d), the removal rates for all the samples increase sharply until at ˜150 min and then they increase slowly reaching their own equilibriums. The equilibration removal efficiency of Hg was 96.8% for porous carbon/Fe-Mn, compared to 85.0% for porous carbon/Fe and 81.1% for porous carbon. It is worthy to note that the adsorption performance of the porous carbon/Fe-Mn composite was much higher than that of porous carbon. Fe-Mn composite improves the adsorption
3.4. Mechanisms for high Hg2+ removal performance Porous carbon/Fe-Mn exhibits not only a fast adsorption rate but a high adsorption capacity. The Hg (II) amount adsorbed by unit mass of adsorbent at the equilibrium is defined as qe for measurement of the equilibrium adsorption capacity (mg g−1). The equilibration adsorption capacity of the porous carbon/Fe-Mn composite was calculated to be 9.8 mg g−1. In Fig. S2(a), adsorption kinetic tests examine the efficiency of Hg removal by the porous carbon/Fe-Mn composite. The removal rate increases significantly at the initial stage. During the first 30 min, Hg removal efficiency of 91.8% is reached as most accessible adsorption sites are occupied. After 3 h removal efficiency reaches a stable level, and 12 h later the adsorption sites are saturated. Data processing of the graph yields a maximum Hg removal efficiency of 96.8%. Accordingly, the commonly used pseudo-first order model and pseudo-second order model were adopted to simulate the kinetics data. The formulas are 38
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Fig. 5. XPS spectra of carbon/Fe-Mn for (a) C1 s, (b) Fe2p, (c) Mn2p and (d) Hg4f.
In these equations, qe and q correspond to the adsorption capacity respectively at an equilibrium and at time; K1 and K2 are the adsorption rate constants of pseudo-first order and pseudo-second order models; and R2 is the resultant fitting parameter. Apparently, the pseudo-second order model (R2 = 0.999) fits better. Hence, the rate limiting step was the adsorption process instead of the diffusion process [38], and the Hg removal mechanism was mainly ligand exchange and surface complexation between porous carbon/Fe-Mn and Hg. Adsorption isotherms were investigated to further identify the
expressed as Eqs. (1)–(2). Pseudo-first order model: ln(qe-q)= lnqe-K1t
(1) −1
Where qe = 9.8 (mg g ), K1 = 0.005 (g mg Pseudo-second order model:
−1 ⋅
min
−1
2
−1
2
) and R = 0.944
t/q = 1/(K2qe2)+ t/qe Where qe = 9.8 (mg g
(2)
−1
), K2 = 0.083 (g mg
−1 ⋅
min
) and R = 0.999
Fig. 6. (a) Hg removal with various adsorbent concentration (pH 6, Hg2+ 1 mg L−1, 24 h at 25 ± 2℃). (b) Hg removal with various pH (adsorbent 80 mg L-1, Hg2+ 1 mg L−1, 24 h at 25 ± 2℃). (c) Hg removal with various Hg2+ concentration (pH 6, adsorbent 80 mg L−1, 24 h at 25 ± 2℃). (d) Hg removal rate by different adsorbent materials (pH 6, adsorbent 80 mg L−1, Hg2+ 1 mg L−1, at 25 ± 2℃).
39
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pseudo-second order model, and the adsorption isotherms are consistent with the Freundlich model. Furthermore, the equilibration behavior is multilayer adsorption on heterogeneous surface. This research opens new avenues for the development of cost-effective, environmentally-friendly materials for removal of Hg ions and other heavy metal ions from wastewater.
adsorption mechanism and analyze how Hg interacted with porous carbon/Fe-Mn. The adsorption isotherm describes the amount of Hg adsorbed by unit mass of adsorbent at a constant temperature as a function of Hg concentration at equilibration. In Fig. S2 (b)-(e), Langmuir and Freundlich isotherm models are used to simulate the experimental data in table S1. The Langmuir isotherm model is applicable to monolayer adsorption on the homogeneous surface, with all binding sites equal and adsorption energy constant [39]. The Freundlich isotherm model describes multilayer adsorption on the heterogeneous surface, whose binding sites have different adsorption energies [40]. The linear forms of the two models are given as Eqs. (3)–(4). Langmuir isotherm model:
Ce 1 C = + e qe KL⋅qmax qmax
The authors gratefully acknowledge schematic diagram design work provide by Mr. Rong Jiang at School of Design of Shanghai Jiao Tong University and financial support for this research from the National Key R&D Program of China (2016YFA0202900, 2016YFC1402400), National Natural Science Foundation of China (51672173, 51801121), Shanghai Science and Technology committee (17JC1400700, 18ZR1421000, 18520744700), Science and Technology Planning Project of Guangdong Province (2016A010103018), Program of Guangdong Academy of Sciences (2017GDASCX-0503), Shanghai Research institute of criminal science and technology (2016XCWZK15). We also thank the Shanghai Synchrotron Radiation Facility for the measurements.
(3) −1
Where qmax = 51.8 (mg g ), KL = 9.280 (L mg Freundlich isotherm model:
ln qe = ln kF +
Acknowledgments
1 ⋅ln Ce n
−1
2
) and R = 0.502
(4) −1
2
Where 1/n = 0.898, Kf = 304.3 (L g ) and R = 0.989 In these equations, Ce is the equilibration concentration of Hg (mg L−1); qmax is the maximum adsorption capacity; KL and Kf are the Langmuir and Freundlich constants relating to the adsorption energy; and 1/n is the constant related to the adsorption intensity. The Freundlich isotherm model (R2 = 0.989) is more favorable, implying that the adsorption process was multilayer adsorption on the heterogeneous surface. In summary, the process of Hg removal covers both physical and chemical adsorption. On and near the surface of the adsorbent, the dominant mechanism would be chemical adsorption including ligand exchange and surface complexation. Further away from the surface, physical adsorption such as electrostatic attraction in the multilayer adsorption process would take a major role. The comparison of different adsorbents reported with our sample was summarized in Table S2. Activated carbon as adsorbent works through physical adsorption; Mn-Fe oxide and modified Fe oxide work through chemical adsorption; magnetic powered activated carbon (MPAC) and GO/Fe-Mn composite involve both physical and chemical adsorption. The results of our work are outstanding in both Hg(II) removal efficiency and adsorption capacity. The G-RHC/Fe-Mn composite possessed excellent Hg(II) removal efficiency of 96.8%, higher than those of reported activated carbons (82%), in-situ formed Mn-Fe oxides (80%), MPAC (96.3%) and GO/Fe-Mn composite (89%). The Hg(II) removal efficiency of modified Fe oxide is higher but its adsorption capacity is much lower than our sample. Our sample displayed a considerable Hg(II) adsorption capacity of 9.81 mg g−1. This value is much higher than those of reported activated carbons (0.87 mg g−1), Mn-Fe oxides (0.6 mg g−1), modified Fe oxide (0.59 mg g−1) and MPAC (0.096 mg g−1). The adsorption capacity of GO/Fe-Mn is higher but its removal efficiency is lower than our sample. The adsorption capacity of our G-RHC/Fe-Mn composite is far higher than that of the similar counterpart MPAC, owing to the synergetic effect of the active binary Fe-Mn oxides.
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4. Conclusion A novel hierarchical carbon/Fe-Mn composite readily fabricated from biomass was utilized as an adsorbent for Hg (II) removal. Since the composite contained Fe2O3 and MnO2 as dispersion phase which had oxygen-containing functional groups FeOOH and MnOOH, it was found to readily bond with mercury ions in wastewater through ligand exchange and surface complexation. This combined with the large surface area of 350.8 m2 g−1 creating a synergic effect on Hg ion removal. The composite exhibited fast adsorption rate, high removal efficiency of 96.8%, and considerable adsorption capacity of 9.8 mg g−1, superior to those reported adsorbents. The adsorption kinetics agrees with the 40
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