Cellulose Acetate–PHB Biocomposite from Saccharum officinarum for Ni (II) Adsorption: Equilibrium and Kinetics

Abstract

This research work focused on the development of an adsorbent biocomposite material based on polyhydroxybutyrate (PHB) and cellulose acetate derived from sugarcane (Saccharum officinarum) fibre, through cellulose acetylation. The resulting material represents both an accessible and effective alternative for the treatment and remediation of water contaminated with heavy metals, such as Ni (II). The biocomposite was prepared by blending cellulose acetate (CA) with the biopolymer PHB using the solvent-casting method. The resulting biocomposite exhibited a point of zero charge (pHpzc) of 5.6. The material was characterised by FTIR, TGA-DSC, and SEM analyses. The results revealed that the interaction between Ni (II) ions and the biocomposite is favoured by the presence of functional groups, such as –OH, C=O, and N–H, which act as active adsorption sites on the material’s surface, enabling efficient interaction with the metal ions. Adsorption kinetics studies revealed that the biocomposite achieved an optimal adsorption capacity of 5.042 mg/g at pH 6 and an initial Ni (II) concentration of 35 mg/L, corresponding to a removal efficiency of 86.44%. Finally, an analysis of the kinetic and isotherm models indicated that the experimental data best fit the pseudo-second-order kinetic model and the Freundlich isotherm.

1. Introduction

The concern about environmental pollution and its effects on human health and safety has generated significant interest in recent years in the development of new alternatives for restoring environmental wellbeing, such as the design and implementation of biomaterials [1]. The utilisation of agricultural and agro-industrial waste materials as biomaterials for pollutant adsorption in aquatic environments has become a low-cost and readily accessible option [2].

The continuous development of industry, the discharge of production residues, and improper disposal into water bodies are major sources of heavy metal contamination [3]. Elevated concentrations of heavy metals, such as nickel, chromium, lead, and cadmium in water sources have raised alarms due to their high toxicity and persistence, severely affecting water quality, aquatic fauna, and particularly human health, as they may cause physiological disorders that can ultimately lead to diseases including skin lesions, acute gastroenteritis, cancer, hypertension, and renal failure [4,5].

Nowadays, owing to the ease of obtaining residual biomasses derived from the agricultural industry for novel applications, numerous studies have been conducted on adsorbent biomaterials, most of which are subjected to chemical or physical modification processes to enhance their adsorption capacity [6]. Natural fibres have attracted considerable research interest due to their remarkable properties, including low density, low cost, ready availability, biodegradability, and ease of processing [7]. Additionally, they exhibit notable mechanical and acoustic properties, such as high fracture resistance, and their lignocellulosic nature ensures the presence of various functional groups, which contribute to an improved adsorption process [8].

Among the studies conducted is that of Harripersadth et al., which investigated the adsorption efficiency of sugarcane bagasse in removing Pb (II) and Cd (II) from aqueous solutions. The characterisation of the materials by FTIR and XRD revealed the presence of carbonyl functional groups in the biomaterial. The adsorption equilibrium was examined by varying the initial contaminant concentration to determine the equilibrium and saturation points of the biomaterial during the adsorption process. The pseudo-second-order model was applied to describe the adsorption kinetics of Pb and Cd ions by varying the contact time. The Langmuir and Freundlich isotherm models were employed to mathematically describe the adsorption isotherms, indicating the formation of a monolayer of metal ions on the adsorbent surface. The maximum adsorption capacities of Pb and Cd, based on 1 g of biosorbent, were 31.45 mg/g and 19.49 mg/g, respectively, for sugarcane bagasse, demonstrating good heavy metal adsorption efficiency and establishing it as a promising bioadsorbents [9]. Similarly, Ding et al. synthesised a bacterial composite based on cellulose and PHB (BC/PHB) through the co-cultivation of Gluconacetobacter xylinus and Ralstonia eutropha. The prepared material was evaluated for Cu²⁺ adsorption at pH 5, achieving a removal capacity of up to 37.86 mg/L [10]. In another study, Zhao et al. developed a composite material for the adsorption of Co (II) and Ni (II) ions in solution. The material was prepared using cellulose modified with L-cysteine, sodium alginate, and polyethyleneimine. Their findings showed that the adsorbent was able to remove approximately 90% of the contaminants [11]. Undoubtedly, several studies have demonstrated that the modification of cellulose enhances its physicochemical properties for the adsorption of contaminants in aqueous media. For instance, Matsedisho et al. prepared an adsorbent using cellulose extracted from orange peels, modified with phosphoric acid and sodium hydroxide, for application in the removal of Ni (II) from wastewater. Their results indicated that modification with phosphoric acid yielded the highest adsorption capacity for Ni (II), reaching 37.5 mg/g at pH 5 [12]. While these studies highlight the promising adsorption capacities of such materials, it is important to note that the nature of the metal ions plays a critical role in the removal processes, potentially limiting their interaction with the adsorbent. Moreover, surface modifications can also affect or reduce the availability of active sites. Therefore, although agro-industrial waste-based adsorbent materials have been widely explored as alternatives for the treatment of contaminated water, there remains a significant gap in the literature regarding the preparation of materials combining cellulose acetate (CA) specifically extracted from sugarcane bagasse with polyhydroxybutyrate (PHB) for the removal of heavy metals, such as Ni (II), in aqueous solution. This represents an unexplored approach in the scientific literature. Consequently, the present study proposes the use of sugarcane bagasse as a cellulose source for the development of a biocomposite material consisting of cellulose acetate (CA) and PHB as the polymer matrix, prepared via the casting method. The use of CA is justified by its favourable physicochemical properties and improved compatibility with biopolymers, as well as its surface bearing acetylated functional groups that may enhance interactions with metal contaminants. Additionally, the incorporation of PHB not only contributes to the structural stability of the material but also supports its biodegradable nature and environmental applicability. In the absence of prior reports on this specific combination for Ni (II) removal, this study offers an original contribution that could open new avenues in the development of more efficient and sustainable adsorbents for the treatment of heavy metal-contaminated water. Adsorption and kinetic tests were conducted in a batch system using a nickel sulphate (NiSO₄) solution, with quantification performed by atomic absorption spectrometry (AAS). This research presents a reliable and reproducible process, making a significant contribution to the utilisation of agricultural residues for the fabrication of composite materials capable of removing metallic ions from aqueous solutions.

2. Materials and Methods

2.1. Materials

Natural fibre, such as sugarcane bagasse fibre, was sourced from rejected materials or agricultural residues collected in the Bolívar Department (Colombia). Sodium hydroxide (NaOH) was employed to pretreat the fibres and for pH adjustment, during which hydrochloric acid (HCl) was also used. For the preparation of the biocomposite, acetic acid (C₂H₄O₂), sulphuric acid (H₂SO₄), and acetic anhydride (C₄H₆O₃) were used to synthesise cellulose acetate (CA) from the selected fibre. Subsequently, the product of this reaction was subjected to ultrasonic treatment in which two glass containers were placed, as follows: one containing a mixture of dichloromethane (CH₂Cl₂) and polyhydroxybutyrate (PHB) and the other containing a mixture of cellulose acetate and acetone (C₃H₆O). The resulting mixtures were combined in a Petri dish. For the preparation of the Ni (II) solutions, nickel sulphate (NiSO₄) and distilled water were used.

2.2. Obtaining Cellulose Acetate from Sugarcane Bagasse Cellulose

Figure 1 illustrates the process for obtaining cellulose acetate (CA), where 25 mL of acetic acid was initially mixed with 10 g of cellulose under magnetic stirring for 30 min at 35 °C and 160 rpm. Subsequently, 40 mL of acetic acid and 0.08 mL of sulphuric acid (H₂SO₄) were added to the mixture, which was stirred for 45 min under the same conditions. Thereafter, 40 mL of acetic anhydride and 0.6 mL of sulphuric acid were added and stirring continued for 1.3 h to homogenise the mixture. Finally, 10 mL of distilled water and 20 mL of acetic acid were added to the resulting composition, which was then stirred for 14 h. The mixture was subsequently filtered under vacuum, washed with distilled water until neutral pH was achieved, and dried in a Thermo Scientific oven at 70 °C for 2 h to obtain the CA paste. This paste was then ground to standardise the particle size [13].

The reaction mechanism involved in the acetylation of cellulose consists of the hydroxyl groups present in the cellulose molecular structure reacting with C₄H₆O₃ in the presence of H₂SO₄ as a catalyst. This reaction leads to the formation of acetyl groups, converting cellulose into cellulose acetate (CA), as illustrated in Figure 2.

2.3. Synthesis of the Composite Biomaterial (CA/PHB)

For the preparation of the CA/PHB composite biomaterial, 1.8 g of polyhydroxybutyrate (PHB) and 50 mL of dichloromethane (DCM) were added to a glass container, while 5 g of previously obtained cellulose acetate (CA) were placed in another container with 50 mL of acetone. Both mixtures were subjected to ultrasonic treatment at 60 °C for 40 min. Subsequently, 50 mL of the CA/acetone solution and 17 mL of the PHB/DCM solution were simultaneously poured into a Petri dish, as illustrated in Figure 3. The proportion used between CA and PHB was selected based on preliminary studies conducted by the research group, aiming to achieve a balance that ensured good structural integrity, adequate mixture homogeneity, and a sufficient presence of active functional groups to promote the adsorption process. Finally, the mixture was left in a fume hood until the volatile compounds evaporated, yielding the desired biomaterial [14,15].

2.4. Determination of the pH at the Point of Zero Charge (pHpzc)

The pHpzc of the CA/PHB biomaterial was determined to identify the pH at which the composite exhibits optimal contaminant retention capacity [6]. For this purpose, the following two 0.1 M solutions were prepared: sodium hydroxide (NaOH) and hydrochloric acid (HCl). Subsequently, ten distilled water solutions were prepared, with pH values ranging from 2 to 11. Then, 10 mL of each solution were added to separate test tubes, each labelled according to its initial pH, to which 100 mg of the CA/PHB biocomposite was added. The test tubes were agitated at 180 rpm for 24 h at room temperature. After this period, the supernatant from each test tube was extracted, and the final pH was measured. Each test was performed in duplicate, and the results averaged. The pHpzc value of the biocomposite was determined as the x-axis intercept by plotting the change in pH (∆pH) against the initial pH [16].

2.5. Characterisation of the Biocomposite CA/PHB

The biocomposite was subjected to various characterisation techniques using 1 g samples prepared both before and after the adsorption process, under the optimal conditions established for the evaluation of its physicochemical properties. Fourier transform infrared spectroscopy (FTIR) was performed to identify the functional groups present in the materials using an IRAffinity-1 FTIR spectrophotometer (Shimadzu Corporation, Kyoto, Japan. Serial No. A213749), with a spectral range of 450 to 4500 cm⁻¹. To determine the surface morphology of the biocomposite, scanning electron microscopy (SEM) was conducted using a TESCAN MIRA 3 microscope operated at an accelerating voltage of 15 kV. The samples were coated with a thin layer of gold to enhance electrical conductivity and improve image resolution. Additionally, thermogravimetric analysis (TGA) was performed to assess the characteristics and composition of the CA/PHB biomaterial, including decomposition and evaporation rates, oxidation, and overall purity. The analysis was carried out using a TA INSTRUMENTS thermogravimetric analyser (Serial No. 0600-11099, Model: SDTQ600), under a nitrogen atmosphere, with a controlled heating rate applied to the dried samples [17].

2.6. Evaluation of the Adsorption Capacity of the CA/PHB Biocomposite

For the evaluation of Ni (II) adsorption capacity, three solutions were prepared in test tubes with varying concentrations of 15 ppm, 25 ppm, and 35 ppm, while maintaining constant pH and adsorbent dosage, as detailed in

Table 1. Value range to be evaluated during experimentation.

Variables	Range
	−1	0	1
Initial contaminant concentration (mg/L)	15	25	35
Amount of adsorbent (g)	-	0.03	-

. For pH adjustment, 20 mL of each solution was taken and adjusted to the desired pH, based on the previously determined pHpzc. Subsequently, 0.03 g of the biocomposite was weighed, and 5 mL of each solution was added to separate containers. The containers were then placed in a shaker at 40 rpm for 24 h. After mixing, the samples were analysed using atomic absorption spectrometry (AAS) to determine the Ni (II) removal capacity at each initial concentration.

2.7. Identification of Adsorption Mechanisms

For the identification of the adsorption mechanisms, seven containers were prepared, each containing 0.03 g of CA/PHB, which was mixed with 5 mL of nickel (II) sulphate solution at the optimal concentration previously determined by AAS analysis, with the pH adjusted accordingly. The containers were placed in a shaker, and the mixing times were varied as follows: 5 min, 10 min, 15 min, 30 min, 1 h, 2 h, and 24 h. After each mixing period, the samples were filtered and analysed by AAS. The resulting data were then fitted to the pseudo-first-order, pseudo-second-order, Elovich, and intraparticle diffusion kinetic models, corresponding to Equations (1), (2), (3), and (4), respectively [18,19,20].

$$ln\left(q_e-q_t\right)=ln{q}_e-k_{1}t$$

(1)

where $q_e\mathrm{a}\mathrm{n}\mathrm{d}{q}_t$ are the adsorption capacities in each time t $\left(\mathrm{m}\mathrm{g}/\mathrm{g}\right)$. $k_1$ represents the pseudo-first-order constant (mg/g min).

$${\displaystyle \frac{t}{q_t}}={\displaystyle \frac{1}{k_2{q_e}^2}}+{\displaystyle \frac{t}{q_e}}$$

(2)

where $k_2$: pseudo-second-order constant (mg/g min).

$$q_t={\displaystyle \frac{1}{\beta }}\log \left(\alpha \beta t+1\right)$$

(3)

where $\alpha$: Elovich constant (mg/g min), $\beta :$ Elovich exponent.

$$q_t=k_3\sqrt{t}$$

(4)

where $k_3$: diffusion constant.

2.8. Equilibrium in the Removal of Ni (II)

Equilibrium analysis was conducted by examining the experimental data obtained from the previous procedures and fitting these data to the Langmuir, Freundlich, and Dubinin–Radushkevich isotherm models, corresponding to Equations (5), (6), and (7), respectively [21,22,23].

$$q_e=q_{max}\times {\displaystyle \frac{b{C}_{eq}}{1+b{C}_{eq}}}$$

(5)

where $q_{max}$: amount of metal needed to form a monolayer on the surface (mg/g), $b$: asortion/desorption ratio (L/mg).

$$q_e=k_F\times {C_{eq}}^{\left({\displaystyle \frac{1}{n}}\right)}$$

(6)

where $K_F$: Freundlich equilibrium constant (mg L/g²), $n$: adsorption intensity.

$$q_e=q_{DR}\times e^{-k_{DR}{\epsilon }^2};\epsilon =RT\times ln\left(1+{\displaystyle \frac{1}{C_{eq}}}\right)$$

(7)

where $\epsilon$: temperature-based Polanyi potential, $k_{DR}$: Dubinin–Radushkevich constant related to adsorption energy (mol²/kJ²), $R$: ideal gas constant (J/mol·K), $T$: temperature (K).

3. Results and Discussion

3.1. Synthesis of the Biocomposite from Cellulose Acetate and Polyhydroxybutyrate (CA/PHB)

To synthesise CA/PHB, 10 g of cellulose was initially subjected to an acetylation process to obtain cellulose acetate (CA), following the established methodology. The amount of CA obtained after acetylation was 9.45 g, corresponding to a yield of 94.5%. Subsequently, 1.8 g of polyhydroxybutyrate (PHB) was mixed with 5 g of CA previously dissolved in 50 mL of dichloromethane and acetone, respectively, and treated at 60 °C for 40 min, as specified in the methodology, resulting in a CA:PHB weight ratio of 2.78:1. Two batches of the biomaterial were fabricated, yielding 9.38 g of CA/PHB after grinding in a blade mill. The yield obtained in the synthesis process indicates that it was successfully conducted from start to finish. The resulting material was a homogeneous powder, attributed to the properties of the cellulose and biopolymer used, as shown in Figure 4. These results are consistent with those previously reported in the literature; for instance, Villabona-Ortíz et al. [24] reported an average yield of 89.2% for CA obtained from coconut fibre, while Guerrero [25] obtained CA from Festuca arundinacea residues with an average yield of 92.08%.

3.2. Determination of the pH at the Zero Load Point (pHcc) for CA/PHB

The pH at the point of zero charge (pHpzc) was determined to identify the pH at which the biocomposites surface charge is neutral. The experiment was conducted by varying the pH from 2 to 11 over a 24-hour period. The results, shown in Figure 5, indicate a pHpzc value of approximately 5.6, meaning that this is the pH at which the surface charge of the CA/PHB biomaterial is zero. This implies that at pH values below the pHpzc, the surface of the biomaterial is positively charged, thereby favouring the adsorption of negatively charged contaminants through electrostatic attraction. Conversely, at pH values above the pHpzc, the surface carries a negative charge, promoting the adsorption of positively charged contaminants. Given that Ni (II) exists in a +2 oxidation state, its adsorption is thus favoured at pH values above 5.6 [26].

3.3. Evaluation of Adsorption Capacity

Following the methodology described above, a pH value of 6 was selected as the reference and starting point for the Ni (II) adsorption tests, based on the pHpzc determination and supported by various studies reported in the literature that identify this pH as optimal for nickel adsorption using sugarcane biomass (see

Table 2. Optimal pH conditions for heavy metal adsorption tests reported in the literature implemented sugarcane-based biomaterial.

Metal Ion	Optimum Conditions	Reference
Cd2+	pH = 6	[28]
Pb2+, Ni2+	pH = 6	[29]
Cd2+	pH = 6–7	[30]
Pb2+, Cu2+, Cd2+, Ni2+ y Cr3+	pH = 5	[31]
Ni2+	pH = 6	[32]
Pb2+	pH = 6	[33]

). Additionally, a dose of 30 mg of the CA/PHB biocomposite was used, as previous experiments conducted by the Chemical Engineering research group at Universidad de Cartagena demonstrated that this dosage consistently yielded optimal adsorption results [6,27].

Based on the above, the adsorption capacity of the CA/PHB biocomposite was evaluated by varying the initial concentration of the contaminant. Three solutions with concentrations of 15, 25, and 35 ppm were prepared. For each concentration, three samples consisting of 5 mL of solution and 30 mg of biocomposite were placed under constant agitation for 24 h, then filtered and analysed using atomic absorption spectroscopy (AAS). The results, presented in Figure 6, indicate that the biocomposite achieves the highest removal at a concentration of 35 mg/L, with an adsorption capacity of up to 5.042 mg/g and an efficiency of 86.44%.

3.4. FTIR Analysis of the Biomaterial Before and After the Ni (II) Adsorption Process

FTIR analysis before and after the Ni (II) adsorption process (Figure 7) reveals changes in the functional groups present on the surface of the material, suggesting mechanisms of interaction with the metal ions. The CA/PHB biocomposite prior to Ni adsorption (Figure 7a) exhibits subtle peaks between 3650 and 3450 cm⁻¹, indicative of N–H amines. Additionally, peaks corresponding to C–H bonds appear in the range 3000–2900 cm⁻¹, alongside a prominent peak at 1750 cm⁻¹ attributed to the C=O bond. Further peaks at 1370, 1250, and 1075 cm⁻¹ also indicate C–H bonds, reflecting the molecular stability of the material. At lower frequencies, aromatic C–H ring vibrations are observed between 900 and 450 cm⁻¹. In the FTIR spectrum of the biocomposite following Ni ion adsorption (Figure 7b), many features similar to (Figure 7a) are maintained; however, O–H functional groups emerge between 3950 and 3700 cm⁻¹, and stronger C–H peaks appear within 3000–2900 cm⁻¹. Amine-related peaks are evident at 1420 cm⁻¹, which contribute to neutralising acidity in the adsorbed solutions and maintaining pH balance during Ni adsorption. Finally, the persistence of aromatic C–H ring vibrations at lower frequencies suggests the biocomposite retains its structural stability [34,35]. These changes suggest that both O–H and N–H groups participate in the interaction with Ni (II) ions, through coordination bonds or electrostatic interactions. In addition, the possible involvement of the carbonyl group (C=O), which remains present after adsorption, may also contribute to the process via dipole–ion interactions, whereby the electric field generated by the metal ion induces attraction with the dipole of the functional group. The presence of these spectral changes in the FTIR analysis after adsorption—such as the strengthening or shifting of characteristic bands—supports the idea that these functional groups are not only present but actively participate in the contaminant retention mechanism [36]. Taken together, these results indicate that the adsorption of Ni (II) occurs primarily through specific interactions between the metal ions and the active functional groups on the surface of the bioadsorbents.

3.5. TGA Analysis of CA/PHB Before and After Ni (II) Adsorption Process

To investigate the thermal properties and stability of the biocomposite before and after the adsorption process (Figure 8), thermogravimetric analysis (TGA) was conducted over a temperature range of 0 °C to 600 °C, with a heating rate of 10 °C/min and a nitrogen flow rate of 20 cm³/min. In Figure 8a, three distinct stages of weight loss can be observed. The first stage, occurring between 23 and 57 °C, shows an approximate 5% weight loss, attributed to the evaporation of moisture from the sample [37]. Following this, the biomaterial remains thermally stable up to 260 °C. The second stage, between 259 and 377 °C, involves a weight loss of around 80%, corresponding to the thermal degradation of cellulose and other organic components in the biocomposite [38]. The third stage, from 377 to 600 °C, shows an additional weight loss of approximately 10%, which is attributed to the degradation of residual compounds and carbonisation of the material [39].

After the adsorption process (Figure 8b), two major degradation stages are evident. The first, between 22 and 85 °C, exhibits a 35% weight loss, which may be attributed to increased water retention in the material—possibly due to the presence of a greater number of available or activated hydrophilic groups (such as O–H) following metal interaction, indicating a modification of the biocomposite surface. The second stage occurs between 275 and 378 °C, with a weight loss of 53%, attributed to the thermal degradation of the organic matrix. Although this stage occurs within a similar temperature range as observed before adsorption, the reduced relative weight loss may suggest that part of the material is now stabilised by the presence of the metal ion, or that certain internal bonds have been reinforced due to Ni (II) incorporation [40].

Taken together, these changes support the idea that the adsorption process was not merely physical but involved chemical interactions that affected the composition and thermal stability of the material—an interpretation that is also consistent with the FTIR findings [41].

3.6. SEM Analysis of the Biocomposite Before and After the Ni (II) Adsorption Process

Figure 9 presents SEM micrographs of the CA/PHB biocomposite before and after the Ni (II) adsorption process at 3000× and 10,000× magnification. In Figure 9a, prior to Ni (II) removal, the surface appears irregular and agglomerated, with well-defined pores, likely due to the powder-like texture of the components during biomaterial synthesis [42]. In Figure 9b, following the adsorption process, several lighter regions can be observed, indicating areas of increased density associated with Ni (II) deposition on the biocomposite surface. Additionally, a distinct white spot suggests the presence of other elements, possibly residues or concentrated metallic sites [43].

3.7. Identification of the Adsorption Mechanisms Using the CA/PHB Biocomposite

To elucidate the adsorption mechanisms, the contact time was varied, seven samples were prepared, all at a concentration of 35 mg/L, which was identified as the concentration at which the highest adsorption capacity was recorded. Each sample was mixed with 30 mg of the CA/PHB biocomposite. The contact times varied from 5 min to 1440 min, covering a total period of 24 h, after which the samples were analysed using the AAS method. It was observed that, although the highest removal percentage occurred in the 15-min sample, the removal efficiency was already notably high from 5 min of contact, indicating excellent performance of the synthesised biocomposite in the adsorption of Ni (II) ions. The results obtained were fitted to the pseudo-first-order, pseudo-second-order, Elovich, and intraparticle diffusion models, using Equations (1)–(4), and are presented in Figure 10.

By analysing the fitting parameters of the kinetic models for Ni (II) adsorption, it can be observed that the models showing the best fit are the Elovich, pseudo-first-order, and pseudo-second-order models, with the latter providing the best mathematical correlation with the experimental data (see

Table 3. Adjustment parameters of Ni (II) adsorption kinetics.

Kinetic Model	Parameter	Non-Linear Adjustment
Pseudo first order	qe	24.7861
	K1	1.2131
	R2	0.99996
	Square Error	0.0158
Pseudo second order	qe	24.8081
	K2	2.2353
	R2	0.99997
	Square Error	0.01341
Elovich	$\beta$	3.7951
	$\alpha$	7.6915
	R2	0.99905
	Square Error	0.4187
Intraparticle diffusion	K3	1.4888
	R2	0.2009
	Square Error	549.6321

). Therefore, it can be inferred that the adsorption surface is heterogeneous and that interactions occur between the biocomposite and the contaminant. Furthermore, the process is governed by the number of active sites available on the surface of the adsorbent. Furthermore, beyond the kinetic behaviour, it is proposed that the adsorption mechanism primarily involves interactions between the functional groups present in the biocomposite—such as hydroxyl (–OH) and carbonyl (C=O) groups—and Ni (II) ions in solution. These interactions may occur through coordination bonds or electrostatic forces, suggesting a predominant chemisorption process. The presence of amino groups (–NH) may also facilitate the formation of surface complexes with the metal ions, promoting more stable adsorption. This interpretation is consistent with the pseudo-second-order kinetic model, which is typically associated with chemically controlled adsorption mechanisms [44,45].

Likewise, to study the behaviour of the experimental data in the adsorption equilibrium and to describe the analytical relation that exists between the contaminant and its initial concentration, the data were fitted to the Langmuir, Freundlich, and Dubinin–Radushkevich models by implementing Equations (5)–(7) (see Figure 11). By observing the fitting parameters shown in

Table 4. Adsorption isotherm parameters.

Isotherm Model	Parameter	Non-Linear Adjustment
Langmuir	qm	35,420.125
	b	2.91609 × 10−5
	R2	0.96579
Freundlich	kf	0.82124
	n	0.85708
	R2	0.99929
Dubinin–Radushkevich	qdr	6.88887
	kdr	9.36952
	R2	0.95484

, which indicate the most adequate nonlinear fit of the experimental data, it can be concluded that the Freundlich isotherm model is the best fit, which indicates that adsorption tends to occur on the surface of the biocomposite as multilayers, meaning that as the contaminant is adsorbed, it is successively arranged in layers [45]. However, the value obtained for the parameter n in the Freundlich model results in a 1/n value slightly above the typical range (0.1–1) that indicates favourable physical and multilayer adsorption, suggesting low favourability of the process on the bioadsorbents [46]. Nevertheless, the high correlation coefficient (R² = 0.99929) suggests that the model still adequately describes the behaviour of the system [22]. This outcome may be associated with more complex interactions between the active sites of the adsorbent and Ni (II), possibly involving some degree of chemical processes or site saturation effects at higher concentrations [47].

Finally, the results obtained throughout this study are compared with other bioadsorbents reported in the scientific literature for the removal of contaminants in solution.

Table 5. Comparative characterisation of chemically modified biocomposites made from different biomasses.

Precursor	Modifying Agent/Reinforcing Matrix	Pollutant Removed	Removal Capacity (mg/g)	Removal (%)	pH	Adsorbent Dosage (mg)	Initial Concentration of the Contaminant (ppm)	Reference
Sawdust cellulose	NaNO2/NaHCO3	Cd2+	99.9–2206.9	44.1–99.9	7.10	20	1–50	[48]
		Ni2+	49.98–956.6	38.2–99.1
Brown marine algae (Phae-ophyceae)	CaCl2; activation with EDC and NHS	Pb2+	369.6	90	4	19	500	[49]
		Cu2+	124.1		5		200
Natural raffia fibres (Raphia farinifera)	CaCl2; activation with Fe3O4	Cd2+	16.34	87.3	6	100	150	[50]
Sugarcane cellulose	Sodium alginate	Methylene blue	4.27	85.33	8	200	10	[51]
Banana rachis (M. oranta)	Bituminous coal	Ni2+	328.7	67.95	4.9	-	25–45	[36]
		Congo Red	478.3	76.65
Sugar cane bagasse (Saccharum officinarum)	PHB	Ni2+	5.042	86.44	6	30	35	Present study

presents adsorption data from previous studies, aiming to evaluate the performance of the biocomposite synthesised in the present research and to provide a useful benchmark for contextualising the efficiency of the proposed material. It is important to note that, among the reported studies, operational variables play a crucial role in adsorption processes. The biocomposite developed in this study demonstrated superior performance, representing a scarcely explored approach that combines cellulose acetate derived from sugarcane bagasse with PHB for the removal of nickel. This opens new possibilities in the design of sustainable and functional materials for the treatment of contaminated water.

The table above shows that the CA/PHB biocomposite studied performs well in removing metal ions from aqueous solutions, which can be the basis for further studies and even large-scale implementations.

4. Conclusions

• The incorporation of the results showed that the synthesis efficiency of cellulose acetate (CA) from commercial cellulose using the acetylation method is high, with a yield percentage of 96.32%. This demonstrates the enormous potential of physicochemical modifications of cellulose in terms of cost and performance.
• It was found that the adsorption capacity q_e of the CA/PHB composite biomaterial was excellent, demonstrating a result of 5.042 mg/g at a concentration of 35 mg/L of Ni (II), resulting in a removal rate of 86.44%. This indicates that the composite biomaterial presents suitable and promising properties for its application in heavy metal adsorption processes.
• The pseudo-second-order adsorption kinetic model was found to provide the best fit to the experimental data, with an R² = 0.99997, suggesting an adsorption mechanism dominated by chemical interactions. In the adsorption equilibrium study, the Freundlich isotherm model was found to provide the best fit to the data.
• The efficiency of the CA/PHB biocomposite is notable, since considering the results of the adsorption kinetics analysis, it can be observed that, from 5 min in terms of contact time with the contaminant, the biocomposite presents adsorption capacities similar to those of the same sample with a contact time of 24 h.
• The characterisations revealed information on the presence of functional groups C–H, C=O, O–H, and C–N, which, considering the FTIR analysis, indicate their involvement in the adsorption process of Ni (II) ions. TGA analysis revealed three stages of degradation, showing a decrease in the thermal stability of the biocomposite after adsorption, suggesting an effective interaction between the material and the metal ions. Meanwhile, SEM revealed a porous surface before adsorption and morphological changes after the process, supporting the retention of Ni (II) on the adsorbent surface.