Oak wood ash/GO/Fe3O4 adsorption efficiencies for cadmium and lead removal from aqueous solution: Kinetics, equilibrium and thermodynamic evaluation

Rasool Pelalak; Zahra Heidari; Seyed Mola Khatami; Tonni Agustiono Kurniawan; Azam Marjani; Saeed Shirazian

doi:10.1016/j.arabjc.2021.102991

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Original article

2021

:14;

202103

doi:

10.1016/j.arabjc.2021.102991

Oak wood ash/GO/Fe₃O₄ adsorption efficiencies for cadmium and lead removal from aqueous solution: Kinetics, equilibrium and thermodynamic evaluation

Rasool Pelalak^{^a,^b}, Zahra Heidari^{^c}, Seyed Mola Khatami^{^d}, Tonni Agustiono Kurniawan^{^e}, Azam Marjani^{^f,^g,⁎}, Saeed Shirazian^{^h,ⁱ}

a Institute of Research and Development, Duy Tan University, Da Nang 550000, Viet Nam

b Faculty of Environmental and Chemical Engineering, Duy Tan University, Da Nang 550000, Viet Nam

c Chemical Engineering Faculty, Sahand University of Technology, P.O. Box 51335-1996, Sahand New Town, Tabriz, Iran

d Faculty of Samen Hojaj, Department of Chemistry, Mashhad Branch, Technical and Vocational University (TVU), Tehran, Iran

e Key Laboratory of the Coastal and Wetland Ecosystems (Xiamen University), China’s Ministry of Education, College of the Environment and Ecology, Xiamen University, Xiamen 361102, China

f Department for Management of Science and Technology Development, Ton Duc Thang University, Ho Chi Minh City, Viet Nam

g Faculty of Applied Sciences, Ton Duc Thang University, Ho Chi Minh City, Viet Nam

h Laboratory of Computational Modeling of Drugs, South Ural State University, 76 Lenin prospekt, 454080 Chelyabinsk, Russia

i Department of Chemical Sciences, Bernal Institute, University of Limerick, Limerick, Ireland

⁎Corresponding author at: Ton Duc Thang University, Ho Chi Minh City, Viet Nam. azam.marjani@tdtu.edu.vn (Azam Marjani)

Received: 2020-12-4, Accepted: 2021-1-5,

Disclaimer:
This article was originally published by Elsevier and was migrated to Scientific Scholar after the change of Publisher.

Abstract

Magnetic Oak wood ash/Graphene oxide (Ash/GO/Fe₃O₄) nanocomposites were designed as a high potent adsorbent in the removal of toxic heavy metals such as Lead (Pb(II)) and Cadmium (Cd(II)) ions from aquatic medium. Characterization of Ash/GO/Fe₃O₄ samples was carried out using FESEM, TEM, EDX mapping, BET/BJH, XRD, FTIR, and VSM methods. The obtained results confirmed the successful synthesis of Ash/GO/Fe₃O₄ nanocomposites. In the adsorption process, almost complete adsorption efficiency of produced Ash/GO/Fe₃O₄ nanocomposite was attained under the optimized conditions (99.67% and 98.68% for Pb(II) and Cd(II) adsorption, respectively). The modeling results of kinetics indicated that the mechanism of Pb(II) and Cd(II) adsorption process well fitted by pseudo-second order equation with a high regression coefficient (99.67%). In addition, the equilibrium data were described well by non-linear Langmuir model with the highest adsorption capacity of 47.16 mg/g and 43.66 mg/g for Pb(II) and Cd(II) ions, respectively, which prove the effective adsorption ability of the magnetic nanocomposite. The spontaneous and exothermic nature of adsorption process was confirmed through thermodynamics analyses. The reusability of synthesized Ash/GO/Fe₃O₄ nanocomposites were demonstrated with negligible decrease in adsorption and high stability up to 8 repetitive adsorption cycles. The mechanism of Pb(II) and Cd(II) adsorption on the Ash/GO/Fe₃O₄ nanocomposite was assessed.

Keywords

Adsorption

Magnetic oak wood ash/graphene oxide

Lead and cadmium ions

Wastewater treatment

Thermodynamics

Equilibrium

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1 Introduction

In recent years, the increment of population and development of human activities by different industries and manufacturing processes resulted in increase in type and amount of the released hazardous materials to the aquatic environment (Alkherraz et al., 2020; Heidari et al., 2021). Among these pollutants, the heavy metals, including cadmium (Cd), arsenic (As), lead (Pb), mercury (Hg), and thallium (Tl) are the most serious hazardous materials which are being produced by electroplating, metallurgy, leather tanning and pesticides industries (Heidari et al., 2020; Liu et al., 2020; Pelalak et al., 2020; Zhao et al., 2018 ). The existence of these toxic and non-biodegradable metals in the aquatic environment affects serious long-term problems to the environment even at low concentrations. The heavy metal ions due to high movement can enter very fast to the ground-water sources and cause dangerous health complications (Lebron et al., 2020; Farooq and Jalees, 2020). Furthermore, the accumulation of the heavy metals such as cadmium and lead which cannot be metabolized by human, may lead to brain damages, kidney malfunction and also negative impacts on the human nervous system (Alkherraz et al., 2020; Bereket et al., 1997). Various treatment approaches have been used for remediation of the heavy metals, consisting coagulation/precipitation (Kalaitzidou et al., 2020), reduction/oxidation (Jia et al., 2020), ion exchange (Peng and Guo, 2020), adsorption (Wieszczycka et al., 2020; Soltani et al., 2020; Pelalak et al., 2020 ) and membrane technology (Pelalak et al., 2018; Soltani et al., 2016). Among the investigated methods, adsorption has been considered as the most efficient technique in terms of low cost, easy operation and low energy requirements (Wieszczycka et al., 2020). Furthermore, adsorption method is capable to remove the trace amount of the contaminants, which is one of the main disadvantages of the other treatment methods (Moreira et al., 2019; Yang et al., 2010). Over the past decade, vast variety of the low cost and available adsorbents have been applied for heavy metals removal such as potato peels (Taha et al., 2011), Cashew nut shell (Coelho et al., 2014), waste leaves biomass of Myrica Esculenta (Joshi et al., 2018), and mustard husk (Meena et al., 2008). One of the most abundant waste materials is the fly ash produced by various sources which have been used as adsorbents in heavy metal removal (Núñez-Delgado et al., 2015; Mohan and Gandhimathi, 2009). Improvement in the adsorption capacity and providing strong functional groups on natural adsorbents can be achieved by production of composite adsorbents. Graphene and graphene oxide (GO) as carbon-based nanomaterials, have attracted a great attention in composite synthesis (Xu et al., 2018). GO has some advantages over graphene, including easy manufacturing, higher electron density and presence of various surface functional groups (Sherlala et al., 2018; Pelalak et al., 2013). Furthermore, the results of studies indicated the high desorption ability of the GO, which enables the recovery of the nanostructures up to 90% (Farooq and Jalees, 2020; Chen et al., 2016). The oxidized form of the graphene contains some edge defects in their structure, which resulted in hydrophilic sp³ hybridization. On the other hand, GO due to small particle size and hydrophilic structure cannot be separated from aqueous solution very easily. Application of the Fe₃O₄ nanoparticles in the structure of GO provides possibility of easy separation of nanoparticles from the solution media (Zhou et al., 2020; Zhang et al., 2019). Furthermore, Fe₃O₄ nanoparticles according to their high adsorption capacity and surface area have been employed in many treatment processes. Therefore, utilization of these nontoxic nanoparticles as active phase on GO could be very effective and also economical (Wang, 2018).

In the present research, the novel magnetic Oak wood ash/GO/Fe₃O₄ nanocomposite was effectively prepared through chemical precipitation method. To the best of authors’ knowledge, the GO/Fe₃O₄ magnetic nanocomposite in the presence of Oak wood ash as a cheap, available and ecofriendly adsorbent has not been used for removal of contaminants. The FESEM, TEM, EDX mapping, XRD, FTIR, and VSM techniques were utilized for physicochemical characterization of samples. The synthesized nanocomposite was used for adsorption of lead and cadmium heavy metal ions from aqueous solution. The effect of various operational parameters including pH, the amount of adsorbent, contact time and the initial ion concentrations were investigated on removal efficiency. The kinetics of the adsorption were studied using pseudo first-order, pseudo second-order, and intraparticle diffusion models. The thermodynamic studies as well as adsorption isotherm modeling were also revealed. Moreover, the exothermic, spontaneity characteristics of process and reusability capacity of the nanocomposite were evaluated. Finally, the mechanism of lead and cadmium adsorption by Ash/GO/Fe₃O₄ nanocomposite was assessed.

2 Materials and methods

2.1

2.1 Materials

Oak wood was gained from agricultural fields in Kohgiluyeh and Boyer-Ahmad province of Iran. All the chemicals were used as received without any modifications, and purchased from German Merck Company. Cadmium and lead ions were used as their nitrate form (Cd(NO₃)₂·4H₂O and Pb(NO₃)₂) in the solution. The stock solution of heavy metal ions (1000 ppm) were prepared by 1.598 g of Pb(NO₃)₂ and 2.744 gr of Cd(NO₃)₂·4H₂O dissolution in 1000 mL of distilled water. For pH adjustment HCl and NaOH solution (0.1–1 M) were used.

2.2

2.2 Synthesis of the nanocomposite adsorbent

2.2.1

2.2.1 Preparing the Ash

The Oak wood used as the ash in this study was obtained from Kohgiluyeh forest in Iran. At first the surface of natural Oak wood was rinsed several times with distilled water in order to remove any surface pollutants. The obtained wood was dried at 378 K for 24 h (Saygili Canlidinç et al., 2017). Then, the dried samples were transferred to a furnace and kept at 973 K (5 °C/min) for 6 h. At the end of process, the product powders were collected and cooled at room temperature.

2.2.2

2.2.2 Synthesis of GO

In this study the improved Hummers’ method was utilized for preparation of GO (Marcano et al., 2010). In brief, 3.0 g graphite flakes and 18.0 g KMnO₄ were added to 400 mL of H₂SO₄/H₃PO₄ solution with ratio of 360:40 mL. The obtained solution was magnetically stirred for 14 h (323 K). After the aging time, the reaction solution was cooled to the room temperature and then immersed into 400 mL of iced water. The synthesis procedure was completed by adding H₂O₂ (30%) to produce a bright yellow color (Yang et al., 2019). The pH of the resulting solution neutralized by washing with distilled water. The final produced GO were dried at 343 K for 24 h.

2.2.3

2.2.3 Synthesis of Ash/GO/Fe₃O₄ nanocomposite

In order to synthesize the magnetic Ash/GO/Fe₃O₄ nanocomposite, firstly about 0.5 g of the synthesized GO was added to distilled water (100 mL) and sonicated for 30 min. Then by adding 1.0 g of Oak wood ash, solution was sonicated for 20 min. The Ash/GO solution was properly mixed with chloride hexahydrate (FeCl₃·6H₂O) and ferrous chloride tetrahydrate (FeCl₂·4H₂O) (2:1 M ratio) solution for 20 min to obtain a homogeneous solution. In the next step, the solution temperature was increased to 358–363 K and about 50 mL of NaOH solution (3 M) was added dropwise as precipitant agent. After 40 min by using a magnet, the synthesized composite was gathered from aqueous solution then rinsed with distilled water to completely neutralize (pH around 7). Finally, the produced Ash/GO/Fe₃O₄ nanocomposites were dried at 378 K (24 h). Scheme 1 has shown a brief synthesis steps and adsorption of Cd(II) and Pb(II) ions on the Ash/GO/Fe₃O₄ nanocomposite.

Synthesis of Ash/GO/Fe3O4 nanocomposite and grafting of Cd(II) and Pb(II) on the silica surface.

2.3

2.3 Nanocomposite characterizations

Morphology of the synthesized oak wood ash, GO, and Ash/GO/Fe₃O₄ nanocomposite was explored by DSM-960A Field Emission Scanning Electron Microscope (FESEM, Zeiss, Germany) equipped with the second generation Energy-Dispersive X-ray Spectroscopy (EDS). The High Resolution Transmission Electron Microscope analysis (HRTEM) was performed by using (JEOL, JEM-F200, Japan). The surface functional groups of the synthesized nanocomposite were surveyed using Fourier Transform Infrared Spectroscopy (FTIR, Bruker, Germany). The crystalline structure of the synthesized samples was investigated using X-ray diffraction (XRD) applying an Asenware (AW-XDM300, China). The Vibrating Sample Magnetometer analysis (Meghnatis Daghigh Kavir Co, Iran) was used for recording the magnetic properties of the synthesized Ash/GO/Fe₃O₄ nanocomposite. The pH value at point of zero charge (pH_pzc) was measured from plots between ΔpH (pH_final – pH_initial) Vs. pH_initial using salt addition technique defined by Prahas et al. (2008).

2.4

2.4 Pb(II) and Cd(II) adsorption experiments

2.4.1

2.4.1 Adsorption tests

The batch experiments for of Pb(II) and Cd(II) removal from aquatic media were performed by using the synthesized Ash/GO/Fe₃O₄ nanocomposite. In a typical test, 100 mL of the reaction solution containing definite concentrations of heavy metal ions (5–70 mg L⁻¹) and adsorbent (0.25–5 g L⁻¹) was prepared. After adjusting the pH and temperature at specified values, the tests were started and samples were taken at different times of treatment process. At the end of the reaction time, the samples were separated from the liquid phase by using a magnet. The concentrations of ions were measured by a flame atomic absorption spectroscopy (FAAS, Shimadzu AAS-6300 model) analyzer. The effect of the pH parameter on the efficiency of the Pb(II) and Cd(II) adsorption process was explored by changing the pH of the solution in the range of 2–9, in the presence of 1 g L⁻¹ of the magnetic nanocomposite at room temperature for 60 min. The removal efficiency of the adsorption (R(%)) and adsorption capacity of the adsorbent (q_e, mg g⁻¹) were evaluated according to Eqs. (1) and (2), respectively:

(1)

R (%) = \frac{C_{i} - C_{e}}{C_{i}} \times 100

(2)

q_{e} = \frac{C_{i} - C_{e}}{W} \times V

In these equations, C_i and C_e are the initial and final concentration (mg L⁻¹) of ions in the aqueous media. The amount of the adsorbent nanocomposite (g) and solution volume (L) are presented by W and V, respectively.

2.4.2

2.4.2 The nanocomposite desorption and reusability tests

To perform the desorption test, 1 g L⁻¹ of the prepared Ash/GO/Fe₃O₄ magnetic adsorbent was mixed with 100 mL of metal ions solution (5 mg L⁻¹) at optimized conditions. After adsorption experiments, the adsorbed magnetic nanocomposite was removed from the aqueous phase and rinsed in order to eliminate the deposited ions (Cd(II) and Pb(II)). Then the separated nanocomposite was mixed with 20 mL HNO₃ solution (0.7 M) at different treatment times (30–150 min). Finally, the samples were collected from liquid solution and the Pb(II) and Cd(II) concentrations were measured. The desorption efficiency of the nanocomposite was calculated from Eq. (3):

(3)

% of D e s o r p t i o n = (\frac{amount o f P b (I I) a n d C d (I I) d e s o r b e d}{amount o f P b (I I) a n d C d (I I) a d s o r b e d}) \times 100

3 Results and discussion

3.1

3.1 Ash/GO/Fe₃O₄ nanocomposite characterization

The presence of surface functional groups on the prepared Ash, GO and Ash/GO/Fe₃O₄ and also the changes on the surface functional groups of the nanocomposite after adsorption process was determined using FTIR. Fig. 1(a) shows the FTIR spectrum of the ash, GO and synthesized nanocomposite (before and after adsorption process). The broad adsorption peaks around 3379, 3421 and 3400 cm⁻¹ were observed for GO, ash and Ash/GO/Fe₃O₄ nanocomposite, which were related to O—H stretching vibrations of the H₂O molecule (Farooq and Jalees, 2020; Pelalak and Heidari, 2014; Pelalak et al., 2020 ). The observed peaks at about 2900–2874 cm⁻¹ could be related to the C—H bonds. The adsorption peaks at 1050, 1220, 1573, and 1721 cm⁻¹ in the FTIR spectrum of the GO were corresponded to the C—O—C, C—OH, C⚌C from unoxidized sp² CC bonds, and C⚌O stretching vibration due to carboxylic acids, respectively (Pelalak et al., 2013; Çiplak et al., 2015; Chaiyakun et al., 2012 ). In the structure of the produced Ash, the adsorption peaks at the 1438, 871, 710 and 606 cm⁻¹, which were related to the O—C—O, Si—O, Si—O—Si and Al—O—Si stretching vibration, respectively (Mosoarca et al., 2020; Heidari et al., 2020). The adsorption bonds related to the GO and Ash was discovered in the FTIR spectrum of the Ash/GO/Fe₃O₄ composite, which indicated the successful synthesis of the magnetic nanocomposite. An intense peak at wavenumber of the 578 and 582, cm⁻¹ was detected in the FTIR spectrum of the Ash/GO/Fe₃O₄ which could be related to the Fe—O stretching vibration of the Fe₃O₄ (Lopez et al., 2010; Pelalak et al., 2021). It is observed from FTIR spectrum of the Ash/GO/Fe₃O₄ adsorbent, ions adsorption created some differences in the strength and location of the adsorption peaks which can be due to the physical adsorption of the ions.

(a) The FTIR spectrum of the prepared Ash, GO and Ash/GO/Fe3O4 nanocomposite, (b) the XRD pattern of the prepared Ash, GO and Ash/GO/Fe3O4 nanocomposite, and (c) the VSM analysis of the Ash/GO/Fe3O4 nanocomposite.

The crystalline structure of the prepared Ash, GO and nanocomposite was assessed using XRD analyses and reported in Fig. 1 (b). For GO, the diffraction peaks located at 2θ values of 10.5° and 42.5° were corresponded to the characteristics (0 0 1) and (1 0 1) planes, respectively (Ganesan et al., 2020; Shi et al., 2020). The presence of quartz and calcite in the structure of the produced Ash was confirmed at 2θ values of 23.3°, 29.6°, 36.2°, 39.65° and 43.4°, respectively (Akanyeti et al., 2020). The observed diffraction peaks located at 2θ values of 35.65°; 53.75° and 62.9° in the XRD pattern of the Ash/GO/Fe₃O₄ nanocomposite were related to the characteristics of (3 1 1), (4 4 2) and (4 4 0) planes, respectively, indicating the successful loading of the Fe₃O₄ nanoparticles in the nanocomposite structure (Foroutan et al., 2018). The XRD spectra of the generated Ash and nanocomposite proved the crystalline structure.

The magnetic behavior of the magnetic Ash/GO/Fe₃O₄ nanocomposite was determined using Vibration Magnetometer Analysis (VSM) in the range of −10,000 Oe to 10,000 Oe as shown in Fig. 1(c). The magnetic saturation for the Ash/GO/Fe₃O₄ magnetic composite was achieved as 36.842 emu/g. The obtained VSM results confirmed the superparamagnetic behavior of the produced nanocomposite, which enables easy separation of the used nanocomposite from the aqueous media using a magnet. The inset of Fig. 1(c) shows that Ash/GO/Fe₃O₄ composite can be attracted by an external magnet, and the clear solution can be easily removed by pipette or decanting.

The surface characteristics, including surface area, pore volume and mean pore diameter of the Ash, GO and Ash/GO/Fe₃O₄ nanocomposite were determined using BET analysis which were obtained from N₂ adsorption/desorption analysis (Fig. 2(a)–(c)). As the obtained isotherms show according to IUPAC (International Union of Pure and Applied Chemistry) classification the type IV isotherm with the hysteresis loop can be detected. The hysteresis loops of adsorbent suggest the H4 form demonstrates narrow slit-like pores. The S_BET for the prepared Ash, GO and Ash/GO/Fe₃O₄ samples was obtained as 7.7, 13.2 and 85.8 m²/g, respectively. Furthermore, the total pore volume was measured as 0.069, 0.039 and 0.395 cm³/g for Ash, GO and Ash/GO/Fe₃O₄ nanocomposite, respectively. The S_BET and pore volume results highlighted the significant increase in surface properties of the nanocomposite, which are the most important characteristics of an adsorbent. The mean pore size values of the prepared samples proved the porous structure of the produced Ash, GO and nanocomposite with meso-pores (Bonyadi et al., 2019). Moreover, the pH_PZC of Ash/GO/Fe₃O₄ nanocomposite was obtained 5.6 as shown in Fig. 2(d). The pH_PZC for a given surface is usually described as the pH at which that surface charge is neutral. At pH values above and below the pH_PZC value, the surface of samples becomes negatively and positively charged, respectively.

The nitrogen adsorption/desorption isotherms of the (a) ash, (b) GO (c) ash/GO/Fe3O4, and (d) the pHPZC of Ash/GO/Fe3O4 nanocomposite.

The SEM and EDX analysis of the Ash, GO and Ash/GO/Fe₃O₄ nanocomposite which was conducted to investigate the morphological and structural properties of the prepared samples were reported in Fig. 3. The SEM (Fig. 3(a)–(d)) and EDX (Fig. 4(a)–(c)) analysis of the GO revealed the layered structure of the GO composing of 66.34 w% of C and 33.66 w% of O. The observations in the SEM images of the produced Ash (Fig. 3(e)–(h)) confirmed the heterogeneous structure and presence of the cracks which would be useful in embedding of the GO nanosheets or Fe₃O₄ nanoparticles. Furthermore, the EDX analysis of the produced oak wood Ash (Fig. 4(d–f)) confirmed the homogenous distribution of the various elements in the structure of Ash. The SEM analysis from the surface of the synthesized magnetic nanocomposite (Fig. 3(i)–(l)) shows the nanoparticles with various sizes, which could be related to the presence of the Fe₃O₄ nanoparticles. The existence of the Fe ions in the EDX analysis of the magnetic nanocomposite (Fig. 4(g)–(i)) indicated the successful embedding of the Fe₃O₄ nanoparticles.

The surface SEM images of the (a–d) GO, (e–h) ash, and (i–l) Ash/GO/Fe3O4 nanocomposite.

The EDS analysis of the (a–c) GO, (d–f) Ash, (g–i) Ash/GO/Fe3O4. nanocomposite, Ash/GO/Fe3O4 nanocomposite after adsorption of the (j–l) Pb(II) and (m–o) Cd(II) ions, respectively.

The change occurred on the nanocomposite surface after adsorption of the Pb(II) and Cd(II) ions was monitored using FESEM (Fig. 4(j) and (m)) and EDX dot (Fig. 4(k) and (n)) and map (Fig. 4(l) and (o)) analysis of the used nanocomposite. According to the obvious results in these figures the presence of heavy metal ions in the structure of the nanocomposite were confirmed. This demonstrates the ability of the nanocomposite for heavy metal adsorption from solution.

Fig. 5(a) and (b) show the TEM micrographs of the as-prepared graphene oxide and the Ash/GO/Fe₃O₄ nanocomposite, respectively. It can be inferred from the outcomes that precipitation method provides a uniform distribution of Fe₃O₄ particles without agglomeration, on the surface of support. Moreover, the size of synthesized samples were very small and less than 20 nm. As can be seen, the hexagonal arrays of Fe₃O₄ as well as parallel channels of mesopores are clearly observed in Fig. 5(c). The obtained results from TEM analysis were in accordance with the FESEM and BET results.

The TEM analysis of the (a) GO, and (b, c) the Ash/GO/Fe3O4 nanocomposite.

3.2

3.2 Influence of the operational parameters on the Pb(II) and Cd(II) ions adsorption by Ash/GO/Fe₃O₄ nanocomposite

3.2.1

3.2.1 Impact of the pH and concentration of the nanocomposite

The pH value of solution is an important parameter in ions adsorption, since this parameter influences the surface charge of the adsorbent. Fig. 6 shows the heavy metal ions removal from aqueous solution as a function of solution pH. As can be seen the ions removal in extremely acidic medium was low. This can be due to the protonated surface functional groups of the nanocomposite resulted in the repulsion forces between the Pb(II) and Cd(II) ions. At pH values lower than pH_PZC the positively charged surface of adsorbent inhibit the Pb(II) and Cd(II) adsorption, consequently reduce the adsorption capacity of the nanocomposite (Foroutan et al., 2020). The removal of Pb(II) and Cd(II) ions was increased rapidly from pH 2 to 6. About 43% of Pb(II) and 50% of Cd(II) ions was removed at pH value of 2, while almost total removal of ions was observed at pH = 6. By increasing the pH value the adsorbent surface becomes more deprotonated hence the Pb(II) and Cd(II) adsorption will be increased. At pH values higher than pH_PZC the surface will be negatively charged which improves the cationic ions adsorption. The formation of the strong attractions between the present OH⁻ ions on the surface of the nanocomposite and Pb(II) and Cd(II) ions, led to enhanced adsorption rate up to pH = 6. However, in higher pH values the removal efficiency of heavy metal ions was gradually decreased, which can be related to the formation of complex between adsorbent particles and metal ions which restricted the adsorption of the ions (Huang et al., 2020; Kataria and Garg, 2018). Also, in higher pH values the free Pb(II) and Cd(II) ions can begin to precipitate and form Pb(OH)₂ and Cd(OH)₂ (Wang et al., 2013; Lv et al., 2005) which affect the adsorption process. According to the above mentioned, pH = 6 was selected as the optimum pH value for the following experiments.

the effect of (a) the pH of solution, and (b) the adsorbent dosage on the Pb(II) and Cd(II) ions removal using Ash/GO/Fe3O4 nanocomposite.

The concentration of the adsorbent is one of the key variables affecting the efficiency of process and adsorption capacity. In this section, the impact of the various amounts of the synthesized Ash/GO/Fe₃O₄ varying from 0.25 to 5 g L⁻¹ was studied on the adsorption efficiency of the Pb(II) and Cd(II) ions. As shown in Fig. 6(b), the rate of the adsorption process was increased as a result of the increase in nanocomposite loading up to 1 g L⁻¹. This increment could be related to the enhancement in the number of the active adsorbent sites for ions adsorption (Abshirini et al., 2019). Further increase in concentration of the magnetic nanocomposite more than 1 g L⁻¹ leads to the decreased adsorption rate of process. The observed reduction in the ions adsorption could be related to the decline of ions concentration in the solution and also the reduced driving force of the mass transfer from liquid to solid phase. Furthermore, the accumulation of the adsorbent particles at higher concentrations resulted in reduction of active surface for adsorption process (Savari et al., 2020). Based on the obtained results, the adsorbent concentration of 1 g L⁻¹ was considered as the optimal concentration.

3.2.2

3.2.2 The influence of process time and kinetic studies

Fig. 7 shows the adsorption time dependency for Pb(II) and Cd(II) adsorption on Ash/GO/Fe₃O₄ nanocomposite at contact times between 5 and 150 min. The presented results in Fig. 7, confirmed that adsorption of both ions on the surface of the Ash/GO/Fe₃O₄ nanocomposite occurred at two steps. The first step of the adsorption process occurred faster between 5 and 60 min of reaction. This could be because of the free adsorption sites on the Ash/GO/Fe₃O₄ surface and higher concentration gradient between liquid and solid phase (Fig. 7(a)). On the opposite way by increasing the reaction time more than 60 min, the adsorption rate of the ions on the surface decreased which was corresponded to the saturation of the nanocomposite surface by the adsorbed metal ions. Furthermore, the repulsion effect between the surface adsorbed ions and the other ions occurred in solution and consequently decreased the adsorption efficiency (Chidi and Kelvin, 2018).

(a) The contact time dependency, plots of (b) pseudo first order, (c) pseudo second order and (d) intraparticle diffusion kinetic model, and for Pb(II) and Cd(II) heavy metal adsorption on Ash/GO/Fe3O4 nanocomposite (pH = 6, initial concentration of ions = 10 mg L−1, adsorbent dosage = 1 g L−1, T = 25 °C and stirring speed of 600 rpm).

The kinetics of the heavy metal ions adsorption on the nanocomposite were investigated using pseudo first (PFO) and second (PSO) order and intraparticle diffusion (IPD) kinetic models. The PFO, PSO and IPD models can be calculated from Eqs. (4)–(6), respectively:

(4)

{ln(q}_{e} - q_{t}) = {lnq}_{e .cal} - K_{1} t

(5)

\frac{t}{q_{t}} = \frac{1}{K_{2} q_{e.cal}^{2}} + \frac{t}{q_{e.cal}}

(6)

q_t = K_id t^1/2 + C in this equation, q_e and q_t represent the adsorbed ions amount (mg g⁻¹) at equilibrium and time t of process (min), respectively. The kinetic rate constants of the IPD, PFO and PSO were provided by k_id (mg/g min ^−1/2), K₁ (min⁻¹) and K₂ (g mg⁻¹min⁻¹), respectively. C (mg g⁻¹) is the intercept showing the thickness of boundary layer effect. Table 1 reports the calculated values from Eqs. (4)–(6). The R² value for PFO and PSO models were obtained as 96.39% and 99.65%, respectively, indicating more susceptibility of the PSO model to predict the kinetic properties of the Cd(II) and Pb(II) adsorption by magnetic nanocomposite which is in accordance with the data shown in Fig. 7(b) and (c). Furthermore, the obtained results indicated that titled ions adsorbed both physically and chemically on the surface of the synthesized Ash/GO/Fe₃O₄ nanocomposite (Thirumoorthy and Krishna, 2020). Moreover, in order to study the diffusion mechanism of the ion adsorption on the nanocomposite the IPD model was used (Ozdes et al., 2014). The kinetic parameters and regression coefficients of IPD model were calculated from plot of q_t versus t (Fig. 7(d)). In Eq. (6) if C is zero and linear plot passing through the origin it can be said that the IPD mechanism is only rate limiting one. If C value is not zero, the separation is controlled by film diffusion. As can be seen, IPD model was obtained nonlinearly and it can be concluded that in the adsorption of heavy metal ions on the Ash/GO/Fe₃O₄ nanocomposite contained more than one type of effective mechanism. The first sharper portion represents the external surface adsorption (film diffusion). This section suggests that Pb (II) and Cd (II) transferred towards film or boundary layer of nanocomposite and this step takes place at high speed. Second linear section represents the penetration of heavy metal ions into the nanocomposite layers. In this portion the adsorption rate is very slow and the intraparticle diffusion is rate-controlled. Comparing the calculated values of k_id of first and second stages it can be said that contrary to the initial high adsorption rate, the rate of process diminished over time indicating that the rate limiting step is the second stage.

Table 1 The kinetic models parameters for Pb(II) and Cd(II) adsorption on Ash/GO/Fe₃O₄ nanocomposite.

Kinetic model	Pb(II)	Cd(II)
Pseudo-first order
q_{e cal}	10.30	16.410
K₁ (min⁻¹)	0.0656	0.072
R²	0.9639	0.953
Pseudo-second order
q_e.cal	10.830	11.185
K₂ (g/mg min⁻¹)	0.0093	0.005
R²	0.9965	0.993
Intraparticle diffusion
Stage (I)
K_id (mg/g min^−1/2)	1.186	1.257
R²	0.980	0.990
C (II)	1.119	0.057
Stage (II)
K_id (mg/g min^−1/2)	0.026	0.015
R²	0.901	0.924
C (I)	9.818	9.780

3.2.3

3.2.3 Influence of the initial Pb(II) and Cd(II) concentration and isotherm studies

In order to evaluate the influence of initial concentration of contaminants, the adsorption efficiency was explored in different concentrations of 5–70 mg L⁻¹ and the outcomes are reported in Fig. 8. As evidenced by the results, the adsorption process was favored at lower concentrations of the ions. The increment in the Pb(II) and Cd(II) concentration led to reduction in adsorption rate. The adsorbed ions occupied the active sites and therefore a repulsion force is created between the adsorbent surface and the presented ions by increasing the initial concentrations of contaminant. In fact, the limited number of the adsorbent particles toward large number of the available resulted in decreased removal efficiency. It is evident from Fig. 8(a) and (b), that the higher ions concentration improved the capacity of the nanocomposite for adsorption. It can be said that the difference between concentration of the Pb(II) and Cd(II) ions on the surface of the nanocomposite and metals provide a driving force at higher loading content of the metal ions, which elevates the adsorption rate. The higher concentration gradient of ions caused to the formation of attraction between metal ions with surface of the Ash/GO/Fe₃O₄ nanocomposite (Vimala and Das, 2009). The adsorption isotherms reveal important information about adsorption capacity, adsorption mechanism between the contaminant and the adsorbent and the contaminant distribution between the adsorbent and the solution (Arabkhani and Asfaram, 2020). Herein, the adsorption behavior of the ions adsorption on the synthesized Ash/GO/Fe₃O₄ magnetic composite was studied. Langmuir and Freundlich isotherm models were used according to Eqs. (7) and (8), respectively and the obtained results are exhibited in Fig. 8(c) and (d).

(7)

\frac{C_{e}}{q_{e}} = \frac{C_{e}}{q_{m a x}} + \frac{1}{q_{m a x} K_{L}}

(8)

l o g q_{e} = l o g K_{f} + \frac{1}{n} l o g C_{e}

where q_max and K_L are the adsorbent maximum adsorption capacity (mmol/g) and the constant value of the Langmuir model (L/mmol), respectively. K_F and n in Eq. (8) present the experimental constants of Freundlich model. For analyzing the performance of the isotherm models, the statistical factors including coefficient of determination (R²), root-mean-square error (RMSE) and mean-absolute error (MAE) were calculated to assess the model’s accuracy. The equations are expressed as:

(9)

R^{2} = \frac{\sum_{i = 1}^{n} {({\bar{q}}_{\exp} - q_{cal})}^{2} - \sum_{i = 1}^{n} {(q_{\exp} - q_{cal})}^{2}}{\sum_{i = 1}^{n} {({\bar{q}}_{\exp} - q_{cal})}^{2}}

(10)

RMSE = {[\frac{\sum_{i = 1}^{n} {(q_{\exp} - q_{cal})}^{2}}{n}]}^{1 / 2}

(11)

MAE = (\frac{1}{n}) \sum_{i = 1}^{n} |q_{\exp} - q_{cal}|

The effect of the initial Pb(II) and Cd(II) concentration on (a) the removal rate, and (b) the adsorption capacity of the Ash/GO/Fe3O4, the isotherms plots for (c) Langmuir and (d) Freundlich for Pb(II) and Cd(II) heavy metal adsorption on Ash/GO/Fe3O4 nanocomposite.

Table 3 reports the calculated constant and variable parameters of the both models for Pb(II) and Cd(II) adsorption. As can be seen in Table 2, the values of the RMSE and MAE for the Langmuir model for Pb(II) ions were achieved as 3.26 and 2.58, respectively. These parameters for Freundlich in Pb(II) removal were obtained as 4.97 and 4.37, respectively. Moreover, the regression coefficient for Langmuir and Freundlich models were attained as 0.99 and 0.81 for Pb(II), and 0.99 and 0.96 for Cd(II) ions, respectively. These outcomes demonstrate that the Langmuir model more properly predict the process behavior compared to the Freundlich model for both studied ions. It can be said that the homogeneous surface of the Ash/GO/Fe₃O₄ nanocomposite were responsible for adsorption process. The calculated R_L and n parameters are in the desired range. The highest adsorption capacity of the Pb(II) and Cd(II) on the nanocomposite surface were obtained as 47.16 and 43.66 mg/g, respectively, which prove the effective adsorption ability of the generated magnetic nanocomposite.

Table 2 The Langmuir and Freundlich isotherms constants Pb(II) and Cd(II) heavy metal ions adsorption on Ash/GO/Fe₃O₄ nanocomposite.

Models	Parameters	Pb(II)	Cd(II)
Langmuir	q_m(mg/g)	47.16	43.66
	K_L (L/mg)	0.98	0.81
	R_L	0.014–0.168	0.017–0.197
	R²	0.99	0.99
	RMSE	3.26	2.62
	MAE	2.58	2.18

Freundlich	n	2.58	2.78
	K_f (mg/g (L/mg)^1/n)	16.21	15.75
	R²	0.81	0.96
	RMSE	4.97	4.32
	MAE	4.37	2.69

Table 3 The calculated thermodynamic parameters for adsorption of the Pb(II) and Cd(II) ions on the surface of Ash/GO/Fe₃O₄ nanocomposite.

Pollutant	T (^oC)	ΔG° (KJ/mol)	ΔH° (KJ/mol)	ΔS° (J/mol K)
Pb(II)	25	−10.44	−59.54	−163.51
	30	−10.43
	35	−9.09
	40	−8.61
	45	−7.63
	50	−6.44

Cd(II)	25	−9.00	−46.45	−126.32
	30	−8.41
	35	−7.16
	40	−6.55
	45	−6.23
	50	−6.06

3.2.4

3.2.4 Impact of the adsorption temperature and thermodynamic studies

Temperature is another influencing key variable on the remediation efficiency of the process. For this purpose, the impact of the operating temperature was studied on the removal efficiency of Pb(II) and Cd(II) ions by Ash/GO/Fe₃O₄ and reported in Fig. 9(a). The adsorption tests were performed at pH = 6, [Ash/GO/Fe₃O₄] = 1 g L⁻¹, [Pb(II)] = [Cd(II)] = 10 mg L⁻¹ and process time = 60 min with agitation speed of 600 rpm. It is obvious from the results that increasing the temperature adversely affected the rate of the adsorption process, confirming the exothermic nature of the process. The observed decline in the process efficiency could be linked to the changes and shrinkage of the adsorbent particles at higher temperature, which lead to decrease the number of the active sites. Furthermore, the adsorbed ions on the surface of the nanocomposite particles could be detached from surface by increasing temperature (Shafiee et al., 2019). The highest removal of 98.54 and 97.42% was obtained for Pb(II) and Cd(II) ions, respectively, at temperature of 25 °C. The thermodynamic aspect of the adsorption of the Pb(II) and Cd(II) heavy metal ions on the synthesized Ash/GO/Fe₃O₄ nanocomposite was investigated to evaluate the endothermic or exothermic nature of the process in the temperature range of the 25–50 °C. Thermodynamic parameters including Gibb’s free energy (ΔG), enthalpy (ΔH), entropy (ΔS) are obtained through Eqs. (12) and (13):

(12)

Δ G^{o} = - {RT ln K}_{c}, K_{c} = \frac{q_{e}}{C_{e}}

(13)

{ln K}_{c} = \frac{Δ S^{o}}{R} - \frac{Δ H^{o}}{RT}

in these Eqs. R is the constant of ideal gas, T is the absolute temperature (K) and k_c is the equilibrium constant. Fig. 9(b) shows the linear relation between ln K_c against 1/T and the calculated ΔG°, ΔH° and ΔS° are reported in Table 3. The obtained negative values for both Cd(II) and Pb(II) ions showed that the ions adsorption on the magnetic Ash/GO/Fe₃O₄ nanocomposite thermodynamically effective and spontaneous. The exothermic nature of the adsorption of Pb(II) and Cd(II) ions on the nanocomposite was proved by the ΔH^o values which obtained as −59.548 and −46.459 KJ/mol for Pb(II) and Cd(II) ions, respectively (Esvandi et al., 2020).

(a) The changes in the removal efficiency of the Pb(II) and Cd(II) ions by varying the adsorption temperature, and (b) the adsorption thermodynamics plots for Pb(II) and Cd(II) ions (pH = 6, initial concentration of ions = 10 mg L−1, adsorbent dosage = 1 g L−1 and stirring speed of 600 rpm).

Besides, the values of ΔS° for Pb(II) and Cd(II) ions at the mentioned temperatures were obtained as −163.519 and −126.323 J/mol K. These negative values confirmed that the irregularity or randomness is reduced in the liquid-solid interface.

3.3

3.3 Nanocomposite desorption and reusability studies

The desorption capability of the used Ash/GO/Fe₃O₄ nanocomposite was assessed by washing the surface of adsorbed nanocomposite by HNO₃ solution with concentration of 0.7 M at treatment times between 30 and 150 min. The results presented at Fig. 10(a), indicate that an increment in treatment time led to decrease in the desorption efficiency. The optimum treatment time for desorption was obtained as 120 min and the increase in desorption time lead to insignificant change in the efficiency of desorption process. The observed insignificant changes with increased time up to 120 min could be related to formation of the intense bonds between the Pb(II) and Cd(II) ions to surface sites of the magnetic nanocomposite. The reusability test of the nanocomposite was conducted within eight repetitive cycles and the outcomes are depicted in Fig. 10(b). As can be seen, with increasing the number of cycles the capacity of the nanocomposite for ions adsorption from aquatic media was reduced. The detected decline could be related to irreversible adsorption of the ions on the active sites of the Ash/GO/Fe₃O₄ nanocomposite and/or destruction of active binding sites of adsorbent during adsorption and desorption tests. It should be mentioned that the observed decrease in the efficiency of the synthesized nanocomposite was insignificant (about 5% after 8 cycles), indicating the good performance of the nanocomposite for adsorption of the heavy metals.

(a) The desorption tests, and (b) reusability of the used Ash/GO/Fe3O4 nanocomposite.

3.4

3.4 Comparison the removal efficiency of synthesized Ash/GO/Fe₃O₄ nanocomposite for Pb(II) and Cd(II) ions from aqueous solution with other research studies

The adsorption capacity of different adsorbents in Pb(II) and Cd(II) ions removal from aqueous solution were summarized in Table 4. As can be observed, the synthesized Ash/GO/Fe₃O₄ nanocomposite showed adsorption capacity of 47.16 and 43.66 mg/g for Pb(II) and Cd(II) ions, respectively, which is higher compared to other reported studies.

Table 4 The adsorption capacity of various adsorbents for Cd (II) and Pb(II) ions.

Adsorbent	pH	Metals	Q_max (mg/g)	Ref.
Activated carbon	6.0	Pb(II)	22.8	Kobya et al. (2005)
Graphene oxide	5.6	Pb(II)	36.0	Lee and Yang (2012)
Magnetic oak wood char	5.0	Pb(II)	10.1	Mohan et al. (2014)
Fly ash	6.4	Pb(II)	20.55	Alinnor (2007)
Fe₃O₄ nanoadsorbents	5.5	Pb(II)	36.0	Nassar (2010)
Magnetic oak bark char	5.0	Pb(II)	30.2	Mohan et al. (2014)
Magnetic ChNTs	5.5	Pb(II)	23.7	Yu et al. (2015)
Olive branches AC	5.0	Pb(II)	41.3	Alkherraz et al. (2020)
Waste of Myrica esculenta	6.0	Pb(II)	8.3	Joshi et al. (2018)
Mustard husk	6.0	Pb(II)	30.4	Meena et al. (2008)
Ash/GO/Fe₃O₄	6.0	Pb(II)	47.1	This study
Activated carbon	5.0	Cd(II)	30.6	Kobya et al. (2005)
Magnetic oak wood char	5.0	Cd(II)	2.9	Mohan et al. (2014)
Magnetic oak bark char	5.0	Cd(II)	7.4	Mohan et al. (2014)
Magnetic ChNTs	5.5	Cd(II)	23.8	Yu et al. (2015)
Olive branches AC	5.0	Cd(II)	38.17	Alkherraz et al., (2020)
Waste of Myrica	6.0	Cd(II)	5.61	Joshi et al. (2018)
Mustard husk	6.0	Cd(II)	42.8	Meena et al. (2008)
Ash/GO/Fe₃O₄	6.0	Cd(II)	43.6	This study

3.5

3.5 Adsorption mechanism

Mechanism of metal ions adsorption on the surface of adsorbent depends on several factors including the surface chemistry of solid adsorbent, functional groups, surface charge and solution pH. Adsorption of metal ions on the adsorbent can occur through electrostatic interaction, precipitation of ions, ion-exchange and/or formation of complex. Here, the mechanism of metal ions adsorption is evaluated using results of characterization analyses and the pH of solution. The results of FTIR revealed that some functional groups such as —OH, —COO, and C—O were available in the structure of the nanocomposite. These functional groups in extremely acidic pH values (pH < pH_PZC) can be converted to —OH₂⁺ and —COOH species therefore ionization reduced. In this condition, surface became protonated and repulsive electrostatic forces were formed between the nanocomposites surface and the heavy metal ions (Fig. 11) (Zou et al., 2019). Hence, lower removal efficiency of heavy metal ions was achieved in this condition. At higher pH values the nanocomposite surface become more deprotonated and heavy metal ions encouraged to be adsorbed. The —COO and —OH groups were converted to —COO⁻ and —O⁻ groups on the nanocomposite surface. The negative charged surface react with Pb(II) and Cd(II) and consequently removal efficiency was increased. Possible reactions and mechanism of metals adsorption in basic solution are shown in Fig. 12. The hydroxyl and oxygen groups on the surface of nanocomposite provide electrostatic interactions with cationic ions (Pb(II) and Cd(II) ions) at pH value of 6 (Fig. 12(a)). As mentioned before, in higher pH values the removal efficiency of heavy metal ions was gradually decreased, which can be related to the formation of complex between adsorbent particles and metal ions which restricted the adsorption of the ions (Huang et al., 2020; Kataria and Garg, 2018) (Fig. 12(b)). Also, by increasing the solution pH values the metal ions began to precipitate and form Pb(OH)₂ and Cd(OH)₂ (Wang et al., 2013; Lv et al., 2005) which affect the adsorption process. Similar results have been reported by other studies on the adsorption of heavy metal ions on ash (Mohan et al., 2007).

The mechanisms of Cd(II) and Pb(II) adsorption onto Ash/GO/Fe3O4 nanocomposite.

Possible schematic mechanism of Cd(II) and Pb(II) interaction with nanocomposite in basic solution.

4 Conclusion

In this study, the as-synthesized magnetic Oak wood Ash/GO/Fe₃O₄ nanocomposite was used as an efficient adsorbent for removal of hazardous Pb(II) and Cd(II) heavy metal ions from aquatic media. The chemical precipitation technique was applied in order to synthesize the Oak wood Ash/GO/Fe₃O₄ nanocomposite. The chemical, structural and magnetic features of the synthesized samples were investigated by FESEM, TEM, XRD, BET/BJH, FTIR and VSM techniques. The results of the XRD and VSM analysis indicated the crystalline structure and paramagnetic behavior of the produced Oak wood Ash/GO/Fe₃O₄ nanocomposite, which enables the easy separation from the aquatic solution. The surface area and mean pore diameter of the produced nanocomposite were attained as 85.806 m² g⁻¹ and 18.445 nm, respectively, indicating high capability of the samples for adsorption. The impact of the various operating variables was investigated on the adsorption efficiency including the pH, treatment time, adsorbent content, and the initial Pb(II) and Cd(II) concentration. The results of the isotherm studies showed that Langmuir model can properly predict the adsorption equilibrium data. The highest adsorption percentage of the Ash/GO/Fe₃O₄ for Pb(II) and Cd(II) ions were obtained as 99.67 and 98.68%, respectively. The kinetic observation indicated the pseudo second order behavior of the adsorption process. Finally, the reusability test of the as-prepared nanocomposite showed the insignificant changes in the adsorption capacity of the adsorbent after 8 repetitive cycles.

Acknowledgements

This work was supported by the Government of the Russian Federation (Act 211, contract 02.A03.21.0011) and by the Ministry of Science and Higher Education of Russia (grant FENU-2020-0019).

Declaration of Competing Interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

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Introduction

Materials and methods

Materials
Synthesis of the nanocomposite adsorbent
Nanocomposite characterizations
Pb(II) and Cd(II) adsorption experiments

Results and discussion

Ash/GO/Fe3O4 nanocomposite characterization
Influence of the operational parameters on the Pb(II) and Cd(II) ions adsorption by Ash/GO/Fe3O4 nanocomposite
Nanocomposite desorption and reusability studies
Comparison the removal efficiency of synthesized Ash/GO/Fe3O4 nanocomposite for Pb(II) and Cd(II) ions from aqueous solution with other research studies
Adsorption mechanism

Conclusion

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