Lignin-based Magnesium Hydroxide Nanocomposite. Synthesis and

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Lignin-based Magnesium Hydroxide Nanocomposite. Synthesis and Application for the Removal of Potentially Toxic Metals from Aqueous Solution Nikolai Ponomarev, Olga Pastushok, Eveliina Repo, Bhairavi Doshi, and Mika Sillanpää ACS Appl. Nano Mater., Just Accepted Manuscript • DOI: 10.1021/acsanm.9b01083 • Publication Date (Web): 14 Aug 2019 Downloaded from pubs.acs.org on August 15, 2019

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ACS Applied Nano Materials

Lignin-based Magnesium Hydroxide Nanocomposite. Synthesis and Application for the Removal of Potentially Toxic Metals from Aqueous Solution Nikolai Ponomarev 1*, Olga Pastushok 1, Eveliina Repo 2, Bhairavi Doshi 1, Mika Sillanpää 1 1

Department of Green Chemistry, School of Engineering Science, Lappeenranta-Lahti University

of Technology LUT, Sammonkatu 12, FI-50130 Mikkeli, Finland 2

Department of Separation and Purification Technology, School of Engineering Science,

Lappeenranta-Lahti University of Technology LUT, P.O. Box 20, FI-53851 Lappeenranta, Finland TOC GRAPHICS

ACS Paragon Plus Environment

ACS Applied Nano Materials 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

Abstract The pollution of water by potentially toxic metals or so-called heavy metals is the most severe form of environmental impact. Nanocomposites are considered promising materials for the removal of potentially toxic metals from aqueous solution through the adsorption or ion-exchange. In order to produce high-performance adsorbent for the removal of Ni2+, Cd2+ and Pb2+, ligninMg(OH)2 nanocomposite (LH-MH) was developed utilising hydrolytic lignin waste with possible regeneration after metal uptake. The LH-MH was prepared using hydrolytic lignin, magnesium chloride and NaOH solution by a thermally-assisted method. The structure was studied using XRD, TEM and SEM, while chemical composition was evaluated by FTIR, TGA/DTA and EDS mapping. Adsorption was described by the Langmuir and Sips models, whereas kinetics was elucidated by pseudo-first-order (PFO) and pseudo-second-order (PSO) models. The obtained material demonstrated a nanocomposite structure indicating well-distributed nano-Mg(OH)2 onto the lignin polymer matrix. The nanocomposite demonstrated superior removal of Ni2+, Cd2+ and Pb2+. The mechanism of adsorption was investigated indicating ion exchange between toxic metals and Mg2+. The obtained adsorbent was successfully regenerated using combined treatment by HClMgCl2-NaOH. Keywords: lignin, brucite, nanocomposite, adsorption, ion exchange, heavy metals 1. Introduction Due to the significant environmental impact of potentially toxic metals present in waters, various techniques for their removal such as precipitation and co-precipitation, coagulation, electrochemical treatment, membrane separation and ion-exchange have been applied 1. However, the relatively high capital and/or operating costs of the listed processes have initiated the development of cost-efficient adsorption techniques applicable for the uptake of toxic metals 2.


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Activated carbon, zeolites, clays, etc. were found to be effective adsorbents for potentially toxic metals, but limited adsorption capacities led to the research of novel adsorption materials 3. Recently, nanoparticles have been widely applied for adsorption

4

but, despite the adsorption

performance of nanoparticles, there are some critical drawbacks such as high initial costs, complicated separation and toxic properties leading to limited application in water treatment 5. In recent years, the application of nanocomposites for toxic metal removal has attracted substantial attention. An association of nanoparticles onto the polymer matrix of nanocomposite eliminates disadvantages and saves the attractive properties of nanoparticles. For example, graphene oxides and carbon nanotubes have been used in the synthesis of nanocomposites, which have then demonstrated good adsorption performance towards toxic metals

6,7.

However, expensive

production costs and the hazardous properties of the aforementioned nanomaterials make their utilisation less attractive 8. In order to overcome the environmental challenges of nanocomposites, cellulose was used as a sustainable raw material. For example, the cellulose-carbonated hydroxyapatite, nanocomposite demonstrated impressive removal of Ni2+ and Cd2+ 9. However, the production of cellulose-based nanocomposites requires the use of cellulose solvents such as ionic liquids, DMSO, alkali-urea solutions, etc., which negatively impact on the environment 10. Moreover, cellulose is not a side-product and the use of this polymer does not solve the problem of waste utilisation. Therefore, the development of novel environmentally-friendly nanocomposites based on cost-efficient or waste materials is a worth studying. One of the promising raw materials for adsorbents or nanocomposites synthesis is lignin. Lignin is the next abundant plant polymer after cellulose. It appears in arboreous plants and is located in the cell wall of the plant. Lignin is a by-product of the pulp and hydrolysis industries. There are around 95 million tonnes of waste hydrolytic lignin in the post-Soviet Union area,



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causing environmental problems and requiring utilisation 11. The availability of hydrolytic lignin coupled with the required utilisation promote lignin as the ideal raw material candidate for nanocomposite. Compared with other materials, lignin, with its irregular structure with phenolic and carboxylic functional groups, make it good for heavy metal removal 12. For instance, lignin clay composite and lignin-based nano-trap have been successfully employed for the uptake of potentially toxic metals

13,14.

Lignin was applied to some (nano)composites including lignin-

modified silica nanoparticles 15, magnetic lignin composite

16,17

and lignosulphonate interleaved

layered double hydroxide 18. Magnesium hydroxide or brucite has the basic layered unit with C6-type hexagonal symmetry. Nano-Mg(OH)2 has exhibited significant removal performance towards heavy metals through ion-exchange 19. As reported earlier, nano-sized Mg(OH)2 onto a cellulose matrix can be easily synthesized using a straightforward and environmentally-friendly method

10.

Moreover,

Mg(OH)2 is a safe reinforcement for lignin since the WHO limit for the possible leached concentration of Mg ions is quite high (450 mg/L) 20. The aim of this study is to obtain a cost-efficient and industrial applicable material for potentially toxic metal removal using sustainable techniques utilising industrial wastes. Hydrolytic lignin being a waste product of the hydrolysis industry was selected and further reinforced by Mg(OH)2 for the effective removal of Ni2+, Cd2+, and Pb2+. To overcome the typical disadvantages of nanoparticles, nano-Mg(OH)2 was dispersed on a lignin matrix using a simple thermally-assisted method without hazardous chemicals. In addition, the regeneration of the obtained nanomaterial became possible due to an ion-exchange removal mechanism.


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2. Materials and methods 2.1 Raw materials and chemicals Polyphepan®, a commercial industrial coniferous hydrolytic lignin was used as received without further purification

21.

MgCl2∙6H2O, NaOH and HCl were of analytical purity and

purchased from Sigma-Aldrich. Deionised water was used for solution preparation. 2.2 Synthesis of lignin-based nanocomposite The synthesis route and proposed structure of obtained material are illustrated in Scheme 1. Sodium hydroxide solution containing 20.0 g of NaOH and 180.0 g of deionized water was mixed with 55.5 g of wet polyphepan® (20.0 g of dry matter) and allowed to swell for 1 hour under magnetic stirring. A certain volume (105.0 mL) of 1 mol/L MgCl2 solution was dropwise added to prepared polyphepan® suspension under magnetic agitating, that equals 1.0 g/g of Mg(OH)2to-lignin mixing ratio (R). The received mixture was heated under a reflux condenser at 65 ºC for 16 hours. The obtained slurry was isolated from the mixture by centrifugation. The received gel after centrifugation was mixed with 200 mL of deionized water and the material was coagulated by the 10% HCl from the obtained suspension. Coagulated product was washed in a Buchner funnel with hot water (70 ºC) several times until neutral pH was reached. The wet sludge from the Buchner funnel was collected into an extruder (Supplementary Information: Fig. S1) for granule production. The granulated product was dried at 80 ºC for 8 hours. The size of the obtained granules was 1-3 mm (Supplementary Information: Fig. S2).



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2.3 Characterization studies of lignin-based nanocomposite Fourier transform infrared spectroscopy (FTIR) type Bruker Vertex 70 was applied to analyze the surface functional groups of the LH-MH. The crystallinity of the product was determined by X-ray powder diffraction (XRD) on a PANanalytical X-ray diffractometer operating at 40 kV with Co Kα radiation in 2θ range from 10º to 120º. To obtain spectra with peaks, which are represented in Cu 2θ degrees, the values of spectra recorded in Co 2θ degrees were recalculted using Bragg law 22: 𝜆𝐶𝑢 = 2𝑑 ∙ 𝑠𝑖𝑛𝜃𝐶𝑢

(1)

where 𝑑 – lattice parameter (Å); 𝜆𝐶𝑢 and 𝜃𝐶𝑢 are the X-ray wavelength (1.5406 Å) and the Bragg diffraction angle (rad) for Cu radiation, respectively. Thermal decomposition of the material was studied with thermogravimetric analyses (TGA) and differential thermal analysis (DTA) on a simultaneous thermal analyzer NETZSCH STA 449C at a heating rate of 10 ºC min-1 from 25 ºC to 800 ºC. Surface areas and pore sizes were determined by Tristar® II Plus. The information on pore size distribution and surface area was derived from adsorption-desorption N2 isotherm using the non-local density functional theory (NLDFT) 23 and Brunauer-Emmett-Teller (BET) 24 model, respectively. The total volume of pores was calculated at a relative pressure p/p0=0.99. Transmission electron microscopy (TEM) micrographs were made with a Hitachi H-7600 transmission electron microscope using 100 kV of acceleration voltage. Scanning electron microscopy (SEM) images were taken with a Hitachi S-4800 scanning electron microscope using 30 kV of acceleration voltage, which was equipped with an energy dispersive X-Ray spectrometer (EDS) operated at 30 kV. The pH-zero point of charge (pHzpc) was studied by the method reported earlier 25. The pH of 0.01 mol/L sodium chloride solution contained 200 mg of the LH-MH was adjusted from 2-12 by dropwise addition of 0.1 mol/L HCl or NaOH. After 48


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hours the pH of the solutions with and without material was determined and the point of intersection of initial pH (pHinitial) versus final ΔpH (pHinitial-pHfinal) was estimated as the pH-zero point of charge of the studied material. 2.4 Bed adsorption experiments An appropriate amount of Ni2+, Cd2+ and Pb2+ nitrate salts were dissolved in deionized water to prepare cation stock solutions of 0.1 mol/L concentration. All the tests were performed in polypropylene tubes containing 2 g/L of material with 0.5 mmol/L Ni2+, Cd2+ or Pb2+ solution on a rocker shaker (CAT ST5) with 420 x 350 mm platform at constant oscillations per minute (70 min-1) for a certain time at varied temperature. In a typical experiment, the adsorption was performed at pH 6.5 for 16 hours at 23 °C. The plasma and nebuliser gas flow were 12.0 and 0.7 L/min, respectively. Sample uptake pump speed was 12 rpm at a rinsing time of 20 s between measurements. The wavelengths used for the analysis of elements were Ni = 231.6 nm; Cd = 226.5 nm; Pb = 220.4 nm.

The pH of initial solutions was controlled by addition of 0.1 mol/L NaOH

or HCl to evaluate the effect of pH on cations uptake. Because of possible precipitation of Ni2+, Cd2+ and Pb2+ as corresponding hydroxides at high pH values (pH>8) during adsorption the pH range of 2 to 7 was selected. For isotherm and kinetic studies, the experiments were performed with a varied initial concentration range (0.1-20 mmol/L) and for a varied adsorption time (5-1440 min). The effect of competing ions was studied for Ni2+, Cd2+ and Pb2+ at the simultaneous presence in 40 mL aqueous solution containing 0.5 mmol/L of Na+, K+, Mg2+ and Ca2+ of corresponding nitrates. The optimal ratio of Mg(OH)2 in LH-MH was investigated by varying of 1.0 mol/L MgCl2 volume added to polyphepan® suspension: 105.0, 53.0, 26.0 and 10.5 mL corresponding to the following MgCl2 amount: 10.0; 5.0; 2.5 and 1.0 g, respectively. A blank test



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with no adsorbent in a polypropylene tube was performed and adsorption on the tube walls was negligible. Experimentally determined adsorption capacity (𝑞𝑒𝑥𝑝) of the LH-MH (mmol/g) was defined as follows: 𝑞𝑒𝑥𝑝 =

𝐶𝑖 ― 𝐶𝑒 𝑚

𝑉

(2)

where 𝐶𝑖 – initial concentration of the cations (mmol/L); 𝐶𝑒 - concentrations of the cations after removal (mmol/L); 𝑚 - the mass of the granulated material (mg); 𝑉 - the volume of the solution (mL). 2.5 Regeneration studies At first, the LH-MH was loaded by the cations by agitating 80 mg of the material with 40 mL of 0.4 - 0.6 mmol/L metal solutions in the same manner as for adsorption studies. After the equilibrium time was attained, the loaded adsorbent was collected from the solution and the residual metal solution was decanted from the granules of the adsorbent. The material was regenerated in two different ways. In the first way, the cations were desorbed using 0.1, 0.01 and 0.001 mol/L hydrochloric acid. In the second way, the nanocomposite after desorption by 0.01 mol/L HCl was regenerated by saturation of 1.0 mol/L MgCl2 with the following saturation of 0.05 mol/L NaOH and washing of granules by DI water several times until neutral pH. The idea of the second way of regeneration was to recover the leached Mg(OH)2 in the LH-MH after desorption of adsorbed products. The efficiency of regeneration (𝑅) was defined with the following formula (%):


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𝑅=

𝑞𝑟 𝑞0

× 100

(3)

where 𝑞0 – adsorption capacity before regeneration (mmol/g); 𝑞𝑟 - adsorption capacity after regeneration (mmol/g). 2.6 Kinetics and isotherms models The non-linear equations were applied for the modelling of adsorption kinetics and adsorption isotherms. The typical remarks and features of adsorption studies were accounted for 26.

The application of non-linear equations is justified by the elimination of error distributions

caused by the linearization technique

27.

Non-linear curve fit and statistical validations such as

regression coefficient R2, F-ratio and sum of residual square (SRS) were performed using OriginPro 2018 software. The mathematical models used for the non-linear curve fit are described by equations 4-7.

Scheme 1. Dropwise addition of MgCl2 to lignin-NaOH suspension (a); refluxing (b); filtration (c); obtained granules (d) and proposed structure of obtained LH-MH (e).



3. Results and discussion 3.1 Characterization studies

Figure 1. FTIR (a); XRD (b); BET N2 adsorption-desorption isotherm (c); NLDFT pore size distribution (d) and TGA/DTG (e) of LH-MH at R = 1.0 g/g.

To study the chemical composition of the LH-MH, FTIR studies were accomplished (Fig. 1a). The wide peak at 3330 cm-1 is attributed to hydroxyl groups of the material and absorbed water 10. The peak at 2935 cm-1 is indexed to C-H stretching vibration of aromatic methoxy, which is a side chain either of methyl and methylene groups. Peaks at 1594, 1500 and 1450 cm-1 are ascribed to common lignin groups and assigned to C-H bond deformation and aromatic vibration


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of benzene ring; 1265 cm-1 and 1211 cm-1 are bands of aromatic phenyl and C=O; the insignificant peak at 850 cm-1 is assigned to C-H deformations 28. After the modification of lignin using Mg(OH)2, the following changes were observed. The strong sharp peak appearing after lignin modification at 3697 cm-1 is attributed to brucite lattice vibration

29.

The intensity and sharpness of the peak indicate the existence of Mg(OH)2 in the

crystalline form in the nanocomposite. The broad peaks observed from 400 to 800 cm-1 also come from Mg-OH vibrations and O-Mg-O stretching 10. Since NaOH was used in the synthesis, there are some changes in lignin groups, which can be assumed to be due to possible interaction in alkaline media used in the synthesis. In particular, reactions of destruction and condensation are intrinsic for lignin in the presence of alkali 30,31. The peak of aromatic C-Hn stretching at 1709 cm-1 is not observed for LH-MH compared to raw lignin 32. The peaks at 1150 and 1115 cm-1, which are ascribed to guaiacylpropane unit, were also changed. The crystalline structure of the LH-MH was investigated with XRD analyses (Fig 1b). All of the peaks are assigned to the well-crystalline phase of Mg(OH)2 (ICSD 98-007-9031) 10. The sharpness and intensity of the clear and well-distributed diffraction peaks in the XRD spectra demonstrate the presence of well-crystalline phase of reinforcement in the nanocomposite. The average crystal size of Mg(OH)2 is 9.94 nm calculated by the Sherrer formula 33. The peaks of lignin are not observed in the spectra of LH-MH because of the amorphous structure. The XRD spectra of lignin exhibit an amorphous plateau and weak diffraction peak of residual silica (2θ = 22.4°) in raw lignin that disappears in LH-MH after alkaline treatment 34. According to the XRD results, the nanocomposite has the structure with nanosized reinforcements of pure crystalline Mg(OH)2.



To evaluate the porous texture of lignin and LH-MH, the N2 adsorption-desorption isotherm was plotted (Fig. 1c). The isotherms, hysteresis and pore size distribution are described according to the IUPAC report 35. Lignin exhibits Isotherm Type III non-porous materials, while isotherm IV(a), observed for LH-MH with a characteristic plateau starting from 0.7 p/p°, is attributed to mesoporous materials. Hysteresis is observed for both isotherms indicating capillary condensation. However, in the case of LH-MH, hysteresis is more pronounced and could be indexed to H2(b) type assuming pore blocking in the narrow pore necks of mesoporous material. NLDFT pore size distribution (Fig. 1d) is in a good agreement with isotherm types, indicating the most significant peak at 50 nm for lignin and 10 nm for LH-MH, which are attributable to the size of macropores and mesopores, respectively. The BET surface area of non-modified lignin was 9.4 m2/g while, for LH-MH, it increased to 35.2 m2/g due to the possible presence of mesopores. The total volume of pores is 0.056 cm3/g and 0.072 cm3/g for lignin and LH-MH, respectively. The thermal stability of lignin and LH-MH was studied by thermogravimetry analysis (TGA). Fig. 1e shows both the TGA and DTG curves of lignin and LH-MH. The peaks on the DTG curves at 58 ºC and 87 ºC are attributable to the desorption of water in the material. A cleavage of C-C linkages occurred at 308 ºC for lignin and 351 ºC for LH-MH 30. The intense peak at 396 ºC arises from the thermal degradation of Mg(OH)2 in the LH-MH sample 36. A broad peak at 390-460 ºC is observed for the lignin sample and attributed to complete decomposition or condensation of the aromatic ring while for LH-MH, the aryl-aryl cleavage more likely occurred at 496 ºC 30. The degradation of C-C bonds and aromatic rings at higher temperatures for LH-MH compared to lignin probably arises from the interactions between lignin and Mg(OH)2. To observe the morphology of the material, LH-MH was examined with TEM and SEM (Fig. 2). Plate-like, rod-like and characteristic hexagonal intercalations of brucite are clearly


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observed in the TEM images. The measured crystal sizes of the plate-like and hexagonal particles are around 30-50 nm. The lengths of the rod-like brucite crystals are approximately 50-135 nm and the widths between 10 and -20 nm. The SEM images (Fig. 2 c,d) show that the brucite nanocrystals were totally integrated into the polymer matrix and were not mechanically mixed. The nanostructures of Mg(OH)2 were formed into the lignin matrix and LH-MH was obtained because of the suggested formation of hydrogen bonds between hydroxyls of lignin and magnesium hydroxide. In the magnified SEM micrograph Fig. 2d the porous structure of the synthesized material is clearly observed. The morphology of the LH-MH is the same as for previously reported cellulose-Mg(OH)2 nanocomposite except that embedded and spherical brucite nanoparticles are not observed 10. TEM and SEM micrographs demonstrate successful synthesis of LH-MH reinforced by Mg(OH)2 nanostructures. The results of TEM and SEM very well match the prediction of XRD crystalline size and BET mesoporous structure of plate-like particles.



Figure 2. TEM (a-b):1 – plate-like, 2 – rod-like, 3 – hexagonal intercalations of Mg(OH)2 and SEM (c-d) images of the LH-MH at R = 1.0 g/g.

EDS elemental spectrum (Fig. 3a) and mapping (Fig. 3b-e) demonstrate the elemental composition of the LH-MH. The results show the presence of corresponding elements: carbon, magnesium, oxygen and silver in the synthezied material. Mentioned elements are assigned to the LH-MH except for silver that comes from the preparation technique of the samples for the SEM analyses. EDS spectrum substantiates the right composition of the LH-MH. The intensity of the magnesium peak indicates the existence of this element in the nanocomposite. An even allocation of the magnesium nano-sized structures in the polymer matrix is demonstrated by the EDS


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elemental mapping (Fig. 3b-e). Other elements of LH-MH are also evenly distributed in the material and verify the composition and structure corresponding to the LH-MH. Based on the characterisation methods, the right composition and structure of the LH-MH were confirmed and the results of these methods matched each other well. The latter part of this article focuses on the application of the material to cation removal.

Figure 3. (a) EDS spectrum and elemental mapping of the LH-MH at R = 1.0 g/g: (b) SEM image; (c) C; (d) Mg; (e) O.



3.2 Adsorption studies 3.2.1

Effect of initial pH

Figure 4. pHzpc of the lignin and LH-MH at R = 1.0 g/g (a), effect of pH on cations removal (b).

The effect of pH on metal removal could be explained based on the point of zero charge (pHzpc) of the adsorbent. Figure 4a shows the positive charge of lignin and LH-MH above pHzpc and negative charge below pHzpc. Due to electrostatic interaction the uptake is favorable below pH = 6.7 for lignin and pH = 10.6 for LH-MH, respectively. However, at pH > 7 the cations are prone to precipitate as corresponding hydroxides (Fig. 4b). At acidic pH < 4, the charge is significantly lower, which explains the reduced metals uptake at pH < 4 (Fig. 4a, b) for lignin and LH-MH. The uptake of cations for LH-MH reached maximum at lower pH values compared to lignin because of its higher negative charge (ΔpH) at pH below pHzpc. 3.2.2 Effect of contact time and kinetic models The influence of contact time on the removal of Ni2+, Cd2+ and Pb2+ by the LH-MH was investigated using single component model solutions. The effect of contact time and plotted kinetic models are shown in Fig. 5a-c. More than 80% of uptake occurred during the first 240 min.


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However, the complete steady state of Ni2+, Cd2+ and Pb2+ was reached in 360 min. The results of contact time were analyzed with the PFO 37 and PSO kinetic models 38, which are commonly used to simulate the kinetics of adsorption. The PFO kinetic model is expressed by the following nonlinear equation: 𝑞𝑡 = 𝑞𝑒(1 ― 𝑒 ― 𝑘1𝑡)

(4)

and the PSO model as follows: 𝑞𝑡 =

𝑞2𝑒 𝑘2𝑡 1 + 𝑞𝑒𝑘2𝑡

(5)

where 𝑞𝑡 (mmol/L) is the capacity of adsorption at a certain time 𝑡 (min), 𝑞𝑒 (mmol/L) is the equilibrium capacity of adsorption, and 𝑘1 (1/min) and 𝑘2 (g/mmol min) are the adsorption rate constants of the PFO and PSO models, respectively. Table 1 represents the kinetic parameters and comparison with the values of experimentally defined equilibrium capacities of adsorption 𝑞𝑒,𝑒𝑥𝑝. According to Table 1 and Fig. 5a-c both kinetic models were well-fitted with the experimental findings and can be applied for the kinetic evaluation of Ni2+, Cd2+ and Pb2+ adsorption by the LH-MH. The values of statistical parameters such as regression coefficient R2, F-ratio, and the sum of residual square (SRS) are higher for PFO than PSO in case of Cd2+ and Pb2+, indicating a good adequacy of experimental data with the PFO model (Table 1). Moreover, the calculated values of 𝑞𝑒 of PFO were closer to the experimentally obtained 𝑞𝑒,𝑒𝑥𝑝 compared to the PSO model. In the case of Ni2+, the values of statistical validations (R2, F-ration and SRS) indicate a better fit for PSO than PFO model, but the PFO model predicts more realistic equilibrium adsorption capacity 𝑞𝑒, which is clear also based on Fig. 5 and Table 1.



Based on these findings, the removal of the cations by the nanocomposite occurs mainly through chemisorption. The removal mechanism is discussed further in Section 3.2.6.

Figure 5. Effect of contact time and kinetic models of (a) Ni2+; (b) Cd2+; (c) Pb2+ and adsorption isotherms of (d) Ni2+; (e) Cd2+; (f) Pb2+


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Table 1. Kinetic parameters and statistical validations of experimental data. Cation

qe,exp

PFO

PSO

(mmol/g)

qe

k1

(mmol/g)

(1/min)

R2

SRS

F-

qe

k2

ratio

(mmol/g)

(g/mmol

R2

SRS

Fratio

min) Ni2+

Cd2+

Pb2+

0.233 ±

0.229 ±

0.011

0.004

0.244 ±

0.253 ±

0.009

0.007

0.236 ±

0.233±

0.010

0.005

0.011

0.9940

0.260

2445

0.259 ±

0.053

0.9968

0.260

4597

0.049

0.9562

0.310

319

0.069

0.9844

0.290

1015

0.004 0.011

0.9848

0.315

923

0.283 ± 0.002

0.014

0.9880

0.288

1329

0.260 ± 0.008



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3.2.3 Adsorption capacities and isotherm modelling The adsorption isotherms of the studied metals are shown in Fig. 5d-f. The obtained maximum adsorption capacities were between 0.94-1.24 mmol/g Pb2+ having the highest and Cd2+ the lowest value. The highest uptake of Pb2+ has earlier been assigned to its lowest hydration number, which makes it easier to be attached on the surface 39. For the evaluation of the interactive behaviour between the adsorbent and adsorbate, Langmuir 40 and Sips 41 models of isotherm were used in order to compare theoretical and experimental results of the adsorption equilibrium. The Langmuir model implies monolayer adsorption on the homogeneous adsorbent surface with identic energy adsorption sites 42. The Langmuir model is described by the following formula: 𝑞𝑒 =

𝑞𝑚𝐾𝐿𝐶𝑒 1 + 𝐾𝐿𝐶𝑒

(6)

where 𝑞𝑒 (mmol/g) and 𝐶𝑒 (mmol/L) are the capacity of adsorption and concentration at equilibrium, respectively; 𝑞𝑚 is the maximum amount of cations adsorbed by the material (mmol/g); 𝐾𝐿 (L/mmol) is the Langmuir constant of equilibrium denoting to the steady state capacity of adsorption and adsorption energy of the cations and material. Fig. 5d-f and Table 2 represent that the Langmuir model well-matched with the experimental findings. The values of 𝑞𝑚 determined experimentally and defined by the Langmuir equation are in a good agreement. The equilibrium Langmuir 𝐾𝐿 constants differ by two or three orders for cations that indicates different bonding energies of each metal. High values of the Langmuir constant for Cd2+ and Pb2+ are not well correlated with rather slow adsorption. The mentioned contradictions could be interpreted by the imperfection of the Langmuir isotherm for a more


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complicated adsorption mechanism than monolayer adsorption. The surface heterogeneity or interactions between adsorbate and adsorbent are not accounted for in the Langmuir model. The Sips adsorption model is gained by introducing the Freundlich equation 43 suitable for heterogeneous surface into the Langmuir model. The Sips isotherm is also called the LangmuirFreundlich isotherm. The Sips model is calculated by the following formula: 𝑞𝑒 =

𝑞𝑚(𝐾𝑆𝐶𝑒)𝑛𝑆 1 + (𝐾𝑆𝐶𝑒)𝑛𝑆

(7)

where 𝐾𝑆 (L/mmol) is the Sips constant assigned to the adsorption energy. The Sips model takes the heterogeneity of the system into account by an empirical constant 𝑛𝑆 that comes from the Freundlich model. When the value of 𝑛𝑆 is closer to zero, heterogeneous adsorption occurs and the Sips isotherm transforms into the Freundlich isotherm. If the 𝑛𝑆 value is equal or close to one, the Sips equation reduces to the Langmuir equation and indicates surface homogeneity 2. Therefore, the Sips model can describe both homogeneous and heterogeneous adsorption. The plotted Sips isotherms are represented and graphically compared with the Langmuir model and experimental data in Fig. 5d-f. In addition, adsorption parameters are presented in Table 2. The values of 𝐾𝑆 vary significantly, indicating unequable energies of adsorption between the adsorbent and adsorbate. The values of the Sips model constant are relatively small and comparable with slow adsorption kinetics. The empirical constant 𝑛𝑆 are deviating from the unity, indicating heterogeneous adsorption process. The results of the statistical validation are presented in Table 2. According to the values of regression coefficient R2, F-ratio and SRS, the experimental data shows the best accordance with the Sips isotherm compared to the Langmuir isotherm. The Sips model preferably describes the



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equilibrium of adsorption than the Langmuir model because the heterogeneity is accounted for and, the statistical validations and adsorption constants are well-matched with the kinetic behaviour. However, the scattering of experimental points makes the straightforward interpretation of these results difficult and more weight should be put on the experimental findings presented in the adsorption mechanism section. Table 2. Langmuir and Sips isotherm parameters and statistical validations of experimental data. Adsorbate

Ks R2

qm,exp

qm

KL;

(mmol/g)

(mmol/g)

(L/mmol)

SRS

Fratio

ns (Sips)

Langmuir model Ni2+

1.17 ± 0.09

1.05 ±

5.65

0.9008

7.28

274

-

81.22

0.9538

8.37

836

-

135.13

0.8415

7.71

115

-

1.34

0.9899

7.42

1836

0.453

20.71

0.9613

8.39

667

0.737

2.17

0.8982

7.85

122

0.366

0.06 Cd2+

0.94 ± 0.07

0.93 ± 0.03

Pb2+

1.24 ± 0.08

1.09 ± 0.08

Sips model Ni2+

1.17 ± 0.09

1.40 ± 0.09

Cd2+

0.94 ± 0.07

0.96 ± 0.04

Pb2+

1.24 ± 0.08

1.47 ± 0.03


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3.2.4 Effect of competing ions The competing ions may significantly affect the uptake of Ni2+, Cd2+ and Pb2+ from water by the LH-MH. To evaluate the removal behaviour of the cations in the multicomponent system, the comparison with the singlecomponent system was performed. The multicomponent system contained the simultaneous presence of Ni2+, Cd2+ and Pb2+ and the most commonly found cations: Na+, K+, Ca2+ and Mg2+ (0.5 mmol/L of each) in wastewaters to increase the ionic strength and simulate the conditions of real wastewaters. Fig. S3 (Supplementary Information) demonstrates the achieved adsorption capacities for singlecomponent system of individual cations and multicomponent system with competing ions. The cations of Ni2+, Cd2+ and Pb2+ were removed from the initial solutions by the LH-MH despite the absence or existence of competing ions. Therefore, the removal efficiency of the adsorbent does not depend on the presence of other cations and competition is negligible. 3.2.5 Optimal Mg(OH)2 amount The presence of magnesium hydroxide in nanocomposite is supposed to be essential for the removal of cations. Therefore, the optimal amount of magnesium hydroxide in LH-MH for the cations removal was investigated. On the Figure S4 (Supplementary Information) the characteristic FTIR peak of Mg(OH)2 at 3697 cm-1 is increased with increasing of MgCl2 added to the lignin suspension demonstrating an increased amount of Mg(OH)2 in the material. Hence, it was suggested that magnesium chloride proportionally converted to magnesium hydroxide in the adsorbent and the optimal value of the added amount of MgCl2 could be used as optimal Mg(OH)2 amount in the material. Table S1 (Supplementary Information) shows the amount of added MgCl2



to the suspension of lignin. With the increasing amount of MgCl2 that was converted to Mg(OH)2 (Fig.S4, Supplementary Information) in the LH-MH, the removal of cations also increased. When the added amount of MgCl2, was 5-10 g the uptake of cations was nearly 100%. 3.2.6

Removal mechanism Fig. 6a demonstrates the FTIR spectra of LH-MH before and after uptake of Ni2+, Cd2+ and

Pb2+. The intensities of characteristic Mg(OH)2 peak at 3697 cm-1 of the adsorbent were decreased after adsorption of nickel, cadmium and lead compared to the material before adsorption, indicating a possible interaction between cations and magnesium hydroxide during adsorption 1. The observed peaks at 1405 cm-1 and 681 cm-1 (Fig. 6a, spectra 4) of 𝐶𝑂23 ― could be ascribed to the formation of hydrocerussite 2PbCO3·Pb(OH)2 44 after adsorption of Pb2+ by the LH-MH. The changes of peaks attributed to lignin in the material after adsorption of cations were not observed in FTIR spectra. Most of the functional groups of lignin, which might be involved in the adsorption of cations, have already been occupied by magnesium hydroxide. Moreover, the changes in the characteristic brucite peak and the formation of hydrocerussite demonstrate that the cation exchange with Mg(OH)2 was a more favourable adsorption mechanism than chemisorption by the lignin functional groups. The XRD spectra of the LH-MH before and after metal adsorption is shown in Fig. 6b. All of the diffraction peaks of the material after adsorption of Ni2+ (Fig. 6b, spectra 2) could be attributed to Ni(OH)2 (ICSD 98-016-9978) and residual Mg(OH)2 (ICSD 98-007-9031). The º2θ positions of the peaks assigned to nickel and magnesium hydroxides are similar due to the same hexagonal brucite type crystal structure, but rather different shapes and intensities of the peaks demonstrate the formation of new adsorbed Ni(OH)2 phase. The XRD spectra of the nanocomposites after Cd2+ and Pb2+ adsorption (Fig. 6b, spectra 3,4) are different to that of pure


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nanocomposite. The XRD peaks for Cd2+ on the surface can be attributed to Cd(OH)2 (ICSD 98002-3415), while the phase adsorbed Pb2+ could be indexed to 2PbCO3·Pb(OH)2 (ICSD 98-0280932), because the formed lead(II) hydroxide immediately interacts with CO2 with further formation of hydrocerussite. In addition, a residual Mg(OH)2 phase (ICSD 98-007-9031) was detected in the both nanocomposite samples after adsorption of Cd2+ and Pb2+. The results are quite well matched with the findings reported in literature 1.

Figure 6. The FTIR (a) and XRD (b) spectra of the LH-MH: 1 – before adsorption, 2,3 and 4 – after adsorption of Ni2+, Cd2+ and Pb2+, respectively. The leached amount of Mg2+ versus adsorbed amount of Me2+ (Me2+ = Ni2+, Cd2+, Pb2+) after adsorption (c). The XRD spectra (d) of the LH-MH: 1 – before adsorption and regeneration, 2,3 and 4 – after adsorption of Pb2+ and regeneration using 0.01 mol/L HCl, 0.001 mol/L HCl and combined regeneration (HCl, MgCl2, NaOH), respectively.

The concentration of leached Mg2+ after adsorption of each cation was detected and the results versus the adsorbed amount of Me2+ (Me2+ = Ni2+, Cd2+, Pb2+) are presented in Fig. 6c. One can see that magnesium cations were leached from the LH-MH to the solution equivalently to the



Page 26 of 47

adsorbed amount of Me2+. Despite some leached Mg2+ amount to the treated water during adsorption, its concertation is significantly below the WHO drinking water limit for magnesium (450 ppm) making LH-MH suitable and safe adsorbent for the cations removal. The results demonstrate that Mg2+ cations were exchanged with Me2+ indicating the cationexchange mechanism of adsorption. The relationship is nearly linear and the following chemical reactions of ion-exchange during adsorption process could be suggested: (8)

Mg(OH)2 + Me2+ → Me(OH)2↓ + Mg2+ The formation of hydrocerussite including intermediate dissolved CO2 capture: 3Mg(OH)2 + 3Pb2+ +2CO2·H2O → Pb3(CO3)2(OH)2↓ + 3Mg2+ + 4H2O

(9)

The assumed cation-exchange could be also explained by the differences of the solubility products (Ksp) between Mg(OH)2 and corresponding hydroxides of cations. Table 3 demonstrates the differences between Ksp(Mg(OH)2) and Ksp(Me(OH)2) indicating solubility product of Me(OH)2 far below compared to the solubility product of magnesium hydroxide making the surface precipitation of corresponding hydroxides of Ni2+, Cd2+ and Pb2+ possible. Table 3. Solubility products (Ksp) of corresponding hydroxides 45. Name

Formula

Ksp

Magnesium hydroxide

Mg(OH)2

5.61·10-12

Nickel hydroxide

Ni(OH)2

5.48·10-16

Cadmium hydroxide

Cd(OH)2

7.20·10-15

Lead(II) hydroxide

Pb(OH)2

1.43·10-20

The thermodynamics of adsorption was studied to additionally evaluate the adsorption mechanism. To study thermodynamics, the adsorption was performed at different temperatures


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such as 298, 313, 325 and 343 K. The Gibbs free energy ∆𝐺° is determined by the following equation: ∆𝐺° = 𝑅𝑇𝑙𝑛𝑘𝑑

(10)

where 𝑅 – is a gas constant (J/mol∙K); 𝑇 – temperature (K), and 𝑘𝑑 – is a distribution coefficient: 𝑘𝑑 =

𝑞𝑒

𝐶𝑒

(11)

where 𝑞𝑒 – is adsorption capacity (mmol/g), and 𝐶𝑒 – equilibrium concentration (mmol/L). Changes of enthalpy ∆𝐻° (kJ/mol) and entropy ∆𝑆° (J/K∙mol) are calculated using van’t Hoff equation:

𝑙𝑛𝑘𝑑 = ―

∆𝐻° ∆𝑆° + 𝑅𝑇 𝑅

(12)

Enthalpy ∆𝐻° and entropy ∆𝑆° are derived from the slope and intercept of the van’t Hoff plot, respectively (Fig. S5, Supplementary Information). The results of calculated thermodynamic parameters are shown in Table 4. The negative values of Gibbs free energy ∆𝐺°indicate spontaneous and possible adsorption of Ni, Cd and Pb on LH-MH. The positive values of enthalpy ∆𝐻° exhibit endothermic adsorption that increases with temperature. The values of entropy ∆𝑆° are also positive, assuming an enhanced degree of randomness at the solid-liquid interface, which may be attributed to the release of water during adsorption. The relationship of |𝑇∆𝑆°| >|∆𝐻°| is fulfilled for all temperatures, indicating that entropic changes are more dominant than enthalpic. The aforementioned thermodynamic changes are in good agreement with the previously studied



Page 28 of 47

chemisorption of Cr(IV) onto cellulose-hydroxyapatite nanocomposite, which further assumes the chemisorption mechanism of cations on LH-MH 9. Table 4. Thermodynamic parameters. Cation

Ni(II)

Cd(II)

Pb(II)

T (K)

Kc

R2

ΔG°

ΔH°

ΔS°

(kJ/mol)

(kJ/mol)

(J/mol∙K)

298

0.219

-3.77

313

0.210

-4.07

325

0.186

-4.55

343

0.166

-5.12

298

1.090

-0.21

313

3.584

-3.32

325

8.218

-5.70

343

12.358

-7.17

298

1.576

-1.13

313

1.854

-1.61

325

2.645

-2.63

343

3.223

-3.34

5.52

30.92

0.9591

47.19

160.21

0.9520

14.47

51.99

0.9651

EDS elemental spectra (Fig. S6a-S8a, Supplementary Information) and mapping (Fig. S6c,d-S8c,d) demonstrate the elemental composition of LH-MH after adsorption of Ni, Cd and Pb. EDS spectra show significant peaks of Ni, Cd and Pb indicating their presence on the adsorbent. EDS mapping (Fig. S6c-S8c) shows evenly distributed metal intercalations of a similar size to Mg(OH)2 nanoparticles (Fig. 3). EDS mapping of oxygen (Fig. S6d-S8d) exhibits the same


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positions of elements as well as metal intercalations, assuming the formation of corresponding hydroxides (chemical reactions (8) and (9)). Consequently, the mechanism of Ni2+, Cd2+ and Pb2+ removal from water solution using LH-MH was proposed to be a combination of cation-exchange and surface precipitation with the formation of the corresponding hydroxides. Moreover, despite a relatively small surface area (discussed above), the LH-MH demonstrated a high adsorption capacity further indicating the ionexchange mechanism. The influence of lignin in LH-MH on the adsorption process was not observed and assumed to be negligible, while the cation exchange between cations and Mg(OH)2 was the governing mechanism of cation removal, which was confirmed by analytical techniques (FTIR, XRD, SEM-EDS) and thermodynamic studies. 3.3 Regeneration studies The LH-MH was regenerated in two different ways. In the first way, the material was regenerated using hydrochloric acid with various concentrations. These concentrations were selected to minimize dissolution of magnesium hydroxide as a key component in the material. The nanocomposite was regenerated by hydrochloric acid with a concentration of 0.1 mol/L and after the first cycle, regeneration efficiency was dramatically decreased because at this concentration of HCl, magnesium hydroxide might be significantly leached from the adsorbent. Therefore, the application of 0.1 mol/L HCl as desorbent is not reasonable and the first way of regeneration was performed using 0.01 and 0.001 mol/L HCl. Figure S9 (Supplementary Information) demonstrates the results of the regeneration. The regeneration efficiency decreased after 5 cycles using 0.01 mol/L HCl, while the decline of regeneration using 0.001 mol/L was less steep.



Fig. 6d demonstrates XRD spectra before adsorption-regeneration (spectra 1) and after adsorption of Pb2+ with further regeneration using different methods (spectra 2-4). It was proposed that Mg(OH)2 was leached using 0.01 mol/L HCl and regeneration efficiency was decreased because of a reduced amount of magnesium hydroxide. An absence of Mg(OH)2 is observed in XRD spectra (Fig.6, spectra 2) indicating assumed leaching. In addition, no peaks attributed to the adsorbed product, more specifically 2PbCO3·Pb(OH)2, were found. In this way, the efficiency of 0.01 mol/L HCl as a desorbent was confirmed, but regeneration is rather challenging. Magnesium hydroxide was suggested to be insoluble in 0.001 mol/L HCl and corresponding peaks of Mg(OH)2 (ICSD 98-007-9031) were observed in Fig. 6d (spectra 3) after regeneration. On the other hand, the adsorbed product of Pb2+ was not desorbed and corresponding diffraction peaks indexed to 2PbCO3·Pb(OH)2 (ICSD 98-028-0932) were found. Accordingly, the regeneration efficiency was somewhat better because of saved Mg(OH)2, but still unsatisfactory because of adsorbed products. It is complicated to select a proper desorbent able to desorb adsorbed metals and at the same time maintain a sufficient amount of Mg(OH)2 in the LH-MH, because magnesium hydroxide is more soluble than the observed products formed during adsorption (Ksp(Mg(OH)2) > Ksp(Me(OH)2)). In this way, it was suggested that regeneration of the material could be carried out in two steps: 1st step – desorption of adsorbed metal species from the material; 2nd step – recovery of Mg(OH)2 from the adsorbent. For the first step, the 0.01 mol/L HCl was selected as an effective desorbent, based on the results above (Fig.6d, spectra 2), then the recovery was performed using 1.0 mol/L MgCl2 for saturation of granules by Mg2+ cations with the following addition of 0.05 NaOH mol/L to the decanted granules for the Mg(OH)2 formation onto the polymer matrix. The saturation agent (MgCl2) must be of sufficient concentration to provide enough ions for


Page 30 of 47

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magnesium hydroxide formation while the concentration of NaOH must be enough to create alkali media and at the same time not degrade the lignin polymer matrix. In Fig.6d (spectra 4), the peaks attributed to the adsorbed product 2PbCO3·Pb(OH)2 after removal of Pb2+ were not found indicating successful desorption, whereas peaks indexed to magnesium hydroxide were clearly observed, identifying the fortunate recovery of Mg(OH)2 in the LH-MH after 5 cycles of adsorption-desorption/regeneration. In addition, the regeneration efficiency was satisfied and did not significantly decrease for each Ni2+, Cd2+ and Pb2+ using sequential HCl-MgCl2-NaOH regeneration (Figure S9, Supplementary Information) after 5 cycles. In general, the LH-MH was successfully regenerated based on the knowledge of the adsorption mechanism, and the introduced assumptions are well-matched and confirm with the experimental findings. 3.3 Comparison with other materials The comparison of adsorption performance with lignin and other materials including commercial activated carbon is shown in Table S2. The studied material is superior to commercial activated carbons and raw lignin used in synthesis

46.

The studied LH-MH demonstrated

comparable adsorption capacity with Si/lignin hybrid for the uptake of Ni2+ or Cd2+ and lignin grafted nanotubes for Ni2+

12.

Since only cost-efficient and environmentally-friendly starting

materials were used in the production of LH-MH, the studied material is beneficial for the removal of potentially toxic metals. Moreover, the lignin-Mg(OH)2 can be regenerated making its application economically reasonable.



Conclusion The novel granulated LH-MH was successfully synthesized in a green way using affordable raw material and cost-efficient, environmentally friendly chemicals. The chemical structure corresponding to LH-MH reinforced by Mg(OH)2 was confirmed by FTIR, TGA/DTA and EDS elemental analyses. The results of XRD, TEM and SEM illustrated the morphology and crystalline structure of the nanocomposite. The data of characterization methods were well-matched between each other. The material demonstrated superior adsorption performance of Ni2+, Cd2+ and Pb2+ in singlecomponent and multicomponent solutions. The experimental findings were modelled with the Langmuir and Sips models and the Sips isotherm showed the best accordance with the equilibrium data. The equilibrium of adsorption was achieved relatively fast. Among the PFO and PSO kinetic models, the experimental data were well described by the both models. The ionexchange mechanism between nickel, cadmium or lead and magnesium was assumed and confirmed, indicating the negligible role of lignin polymer matrix in adsorption. The granules of the LH-MH were satisfactorily regenerated using combined regeneration including sequential usage of HCl, MgCl2 and NaOH. The results of this research demonstrate that cost-efficient, easily granulated LH-MH can be successfully employed for the uptake of potentially toxic metals from water stocks. Associated content Supporting information. Photographs of LH-MH, effect of competing ions, FTIR of LH-MH with various MgCl2 amount, the van’t Hoff plots, EDS spectrum and elemental mapping of the LH-MH after adsorption


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Author information Corresponding Author *Email: [email protected], [email protected] ORCID Nikolai Ponomarev 0000-0003-0642-5870 Bhairavi Doshi 0000-0002-8355-533X Acknowledgements We thank Varsha Srivastava and Liisa Puro from Lappeenranta-Lahti University of Technology (LUT University) for their help with analyses of the studied materials with TEM and TGA/DTA, respectively. We appreciate Simo Torniainen and Aapo Nylén of the Laboratory of Materials at South-Eastern University of Applied Sciences (XAMK, Mikkeli) for making SEM micrographs with EDS mapping. References (1)

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Scheme 1. Dropwise addition of MgCl2 to lignin-NaOH suspension (a); refluxing (b); filtration (c); obtained granules (d) and proposed structure of obtained LH-MH (e).



Figure 1. FTIR (a); XRD (b); BET N2 adsorption-desorption isotherm (c); NLDFT pore size distribution (d) and TGA/DTG (e) of LH-MH at R = 1.0 g/g. 190x146mm (300 x 300 DPI)


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Figure 2. TEM (a-b):1 – plate-like, 2 – rod-like, 3 – hexagonal intercalations of Mg(OH)2 and SEM (c-d) images of the LH-MH at R = 1.0 g/g. 164x130mm (300 x 300 DPI)



Figure 3. (a) EDS spectrum and elemental mapping of the LH-MH at R = 1.0 g/g: (b) SEM image; (c) C; (d) Mg; (e) O. 139x107mm (600 x 600 DPI)


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Figure 4. pHzpc of the lignin and LH-MH at R = 1.0 g/g (a), effect of pH on cations removal (b). 161x78mm (300 x 300 DPI)



Figure 5. Effect of contact time and kinetic models of (a) Ni2+; (b) Cd2+; (c) Pb2+ and adsorption isotherms of (d) Ni2+; (e) Cd2+; (f) Pb2+


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Figure 6. The FTIR (a) and XRD (b) spectra of the LH-MH: 1 – before adsorption, 2,3 and 4 – after adsorption of Ni2+, Cd2+ and Pb2+, respectively. The leached amount of Mg2+ versus adsorbed amount of Me2+ (Me2+ = Ni2+, Cd2+, Pb2+) after adsorption (c). The XRD spectra (d) of the LH-MH: 1 – before adsorption and regeneration, 2,3 and 4 – after adsorption of Pb2+ and regeneration using 0.01 mol/L HCl, 0.001 mol/L HCl and combined regeneration (HCl, MgCl2, NaOH), respectively.


Lignin-based Magnesium Hydroxide Nanocomposite. Synthesis and

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