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Research Article pubs.acs.org/journal/ascecg

Construction of Layered Double Hydroxides/Hollow Carbon Microsphere Composites and Its Applications for Mutual Removal of Pb(II) and Humic Acid from Aqueous Solutions Shuyi Huang,† Shuang Song,† Rui Zhang,† Tao Wen,*,† Xiangxue Wang,† Shujun Yu,† Wencheng Song,† Tasawar Hayat,‡,§ Ahmed Alsaedi,‡ and Xiangke Wang*,†,‡,∥ †

College of Environmental Science and Engineering, North China Electric Power University, Beijing 102206, P. R. China NAAM Research Group, Faculty of Science, King Abdulaziz University, Jeddah 21589, Saudi Arabia § Department of Mathematics, Quaid-I-Azam University, Islamabad 44000, Pakistan ∥ Collaborative Innovation Center of Radiation Medicine of Jiangsu Higher Education Institutions, School for Radiological and Interdisciplinary Sciences, Soochow University, Suzhou 215123, P. R. China ‡

S Supporting Information *

ABSTRACT: Development of a highly efficient material with large specific area was scientifically and technologically important for simultaneous elimination of inorganic and organic pollutants from wastewater. In this work, layered-double-hydroxides-coated (LDHs-coated) hollow carbon microsphere composites (HCMSs; LDHs/HCMSs) were fabricated using carbon spheres as templates via a hydrothermal method and were then applied for mutual removal of Pb(II) and humic acid (HA) under various experimental conditions, i.e., pH, ionic strength, contact time, addition sequences, and coexisting ions. The results indicated that the copresence of Pb(II) and HA facilitated single target pollutant [Pb(II) or HA] adsorption at pH < 8.0. The adsorbed HA was responsible for the improvement of Pb(II) adsorption, primarily as a result of the formation of HA-Pb-LDHs/ HCMSs ternary complexes. Competitive adsorption of Cu(II), Co(II), Pb(II), and Ni(II) ions on LDHs/HCMSs was also investigated, and the LDHs/HCMSs had the affinity in the order of Co(II) < Ni(II) < Cu(II) < Pb(II). Interestingly, in binary/quaternary systems, the results showed that the presence of Cu(II), Co(II), and Ni(II) ions exhibited slight inhibition on Pb(II) adsorption. However, the total amounts of heavy metal ions adsorbed on LDHs/ HCMSs increased with the increase of the heavy metal ions. Results of X-ray photoelectron spectroscopy and Fourier transformed infrared spectroscopy indicated that outer-sphere surface complexation mainly dominated the adsorption of Pb(II) on LDHs/HCMSs, while the adsorption of HA was attributed to surface complexation of the disassociated HA on LDHs/ HCMSs. The findings highlighted the novel synthesis of LDHs/HCMSs and its potential application for simultaneous removal of different metal ions and natural organic contaminants from wastewater in environmental pollution cleanup. KEYWORDS: LDHs/HCMSs, Humic acid, Pb(II), Simultaneous removal, Surface complexation

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INTRODUCTION The treatment of wastewater contaminated by toxic heavy metal ions and natural organic pollutants has become one of the most onerous challenges for the sustainable development of human society because of their detrimental effects on environmental protection and human health. Among the toxic heavy metal ions, Pb(II) ions, arising from metal processing, leaded gasoline, lead-based paint, and battery manufacturing industries, might cause detrimental effects on the intellectual development of the human central nervous function and semipermanent brain.1 Because of the important biological activity and its adverse influence on human health even at low concentration, Pb(II) ions should be removed from aqueous solutions. At the same time, Pb(II) ions inevitably coexist with natural organic matters in aqueous solutions.2−4 For natural © 2017 American Chemical Society

organic pollutants, humic acid (HA), a subclass of humic substances which widely exist in nature and are not easily decomposed by microorganisms, has a strong influence on the physicochemical behavior of metal ions in natural aqueous systems. Various functional groups (e.g., amine, phenol, hydroxyl, and carboxylic groups) of HA could provide considerable sites for binding metal ions and organic molecules, resulting in significant physicochemical behavior changes of the contaminants in the environment. Moreover, the HA molecules can react with chlorine and produce some hazardous substances such as trihalomethanes and haloacetic acids during the process Received: May 31, 2017 Revised: October 9, 2017 Published: October 20, 2017 11268

DOI: 10.1021/acssuschemeng.7b01717 ACS Sustainable Chem. Eng. 2017, 5, 11268−11279

Research Article

ACS Sustainable Chemistry & Engineering

Figure 1. Schematic illustration of the approach used for the synthesis of LDHs/HCMSs. SEM images of (b) as-prepared bare carbon spheres and (c) LDHs/HCMSs, and the corresponding TEM images of (d) bare carbon spheres and (e) LDHs/HCMSs. (f) Elemental mapping of C, O, Al, Mg, and Pb of Pb(II)-laden LDHs/HCMSs.

LDHs are generally represented by the formula [M2+1−xN3+x(HO−)2]x+ [(Xn−)x/n yH2O]x−, where the M2+ and N3+ represent divalent metal cations (Ca2+, Mg2+, Co2+, Cu2+, Ni2+, Zn2+, Mn2+) and trivalent metal cations (Al3+, Co3+, Fe3+, Cr3+, Mn3+), respectively, and Xn− is the intercalating anions. LDHs have large interlayer space with considerable anions for ion-exchange, and the positively charged layers would make them excellent adsorbents for efficient removal of negatively charged contaminants or the ion-exchange of cations through the substitution with M2+ ions in LDHs.14−16 Hence, LDHs were regarded as a kind of essential adsorbent for cations/anions and organic compounds. Meanwhile, the abundant functional groups of carbon-based nanomaterials are useful for the binding of toxic organic pollutions and heavy metal ions.17−19 Considering the above-mentioned properties, one can introduce LDHs to a carbon microsphere surface to construct the calcined LDHs hollow carbon microsphere hierarchical nanostructure materials with large surface area and stable hollow structure via an environmentally friendly approach (Figure 1a). The calcined LDHs/hollow carbon microsphere composites (CLDHs/HCMSs) can be converted to LDHs/hollow carbon microsphere composites (LDHs/ HCMSs) with the property known as “memory effects”,20 and then, the LDHs/HCMSs can be applied as the efficient

of the chlorination of drinking water, which might cause a seriously harmful influence on human health.5 Therefore, it is of great importance to develop advanced technologies and materials to attenuate Pb(II) and HA concentrations in the natural water. A variety of technologies have been applied to minimize the amount of Pb(II) and HA in aqueous solutions, such as reverse osmosis,6 biosorption,7 electrochemistry,8 photodegradation,9 chemical coagulation,10 ion-exchange,11 and adsorption.12 Nevertheless, most methods have been found to be limited by high investment, poor efficiency, and secondary pollution and are thereby greatly restricted in their application in environmental pollution remediation. For example, ionexchange exhibited high efficiency, but it was limited in practical applications because of the high cost required in the regenerative process. Compared to these methods, the adsorption technique has the advantages of trouble-free operation, low cost, and wide adaptability. Therefore, the adsorption technique was regarded as one of the most important considerable technologies in industrial applications and has been widely used in the enrichment of different kinds of pollutants from large volumes of aqueous solutions. Layered double hydroxides (LDHs), a new class of inorganic layered materials, have received multidisciplinary attention because of their unique physicochemical properties.13 The 11269


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Figure 2. EDX of (a) LDHs/HCMSs and (b) Pb(II)-laden LDHs/HCMSs. (c) FTIR spectra of the samples before and after Pb(II) and HA adsorption. (d) TG curve of the LDHs/HCMSs. (e) Nitrogen adsorption−desorption isotherms and (f) pore size distribution of LDHs/HCMSs. acid (HCl) were purchased from Sinopharm Chemical Reagent Beijing Cl., Ltd. (Beijing, China). The humic acid (HA) sample was obtained from Heilongjiang Province, China, and the HA constituents were approximately 0.50% S, 3.53% H, 4.22% N, 31.31% O, and 60.44% C. Synthesis of LDHs/HCMSs. The carbon spheres were first synthesized through a traditional hydrothermal method.21 Briefly, 21 g of glucose was dissolved in 70 mL of deionized water; then, the above solution was hydrothermally treated at 180 °C for 5 h. After the reaction, the black products were rinsed using ethanol and deionized water several times and collected by centrifugation. Ultimately, the powders were gathered after vacuum drying at 60 °C for 12 h. The hollow carbon microsphere coated with Mg-Al-LDHs was obtained through an alkaline hydrothermal treatment using carbon microspheres as the template. In a typical synthesis, 0.36 g of carbon spheres was dispersed in 70 mL of Al(NO3)3·9H2O (0.37 g) and Mg(NO3)2·6H2O (0.50 g) solution by stirring (IKA: C-MAG HS7, 500 rpm) for 12 h to ensure that Mg2+ and Al3+ can substantially contact the carbon microspheres. Then, 1.25 g of urea was stirred with this suspension for 1 h. Then, the black homogeneous suspension was transferred to a Teflon-lined autoclave and maintained at 90 °C for 4 h, followed by filtrating and rinsing with Milli-Q water and ethanol 3

adsorbents for the elimination of Pb(II) and HA from wastewater. This research contributed to the understanding of HA and Pb(II) uptake to LDHs/HCMSs. The physicochemical properties of LDHs/HCMSs were carefully characterized by Fourier transformed infrared (FTIR) spectroscopy, thermogravimetric (TG) analysis, X-ray photoelectron spectroscopy (XPS), Brunauer−Emmett−Teller (BET), and ζ potential analysis. The mutual effects of Pb(II) and HA on LDHs/HCMSs were systemically investigated. Potential adsorption of coexisting heavy metal ions (i.e., Cu2+, Ni2+, Co2+, Pb2+) was also studied to have an insight into the interaction of HA and Pb(II) on LDHs/HCMSs.

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EXPERIMENTAL SECTION

Materials and Chemicals. All reagents used in this experiment were purchased in analytical grade. Glucose (C6H12O6·H2O), magnesium nitrate hexahydrate [Mg(NO3)2·6H2O], urea, aluminum nitrate nonahydrate [Al(NO3)3·9H2O], sodium hydroxide (NaOH), chloroacetic acid (C2H3ClO2), nitric acid (HNO3), and hydrochloric 11270



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times. At last, Mg-Al-LDHs/carbon microsphere composites (Mg-AlLDHs/CMSs) were obtained after drying at 60 °C for 12 h. The MgAl-LDHs/CMSs were heated to 400 °C at the heating rate of 2 °C min−1 and then annealed at 400 °C for 4 h to obtain CLDHs/HCMSs. The schedule for the synthesis of CLDHs/HCMSs is described in Figure 1a. The abundant adsorption sites (e.g., OH and COOH groups) of carbonaceous microspheres would be responsible for coordination with Mg2+ and Al3+. Characterization. For detection of the interior structure and the change of elements, transmission electron microscopy (TEM) and energy-dispersive X-ray (EDX) spectroscopy elemental mapping were acquired on an FEI Tecnai G2 F20 electron microscope at 200 kV. The scanning electron microscopy (SEM) technique was applied to investigate the surface morphology using a Hitachi S4800 microscope. The X-ray photon spectroscopy (XPS) spectra were collected on the ESCALAB 250 Xi XPS of Thermo Scientific to achieve the detailed information on the adsorption mechanism. The ζ potential value was measured on a Zetasizer (Nano-ZS) instrument from Malvern Instruments. The thermogravimetric (TG) measurement was performed using a TG-50 thermal analyzer under air at a heating rate of 10 °C min−1. The Fourier transformed infrared (FTIR) spectra were collected on the Nicolet Magana-IR 750 spectrometer over the range 400−4000 cm−1. The N2−Brunauer−Emmett−Teller (BET) specific surface area test was measured using the Micromeritics ASAP 2010 instrument at 77 K. It is worth noting that the samples were heated to 80 °C and maintained at the temperature of 80 °C for 24 h before each measurement. Batch Adsorption Experiments. The batch adsorption measurements were conducted at 298 K and pH 5.0 ± 0.05 in polyethylene tubes. A 0.06 g portion of LDHs/HCMSs powders was dispersed into 100 mL of Milli-Q water, and a 0.6 g L−1 stock adsorbent concentration was achieved. For Pb(II) and HA adsorption, the LDHs/HCMSs stock solution, the NaNO3 background electrolyte solution, Milli-Q water, and the Pb(II) (200 mg L−1) or HA (400 mg L−1) solutions were added into the polyethylene tubes in sequence to obtain the desired different component concentrations. The desired HA and Pb(II) solutions were added into the polyethylene tubes simultaneously when the mutual removal of HA and Pb(II) on the LDHs/HCMSs was investigated. After the pH value of the suspensions in each tube was adjusted by adding 0.1 and/or 0.01 M HNO3 and/or NaOH, the suspensions were vigorously stirred at 298 K for 24 h. After that, the above suspensions were centrifuged at 10 000 rpm for 30 min to separate the solid and liquid phases. The HA concentration in the supernatant was measured by spectrophotometry at the wavelength of 256 nm. The adsorption percentage and the adsorption capacity of Pb(II) and HA on solid phases were calculated from eqs 1 and 2:

adsorption =

qe =

(C0 − Ce) × 100% C0

V (C0 − Ce) m

Research Article

RESULTS AND DISCUSSION

Characterization of LDHs/HCMSs. The morphologies and microstructures of as-prepared carbon microspheres and LDHs/HCMSs were measured by SEM and TEM images. From Figure 1b,d, the uniform carbon microspheres showed smooth surfaces with the average size of approximately 150 nm. After being coated with LDHs and pyrolysis of the composites, the LDHs/HCMSs displayed an interconnected 3D framework with many junctions and exhibited a rough surface, suggesting that LDHs were coated on the surface of carbon spheres successfully (Figure 1c,e). As anticipated, the TEM image further confirmed the formation of the hollow structure. The EDX element mapping (Figure 1f) showed the homogeneous distribution of C, O, Mg, Al, and Pb elements in the LDHs/ HCMSs. After Pb(II) adsorption, the LDHs/HCMSs kept the original morphologies. The EDX spectrum (Figure 2a,b) also confirmed the existence of the aforementioned elements, indicating the homogeneous distributions of LDHs coating on the surfaces of hollow carbon. The FTIR spectra of LDHs/HCMSs samples before and after HA/Pb(II) adsorption were shown in Figure 2c. The intense broad peak at 3440 cm−1 resulted from the stretching vibration of structural O−H groups in association with the interlayer H2O molecules, LDH layers hydrogen bonding, and carboxyl groups.20 The adsorption band at 1423 cm−1 was attributed to the carbonate anion stretching vibrations, indicating that some CO 3 2− ions existed in LDHs/ HCMSs.22,23 The peaks in the range 400−800 cm−1 were assigned to the lattice vibrations of O−M−O, M−O−M, and M−O lattice vibrations (M refers to Mg and Al).24 These peaks confirmed the successful modification of carbon spheres by MgAl-LDHs. Thermogravimetric (TG) analysis was employed to estimate the thermal stability of LDHs/HCMSs. As illustrated in Figure 2d, the thermal evolution of LDHs/HCMSs underwent three weight loss steps as follows: The first weight loss (∼7 wt %) from room temperature to 150 °C was attributed to the loss of the surface-adsorbed H2O. Then, the second weight loss step (∼8 wt %) in the region 150−380 °C arose from the decomposition of brucite-like layers, intercalated CO32− anions, and oxygen-containing groups.23 Finally, the third weight loss (∼8 wt %) in the region 400−550 °C was assigned to the destruction of the carbon skeleton. This was in accordance with the result of the EDX spectrum. Thus, one can speculate that the carbon skeleton in LDHs/HCMSs was not effectively removed in the synthesis process and maintained the hierarchical hollow structure. The N2−BET specific area of the LDHs/HCMSs was examined via N2 adsorption−desorption measurements. As plotted in Figure 2e, the adsorption isotherm can be classified as the mixed characteristics of types II and IV, and the hysteresis loop was found in the range 0.8−1.0 P/P0, implying the coexistence of nanoporous, mesoporous, and macroporous structures in LDHs/HCMSs, which was further supported by the distribution of pore sizes (Figure 2f). The LDHs/HCMSs displayed a narrow pore size distribution centered at the diameter of 2.7 nm. Furthermore, the Brunauer−Emmett− Teller (BET) surface area of the LDHs/HCMSs was calculated to be 231 m2 g−1. The large surface area and abundant functional groups made the LDHs/HCMSs an effectual material for the preconcentration of HA or Pb(II) ions from a large volume of solutions.

(1)

(2)

where Ce (mg L−1) and C0 (mg L−1) were the equilibrium concentration after adsorption and the initial concentration, respectively, m (mg) was the dosage of LDHs/HCMSs, and V (mL) was the total volume of the suspension. The effect of the addition sequences on HA and Pb(II) adsorption to LDHs/HCMSs was further investigated. Specifically, the addition sequences of HA and Pb(II) were described as follows: (1) In the (LDHs/HCMSs-Pb)-HA system, LDHs/HCMSs, Pb(II), and NaNO3 were equilibrated for 24 h before contacting HA. (2) In (LDHs/ HCMSs-HA)-Pb system, LDHs/HCMSs, HA, and NaNO3 were equilibrated for 24 h before contacting Pb(II). The batch experiments were performed in duplicate to observe the reproducibility, and all experimental results were the average value of triplicate measurement within relative errors of ∼5%. 11271


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Figure 3. (a) XPS survey of LDHs/HCMSs before and after Pb(II) and HA adsorption. XPS high-resolution spectra of (b) Pb 4f, (c) C 1s, (d) O 1s, (e) Mg 1s, and (d) Al 2p for LDHs/HCMSs before and after Pb(II) and HA adsorption.

531.1, and 529.9 eV, which represented the binding energies of the different oxygen form in OH−, O−, and chemical or physical adsorbed H2O, respectively.26 More detailed discussion of the XPS spectra of LDHs/HCMSs after Pb(II)/HA adsorption is described in the following adsorption mechanism section. Effect of pH. The influence of solution pH on Pb(II) and HA adsorption to LDHs/HCMSs is presented in Figure 4a. One can see that the adsorption of Pb(II) was deeply dependent on pH and reached the maximum amount at pH ∼ 6.5. With the increase of pH from 3.0 to 6.0, the adsorption efficiency of Pb(II) on LDHs/HCMSs increased dramatically from ∼10% to ∼75%, maintained the high level in the pH range 6.0−8.0, and then decreased with the increase of pH (pH > 8.0). The variation of pH can either promote or suppress the efficiency of Pb(II) adsorption on LDHs/HCMSs, which may be attributed to the relative distribution of Pb(II) species in solution and the surface properties of LDHs/HCMSs. As shown in Figure 4c, the zero point of charge (pHzpc) of LDHs/

The XPS spectra of LDHs/HCMSs before and after HA and Pb(II) adsorption are presented in Figure 3. In the survey spectra (Figure 3a), four apparent peaks of C 1s, O 1s, Mg 1s, and Al 2p illustrated that carbon, oxygen, magnesium, and aluminum were the predominant elements of the LDHs/ HCMSs.12 For comparison, the Pb(II) peak was found in the survey spectrum of Pb(II)-laden LDHs/HCMSs. The highly resolved C 1s spectrum of LDHs/HCMSs (Figure 3c) can be deconvoluted into three components, i.e., the carboxylate carbon (OCO, 289.2 eV), the carbon in CO (285.8 eV), and nonoxygenated C (284.5 eV).25 The specific peak at 289.2 eV showed a wider peak area rather than other peaks, indicating that the dominated oxygen-containing functional group was OCO. The carboxylate carbon signal further implied the presence of intercalated CO32− in the layer of LDHs/HCMSs, which was also confirmed by the FTIR result. It was interesting that the O 1s spectrum of LDHs/HCMSs (Figure 3d) could be deconvoluted into three peaks at 532.5, 11272


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Figure 4. Effects of pH on the (a) Pb(II) and (b) HA adsorption on LDHs/HCMSs. T = 298 K; m/V = 0.1 g L−1; C[Pb(II)]initial = 20 mg L−1; C(HA)initial = 20 mg L−1; and I = 0.01 M NaNO3. (c) ζ potential of LDHs/HCMSs. (d) Distribution of Pb(II) species in aqueous solution in the presence of 0.01 M NaNO3.

HCMSs was about 9.6, meaning that the LDHs/HCMSs’ surface charge was positive at pH < 9.6 and remained negative at pH > 9.6. The surface of LDHs/HCMSs contained abundant functional groups (i.e., hydroxyl and carboxylic groups) which could be protonated at low pH values ([SOH + H+ ⇔ SOH2+]) or deprotonated at high pH values ([SOH ⇔ SO− + H+]) (SOH represents the surface functional groups). From the hydrolysis constants of Pb(II) {i.e.; log K1 = 6.48; log K2 = 11.16; log K3 = 14.16 (Kn = [Pb(OH)n2−n]·[H+]/[Pb2+])}, the Pb(II) ions presented the species of Pb2+, Pb(OH)+, Pb(OH)3−, and Pb3(OH)42+ at variable pH values, and the relative distribution of Pb(II) species was computed by Visual MINTEQ version 3.0 software (Figure 4d).12 Pb2+ was the dominant species in solution at pH < 6.0. At low pH values, the low sorption of Pb2+ was attributed to the strong electrostatic repulsion between Pb2+ ions and the positive surface charge of adsorbents. With the increase of solution pH, the electrostatic repulsion between Pb2+ ions and LDHs/HCMSs became weak, and thereby the adsorption of Pb2+ increased with solution pH increasing. The main species of Pb(II) were Pb2+, Pb(OH)+, and Pb3(OH)42+ in the pH range 6.0−9.6. The surface charge of LDHs/HCMSs was still positive at pH < 9.6, which suggested that the high removal efficiency of Pb(II) was primarily ascribed to strong surface complexation or the ionexchange with Mg2+ ions in LDHs/HCMSs. As the solution pH further increased, Pb(OH)3− progressively became the main species, and it was constrained to be adsorbed on the negative LDHs/HCMSs because of the electrostatic repulsion. The surface became more negatively charged with the increase of solution pH, which resulted in the decrease of the negative Pb(II) species adsorption on the negatively charged LDHs/ HCMSs because the electrostatic repulsion became stronger at

higher pH values. At the same time, the competition between Pb(OH)3− and OH− also gave rise to the dramatic decrease of Pb(II) adsorption on LDHs/HCMSs. From the batch adsorption results, one can see that the electrostatic repulsion was presented in the whole pH range. However, the LDHs/ HCMSs still exhibited high adsorption ability in the pH range 5−9, suggesting that the elimination of Pb(II) was mostly related to strong surface complexation or the substitution of Mg2+ in LDHs/HCMSs. From Figure 2b, one can see that the atomic ratio of Mg/Al after Pb(II) adsorption was about 0.33, which was lower than the ratio of Mg/Al (1.72) in LDHs/ HCMSs.27 The decrease of the Mg/Al ratio suggested that Mg2+ participated in the ion-exchange of Pb(II) in the adsorption process. The pH-dependent adsorption of HA on LDHs/HCMSs was also investigated. From Figure 4b, it was found that the adsorption percentage of HA on LDHs/HCMSs decreased from ∼96% to ∼30% with the increase of pH from 3.0 to 11.0. The pHzpc of HA was ∼2.0, suggesting that the charge of HA molecules was negative at pH > 2.0.28 Therefore, high adsorption of HA on LDHs/HCMSs was attributed to the electrostatic attraction between the positively charged LDHs/ HCMSs and negatively charged HA. Generally, large HA molecules probably “huddle up” at low pH values, and the HA molecules have a smaller volume because of the low charge development, meaning more binding sites on the surface of LDHs/HCMSs can interact with HA molecules and form more surface complexes,29 which also results in the high adsorption of HA at low pH values. However, HA contained a great amount of carboxyl and phenolic groups, suggesting that HA became negative at high solution pH because of the deprotonation reactions of oxygen-containing functional groups. The adsorption of HA on LDHs/HCMSs was a 11273


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Figure 5. Adsorption of (a) Pb(II) and (b) HA on LDHs/HCMSs as a function of initial pH and different ionic strength. T = 298 K; m/V = 0.1 g L−1; C[Pb(II)]initial = 20.0 mg L−1; and C(HA)initial = 20.0 mg L−1.

Figure 6. Adsorption isotherms of (a) Pb(II) and (b) HA, and the mutual effects of Pb(II) and HA adsorption on LDHs/HCMSs. pH = 5.0; T = 298 K; m/V = 0.1 g L−1. Symbols denote experimental data. The solid lines represent the Langmuir model simulation, and the dashed lines represent the Freundlich model. C[Pb(II)]initial = 20.0 mg L−1; and C(HA)initial = 20.0 mg L−1.

repulsion.31 Consequently, the competitive adsorption of Pb(II) with the dissolved HA on LDHs/HCMSs and the electrostatic repulsion among Pb(OH)3−, HA molecules, and LDHs/HCMSs resulted in the quick decrease of adsorption of Pb(II) at the high solution pH condition.32 Effect of Ionic Strength. The removal percentages of Pb(II) on LDHs/HCMSs as a function of solution pH under different NaNO3 concentrations were given in Figure 5a. The removal of Pb(II) increased with ionic strength decreasing from 0.1 to 0.001 M. Generally, the ionic strength can affect Pb(II) adsorption on LDHs/HCMSs by influencing the interface potential and double-layer thickness. The ionic strengthdependent adsorption showed that ion-exchange or outersphere surface complexation contributed mainly to the adsorption of Pb(II) on LDHs/HCMSs.33 This phenomenon might be ascribed to three possible reasons: (1) The Pb(II) ions combined with more H2O molecules at low ionic strength rather than at high ionic strength. The high ionic strength could reduce the activity coefficient of Pb(II) ions in aqueous solutions and restrict their transfer from solution to the surface of LDHs/HCMSs. (2) Pb(II) ions interacted with LDHs/ HCMSs by electrostatic reaction and formed electrical doublelayer complexes with LDHs/HCMSs. In other words, the concentration of NaNO3 affected the ion-exchange of Pb2+ with Mg2+ on the surface of LDHs/HCMSs. (3) Ionic strength affected particle aggregation through influencing the electrostatic interactions. In other words, high ionic strength can reduce the electrostatic repulsion and then increase the particle

combined effect of electrostatic attraction between LDHs/ HCMSs and HA molecules and electrostatic repulsion between HA molecules themselves, which led to the gradually expanding volume of HA molecules with pH increasing. Thus, both the increasing solubility of HA molecules and the negative surface charge of LDHs/HCMSs at high pH were responsible for the rapid decrease of HA removal due to the electrostatic repulsion. Figure 4a,b also shows the mutual effect of Pb(II) and HA adsorption to LDHs/HCMSs. It was found that Pb(II) adsorption on LDHs/HCMSs in the presence of HA was improved at pH < 9 and achieved the maximum adsorption at pH ∼ 5.5, which was lower than the pH value in the absence of HA. Pb2+ ions can not only be substituted by Mg2+ in LDHs/ HCMSs but also form strong complexes with surface-adsorbed HA at low pH values, resulting in the formation of Pb-HALDHs/HCMSs ternary complexes. The surface-adsorbed HA molecules can serve as a “bridge” between Pb(II) and LDHs/ HCMSs.30 At pH < 5.5, the surface of the HA molecule became more negative with increasing pH, indicated the easier formation of strong complexes with pH increasing due to the increasing electrostatic attraction of the HA molecule and Pb2+. However, Pb(OH)+ progressively became the main species. At pH > 5.5, the adsorption percentage of Pb(II) decreased with the decreasing electrostatic attraction of Pb(II) and HA, but was still higher than it was in the absence of HA. At pH > 9, Pb(II) was mainly present as negatively charged Pb(OH)3− species in solution, of which it was hard to form complexes with the negatively charged HA molecules because of electrostatic 11274


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Table 1. Parameters Calculated from the Langmuir and Freundlich Models for Pb(II) and HA Adsorption on LDHs/HCMSs Langmuir Pb Pb + HA HA HA + Pb

Freundlich

qmax (mg g−1)

KL (L mg−1)

RL 2

KF (mg1−n Ln g−1)

n

R2

205.68 291.46 300.46 467.54

1.94 1.04 1.77 1.16

0.986 0.981 0.942 0.968

122.69 138.54 203.18 251.92

6.460 4.320 8.966 5.372

0.910 0.919 0.909 0.853

5.0 and 298 K, which was also significantly higher than that of HA (qmax = 300.46 mg g−1) without Pb(II). The results demonstrated that the surface-adsorbed Pb(II) enhanced HA adsorption through the formation of Pb(II)-HA complexes on LDHs/HCMSs. The adsorbed HA molecules on the solid surface could also act as a “bridge” between Pb(II) and LDHs/ HCMSs and thereby enhanced Pb(II) adsorption to LDHs/ HCMSs. The results were crucial for the real applications of LDHs/HCMSs in the simultaneous elimination of HA and Pb(II) from the natural water system. In the natural environment, the copresence of heavy metal ions and organic pollutants could be simultaneously removed through the formation of strong surface complexes on LDHs/HCMSs. Moreover, the comparison of adsorption capacities of LDHs/ HCMSs with LDHs-based materials or carbon materials is shown in Table 2.23,29,35−39 It can be observed that the LDHs/ HCMSs with high surface area had remarkable adsorption capability for both Pb(II) and HA.

aggregation, which resulted in the decrease of the available binding sites, and then decreased the adsorption performance of Pb(II) ions on LDHs/HCMSs.34 Figure 5b shows that the adsorption efficiency of HA to LDHs/HCMSs obviously decreased with the increase of the ionic strength. It can be observed that nitrate ions affected HA adsorption by competing with HA molecules to be attracted by positively charged LDHs/HCMSs. Thereby, the main adsorption mechanism of HA adsorption might be the formation of the surface complexation between HA and LDHs/HCMSs. Considering the property that HA has a tendency to precipitation in the presence of metal ions and at high ionic strength, blank experiments in the absence of adsorbent are shown in Figure S1. The precipitation percentages of HA at pH 5.0 were about 5.2%, 10.2%, and 15.3% at 0.001, 0.01, and 0.1 M NaNO3, respectively, which were far lower than the elimination percentages of HA in the presence of LDHs/HCMSs, indicating that the precipitation of HA hardly affected the adsorption experiments, and the precipitated HA might be removed from solution by the formation of LDHs/HCMSs-PbHA ternary complexes. Adsorption Isotherms and Effect of Addition Sequences. Figure 6 shows the adsorption isotherms of Pb(II) ions and HA on LDHs/HCMSs. All the isotherms displayed a tendency that qe (mg g−1) increased rapidly at low initial Pb(II) or HA concentrations and then increased gradually at high-level Pb(II) or HA loading conditions. With the purpose of understanding the interaction mechanism, the Langmuir model and Freundlich model were employed to simulate the experiment data. The Langmuir isotherm model is expressed in eq 3: qe =

Table 2. Comparison of the Adsorption Capacities of LDHsBased Materials or Carbon Materials qmax (mg g−1) adsorbent LDHs LDHs−Fe3O4 multiwalled carbon nanotubes MWCNTs/PAAM LDHs/HCMSs

KLqmax Ce 1 + KLCe

magnetite-GO-LDHs MWCNTs takovite-aluminosilicate nanocomposite oxidized MWCNTs

(3)

The Freundlich isotherm model is described according to eq 4: qe = KFCe1/ n

(4)

surface area (m2 g−1)

Pb(II)

22 70 86 90.95 231

29.71 205.68

74.9 260.99 256

173 27.3 205

197

4.1

HA

ref

225.6 353.82 82

35 35 36

12.52 300.46

29 this work 23 37 38 39

Various studies gave the effect of addition sequences of heavy metal ions and natural organic materials. The adsorptions of Pb(II) and HA in the LDHs/HCMSs-Pb-HA ternary system for different Pb(II)/HA addition sequences were further investigated. As shown in Figure 6, one can see that the addition sequences made no difference on the adsorption of mutual removal of HA and Pb(II) in (LDHs/HCMSs-Pb)-HA and LDHs/HCMSs-Pb-HA systems, which was consistent with the previous study.40 However, the qmax value of Pb(II) adsorption was slightly decreased in the (LDHs/HCMSs-HA)Pb system, indicating that the complexation between HA and LDHs/HCMSs occupied the adsorption and complexation sites of HA and LDHs/HCMSs and thus decreased the amount of effective sites for Pb2+ adsorption. Adsorption Kinetics of Pb(II) and HA. The kinetic adsorption data of Pb(II) and HA on LDHs/HCMSs was

−1

where Ce (mg L ) represented the equilibrium concentration of Pb(II) or HA in aqueous solutions, qmax (mg g−1) was the amount of Pb(II) or HA adsorbed per weight unit of LDHs/ HCMSs after adsorption equilibration, KL (L mg−1) was a Langmuir constant related to the heat of adsorption, and KF (mg1−n Ln g−1) was the Freundlich constant, representing saturation adsorption capability of the adsorbent. The corresponding parameters were tabulated in Table 1. From the correlation coefficient (R2) values, it was clear that the Langmuir model fitted the adsorption isotherms better than the Freundlich model. From the Langmuir model simulation, the qmax value of Pb(II) was 291.46 mg g−1 in the presence of HA, which was much higher than that of Pb(II) (qmax = 205.68 mg g−1) in the absence of HA at pH 5.0 and 298 K. The qmax value of HA was 467.54 mg g−1 in the presence of Pb(II) at pH 11275


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ACS Sustainable Chemistry & Engineering Table 3. Kinetic Modeling of Pb(II) and HA Adsorption on LDHs/HCMSs pseudo-first-order model Pb Pb + HA HA HA + Pb

pseudo-second-order model

k1 (g mg−1 min−1)

qe (mg g−1)

R2

k2 (g mg−1 min−1)

qe (mg g−1)

R2

0.167 0.172 0.566 0.287

125.75 175.79 181.23 192.96

0.998 0.991 0.936 0.998

0.0046 0.0025 0.0258 0.0119

128.21 182.15 182.15 193.80

0.999 0.999 0.999 0.999

Figure 7. Adsorption kinetic curves of (a) Pb(II) and (c) HA on LDHs/HCMSs, and the linear fit of experimental data of (b) Pb(II) and (d) HA obtained using the pseudo-second-order kinetic model with or without the presence of HA or Pb(II). T = 298 K; m/V = 0.1 g L−1; I = 0.01 M NaNO3; pH = 5.0; C[Pb(II)]initial = 20.0 mg L−1; and C(HA)initial = 20.0 mg L−1.

to the abundant active sites and functional groups on LDHs/ HCMSs at the initial reaction time, which were available for the binding of Pb(II) and HA.24 It was apparent that the adsorption of Pb(II) in the presence of HA was rather faster than that of Pb(II) in the absence of HA, whereas little difference was found for the adsorption kinetics of HA in the presence or absence of Pb(II), although the adsorption percentage of HA coexisted with Pb(II) was a little higher than that of HA in the absence of Pb(II) (Figure 7c). Clearly, the R2 values of the pseudo-second-order model (Figure 7b,d) were >0.999, which were higher than those of the pseudo-firstorder model, indicating that the pseudo-second-order model could better describe kinetic adsorption processes. This phenomenon demonstrated that the rate-limiting step might be the chemisorption process involving valence forces through exchanging or sharing electrons between the target pollution and adsorbents.41 Consequently, this implied that the main interaction of Pb(II) and HA with LDHs/HCMSs was dominated by strong surface complexes or chemisorption rather than by mass transport.42 Adsorption Mechanisms. For elucidation of the interaction mechanism of Pb(II)/HA with LDHs/HCMSs, the FTIR spectra of LDHs/HCMSs after Pb(II) or HA adsorption

described by the pseudo-second-order and pseudo-first-order models. The pseudo-first-order model was presented as the following equation: ln(qe − qt ) = ln qe − k1t

(5)

The pseudo-second-order model equation was expressed as t 1 t = + qt qe k 2qe2

(6)

−1

where qe (mg g ) was the amount of HA or Pb(II) adsorbed at equilibrium time, qt (mg g−1) was the amount of HA or Pb(II) adsorbed on LDHs/HCMSs at time t (min), k1 and k2 (mg g−1 min−1) were the pseudo-first-order model rate and the pseudosecond-order rate, respectively. From the linear plots of ln(qe − qt) versus t and t/qt versus t, the values of k1, k2, and estimated qe can be obtained, and the parameters are tabulated in Table 3. Figure 7 shows the kinetic adsorption curves of Pb(II) and HA on LDHs/HCMSs. The adsorption of both target pollutants increased quickly within the first contact time of 30 min, and then achieved adsorption equilibrium after 60 min (Figure 7a,c). This adsorption tendency was probably ascribed 11276


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Jobbagy and Regazzoni47 reported that, at pH 5−9, almost 20% of the magnesium ions were leached out from Mg-AlLDHs, while the dissolution of the aluminum ion was insignificant. The dissociation equations of Mg2+ and Al3+ from LDHs/HCMSs is expressed as

were measured (Figure 2c). After Pb(II) adsorption, the characteristic peak at 3440 cm−1 was slightly increased, suggesting that the adsorption induced the increase of surface OH groups, which might arise from the absorption of Pb(II) hydroxide on the assembly. Notably, the peak shift from 1423 to 1379 cm−1 after Pb(II) adsorption might be ascribed to the precipitation of Pb3(CO3)2(OH)2,43 which was due to the contact between Pb2+ in solution and intercalated CO32− anions in LDHs. For the HA adsorption, an intense adsorption band at ∼1585 cm−1, which was attributed to the aromatic CC vibrations or hydrogen bonded CO of quinones in HA molecules, appeared in the FTIR spectrum of HA-laden LDHs/ HCMSs, suggesting the successful adsorption of HA molecules on the surface of LDHs/HCMSs.31 This showed that the intensity of the band at 1423 cm−1 after HA adsorption became significantly lower, which was related to the ligand exchange reaction between carbonate ions and HA molecules.44 The ionexchange indicated that the LDHs/HCMSs show higher affinity to HA than carbonate. The XPS spectra of LDHs/HCMSs before and after HA/Pb(II) adsorption are given in Figure 3. In the high-resolution of the Pb 4f spectrum (Figure 3b), the characteristic doublet peaks of Pb 4f5/2 and Pb 4f7/2 were observed at 138.2 and 143.2 eV, respectively, which were higher than their binding energy values in aqueous PbO complexes (136.9 and 141.7 eV), and which were in accordance with those in the Pb(II)CO32− binding form, implying the specific binding of Pb(II) on LDHs/HCMSs as the speciation of Pb3(CO3)2(OH)2.23,26 Compared with that of the Al 2p peak (∼73.8 eV) in LDHs/HCMSs (Figure 3f), the binding energy of Al 2p after HA adsorption shifted to 74.3 eV, indicating that the main phase of Al changed from AlO to AlOH in the HA-laden LDHs/HCMSs.45 In comparison with the Al 2p spectrum, a remarkable decrease of Mg 1s intensities (Figure 3e) was observed from Pb(II)-laden (or HA-laden) LDHs/ HCMSs, showing the strong interaction between Pb(II) ions (or HA) with Mg atoms. The EDX result (Figure 2a,b) showed that the atomic ratio of Mg/Al after Pb(II) adsorption was about 0.33, which was lower than the ratio of Mg/Al (1.72) calculated from the preparation method.46 The decrease of Mg content in LDHs/HCMSs after Pb(II) or HA adsorption was in connection with the pKsp values of Mg(OH)2(s), which resulted in the decrease of Mg 1s intensities on Pb(II)-laden (or HA-laden) LDHs/HCMSs. The decrease of Mg 1s intensities on HA-laden LDHs/HCMSs was ascribed to the property of Mg-Al-LDHs that release Mg2+ in aqueous solution. Notably, the decrease of Mg 1s intensity in Pb(II)-laden LDHs/HCMSs was more significant than it was in the HAladen LDHs/HCMSs, indicating that Mg2+ took part in the Pb(II) binding on LDHs/HCMSs, and the mechanism of Pb(II) adsorption was partly due to ion-exchange with Mg2+. For the Pb(II)-laden LDHs/HCMSs, the contents of O−, OH−, and H2O, which were calculated from the peak areas, varied from about 19.21%, 62.45%, and 18.34% (LDHs/HCMSs before Pb(II) adsorption) to about 5.95%, 74.34%, and 19.70% (after Pb(II) adsorption), respectively. It can be found that the specific OH− peak area increased, and the O− peak area decreased, suggesting the generation of Pb(II) hydroxides.26 For the HA-laden LDHs/HCMSs, the peak of OCO shifted to a lower binding energy after HA adsorption, demonstrating that HA could be easily adsorbed on LDHs/ HCMSs’ surface and had a strong ligand exchange reaction with the intercalated CO32−.

Mg(OH)2 (s) → Mg 2 + + OH−

(7)

Al(OH)3 (s) → Al3 + + OH−

(8)

According to the pKsp values of Mg(OH)2(s) (11.2) and Al(OH)3(s) (32.9), the theoretical value of the calculated equilibrium concentration of Mg(II) was 5.6 × 106 M at pH 5.0, while that of Al(III) ions was 2.0 × 10−6 M.48,49 Consequently, the dissociation of Mg from LDHs/HCMSs might be a critical part in the Pb(II) uptake processes. The slight intensity change of Al 2p spectra after the adsorption of HA or Pb(II) also confirmed part-dissociation of Al ions from LDHs/HCMSs. On the basis of the aforementioned discussion, the interaction mechanism of HA and Pb(II) on LDHs/ HCMSs was illustrated in Figure 8.

Figure 8. Schematic diagram of adsorption of Pb(II) and HA on LDHs/HCMSs.

Competitive Adsorption of Pb(II) and Other Heavy Metal Ions. The individual and competitive divalent ion adsorptions [Cu(II), Co(II), Pb(II), and Ni(II)] on LDHs/ HCMSs in the absence and presence of HA were carried out at the initial concentration of 20 ppm. Figure 9a shows that LDHs/HCMSs can simultaneously remove Pb2+, Ni2+, Cu2+, and Co2+ ions from solutions and hold high affinity in the order of Co2+ < Ni2+ < Cu2+ < Pb2+. In comparison, the adsorption of Pb2+, Cu2+, Ni2+, and Co2+ ions on LDHs/HCMSs all increased in the presence of HA (Figure 9b). As compared to that in the absence of other heavy metal ions, the adsorption capacity of Pb(II) was higher than that of Pb(II) in the presence of Cu(II), Co(II), and Ni(II) ions, suggesting that Pb(II) adsorption was slightly inhibited in the presence of other heavy metal ions because of the competitive interaction behavior of other heavy metal ions. However, the sum amount of target ions adsorbed on LDHs/HCMSs increased with the increases of other metal ions. Therefore, the competitive adsorption results confirmed that LDHs/HCMSs exhibited fine applicability for the simultaneous elimination of various metal ions from wastewater, which would be significant for the practical application of LDHs/HCMSs in the elimination of different metal ions from wastewaters in environmental pollution cleanup. 11277


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Figure 9. Individual and competitive adsorption capacities of Pb2+, Cu2+, Co2+, and Ni2+ on LDHs/HCMSs (a) without or (b) with HA (C0 = 20 mg L−1; pH = 5.0; m/V = 0.1 g L−1; T = 298 K; and I = 0.01 M NaNO3).

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(2) Ma, Q.; Cheng, H.; Yu, Y.; Huang, Y.; Lu, Q.; Han, S.; Chen, J.; Wang, R.; Fane, A. G.; Zhang, H. Preparation of Superhydrophilic and Underwater Superoleophobic Nanofiber-Based Meshes from Waste Glass for Multifunctional Oil/Water Separation. Small 2017, 13, 1700391. (3) Bi, H.; Yin, Z.; Cao, X.; Xie, X.; Tan, C.; Huang, X.; Chen, B.; Chen, F.; Yang, Q.; Bu, X.; Lu, X.; Sun, L.; Zhang, H. Carbon fiber aerogel made from raw cotton: a novel, efficient and recyclable sorbent for oils and organic solvents. Adv. Mater. 2013, 25, 5916−5921. (4) Bi, H.; Huang, X.; Wu, X.; Cao, X.; Tan, C.; Yin, Z.; Lu, X.; Sun, L.; Zhang, H. Carbon microbelt aerogel prepared by waste paper: an efficient and recyclable sorbent for oils and organic solvents. Small 2014, 10, 3544−3550. (5) Chang, M. Y.; Juang, R. S. Adsorption of tannic acid, humic acid, and dyes from water using the composite of chitosan and activated clay. J. Colloid Interface Sci. 2004, 278, 18−25. (6) Kurniawan, T. A.; Chan, G. Y. S.; Lo, W. H.; Babel, S. Physicochemical treatment techniques for wastewater laden with heavy metals. Chem. Eng. J. 2006, 118, 83−98. (7) Yan, G.; Viraraghavan, T. Heavy metal removal in a biosorption column by immobilized M. rouxii biomass. Bioresour. Technol. 2001, 78, 243−249. (8) Kongsricharoern, N.; Polprasert, C. Chromium removal by a bipolar electro-chemical precipitation process. Water Sci. Technol. 1996, 34 (9), 109−116. (9) Wang, G. S.; Liao, C. H.; Wu, F. J. Photodegradation of humic acids in the presence of hydrogen peroxide. Chemosphere 2001, 42, 379−387. (10) Zou, Y. D.; Wang, X. X.; Ai, Y. J.; Liu, Y. H.; Li, J. X.; Ji, Y. F.; Wang, X. K. Coagulation Behaviour of Graphene Oxide on Nanocrystallined Mg/Al Layered Double Hydroxides: Batch Experimental and Theoretical Calculation Study. Environ. Sci. Technol. 2016, 50, 3658−3667. (11) Rengaraj, S.; Yeon, K. H.; Moon, S. H. Removal of chromium from water and wastewater by ion exchange resins. J. Hazard. Mater. 2001, 87, 273−287. (12) Zou, Y. D.; Wang, X. X.; Ai, Y. J.; Liu, Y. H.; Ji, Y. F.; Wang, H. Q.; Hayat, T.; Alsaedi, A.; Hu, W. P.; Wang, X. K. β-Cyclodextrin modified graphitic carbon nitride for the removal of pollutants from aqueous solution: experimental and theoretical calculation study. J. Mater. Chem. A 2016, 4, 14170−14179. (13) Li, L.; Ma, R.; Iyi, N.; Ebina, Y.; Takada, K.; Sasaki, T. Hollow nanoshell of layered double hydroxide. Chem. Commun. 2006, 2, 3125−3127. (14) Das, N. N.; Konar, J.; Mohanta, M. K.; Srivastava, S. C. Adsorption of Cr(VI) and Se(IV) from their aqueous solutions onto Zr4+-substituted ZnAl/MgAl-layered double hydroxides: Effect of Zr4+ substitution in the layer. J. Colloid Interface Sci. 2004, 270, 1−8. (15) You, Y.; Vance, G. F.; Zhao, H. Selenium adsorption on Mg-Al and Zn-Al layered double hydroxides. Appl. Clay Sci. 2001, 20, 13−25. (16) You, Y.; Zhao, H.; Vance, G. F. Surfactant-enhanced adsorption of organic compounds by layered double hydroxides. Colloids Surf., A 2002, 205, 161−172.

CONCLUSION In summary, we herein reported the synthesis of the hierarchical structure LDHs/HCMSs using a hydrothermal method. The adsorption behavior of LDHs/HCMSs toward the simultaneous adsorption of HA and Pb(II) was systematically investigated. The results indicated that the LDHs/HCMSs had superior capacities for the simultaneous removal of HA/ Pb(II) ions from solutions. The findings presented herein demonstrated that LDHs/HCMSs were promising materials for the efficient elimination of different kinds of organic/inorganic substances in wastewater purification.

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ASSOCIATED CONTENT

S Supporting Information *

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acssuschemeng.7b01717. Effect of pH and ionic strength on the precipitation of HA in the absence of adsorbent, and the precipitation percentage of HA at different ionic strengths (PDF)

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AUTHOR INFORMATION

Corresponding Authors

*Fax: +86-10-61772890. Phone: +86-10-61772890. E-mail: [email protected]. *E-mail: [email protected]. ORCID

Tao Wen: 0000-0001-9089-3196 Xiangke Wang: 0000-0002-3352-1617 Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS The financial support from NSFC (21607042, 21577032), the National Special Water Programs (2015ZX07204-007, 2015ZX07203-011), the Fundamental Research Funds for the Central Universities (JB2015001, 2017MS045), the Priority Academic Program Development of Jiangsu Higher Education Institutions, and the Jiangsu Provincial Key Laboratory of Radiation Medicine and Protection are acknowledged.

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REFERENCES

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