Preparation of ZnMgAl-Layered Double Hydroxide and Rice Husk Biochar Composites for Cu(II) and Pb(II) Ions Removal from Synthetic Wastewater

Abstract

The efficiency of a new composite material of the layered double hydroxide (LDH) of ZnMgAl and rice husk biochar (RHB) for the removal of Cu(II) and Pb(II) ions from synthetic wastewater was investigated in this study. The images of the scanning electron microscope showed extremely fine crystalline LDH particles decorated on the rough surface of the RHB, while the successful formation of the composite adsorbent (LDH/RHB) was confirmed by the corresponding energy dispersive X-ray and the Fourier-transform infrared spectroscopy. An equilibrium contact time of 30 and 15 min for Cu²⁺ and Pb²⁺, respectively, was proposed for the optimum performance of the batch adsorption process. The dose of the LDH/RHB adsorbent was optimized at 0.4 g L⁻¹ yielding maximum adsorption capacities of 117 and 124 mg g⁻¹ for Cu²⁺ and Pb²⁺, respectively, with corresponding maximum removal efficiencies of nearly 94% and 99%. A solution pH of 6.0 yielded optimum results with an increasing trend in adsorption capacities and percentage removal by changing the solution pH from 2.0 to 7.0. Based on the best fit of the pseudo-second-order kinetic model to the experimental data, chemisorption was suggested to be the controlling mechanism of adsorption. The fitting of the Langmuir model suggested a monolayer sorption of Cu²⁺ and Pb²⁺, and the application of the Dubinin–Radushkevich isotherm proposed physical adsorption.

1. Introduction

Rising concentrations of toxic pollutants and chemicals in surface and other available freshwater reservoirs have become a major concern worldwide simply because of the expansion of industrial and anthropogenic activities [1]. Heavy metals with higher concentrations, such as copper (Cu) and lead (Pb), are of particular concern because these are predominantly used in industry and are extremely detrimental to human health and aquatic creatures [2]. In humans, acute Pb poisoning has devastating effects on the kidneys, brain, liver, and central nervous system. Copper is undoubtedly a vital trace element for all living things, but its presence in large amounts could be harmful to humans and may result in nausea, vomiting, and abdominal discomfort [3,4]. Thus, it is imperative that efficient methods and legislation must be developed to regulate the untreated discharge of industrial wastewater into freshwater reservoirs to safeguard humanity from these possible harmful impacts.

The development of innovative technologies to purify industrial effluent before it is released into waterways is required. Microbial (microalgae) fuel cell technology, membrane technology, and solar irradiation are some of the new wastewater treatment technologies that have yet to mature [5,6,7]. Reverse osmosis, chemical precipitation, ion exchange, and adsorption are just a few well-known technologies that have been frequently used in order to eradicate heavy metals from industrial wastewater [8,9]. Thus far, adsorption is the best and most practical technology due to its fast action, cost-effectiveness, simple design, low energy consumption, and excellent ability to eliminate a variety of hazardous contaminants from industrial effluent [10,11,12]. On the other hand, adsorption has some disadvantages; for example, it requires regeneration of the adsorbent and weak selectivity. In the adsorption process, the selection of adsorbent plays an important role in achieving maximum efficiency [13]. In recent years, biochar (BC), activated carbon, silica-based sorbents, metal–organic frameworks (MOFs), clay composites, natural polymer, resins, montmorillonite, nanomaterials, and layered double hydroxides (LDHs) have provided the best results for Cu(II) and Pb(II) ions remediation in wastewater [14,15,16,17,18,19,20,21,22,23,24,25,26].

The term “biochar” refers to a carbon-rich porous solid material made from biomass waste through pyrolysis in a reactor at moderate temperature (e.g., 400–700 °C) in the absence or presence of minimal oxygen [27]. The yield and quality of BC depend on the type of feedstock used for pyrolysis conditions due to the catalytic effects during pyrolysis. Agricultural waste biomass is one of the most valuable and renewable sources for BC production. This is because agricultural waste biomass has a higher lignin-to-cellulose ratio, which results in a high carbon byproduct [28]. For instance, rice husk is a frequently available byproduct of the rice milling business, accounting for around 20% of the whole mass of rice products [29,30]. As there is currently no appropriate use for this waste, it could be a good raw material for BC production. Sanka et al. [31] prepared BC from rice husk and claimed 90% removal of Pb(II) and other heavy metals. However, the adsorption performance of BC is still limited because of the low specific surface area and porosity [32]. Consequently, modifying and functionalizing pristine BC into superior composite materials with unique structures and surface characteristics is, thus, a key approach to diversifying its uses [33]. In the process of pursuing this goal, BC is currently being used as a substrate in composites with other superior adsorbents such as carbon nanotube, graphene, metal–organic framework, and layered double hydroxide (LDH), substantially tripling the adsorption capacity that was previously accessible individually [34,35].

The LDHs are a class of materials that are similar to hydrotalcite, an anionic clay with a positively charged host layer, and the counter anion located in the interlayer region, with the general formula [M²⁺_1−xM³⁺_x (OH⁻)₂]^x+ (Aⁿ⁻)_x/n. mH₂O, where M²⁺ and M³⁺ represent divalent and trivalent cations, respectively; Aⁿ⁻ denotes a guest interlayer charge balancing anion; and x symbolizes the proportion of trivalent metal ions, which might vary depending on the type of application [3]. LDHs are highly porous, possess unique supermolecule structures, have large specific surface areas, can withstand high temperatures, has controllable element composition, and are efficient ion exchangers [36]. Moreover, LDH-based catalysts have widely been used in the fields of environmental remediation because of their advantageous properties, such as low cost, long-lasting stability, high catalytic performance, and distinctive structural characteristics (such as intercalation, and the ability to be coupled with other useful materials) [37]. There are many synthesis methods for LDH production, each of which can modify the properties of the final product to make it more or less suitable for specific applications [38]. The most common technique is the co-precipitation method. However, poor porosity development and a high leaching rate in the reaction process have limited the production of LDHs, and it is utilized as a contaminants removal adsorbent [34]. Consequently, it is essential to discover methods of treating LDH deficiency in order to advance the development and increase their applications. Therefore, to fill this gap, porous BC is an ideal carrier matrix for the appropriate coating of LDHs, providing a large reactive surface area and preventing their aggregation [39]. As a result, combining LDHs with BC is a win–win approach for both LDHs and BC in terms of property enhancement [34].

In recent years, composite materials based on BC and LDHs have opened up new areas of investigation into optimizing and accelerating the removal of pollutants. Several studies have been conducted on the mechanism of heavy metal adsorption by BC-supported LDHs composites. For example, Tan et al. [40] prepared a composite of Zn-Fe-LDH and kiwi branch BC (KB/Zn-Fe) for the mitigation of Pb(II) ions from aqueous solution and claimed an adsorption capacity of 161.29 mg g⁻¹, compared to 36.76 mg g⁻¹ for the original BC. Jia et al. [41] synthesized a MgFe-LDH and magnetic BC for the removal of Pb(II) from an aqueous solution, and they observed a maximum adsorption capacity of 476.25 mg g⁻¹. Wang et al. [4] produced MnAl-LDH and BC composite for the attenuation of Cu(II) ions, and they observed an adsorption capacity of 74.04 mg g⁻¹. Khandaker et al. [42] prepared a composite of MgFe-LDH and bamboo waste charcoal for the remediation of Cu(II) ions from wastewater and found an adsorption capacity of 85.47 mg g⁻¹. However, there is a limited number of studies exploring the remediation mechanism of BC-supported LDHs composites for heavy metals removal from wastewater.

Thus, the specific goal of this study was to develop a new composite material of ZnMgAl(LDH) and rice husk BC (RHB) using the co-precipitation and hydrothermal technique and to investigate the removal mechanisms of Cu(II) and Pb(II) ions from synthetic wastewater. A number of batch experiments were performed to probe the effectiveness of the newly created composite material in removing the targeted heavy metals. Moreover, the kinetics and isothermal models were deployed to understand the removal mechanisms and to calculate the adsorption capacity of the composite material. Scanning electron microscopy (SEM), energy dispersive X-ray (EDX) spectroscopy, and Fourier-transform infrared spectroscopy (FTIR) techniques were employed to analyze the morphology and the presence of various functional groups on the surface of composite material.

2. Materials and Procedures

2.1. Chemicals and Batch Testing

Copper nitrate trihydrate (Cu(NO₃)₂·3H₂O) and lead nitrate (Pb(NO₃)₂, as purchased from Tianjin Benchmark, China, were used respectively to prepare the stock solutions of Cu²⁺ and Pb²⁺). The required amount of both salts was added in deionized water for this purpose, which was used further to prepare the dilutions, as required in different batch tests, in obtaining the intended initial concentration of each metal ion. A total amount of 0.1 M of analytical grade HCl and NaOH was used to set and maintain the required pH of the solution.

Two different volumes of sample solutions (50 or 100 mL) containing the required initial concentrations of Cu²⁺ or Pb²⁺ were used for the batch tests, and the desired amount of the adsorbents was added accordingly. The conical flasks containing the tested dilutions were placed inside a temperature-controlled shaker (Wise Cube orbital, Wisd. ThermoStable IS-20, Daihan Scientific Co., Ltd., Wonju, Republic of Korea) operating at 30 °C and 220 rpm. After agitating for a selected contact time, the diluted solution was withdrawn from the shaker and filtered through a 0.45 μm Whatman™ filter. A flame atomic absorption was employed for measuring the residual concentration of each metal after 5 mL of the filtered sample was inserted inside FAAS. The percentage removal and the adsorption capacity (q_t, mg g⁻¹) of either the Cu²⁺ or Pb²⁺ were measured as follows:

$$Percentageremoval(\%)=\left(\frac{C_0-C_t}{C_0}\right)\times 100$$

(1)

$$Adsorptioncapacity=\left(\frac{C_0-C_t}{m}\right)V$$

(2)

where C₀ and C_t (both measured in mg L⁻¹) are the selected initial metal concentration and the residual metal concentration, as measured using FAAS, respectively. The quantity of the adsorbent material (g) and the volume of the tested dilutions (L) is presented by measured in m and V and in Equation (2), respectively.

2.2. Preparation of the Rice Husk Biochar and the Composite Adsorbent

The waste of rice husk was collected locally to make the rice husk biochar (RHB) to be used in preparing the composite adsorbent. Prior to a three-day drying in the open (under sunlight), impurities (dust particles, etc.) were removed from the collected mass by washing multiple times with clean water. The dried mass was crushed in small pieces with an average particle size in the range of 2–3 cm, which were inserted in a muffle furnace (operating at 500 °C at a ramp rate of 5 °C/ min⁻¹) in the absence of oxygen for pyrolysis for nearly five hours. The pyrolyzed product was washed with pure water to remove the ash, dried in an oven for about a day at 100 °C, and subsequently crushed in a ball mill to obtain nano-scale particles of RHB.

The co-precipitation method and hydrothermal technique were used to prepare the composite adsorbent of ZnMgAl(LDH) and RHB. To prepare the LDH of ZnMgAl, 1.071 g (0.06 mol) of zinc nitrate hexahydrate (Fe(NO₃)₃·6H₂O), 1.846 g (0.12 mol) of magnesium nitrate hexahydrate (Mg(NO₃)₂·6H₂O), and 0.675 g (0.03 mol) of aluminum nitrate nonahydrate (Al(NO₃)₃·9H₂O) were added in 60 mL of deionized water inside a beaker. The mixture was constantly stirred for about an hour to ensure that all of the salts were completely dissolved. After complete dissolution, this solution was shifted to another beaker that contained 0.6 g of nano-sized RHB under constant stirring. Afterward, a solution containing NaOH (0.5 M) and Na₂CO₃ (0.125 M) was dripped through the burette until the solution pH reached around 10. The resulting slurry was then stirred for another two hours before transferring to a hydrothermal reactor and kept in an oven for almost 24 h at 180 °C. After cooling down, the slurry solution was transferred to a centrifuge tube operating at 5000 rpm to separate the particles. This procedure was repeated five times to wash the resulting material with deionized water. After washing, the composite material was kept in an oven for 24 h at 75 °C for drying purposes.

The characterization of the RHB, ZnMgAl(LDH), and the composite adsorbent (LDH/RHB) before adsorption and that of the composite adsorbent after adsorption of both heavy metal ions is performed using the SEM, EDX, and the FTIR techniques.

3. Results and Discussion

3.1. Adsorbents’ Properties and the Mechanism of Adsorption

3.1.1. Surface Morphology and Elemental Composition Analysis

Figure 1 shows the findings of SEM and EDX analysis performed on samples of RHB, ZnMgAl(LDH), and LDH/RHB composites. The SEM micrographs of RHB revealed an agglomerated mass, angular, dense, and flake sharply edged particles of the parent biochar, whereas EDX revealed evidence of carbon (51.17%), oxygen (35.36%), and other minerals on BC surface [43]. Similar to this, SEM micrographs of ZnMgAl(LDH) exhibited very prominent hexagonal platelet structures as well as some nanorod-like structures [3,44]. In addition, the distribution of metal elements such as carbon, oxygen, magnesium, zinc, and aluminum was precisely displayed in the EDX spectrums, indicating that the synthesized LDHs had excellent purity [45]. Similarly, highly magnified SEM micrographs of the LDH/RHB composite showed extremely fine crystalline LDH particles decorated on the rough surface of the RHB. To prove that the composite was successfully formed, the corresponding EDX spectrum of the composite displays all bonding elements that were present separately in the RHB and ZnMgAl(LDH) spectra. Therefore, the characterizations indicated that ZnMgAl(LDH) was effectively incorporated into the surface of RHB.

SEM-EDX analysis of the LDH/RHB composite after Cu²⁺ and Pb²⁺ absorption (Figure 2) showed that the composite surface became brighter and very small precipitates appeared on the surface, which could be due to the formation of Pb- and Cu-hydroxides and other metal compounds [46]. This statement was confirmed by the EDX spectra of LDH/RHB–Cu²⁺ and LDH/RHB–Pb²⁺, which showed many variations in elemental composition and the presence of Cu (3.41%) and Pb (47.63%), which may indicate successful absorption of the target metals into the LDH/RHB composite [47,48].

3.1.2. Fourier Transform Infrared Analysis

The presence and changes in functional groups on the surface of RHB, ZnMgAl(LDH), and LDH/RHB composites were investigated using FTIR spectra before and after the adsorption of Cu²⁺ and Pb²⁺, as shown in Figure 3. The wavenumber range was from 4000 to 400 cm⁻¹, where most of the changes in the FTIR spectra between analyzed samples were observed. The spectrum of RHB in Figure 3d showed a broad absorption band around 3300–3650 cm⁻¹ and a distinctive peak at 1602 cm⁻¹, which is associated with the bending and stretching vibration of O-H groups and interlayer water molecules in the tested material structure [44,49]. In addition, the RHB spectrum’s most obvious point is at 1070 and 799 cm⁻¹, which corresponds to Si-O-Si or Si-O, demonstrating the high Si content of RHB [50].

Figure 3c shows the spectra of the LDH/RHB composite, revealing that the distinctive band at about 1359 cm⁻¹ is linked to the stretching vibration mode of NO⁻³ [51,52]. Another prominent set of band shifts was noticed from 1070 to 1001 cm⁻¹ and 799 to 750 cm⁻¹ in the spectra of LDH/RHB composite, which symbolizes the stretching vibration of Si-O-Si or Si-O, but their intensity was significantly reduced compared to that obtained in Figure 3c. A new weak band appeared at 875 cm⁻¹ in the spectra of the LDH/RHB composite, which could correspond to metal-oxygen (M-O or M-O-M) vibrations [53,54]. Hence, the effective synthesis of the LDH/RHB composites is evidenced in these analyses. Furthermore, the FTIR spectrum after the adsorption of Cu²⁺ (Figure 3b) and Pb²⁺ (Figure 3a) onto the LDH/RHB composite indicated a band shift from 1001 to 1002 and 1004 cm⁻¹, respectively. More importantly, nitrate and hydroxyl representing bands may change (low intensity) or vanish, suggesting that these groups were replaced by Cu²⁺ and Pb²⁺, as illustrated in Figure 3a,b, respectively.

3.2. Estimation of the Equilibrium Contact Time and Optimization of the Parameters of Batch Adsorption

The equilibrium contact time was determined by monitoring variations in adsorption capacity and removal efficiencies in order to maximize the performance of the examined adsorption system and for effective adsorbent consumption. Figure 4 shows these changes within a suitable time range (0–300 min) for 50 and 80 mg L⁻¹ of both heavy metal ions using 0.4 g L⁻¹ of the composite adsorbent (LDH/RHB) by maintaining the solution pH at 6.0 ± 0.3. A similar trend of the rapid uptake of each heavy metal ion upon immediate contact with the adsorbent with abundant free active sites was observed irrespective of the initial concentration (50 or 80 mg L⁻¹, as shown in Figure 4a,b, respectively). For Cu²⁺, a steady increase in the adsorption capacity continued until 30 min, while the same was seen for Pb²⁺ until 15 min of the contact time. A different equilibrium contact time, thus, was proposed for Cu²⁺ and Pb²⁺ for the optimal performance of the adsorption system. Due to the unavailability or saturation of the free active sites at the surface of the LDH/RHB adsorbent, no changes in the adsorption capacity or removal efficiency were seen for Pb²⁺ after 15 min, while for 50 mg L⁻¹ of Cu²⁺, both parameters increased slightly up to 60 min of contact time (Figure 4a). The studied adsorbent exhibited slightly higher adsorption capacity and removal efficiency for Pb²⁺ in comparison to Cu²⁺ in addition to the half retention time for optimal performance. For 50 mg L⁻¹ of both heavy metal ions, the optimal values for the adsorption capacity and removal efficiency were 117 mg g⁻¹ and 94%, respectively, for Cu²⁺ at 30 min, whereas for Pb²⁺, these values were recorded as 124 mg g⁻¹ and 99%, respectively at 15 min.

Among the several parameters influencing the effectiveness of the batch adsorption process, the amount of the LDH/RHB adsorbent was customized by identifying the variations in the adsorption capacities and removal efficiencies of both heavy metal ions at various initial doses of the adsorbent (0.05–0.7 g L⁻¹), as shown in Figure 5. Batch tests were performed at an already determined equilibrium contact time of 30 and 15 min for Cu²⁺ and Pb²⁺, respectively, under a maintained solution pH of 6.0 ± 0.3 and 50 mg L⁻¹ as a constant initial concentration for each heavy metal ion. For either heavy metal ion, a steady increase in the adsorption capacity can be seen, reaching maximum values of 117 and 124 mg g⁻¹ for Cu²⁺ and Pb²⁺, respectively, at 0.4 g L⁻¹ of the tested dose of the LDH/RHB adsorbent. The removal efficiency was also maximum (more than 99%) for Pb²⁺ at 0.4 g L⁻¹ of the LDH/RHB adsorbent, while 94% of removal efficiency was recorded for Cu²⁺ at this value which further increased to about 99.9% at the maximum tested dose (0.8 g L⁻¹) of the LDH/RHB adsorbent. The increase in the adsorption capacity was probably due to the availability of more active sites with increasing amounts of the adsorbent to absorb a fixed concentration of the absorbent (50 mg L⁻¹ of each metal ion). After attaining the maximum removal efficiency (about 99%) at 0.4 g L⁻¹ of the LDH/RHB adsorbent as a result of more exchangeable adsorption sites with an increasing amount of the adsorbent, the uptake capacity started decreasing due to unused adsorbent beyond 0.4 g L⁻¹ and hence, rendered it to be the optimum LDH/RHB adsorbent value for optimal performance of the investigated adsorption process.

Variations in the adsorption capacities and the removal efficiencies by varying the initial concentrations of both heavy metal ions in the range of 5–100 mg L⁻¹ are shown in Figure 5. The batch process was optimized in this regard by using 0.4 g L⁻¹ of the LDH/RHB adsorbent under a maintained solution pH of 6.0 ± 0.3 while an equilibrium contact time of 30 and 15 min for Cu²⁺ and Pb²⁺, respectively. As shown in Figure 5, the removal efficiencies remained high and nearly unchanged for up to 50 mg L⁻¹ of both heavy metal ions and decreased by about 50% by increasing the initial concentrations from 50 to 100 mg L⁻¹. The adsorption capacities, however, increased linearly by increasing the initial concentrations of both Cu²⁺ and Pb²⁺ from 5 to 60 mg L⁻¹ and remained unchanged afterward up to the maximum used value (100 mg L⁻¹) in this study. Strong driving forces exist between the surface of the LDH/RHB adsorbent and a high amount of heavy metal ions [55,56], resulting in high adsorption capacities, while low removal efficiencies result due to a fixed dose of the LDH/RHB adsorbent (0.4 g L⁻¹) to attract a high amount of studied metal ions. Considering the changes in both parameters, 50 mg L⁻¹ of both heavy metal ions was considered to be the optimum concentration exhibiting 117 and 124 mg g⁻¹ of adsorption capacities for Cu²⁺ and Pb²⁺, respectively.

Changes in the adsorption capacities and removal efficiencies by varying the solution pH from 2 to 6 are depicted in Figure 5. To optimize the said parameter, 0.4 g L⁻¹ of the LDH/RHB adsorbent was used for 50 mg L⁻¹ of each heavy metal ion by agitating the tested dilution for 30 and 15 min of equilibrium contact time for Cu²⁺ and Pb²⁺, respectively. Low percentage removal and adsorption capacities under acidic solution pH (2–3) were probably due to the high amount of H⁺ presenting more competition to positively charged heavy metal ions to attach to the surface of the LDH/RHB adsorbent. Low H⁺ concentrations at high pH values provide more chances for divalent metal ions to attach to the surface of the LDH/RHB adsorbent [57,58,59], resulting in increased adsorption capacities and percentage removal. Precipitation of the tested solutions occurred at a pH value of 7.0, while a percentage removal of 94% and 99% was recorded at a solution pH value of 6.0 with a maximum adsorption capacity of 117 and 124 mg g⁻¹, respectively, for Cu²⁺ and Pb²⁺.

3.3. Explanation of the Adsorption Data Using Kinetic Models

Various commonly used kinetic models were applied to fit the experimental data obtained through batch tests using 30–80 mg L⁻¹ as initial concentrations of each of the Cu²⁺ and Pb²⁺. The estimation of parameters in linearized fitting was achieved using the slope and intercept of the fitted line, while OriginPro 8.5 Software was used for the same in nonlinear fitting by plotting the adsorption capacity (q_t, mg g⁻¹) against respective retention time (t, min). The original nonlinear expressions of the pseudo-first-order (1st-pseudo, Equation (3)), pseudo-second-order (2nd-pseudo, Equation (4)), the Elovich (Equation (5)) and the intraparticle diffusion of Weber and Morris (ID-WM, Equation (6)) are as follows:

$$q_t=q_e\left(1-\exp \left(-k_{1}t\right)\right)$$

(3)

$$q_t=\frac{{q_e}^2{k}_2·t}{q_e{k}_2·t+1}$$

(4)

$$q_t=\frac{1}{\beta }\ln (1+\alpha \beta t)$$

(5)

$${q_t=K}_{ip}{t}^{1/2}+C$$

(6)

Table 1. Estimation of parameters in kinetic models (nonlinear) at various initial concentrations of Cu²⁺ (against 30 min) and Pb²⁺ (against 15 min) onto 0.4 g L⁻¹ of the composite adsorbent at pH = 6 ± 0.3.

Kinetic Model	Parameter	Initial Cu2+ Concentrations, mg L−1				Initial Pb2+ Concentrations, mg L−1
		30	50	60	80	30	50	60	80
	qe exp (mg g−1)	72.50	117.12	127.50	127.50	74.98	124.17	135.00	135.00
2nd-pseudo	qe cal (mg g−1)	71.48	121.63	129.78	130.47	75.19	124.81	136.24	136.32
	k2 (g mg−1 min−1)	0.030	0.004	0.003	0.002	0.062	0.014	0.011	0.009
	h (mg g−1 min−1)	153.33	57.99	51.03	40.34	350.10	219.95	202.87	172.26
	R2	0.90	0.76	0.89	0.94	0.96	0.92	0.89	0.93
Elovich	α (mg g−1 min−1)	6.60 × 107	1.28 × 103	8.62 × 102	4.20 × 102	2.10 × 1018	1.69 × 107	2.08 × 106	5.65 × 105
	β (g mg−1)	0.30	0.08	0.07	0.06	0.62	0.15	0.12	0.11
	R2	0.76	0.89	0.87	0.87	0.48	0.72	0.74	0.73
ID-WM	Kip (mg g−1 min1/2)	61.54	75.91	78.66	77.34	70.42	105.41	111.97	109.52
	C (mg g−1)	0.87	3.96	4.19	3.05	0.38	1.65	2.08	2.26
	R2	0.38	0.65	0.56	0.49	0.17	0.34	0.37	0.36
1st-pseudo	qe cal (mg g−1)	69.37	115.34	123.69	124.06	74.17	120.94	131.52	131.86
	k1 (min−1)	1.23	0.30	0.23	0.18	1.74	1.00	0.88	0.69
	R2	0.57	0.50	0.73	0.86	0.87	0.64	0.58	0.68

shows the values and comparison of different parameters as calculated in each model for the nonlinear fitting of each model. An amount of 0.4 g L⁻¹ of the LDH/RHB adsorbent was used with a maintained solution pH at 6.0 ± 0.3, while equilibrium adsorption capacities were taken against 30 and 15 min of contact time for Cu²⁺ and Pb²⁺, respectively, to compare the calculated adsorption capacities (q_e, mg g⁻¹). The rate constants of the 1st-pseudo and 2nd-pseudo kinetic models are expressed by k₁ (min⁻¹) and k₂ (mg g⁻¹ min⁻¹) in Equation (3) and Equation (4), respectively. In the Elovich kinetic model (Equation (5)), α (mg g⁻¹ min⁻¹) and β (g mg⁻¹) represents the respective rate constant and the activation energy, respectively, while K_ip (mg g⁻¹ min^1/2) and C (mg g⁻¹) in the ID-WM kinetic model (Equation (6)) represents the respective rate constant and the boundary-layer thickness.

In Figure 6, an illustration of the fitting of nonlinear models is presented for the initial concentrations (50 and 80 mg L⁻¹) of both Cu²⁺ and Pb²⁺ for the 2nd-pseudo and Elovich kinetic models. A perfect-fitting of the linearized 2nd-pseudo model can be predicted for all the tested initial concentrations of each heavy metal ion (results not shown), based on the estimated values of the coefficient of determination (R²) as high as 1.0. Nonlinear fitting of the 2nd-pseudo model also yielded high R² (0.89–0.96, Table 1) for all tested initial concentrations of both heavy metal ions with the exception of 50 mg L⁻¹ of Cu²⁺ (0.77, Figure 6a). A close match of the experimental and calculated adsorption capacities in the linearized as well as nonlinear 2nd-pseudo model (Table 1) also suggests that chemisorption can be thought of as the controlling mechanism [60,61,62] for the studied adsorption process. An average R² of 0.75–0.85 in the linearized and nonlinear Elovich kinetic model (except for 30 mg L⁻¹ of Pb²⁺ in nonlinear fitting, Table 1) also suggests the fitting of the model to the adsorption data. The model’s rate constant (α, mg g⁻¹ min⁻¹) decreased as the initial concentrations of either (Cu²⁺ or Pb²⁺) heavy metal ion increased in both the linearized as well as nonlinear Elovich model. The activation energy (β, g mg⁻¹) increased by increasing the initial concentration of Cu²⁺ or Pb²⁺ in the linearized fitting of the Elovich model (results not shown), whereas an opposite trend (decreasing β, Table 1) can be seen for the nonlinear fitting of the Elovich model as the initial concentration of either Cu²⁺ or Pb²⁺ increased.

Based on R² values, the linearized ID-WM model yielded better fitting results compared with the 1st-pseudo model with the exception of 50 mg L⁻¹ of Pb²⁺ (comparative values between two models), while the nonlinear fitting yielded opposite results with the exception of 50 mg L⁻¹ of Cu²⁺ (Table 1). The rate constant in the nonlinear 1st-pseudo model observed a decreasing trend with an increase in the initial concentrations of respective metal ions, as presented in Table 1. Despite the poorest fit of the linearized 1st-pseudo model (based on R² values), the nonlinear fitting yielded a close match between the experimental and calculated adsorption capacities (Table 1).

3.4. Explanation of the Adsorption Data Using Isotherm Models

The experimental data of the batch testing is further explained, and equilibrium concentrations are analyzed by fitting commonly used isotherm models, which are mentioned in

Table 2. Mathematical expressions of the two- and three-parameter isotherm models and explanation of parameters.

Isotherm Model	Mathematical Expression	Parameters
Langmuir	$q_e=\frac{q_m{K}_L{C}_e}{(1+K_L{C}_e)}$ $R_L=\frac{1}{(1+K_L{C}_o)}$	qm, maximum sorption capacity, mg g−1 KL, Langmuir constant, L mg−1 RL, separation factor coefficient
Dubinin–Radushkevich	$q_e=q_m\exp \left(-K_{DR}{\epsilon }^2\right)$ $\mathsf{\epsilon }=RT\mathrm{l}\mathrm{n}(1+1/Ce)$ E = 1/$\sqrt{2K_{DR}}$	T, absolute temperature, Kelvin R, universal gas constant, 8.314 J mol−1·K−1 E, mean free energy of adsorption, kJ mol−1
Jovanovic	$q_e=q_m[1-\exp \left(k_j{C}_e\right)]$	kj, Jovanovic constant
Redlich–Peterson	$q_e=\frac{K_{RP}{C}_e}{1+\alpha {C_e}^{\beta }}$	α, L mg−1 β (0–1), dimensionless KRP, R–P constant, L g−1
Sips	$q_e=\frac{q_m{K}_S{C_e}^{n_s}}{1+K_S{C_e}^{n_s}}$	ns, degree of heterogeneity, dimensionless KS, energy of adsorption, L g−1

. These included the Langmuir, Dubinin–Radushkevich (D–R), and Jovanovic isotherms as two-parameter models and Redlich–Peterson (R–P) and Sips isotherms as three-parameter models. The absorption of both Cu²⁺ and Pb²⁺ to the LDH/RHB adsorbent is explained using the nonlinear as well as linearized (with the exception of three-parameter models) fitting.

Theoretical adsorption capacities were computed in different models and compared with the experimental values at 30 and 15 min of equilibrium time for 60 mg L⁻¹ of each Cu²⁺ and Pb²⁺, respectively. The initial concentration of the selected metal ions ranged between 5 and 100 mg L⁻¹ for the batch testing using 0.4 g L⁻¹ of the LDH/RHB adsorbent by maintaining the dilution pH at 6.0 ± 0.3. OriginPro 8.5 Software was used to calculate the parameters of each model based on the fit to the plot of the experimental adsorption capacity (q_e, mg g⁻¹) and respective residual concentration (C_e, mg L⁻¹) in nonlinear fitting (

Table 3. Estimation of parameters in nonlinear isotherm models against 60 mg L⁻¹ of both heavy metal ions (solution pH = 6 ± 0.3, contact time = 30 and 15 min for Cu²⁺ and Pb²⁺, respectively, and adsorbent dose = 0.4 g L⁻¹).

Isotherm	Parameter	Cu2+	Pb2+
	qe exp, mg g−1	127.53	135
Langmuir	qm, mg g−1	133.64	131.21
	KL, L mg−1	1.23	125.29
	RL	0.013	0.00013
	R2	0.98	0.98
D–R	qm, mg g−1	124.64	134.71
	KDR, (mol kJ−1)2	1.54 × 10−7	5.3 × 10−9
	E, kJ mol−1	1.80	7.71
	R2	0.95	0.99
Jovanovic	qm, mg g−1	126.29	127.28
	kj, L g−1	−0.82	−4.15
	R2	0.99	0.94
R–P	KRP, L g−1	125.13	213.36
	α, L mg−1	0.73	1.47
	β	1.08	0.98
	R2	0.99	0.99
Sips	qm, mg g−1	130.86	134.82
	KS, L g−1	1.30	2.39
	nS	1.17	0.70
	R2	0.98	1.00

). Figure 7 illustrates the nonlinear fitting of selected models based on the most favorable and commonly employed isotherms.

Linearized as well as nonlinear fitting (Figure 7a) of the Langmuir model yielded a near to perfect-fit to the adsorption of both heavy metal ions onto the LDH/RHB adsorbent with significant high R² (0.97–0.99), as presented in Table 3, supporting monolayer sorption of both heavy metal ions to a fixed number of homogeneous sorption sites on the surface of the LDH/RHB adsorbent. A favorable adsorption is indicated based on the calculated values of the separation factor coefficient (R_L less than 1.0, Table 3) [63] with a quite higher Langmuir constant value for Pb²⁺ compared with Cu²⁺. The calculated adsorption capacities were lower than the experimental values in the linearized approach but matched closely in the case of nonlinear Langmuir isotherm, as presented in Table 3.

The adsorption of Cu²⁺ and Pb²⁺ to the LDH/RHB adsorbent seems to be physical adsorption, as suggested by the calculated E values in D–R isotherm (<8 kJ mol⁻¹, Table 3) using the linearized as well as nonlinear approach. A perfect fit of the model is also evident from the high R² values (0.95–0.99, Figure 7b) for the nonlinear fitting and through a close match of the calculated and experimental adsorption capacities. D–R isotherm exhibited lower K_DR and, in turn, higher mean free energy of adsorption for Pb²⁺ compared with Cu²⁺ (E values in Table 3). The Jovanovic isotherm yielded a close match of the calculated adsorption capacities to the experimental values using the nonlinear fitting with lower values of the model’s constants for Pb²⁺ in comparison to Cu²⁺. A poor fit of the linearized Jovanovic model was seen based on R² values (0.23–0.27, results not shown), while a very good fit of the nonlinear Jovanovic model was observed with high R² values (0.94–0.99, Table 3).

The homogeneous sorption of both Cu²⁺ and Pb²⁺ to the LDH/RHB adsorbent is also proposed as the dominant sorption by the three-parameter model [64,65,66,67] since a near-to-perfect fit of the Sips and R–P isotherms was predicted based on very high R² values (0.98–1.00, Table 3) for both heavy metal ions using the nonlinear approach. The findings of the Sips model suggested a lower degree of heterogeneity and higher heat of adsorption for Pb²⁺ compared with Cu²⁺ (comparison of n_s and K_S values in Table 3) with a close match of the calculated and experimental adsorption capacities. Finally,

Table 4. Comparison of the Langmuir maximum adsorption capacity of ZnMgAl(LDH)/RHB with other composite materials for the adsorption of Cu²⁺ and Pb²⁺.

Composite Materials	Pollutants	Q0 (mg/g)	References
MgAl-LDH/RHB	Cu (II)	104.34	[68]
MnAl-LDH/BC	Cu (II)	74.07	[4]
MgAl-LDH/BC	Pb (II) Cu (II)	294 38.6	[69]
Mg/Al LDH-SHMP	Pb (II)	45.66	[70]
Malate- Mg/Al-LDH	Cu (II)	118	[71]
Fe3O4/GO/MnOx	Cu (II)	62.65	[72]
Pristine-Mg/Al-LDH	Pb (II) Cu (II)	84.7 59.9	[71]
Fe3O4/LDH-AM	Pb (II) Cu (II)	359.5 111.4	[73]
ZnMgAl(LDH)/RHB	Pb (II)	124	This study
ZnMgAl(LDH)/RHB	Cu (II)	117	This study

presents a comparison of the maximum adsorption capacity as obtained in this study using the specific composite adsorbent with other similar compounds reported in previous studies.

4. Conclusions

A new composite material of ZnMgAl(LDH) and RHB using the co-precipitation and hydrothermal technique was developed in this study, and the removal mechanisms of Cu(II) and Pb(II) ions from synthetic wastewater were investigated. Extensive batch tests were conducted, and the kinetics and isothermal models were deployed to understand the removal mechanisms and to calculate the adsorption capacity of the composite material. SEM micrographs of the LDH/RHB composite showed extremely fine crystalline LDH particles decorated on the rough surface of the RHB, while the successful formation of the composite adsorbent was confirmed by the corresponding EDX spectrum of the composite displaying all bonding elements that were present separately in the RHB and ZnMgAl(LDH) spectra. Furthermore, the EDX spectra of the composite adsorbent after the adsorption of studied heavy metal ions showed many variations in elemental composition in addition to the presence of Cu (3.41%) and Pb (47.63%), indicating successful absorption of the target metals into LDH/RHB composites. The effective synthesis of the composite adsorbent was further evident from the FTIR spectrum, which also suggested a change (low intensity) or vanishing nitrate and hydroxyl representing bands with a possible replacement by Cu²⁺ and Pb²⁺.

The rapid uptake of each heavy metal ion upon immediate contact with the adsorbent was observed following a steady increase afterward until 30 and 15 min for Cu²⁺ and Pb²⁺, respectively, thus proposing a different equilibrium contact time for Cu²⁺ and Pb²⁺. Furthermore, the composite adsorbent exhibited slightly higher adsorption capacity and removal efficiency for Pb²⁺ in comparison to Cu²⁺. A steady increase in the adsorption capacity by increasing the amount of the adsorbent was observed with maximum values of 117 and 124 mg g⁻¹ for Cu²⁺ and Pb²⁺, respectively, at 0.4 g L⁻¹ of the tested dose. The corresponding maximum removal efficiency was about 94% and 99% for Cu²⁺ and Pb²⁺, respectively, at 0.4 g L⁻¹ of the LDH/RHB adsorbent. For the changing initial metal concentration, the removal efficiencies remained unchanged for up to 50 mg L⁻¹ and decreased by about 50% by increasing the initial concentrations from 50 to 100 mg L⁻¹. The adsorption capacities observed an increasing linear trend by increasing the initial concentrations from 5 to 60 mg L⁻¹ and remained unchanged afterward up to 100 mg L⁻¹. An increase in the adsorption capacities and percentage removal was observed by changing the solution pH from 2.0 to 6.0, observing a maximum adsorption capacity of 117 and 124 mg g⁻¹ for Cu²⁺ and Pb²⁺, respectively.

With a perfect fitting of the second pseudo model (R² as high as 0.89–0.96) for all tested initial concentrations of both heavy metal ions and of a close match of the experimental and calculated adsorption capacities, the chemisorption was suggested to be the controlling mechanism. Among the isotherm models, a near-to-perfect fit (R² as high as 0.97–0.99) of the nonlinear Langmuir model to the experimental data, monolayer sorption of both heavy metal ions to a fixed number of homogeneous sorption sites on the surface of the LDH/RHB adsorbent can be proposed. The application of the D–R isotherm with high R² values (0.95–0.99) helped to conclude the adsorption of Cu²⁺ and Pb²⁺ to the LDH/RHB adsorbent, which is physical adsorption with estimated E values less than 8 kJ mol⁻¹. The idea of the dominant homogeneous sorption of both heavy metal ions to the LDH/RHB adsorbent is further supported by a perfect-fit (R² values in the range 0.98–1.00) of the Sips and R–P isotherms (three-parameter models) to the experimental data. A close match of the calculated and experimental adsorption capacities was seen using the Sips isotherm with a lower degree of heterogeneity and higher heat of adsorption for Pb²⁺ compared with Cu²⁺.