Accumulation of Vanadium by Nanoscale Zero-Valent Iron Supported by Activated Carbon under Simulation Water Conditions: A Batch Study

Abstract

Vanadium (V(V)) removal from simulation water (SW) was successfully accomplished using nanoscale zero-valent iron that was immobilized by activated carbon (NZVI/AC) which was used as an adsorbent. We investigated the effects of different parameters on V(V) removal, such as pH, dissolved oxygen (DO), common ions and adsorption kinetics for SW. The intraparticle diffusion model fits this study well (R² > 0.9) according to the results of the kinetics investigation which showed that the adsorption of vanadium by NZVI/AC was rapid in the first 12 h and that equilibrium was reached in about 72 h. The amount of V(V) that was removed from the solution increased when it was subjected to pH 2 to pH 8, and this decreased after pH 8. While the effects of other anions and humic acid were negligible, the elimination of V(V) was significantly reduced by using phosphate and silicate. Fe²⁺ and Al³⁺, two common metal cations, improved the V(V) adsorption. High oxygen levels impeded the vanadium elimination, while anoxic conditions encouraged it. Elution with 0.1 M NaOH can be used to renew NZVI/AC in an efficient manner.

1. Introduction

Vanadium is utilized in a variety of disciplines in contemporary industries, although it is most commonly used in the production of steel (92.9%), chemicals (3%), non-ferrous alloys (4%), and 0.1% of batteries [1]. Vanadium-based chemical materials are employed extensively for catalysis in the manufacturing of ceramics, pigments, batteries, and other products [2]. Since 2011, there has been about a 45% rise in the global use of V, and its quantity reached 102.1 kilotons in 2019 and is projected to reach a total of 130.1 kilotons by 2024. A significant amount of damage to human health is caused by the effluent discharge that is carrying a high concentration of V, which takes place through various environmental media, such as such as mining, industrial production, and the use of fertilizers and pesticides have resulted in elevated concentrations of V in soils and natural waters. Vanadium, when it is in food, enters the liver, kidney, testis, spleen, bone and other various organs and tissues. The International Agency for Research on Cancer has listed V₂O₅ as a possible human carcinogen. The negative effects of V on the immune system have been shown by the strong link between viral infections and vanadium air pollution in Czech children and respiratory disorders also [3].

Numerous regional and national governments have taken steps to reduce V pollution in light of the detrimental effects resulting from the excessive exposure of V to human health. V has been proposed to be identified as a priority contaminant of the environment, and which is to be managed by the United Nations Environment Programme (UNEP) [4,5]. In China, the drinking water quality reference index has a 10 μg/L limit for the amount of vanadium that is allowed (“Sanitation Standard for Drinking Water” GB 5749-2022). The scientific community recognizes the need for an efficient and affordable technique for recovering V from water and wastewater. A number of techniques have been developed to achieve the removal of Vanadium from water; for instance, these include adsorption [6], chemical precipitation [7], microbiological treatment [8,9], electrokinetic remediation [10], solvent extraction [11], coagulation [12], photocatalysis [13], and membrane filtration [14]. Adsorption techniques have received the greatest attention among these strategies, in terms of achieving high V-binding capabilities along with having a low energy consumption. Iron-based adsorbents have been used for separating, adsorbing, and recovering Vanadium [15,16,17].

Adsorption techniques have seen significant progress in recent years and hold considerable promise for their use in the treatment of industrial wastewater due to their advantages over other technologies in terms of having high purifying effectiveness and low energy consumption, and their environmental friendliness [18,19,20]. Many adsorbents, including minerals, resins, bio-sorbents, industrial wastes, iron-based adsorbents, and nanomaterials have been proven to be effective for removing V from wastewater, and extensive research has been done on their adsorption behavior and efficacy towards V. A list of different adsorbents is compiled from the literature, and their adsorption capacities for the removal of V in different conditions (e.g., pH, co-existing ions, concentration of dissolved oxygen (DO), contact time, dosage, and temperature) have been presented along with a discussion of the adsorption mechanism. Additionally, the adsorbent’s cost and its environmental impact were taken into consideration when analyzing the elution of the V-loaded adsorbent. Such insights into the current structures and functionalities of them direct the study and so too do the constitution of newer composite materials with good adsorption activity in their respective applications. According to earlier studies, Lewis base or Lewis acid sites can react with nanomaterials possessing a high specific surface structure, thus generating various hydroxyl groups, including bridging and terminal hydroxyl groups [20].

On account of its low standard redox potential, nanoscale zero-valent iron (NZVI), a type of inorganic metallic nanoparticle and an abundant metal, has been used as an effective reductant for the treatment of contaminated water, catalysis or hydrogen storage [18,20]. Yayayuruk [17] used a borohydride reduction technique to fabricate NZVI in ethanol under atmospheric circumstances. With a V concentration that was initially equivalent to 500–4000 mg/L, the NZVI had the maximum adsorption capabilities of acquiring 324.4 mg/g of it. The high magnetic force and surface energy of the exposed NZVI particles, however, cause them to aggregate, which this significantly reduced the efficacy of the removal of the contaminants. Spreading NZVI particles over the pores and surfaces of clay minerals, carbonaceous materials, and biopolymers renders them useful as stabilized substrates. According to Fan et al. [3], charcoal (BC) significantly impeded the aggregation of NZVI particles. More crucially, Fe⁰ and Fe²⁺ could convert the adsorbed pentavalent V to the nontoxic tetravalent V form.

In our earlier research, arsenic, molybdenum, and antimony elimination from drinking water was accomplished by utilizing zero-valent iron at the nanoscale on activated carbon supports. The goal of the current study is to evaluate the effectiveness of vanadium elimination from water while utilizing NZVI/AC. We investigated the adsorption kinetics and effects of different parameters on V(V) removal, such as pH, dissolved oxygen (DO) and common ions.

2. Materials and Methods

2.1. Chemical and Instrumentation

Chemical reagents of analytical grade were used in this investigation and were acquired from Sinopharm chemical reagent Co., Ltd. (Shanghai, China). The sodium metavanadate (V) stock solution (100.0 mg·L⁻¹) utilized in the experiments was obtained using sodium metavanadate (NaVO₃). The solutions were made in deionized (DI) water. V(V) concentration initially was 0.50 mg/L in deionized water (SW).

2.2. Synthesis and Characterization of NZVI/AC

The synthesis and characterization of NZVI/AC were accomplished using the methods described by Zhu [21].

Table 1. Key attributes of NZVI/AC.

Thickness	Diameter	Shape	Total Pore Volume	Fe Content	BET Surface Area
~20 nm	<100 nm	flakes	0.45 cm3/g	~8.2%	821.7 m2/g

summarizes the key characteristics of NZVI/AC.

2.3. Batch Adsorption Experiments

The investigation disregarded the tiny and low V(V) elimination percentages of the regular AC.

A series of studies were conducted by making use of a V(V) solution of 100 mL with a starting concentration of 0.5 mg·L⁻¹ and 1.5 g·L⁻¹ NZVI/AC to describe the parameters influencing the sorption process, without the parameters being optimized. This was done by taking into account the solution’s pH (3, 5, 7, 9, and 11) and the mass of NZVI/AC (0, 1, 1.5, and 2.0 g·L⁻¹) as well as the adsorption kinetics (0–72 h). When NZVI/AC was added to the mixture, the pH was adjusted to neutral using either using diluted NaOH or HCl solutions. Following equilibration for 72 h at 25 °C and the acquiring of pH 6.5 in a shaker, aqueous V(V) measurements were carried out after filtering the solution through a membrane filter of 0.22 μm.

When examining the effect of DO on the vanadium removal rate, (O₂ > 99.9%) or (N₂ > 99.9%) was added to the V(V) solution at 25 °C and pH 6.5, wherein the addition of NZVI/AC was performed for more than 30 min. The solution with the additional NZVI/AC in the anoxic/ moderate oxygen (DO_M)/high oxygen (DO_H) state was simulated.

By allowing 100 mL of 1.0 mg·L⁻¹ solution to come into contact with 1.5 g of adsorbent for 72 h at pH 6.5 and 25 °C, it was possible to determine the effects of various anions (such as carbonate, silicate, oxalate, phosphate, and sulphate) and humic acid on the adsorption of vanadium onto NZVI/AC, with the molar ratio of the added ions to V(V) being 10:1. In addition, 5 mg L⁻¹ of humic acid was supplemented.

A hydride generator and an atomic fluorescence spectrophotometer (AFS-2202E, Haiguang Corp., Beijing, China) were selected for the quantification of V(V) within the mixture solution. The instrumental detection limit was precisely 0.1 µg·L⁻¹. The equation that is mentioned below was solved for computing the residual concentration in the adsorbent (q_t, mg·g ⁻¹).

$$q_t=\frac{{\mathrm{V}(C}_0-C_t)}{W_s}$$

(1)

C₀ denotes the initial V(V) concentration, C_t refers to the V(V) at a time t (in mg·L⁻¹), V gives the volume of solution (mL), and W_s represents the adsorbent weight (g). The percentage of the eliminated V(V) (R%) was calculated using the equation given below.:

$$\mathrm{R} (\%)=\frac{C_0-C_t}{C_0} \times 100$$

(2)

2.4. Desorption of Adsorbed V(V)

After a 72-h adsorption reaction between 0.15 g of adsorbent and 100 mL of 0.5 mg·L⁻¹ V(V) solution, the vanadium-loaded NZVI/AC was removed and rinsed thoroughly with pre-distilled water for the removal of any residual V(V) solution. A 100 mL volume of 0.1 M NaOH with pH 13 was mixed with NZVI/AC and stirred for 12 h.

2.5. Models

An intraparticle diffusion model was employed to study the kinetics of the V(V) adsorption, using the below-mentioned expression [22].

q_t = k_id t^0.5

(3)

k_id denotes the original intraparticular diffusion rate (mg·g⁻¹·h^−0.5) and q_t (mg·g⁻¹) refers to the amount of V(V) adsorbed at time t.

2.6. Analytical Methods

At predetermined time intervals, 5 mL suspension were withdrawn, filtered via a 0.22 μm membrane filter, and acidified before carrying out the batch and column tests. The amount of V(V) was measured by making use of a colorimetric method along with an atomic fluorescence photometer (AFS-8220, Beijing Jitian Instrument Co., Ltd., Beijing, China).

3. Results

3.1. The Iron Decorating on AC/Biochar

In recent years, the research on activated carbon or biochar-supported nano-iron has received much attention.

Table 2. The iron decorating on AC/biochar for removal application.

Carbon from Raw Material	Shape	Total Pore Volume cm3/g	Fe Content	BET Surface Area m2/g	Form of Fe	Average Pore Diameter	Target Removal of Contaminants
Coal [This study]	flakes	0.45	~8.2%	821.7	FexOy	20 nm	Vanadium
Forestry wastes [6]	strip-like, brush hollow/hierarchical structure	0.588	-	116.095	CoFe2O4@BC-LDH	<10	HA
Tea waste [15]	rose flower like pattern	0.201	-	111.215	Fe3O4	1.5–10	$\mathrm{Ni}{2}^{+}, \mathrm{Co}{2}^{+} {\mathrm{NH}}_4^{+}$$, {\mathrm{PO}}_4^{3-}$
Pristine [16]	relatively conspicuous pore structures	0.186	0.9%	431.7	Fe3O4	-	Sr2+
Burley Tobacco Stems [19]	-	0.008	-	4.33	Fe3O4/Fe2O3	-	Cr(VI)
Anaerobically Digested Sewage Sludge [23]	spherical or irregular nodular	0.1494	8.359%	44.75	Fe3O4/FeO	13.358	P

summarizes the characterization and application of activated carbon or biochar-supported nano-iron in recent years.

3.2. Dosage of Adsorbent

Due to its impact on the adsorption sites and surface areas and its financial cost, the adsorbent dosage is a crucial parameter. Due to the simple accessibility of more active exchangeable sites, the efficiency of adsorption typically rises when the adsorbent dosage is increased. However, if the sorbent dose is increased, the adsorption capacity, or total solute adsorption per unit weight of the adsorbent, declines on account of the interference from the adsorbent’s active sites interacting with each other. Thus, by raising the biochar-stabilized nanoscale zero-valent iron (NZVI/BC) dosage to 2.0 g/L, the V(V) removal efficiency increased to 100% (Figure 1), while the removal capacity of V(V) fell from 0.6230 mg/g to 0.2500 mg/g. When the adsorbent dose was increased to a certain level in several investigations, the removal efficiency remained unchanged [18,23,24,25,26]. This is a consequence of aggregation, which led to adsorbent particles self-binding, or the reduced usage of the adsorbent’s surface active sites at greater adsorbent levels [26]. The major goal of this is to prevent aggregates from forming using NZVI/BC. Additionally, the pH of the solution is altered by the use of different dosages of several adsorbents, including ferric oxyhydroxide, synthetic zeolite, and resin [15,18,24].

3.3. Solution pH

Since the pH of the solution has a direct influence on the adsorbent’s surface charges and the V speciation in the water, it is important in the effluent treatment process [27]. The adsorbent’s surfaces are positively charged at pH values that are below pH < pHpzc (pH of zero charge point), which facilitates the V anion’s adsorption. The pH_ZPC of the synthesized NZVI/AC was determined to be pH 7.4, below which the surface is positively charged and favorable for the adsorption of anionic species. The maximal adsorption capabilities for V were frequently obtained at pH 3–9, as shown in Figure 2. The following three causes could be to blame for the poor adsorption capabilities in a solution with a pH < 3 (Figure 2): The electrostatic attraction between cationic V (VO²⁺⁾ and the sorbent surface positive charges is first caused by the existence of cationic V (VO²⁺) as the dominating ionic entity within the solution at pH < 3 [28,29]. Secondly, H⁺ might compete with VO²⁺ for occupying the active site, decreasing the overall amount of vanadium adsorption [30]. Third, the adsorbed vanadium was released into the solution as a result of the adsorbent’s structure being broken at a very low pH [31]. The polynuclear anion species (${\mathrm{V}}_2{\mathrm{O}}_6{\left(\mathrm{OH}\right)}^{3-}$, ${\mathrm{V}}_2{\mathrm{O}}_7^{4-}$, ${\mathrm{V}}_3{\mathrm{O}}_9^{3-}$ and ${\mathrm{V}}_4{\mathrm{O}}_{12}^{4-}$, etc.) are predominant within the solution at a pH value that is ranging from 3.00–9.00 [5,32]. Numerous investigations have suggested that the electrostatic attractive interaction between vanadium and the adsorbent’s positively charged surface groups is majorly responsible for the adsorption of anionic V [5,29]. When the pH > pHpzc, the adsorbents’ surfaces are negatively charged. OH^- ions that are negatively charged compete with V anions or adsorption sites at higher pHs (pH > 9), thereby dramatically reducing the capacity for adsorption [5,20,29]. Additional support adsorption sites were required because at a high pH, V was changed into mononuclear anionic species (${\mathrm{VO}}_4^{3-}$, ${\mathrm{VO}}_2{\left(\mathrm{OH}\right)}^{2-}$, ${\mathrm{VO}}_3{\left(\mathrm{OH}\right)}^{2-}$, etc.) (see Figure 2) [5].

3.4. Impact of DO Concentration

The results of Figure 3 demonstrate that V(V) removal was unaffected by the moderate DO ~7 mg⋅L⁻¹ and anoxic (DO < 0.5 mg⋅L⁻¹) conditions, despite the fact that the V(V) removal rate was slightly reduced from 93.2% to 88.9% in these settings. Nevertheless, the V(V) elimination rate declined from~88.9% to 63.9% at a high DO > 14 mg⋅L⁻¹. This indicates that an elevated DO has a beneficial influence on V(V) removal by nano-zero-valent iron. The DO levels in the raw water that is utilized in industrial operations were often found to be in the moderate DO range. This did not have a major influence on the amount of V(V) that was removed from the water by nano-zero-valent iron, showing that NZVI/AC is capable of removing V(V) from water in an effective manner.

According to the findings of the study, DO serves two purposes within the framework of the zero-valent iron reaction system. The presence of a moderate amount of DO encourages iron corrosion and hastens the elimination of V(V). Nevertheless, an excessive amount of DO speeds up the generation of iron oxides upon the iron’s surface. This results in a more severe passivation upon the surface of the zero-valent iron, which in turn slows down the reaction progress. Thus, it is obvious that DO_H has a passivating effect on zero-valent iron, making the reaction of zero-valent iron considerably more challenging [33,34,35].

3.5. Effect of Coexisting Ions

3.5.1. Inorganic Ions

Anions and cations of varying types are found in drinking water, and these ions and cations have the potential to either positively or adversely influence the V(V) adsorption process. The majority of investigations have depicted that the presence of coexisting inorganic anions and cations typically has an impact on V adsorption through the generation of ternary surface complexes [36], cations bridges [37], and competitive adsorption [5,36,38]. As opposed to the cations that are in the solution, anionic V was less able to bind to adsorbents because of competition from inorganic anions. The inorganic anions (Cl^–, ${\mathrm{NO}}_3^{–}$, ${\mathrm{SO}}_4^{2–}$, ${\mathrm{PO}}_4^{3–}, {\mathrm{SiO}}_4^{2–}$) have been shown to either slightly or significantly impair the ability of V to adsorb on various adsorbents, including ZnCl₂-activated carbon [39], surface-modified palm fruit husk [40], Zr (IV)-loaded orange juice residue [41], chitosan-zirconium (IV) composite [42], and Ti-doped chitosan bead [43]. It is important to note that the valence charge and size of the anions most likely had an impact on how the V oxyanions, adsorption sites, and co-existing anions interacted [44]. The high-priced anions $({\mathrm{SiO}}_4^{2–}$, ${\mathrm{CrO}}_4^{2–}$, ${\mathrm{SO}}_4^{2–}$, ${\mathrm{PO}}_4^{3–},$ ${\mathrm{AsO}}_4^{3–}$) tend to be more competitive in comparison to the low-valence anions (Cl^–, ${\mathrm{NO}}_3^{–}$), thereby resulting in an enhanced impact of the high-valence anions upon the V adsorption [36,42,43]. The anion structure additionally had an important role. The ${\mathrm{SiO}}_4^{2–},{ \mathrm{PO}}_4^{3–}$ and ${\mathrm{AsO}}_4^{3–}$ are categorized as a regular tetrahedral structural analogue of ${\mathrm{VO}}_4^{3–}$, leading to a greater influence on the adsorption of V via competition [36,45].

3.5.2. Organic Ions

Water and wastewater frequently contain natural organic materials (NOMs) like humus. The coexistence of humic acid interfered with the elimination of V(V) as shown by the values that fell range from 97.8% to 81.5% (Figure 4). Islam et al. [46] came to the conclusion that humus (HS) influenced the metal ions adsorption via the formation of aqueous-phase complexes with metal ions, encouraging the dissolution of minerals, and altering both the molecular characteristics as well as the concentration of particular metal-binding sites and the electrostatic properties of the aqueous-mineral interface at a long range.

3.6. Adsorption Kinetics

As depicted in Figure 5, the kinetics of the adsorption of V(V) by NZVI/AC consists of two processes, which are as follows: the initial sorption happens very quickly, and this is followed by adsorption which happens rather slowly, later on. In the first 3.5 h, approximately 63.6% of V(V) was removed from the simulation water. In the subsequent 72 h, an adsorption equilibrium was achieved (at a level of 99.7%). In the subsequent investigations that were carried out with SW, an equilibration time of 72 h was therefore utilized. For the adsorption of metal cations on NZVI/AC, a very similar phenomenon was observed. The initially observed fast adsorption was attributed to the fast transfer of the metal ion to the adsorbent particles’ surface; while the slow adsorption that followed was the outcome of the slow diffusion of the metal ions into the pores of the adsorbent [21].

When analyzing the data of adsorption kinetics, we made use of a great number of models; nevertheless, only the model that is based on intraparticle diffusion provided a good fit (R² > 0.9). A couple of different models were employed to study liquid-phase adsorption kinetic data in an effort to develop adsorption-based treatment systems. However, the intraparticle diffusion model provided the best explanation (R² > 0.99) for the experimental data that were relating to the current study.

According to Weber and Morris, the rate of adsorption is determined by the intraparticle diffusion in an adsorption system, and the quantity of the substrate that is adsorbed (q_t) varies linearly with the square root of time (t^0.5). After gathering all of this information, one can proceed to compute the adsorption rates [47]. In two distinct stages, a plot of q_t against t^0.5 revealed that there was a linear relationship between the two variables (Figure 5). Since this was the case, Equation (3) was applied separately to both of these steps. While V(V) adsorption on the NZVI/AC in the macropores results in the first linear segment of the graph, V(V) diffusion into the meso- and/or micropores results in the second linear segment of the graph. The V(V) particles were quickly adsorbed in the AC macropores or channels, nevertheless, the diffusion of these particles into the meso- and micro-pores was somewhat sluggish since the majority of the pores were blocked. In addition to this, the corrosion of the NZVI surface occurred in addition to the adsorption and diffusion in the corrosion layers. The fact that the K_id values for the first stage were significantly higher than those for the second stage—64.85 times higher in SW—indicates that the stage 1 reaction occurred at a significantly faster rate in comparison to the stage 2 reaction (

Table 3. Rate constants for the kinetic model of the adsorption of V(V) in SW on NZVI/AC.

Parameter	Weber–Morris Diffusion
	1st Step		2nd Step
C0 (0.50 mg/L)	kid1	R2	kid2	R2
Simulation water	0.1550	0.9902	0.00239	0.9937

), which is in agreement with Figure 5. The rate constant of the outer diffusion stage was large which means that the mass transfer resistance in this stage was small. Figure 5 shows that the process of removing V(V) by nZVI/AC is a bilinear correlation, and the intercepts of all the fitted lines are 0, indicating that the internal diffusion is the only factor controlling the process of removing V(V) by nZVI/AC. Fan et al. used the Weber-Morris model to analyze the removal process of Cr(VI) by nZVI/BAC and determined the rate control steps. The removal process can be divided into three dynamic stages: external diffusion, internal diffusion and reaction equilibrium. The V(V) removing process by nZVI/AC is similar to that with Cr(VI) [3].

A rather identical phenomenon was observed in the course of the adsorption of acidic dye [48], ${\mathrm{AsO}}_4^{3-}$ and ${ \mathrm{AsO}}_3^{3-}$ [47], and Cu(II) and Cd(II), on activated palm ash and rice/modified rice husk [49] using NZVI/BC. In comparison to the pseudo-first-order expression, the estimated q_e values and rate constants corroborated remarkably well with experimental q_e values when the pseudo-second-order (PSO) model was used. This was the case when clay-impregnated nanoscale zero-valent iron was employed for the adsorption and degradation of Zn²⁺ and Cu²⁺ from wastewater.

3.7. Removal Mechanisms by Adsorption

Being one of the primary active forces for V adsorption, electrostatic interaction can arise between positively charged adsorbents and anionic V. The influence of solution pH on V adsorption has been used as evidence to confirm the existence of electrostatic interaction during the sorption process [46]. To be more specific, certain surface functional groups upon protonation and deprotonation can cause the surface charge of adsorbents to shift in response to shifts in the pH of the solution. The result that is electrostatic contact between charged functions and anionic V is most directly affected by such a shift. According to Kong et al. [27], the surface hydroxyl groups (${-\mathrm{OH}}_2^{+}$) were protonated when the H⁺ concentration rose, which enhanced the electrostatic interaction between V(V) anions and NZVI/AC. Similar to this, the chelating ions existing on the surface of Ti-doped chitosan beads might protonate to generate Ti⁴⁺ with a positive charge, which would then strongly attract V anions (${\mathrm{H}}_3{\mathrm{HV}}_2{\mathrm{O}}^{7-}$ or ${\mathrm{H}}_2{\mathrm{VO}}_4^{–}$) [43]. In addition to the direct adsorption between V and the charged surfaces, it has been observed that the overall sorption process is also affected by the key role of the electrostatic interaction that is produced by the adsorbed substances on the adsorbent surfaces [30]. For instance, the cationic bridge on the surface of the adsorbent may function as a positive charge (Figure 6). It was proven by Zhu et al. [21] that a bridge may be formed by the Na⁺ in a solution. It has been demonstrated that divalent cations like Mg²⁺ and Ca²⁺ are capable of converting the adsorbent negative sites into positive ones, thereby functioning as a bridge that is capable of attracting V electrostatically [36].

It was also shown that vanadium may be adsorbed by swapping out the OH-coated -Fe₂O₃ nanoparticles [50]. Additionally, the reaction equations for the ion exchange between ${\mathrm{VO}}_4^{3–}$ and ${\mathrm{CO}}_3^{2–}$ on the CLDH are as follows [51].

For the V adsorption process, various adsorption–reduction mechanisms have been identified in recent investigations. The presence of V(III) or V(IV) species and surface group oxidation has been used to demonstrate the reduction possibility using a variety of techniques, such as energy dispersive X-ray spectroscopy, X-ray photoelectron spectroscopy, X-ray diffraction and Fourier transform infrared spectroscopy. There is a possibility of the direct or indirect reduction of V(V) to V(III) or V(IV). Following the binding of the V(V) ions to the adsorbent sites that furnish electrons to assist in their reduction, a direct conversion occurs. The V(V) ions were first electrostatically attracted to and adsorbed onto the NZVI/BC surface as documented by Fan et al. [3], before coprecipitating with the produced Fe³⁺ to create Fe-V precipitates. Finally, Fe⁰ and Fe²⁺ converted 90% of the adsorbed V(V) into V (IV). Similar to this, adsorption–reduction techniques were utilized to illustrate the V(V) adsorption by NZVI@LDH [27]. V oxides were the ultimate forms of V (III) or V (IV) on the NZVI/BC and NZVI@LDH, respectively (VO₂ or V₂O₃). Following is a summary of the redox reaction between NZVI and V(V) [27]:

$${3\mathrm{H}}_2{\mathrm{VO}}_4^{–}+\mathrm{Fe}{0}^{ }+12{\mathrm{H}}^{+}\to 3\mathrm{VO}{2}^{+}+\mathrm{Fe}{3}^{+}+9{\mathrm{H}}_2\mathrm{O}$$

(4)

V(V) that is linked to a site of adsorption undergoes an indirect reduction when electrons from nearby functional groups reduce it. Adsorption sites oxidize upon the process of V(V) reduction to V(IV) or V(III), such as when hydroxyl groups are converted to carboxyl groups, which makes it difficult for adsorbents to be reused for the V(V) adsorption. For instance, the Ti–H₃V₂O₇ or Ti–H₂VO₄ complexes that results from the ligand exchange between the Cl^- ions on the TiCB (Ti-doped chitosan beads) surface and H₃V₂O₇ or H₂VO₄ could be incompletely reduced to V(IV) by the hydroxyl on the chitosan beads’ surface. Additionally, the hydroxyl groups underwent concomitant carboxyl group oxidation [43]. Although V(V) was changed into a less harmful species, for example, V(IV) or V(III), there is still more research that is required regarding the process of recovering V oxide from the adsorbent. The majority of the iron-based materials absorbing V may lead to the creation of inner-sphere complexes. Peacock and Sherman used EXAFS spectroscopy to determine the nature of complexes that are existing on the surface of goethite after they absorbed V, and they predicted that these complexes use the molecular geometries from the beginning [51]. They discovered that the generation of inner-sphere surface complexes were involved in the adsorption of V, which resulted from the sharing of bidentate corners between singly and doubly protonated FeO₆ polyhedra and ${\mathrm{VO}}_4^{3-}$ tertrahedra. These complexes were detected on the inner surface of the sphere. The ${\mathrm{Fe}}_2{\mathrm{O}}_2{\left(\mathrm{OH}\right)}_2^{+}$ and Fe₂O₂VO(OH)⁰ surface complexes on goethite (α-FeOOH) were identifiable. The plausible reactions that are in agreement with the experimentally estimated V adsorption data are as follows:

$${2\mathrm{FeOH}}_2^{+} +{ \mathrm{VO}}_2^{+} ={ \mathrm{Fe}}_2{\mathrm{O}}_2{\left(\mathrm{OH}\right)}_2^{+}+2{\mathrm{H}}^{+}$$

(5)

$$2\mathrm{FeOH}+{\mathrm{HVO}}_4^{2–} ={ \mathrm{Fe}}_2{\mathrm{O}}_2{\mathrm{VO}\left(\mathrm{OH}\right)}^0+2{\mathrm{OH}}^{–}$$

(6)

Another study found that the negatively charged surfaces of clay minerals including kaolinite and montmorillonite, which are generated by deprotonation at pH levels above 4, absorbed V(V) by producing >${\mathrm{SOVO}}_3^{2–}$ [30]. Additionally, the complexation process took place while organic ions were present in the solution [15].

3.8. Regeneration of NZVI/AC

When assessing the potential uses of adsorbents in the real world, one crucial indicator to look at is whether or not they can be reused. NaOH and HCl, in varying strengths, are the eluents that are most frequently utilized. By disabling the electrostatic attractions among the protonated functionalities, a repulsion between the adsorbents and V oxyanions in an acidic and alkaline solution rendered it to be simpler to desorb the V oxyanions from the synthesized adsorbent’s surface. The removal of V oxyanions from the synthetic adsorbents was thus made simpler.

The room temperature exposure of the vanadium-loaded adsorbent to 0.1 M NaOH resulted in the regeneration of the used NZVI/AC. The alkaline solution desorbed all of the adsorbed vanadium in about 12 h. Similar to this, it was claimed that by using strong alkaline solutions, the desorption of >96% of the loaded vanadium on cellulose-based anion exchanger was achieved [52]. By repeatedly submitting the synthesized NZVI/AC to adsorption and desorption with NaOH (0.1 M), it could be determined whether it had degraded.

Table 4. Vanadium removal in individual cycles of repeated adsorption and desorption (0.5 mg/L V(V), 1.0 g/L NZVI/AC, pH 6.5).

Cycle	1	2	3	4	5
Vanadium removal (%)	95.2	97.4	93.1	96.7	95.5

displays the effectiveness of removing vanadium from SW during each cycle. The fact that the performance of the adsorbent did not decrease noticeably after five cycles of use and regeneration suggests that the chemical and mechanical reliability of synthetically supported nano-zero-valent iron for treating vanadium-contaminated drinking water. After five rounds of adsorption and desorption, there didn’t seem to be any decrease in the effectiveness of vanadium removal. In contrast, the partial dissolution of the adsorbent caused by the regeneration process resulted in a 10.9% decrease in the effectiveness of the alumina-based adsorbents [53]. The coated iron (oxy)hydroxide was separated after the elution of the adsorbent during the backwashing process. Additionally, iron-coated sand decreased 13 to 20% of its ability to adsorb arsenic [54].

4. Conclusions

The adsorbent was fabricated by loading nanoscale zero-valent iron onto activated carbon. The batch adsorption studies were used to test the effectiveness of the NZVI/AC adsorbent in removing vanadium from simulation water, and the removal of vanadium was successful with reasonably quick kinetic processes. While sulphate and humic acid had less of an impact, the presence of phosphate or silicate significantly reduced the efficiency of vanadium removal and Fe²⁺ was discovered to increase vanadium adsorption. Following a total of five adsorption-regeneration cycles, it was discovered that the adsorption capacity had not decreased, and the used adsorbent may successfully be eluted and restored by using diluted sodium hydroxide (0.1 M).