Introduction

The presence of contaminants, encompassing pharmaceuticals, within water sources poses an ecological hazard that has become progressively concerning in recent decades (Ezzariai et al. 2018). Recently, the utilization of pharmaceuticals and individuals’ accessibility to diverse drug varieties have escalated, primarily attributed to the proliferation of diseases, advancements in medical sciences, and pharmaceutical innovations worldwide. Consequently, the apprehension regarding the presence of these substances in aquatic ecosystems has intensified. According to literature, pharmaceutical contaminants are detectable in the majority of water bodies, including rivers, lakes, and seas. This observation highlights the problem of biomagnification and the potential development of drug resistance over an extended period (Vaudreuil et al. 2024; Moslehi et al. 2024a). Typically, these compounds are released to the environment via wastewater discharge of effluents from conventional wastewater treatment plants, which has led to increased concern about the presence of pharmaceutical substances in aquatic environments (Moslehi et al. 2024b). Nowadays, more than 100,000 tons of antibiotics are consumed worldwide every year, along with a large amount of wastewater with antibiotics discharged into the ecological environment. The pressure on sewage treatment plants would be increased. Antibiotics in the ecosystem not only endanger the survival of freshwater organisms, but also continue to enrich in the human body, damaging the human microbial environment and increasing the resistance of bacteria.

Antibiotics, including tetracycline (TC) and amoxicillin, are among the cases that contribute to health concerns by inducing bacterial resistance and disrupting the balance of the microbial population (Amarzadeh et al. 2023). These contaminants enter water sources through multiple pathways, including human and animal waste, direct sewage disposal, and the improper disposal of medical, veterinary, and pharmaceutical waste. Tetracycline, with the chemical formula of C22H24N2O8, has a molecular mass of 444.435 g/mol (Rivera-Utrilla et al. 2013) (Fig. S1). This bacteriostatic antibiotic exhibits an impact on gram-positive and gram-negative bacteria, as well as, to a certain extent, mycoplasmas and certain fungi. Its action mechanism involves disrupting the protein synthesis process in bacteria, thereby inhibiting their growth and reproduction (Salman et al. 2022a, b; Aliyu et al. 2022). Tetracyclines accumulate in developing bones and teeth, binding to calcium and resulting in tooth discoloration and gum hypoplasia. The uptake of tetracycline by humans and animals after consumption is limited and inconsistent. Any unabsorbed tetracycline is exerted as unchanged compounds into domestic sewage, eventually leading to environmental contamination (Eniola et al. 2020; Miao et al. 2019).

A range of technologies including advanced oxidation processes, biological treatment, and membrane processes are employed for the elimination of chemical contaminants from water sources and industrial waste. These processes have advantages and disadvantages (Foroutan et al. 2022; Li et al. 2020). For instance, the utilization of technologies of membrane processes (reverse osmosis and nanofiltration) often necessitates substrate regeneration or replacement, as well as the treatment of secondary saline effluents (Kamranifar et al. 2021). Furthermore, microbial processes exhibit slow performance and, in certain instances, incomplete performance compared to chemical reduction. On the other hand, the aforementioned processes may not be justifiable due to factors like high operating expenses, suboptimal efficiency, complexity, and potential adverse environmental ramifications (Hassanzadeh et al. 2019).

Adsorption is a viable method for eliminating antibiotics. In this process, achieving the maximum capacity of contaminant has always been a pivotal objective, and this can only be accomplished by enhancing surface accessibility (Eniola et al. 2020; Mao et al. 2024; Zhao et al. 2024). Consequently, researchers are actively seeking nanostructures capable of significantly amplifying surface area. By reducing the size of material particles, the surface-to-volume ratio can be augmented (Hassanzadeh et al. 2019). Presently, nanoscale technology in the domain of antibiotic removal is primarily focused on the development of magnetic nanocomposites. These materials offer several advantages over traditional adsorbents, such as their inherent selectivity in the removal of target pollutants, robust magnetic responsiveness, eco-friendly nature, stability under varying environmental conditions, cost-effectiveness in fabrication, rapid manufacturing capability, reusability, exceptional performance, and ease of separation (Aliyu et al. 2022). Given the adverse environmental impacts caused by antibiotic residues, and taking into account that the utilization of magnetic nanocomposites for water purification represents cutting-edge research globally, this study employed a green magnetic nanocomposite, namely CuFe2O4 coated with CuS (Eniola et al. 2020; Nasiri et al. 2021a). Green synthesis is an eco-friendly approach that minimizes environmental harm by substituting plant-based materials for chemicals (Król et al. 2019). Different plants can be studied in the field of nanoparticle synthesis. These plants have various therapeutic and medicinal properties that are used as herbal medicine. But nowadays, according to the progress that has been made, scientists have been led to use plants in the field of nanoparticles. Therefore, with this method, the effectiveness of these plants can be increased. Plants contain various compounds such as alkaloids, flavonoids, roponins and other substances that act as reducing agents in the synthesis of nanoparticles. In general, nanoparticles are synthesized by various physical or chemical techniques, but all these methods are divided into two general categories: top-down and bottom-up methods. These methods are expensive and toxic and dangerous for the environment. Therefore, the use of plant extracts can be a suitable alternative to the physical and chemical methods of producing nanoparticles in the form of an environmentally friendly method. Recently, the synthesis of nanoparticles through environmentally friendly routes has become popular among researchers due to its low cost. Alhagi pseudalhagi (camel thorn) is a wild flowering plant of the Fabaceae family, which is abundant in Iran and many parts of the world (Ye et al. 2021). Besides, this plant contains various components like proteins, carbohydrates, flavonoids, phytosterols, saponins, tannins, and phycocyanin in its extract and therefore tried to use in synthesizing of CuFe2O4/Cus nanocomposite in this study. The objective of this investigation was to evaluate the efficiency of the newly synthesized green magnetic nanocomposite (CuFe2O4/CuS) for the adsorption of tetracycline antibiotics from aqueous media. Surface and chemical properties of the nanocomposite, parameters affecting adsorption, kinetic and isotherm modeling, and thermodynamic study have been explored. The main aim of this study was green synthesized, CuFe2O4/Cus nanocomposite initially and then applied as a treatment agent for removal of tetracycline, due to its better compatibility with nature, easier synthesis process and saving time and money.

Experimental

Chemicals

All chemicals and reagents used in the fabrication process and adsorption tests were of analytical grade. The precursors such as iron (III) nitrate (purity:98%) [Fe(NO3)3.9H2O], sodium hydroxide (NaOH) (99.99%), hydrochloric acid (HCl)(37%), copper sulfate (CuSO4)(99.99%), ethylene glycol (C2H6O2)(100%), and sodium thiosulfate (Na2S2O3)(99.99%) were provided from Merck Company. The solutions needed for the adsorption experiments were provided by dissolving tetracycline hydrochloride salt (C22H24O8N2.HCl) (Sigma-Aldrich, purity > 95%) in deionized water (DIW). Other reagents were procured from Sigma-Aldrich Company. The A. pseudalhagi plant was collected from deserts around Birjand City, Iran.

Extraction of plants

Alhagi plants were collected by a plant specialist, and dried leaves were then extracted by percolation with methanol. First, an appropriate amount of the dried plant was immersed in a separating funnel containing a methanol solution for the reaction so that the dried plant was completely covered by the methanol solution. Then, the methanol solution was emptied every twice a day and lasted for 3 days. The solution obtained from the plant extract and methanol was placed in a rotary apparatus; finally, the desired extract was obtained from the Alhagi plant.

CuFe 2 O 4 green magnetic core (GMC) synthesis

To prepare the plant extract, the A. pseudalhagi leaves were thoroughly washed with DIW to remove any surface impurities, then dried under 30 °C for 48 h and crushed into small fragments. At first, 100 g of the dried was immersed in 300 mL methanol into the separating funnel for three days to get a homogeneous solution. Afterward, the solution obtained from the plant extract and methanol was placed in a rotary machine. The desired extract was obtained from the A. pseudalhagi plant by removing methanol from the solution.

For the green synthesis of CuFe2O4, an aqueous solution of iron (III) nitrate (1 mmol) with a fixed temperature of 70 °C was prepared. Then, 0.5 g of A. pseudalhagi plant extract was separately dissolved in 10 mL of DIW and poured drop-wised into the aqueous iron salt solution (Ye et al. 2021; Khormali et al. 2021). After that, 0.5 mmol of copper metal salt was dissolved in 10 mL of DIW and added drop by drop to the solution containing iron salt and the extract of A. pseudalhagi. The mixture pH was adjusted to 14 using the 1N NaOH solution (Barkhor et al. 2024; Moslehi et al. 2024a). Then, the mixture was continuously stirred for 2 h to conduct chemical reactions under the basic conditions. After the reaction, the sediments obtained were separated from the mixture using a magnet owing to its magnetic properties. The collected sediments were washed with DIW and ethanol three times to remove impurities. Then, the rinsed sediments were dried in an oven at 80 °C to remove any residual solvent. Eventually, the dried powder was calcined at 600 °C for 3 h to acquire the desired CuFe2O4 magnetic nonmaterial (Abdelfatah et al. 2024; Ebrahimzadeh et al. 2020).

CuFe 2 O 4 /CuS synthesis

To fabricate CuFe2O4/CuS MNC, 0.15 g of CuFe2O4 GMC was introduced into a beaker containing 20 mL ethylene glycol, and the obtained mixture was sonicated for 30 min. The resulting suspension was then transferred into a flask and put inside an oil bath at 125 °C. Subsequently, 0.8 g of CuSO4 was added to the aforementioned suspension. Following this, a mixture containing 1.9 g of Na2S2O3 and 20 mL of ethylene glycol was poured drop by drop into the flask containing CuFe2O4 GMC and CuSO4. The mixture was refluxed at 140 °C for 90 min. Upon completion of this step, the acquired material was separated using a magnet, washed multiple times with DIW and ethanol, and finally dried at 80 °C for 5 h(li et al. 2019; Majidi et al. 2023a, b). The steps of CuFe2O4/CuS MNC synthesis are schematically exhibited in Fig. 1. The efficiency of nanoparticle production was about 78%, and part of the particles was lost during separation or washing.

Fig. 1
figure 1

Preparation steps of CuFe2O4/CuS magnetic nanocomposites

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Characterization techniques

The crystallite nature of the fabricated samples was evaluated by X-ray diffraction (XRD) using a Philips XRD instrument provided with a Cu-Kα radiation source. Scans were conducted over a wide 2θ range from 10 to 80°. The chemical bonds were characterized via the Fourier-transform infrared (FTIR) technique utilizing a spectrometer (AVTAR 370, Thermo Nicolet Avatar). The shape and surface structure of the materials were assessed using field emission scanning electron microscopy (FESEM, Sigma VP-500, Zeiss) and transmission electron microscopy (TEM, EM10C-100 kV, Zeiss, Germany). The combination of FESEM and TEM imaging provided detailed visualization of the morphological features. To evaluate the surface area and pore properties of the as-prepared nanomaterials, the Brunauer–Emmett–Teller (BET) technique was conducted (Bel, Belsorp mini II). The magnetic properties of the prepared nanostructures were studied by vibrating sample magnetometer (VSM, Lake Shore Cryotronics, 7407, USA).

Adsorption experiments procedure

The TC adsorption on CuFe2O4/CuS MNC was conducted via batch mode at room temperature (25 ± 2 °C). To assess the adsorption efficiency, the purification experiments were performed under the following operational parameters: The initial pH of 3.0–9.0, the adsorbent quantity of 0.025–2 g/L, TC concentrations of 5–100 mg/L, and the temperature of 5–50 °C. The pH level of the solution was determined utilizing a pH meter (HACH, HQ411d) with adjustments made using either HCl or NaOH (0.1 M). At specific time intervals, 3 mL of samples were withdrawn from the reaction flask, separated with the magnet and then it was centrifuged for more certainty. The obtained sample was analyzed using a UV–Vis spectrophotometer (Model: CT06484) at a wavelength of 358 nm to determine the TC concentration. The adsorption efficiency and the equilibrium adsorption capacity (Qe, mg/g) were computed according to Eqs. (1) and (2). To confirm the repeatability of the findings, every phase of the experiment was repeated three times, and the average outcomes were reported (Beyki et al. 2016).

$${\text{Adsorption}} {\text{efficiency}} (\% ) = \left( {1 - \frac{{C_{t} }}{{C_{0} }}} \right) \times 100$$
(1)
$$Q_{{\text{e}}} = \frac{{\left( {C_{0} {-} C_{{\text{e}}} } \right) \times V}}{{{\text{Adsorbent}}\,{\text{mass}}\,({\text{g}})}}$$
(2)

where V reflects the volume of treated water in L, and C0, Ct, and Ce are TC concentration (mg/L) at the initial time, specific time, and equilibrium, respectively.

Calculation of thermodynamic, isotherm, and kinetic parameters

To assess the adsorption nature of TC molecules via CuFe2O4/CuS MNC, the thermodynamic parameters were calculated at different temperatures (5–50 °C) based on the following relations (Eqs. (3) and (4)):

$${\Delta G}^{^\circ }=-\text{RT ln}{K}_{d}$$
(3)

where ∆G° represents the standard Gibbs free energy change (kJ/mol), R denotes the universal gas constant (8.314 J/mol K), T is the absolute temperature (K), and Ko is the distribution coefficient (Ko = Qe/Ce).

The average enthalpy change of TC adsorption on CuFe2O4/CuS MNC was calculated utilizing the Van ’t Hoff equation:

$$\text{ln}{K}_{d}= \frac{{\Delta \text{S}}^{^\circ }}{\text{R}}- \frac{{\Delta H}^{^\circ }}{\text{RT}}$$
(4)

where ∆S° denotes the standard entropy (J/mol K), and ∆H° is the standard enthalpy change (kJ/mol) (Liu et al. 2015). ∆H° and ∆S° were determined from the slope and intercept of the plot of Ln Ko against 1/T, respectively.

To study the adsorption isotherms, the experimental data obtained were fitted to Langmuir and Freundlich models. The Langmuir and Freundlich models were presented in Eqs. (5) and (6), respectively (Ocampo-Pérez et al. 2012; Majidi et al. 2023a, b):

$${\text{Langmuir:}}\quad \frac{{C_{{\text{e}}} }}{{Q_{{\text{e}}} }} = \frac{1}{{Q_{{{\text{max}}}} K_{{\text{L}}} }} + \frac{{C_{{\text{e}}} }}{{Q_{{{\text{max}}}} }}$$
(5)
$${\text{Freundlich:}}\quad {\text{Ln}} Q_{{\text{e}}} = \left( \frac{1}{n} \right){\text{Ln}}C_{{\text{e}}} + {\text{Ln}} K_{{\text{f}}}$$
(6)

where Ce (mg/L) denotes the equilibrium concentration of TC adsorbed on CuFe2O4/CuS MNC, Qe (mg/g) represents the amount of TC adsorbed, and Qmax (mg/g) is the monolayer adsorption capacity of CuFe2O4/CuS MNC. KL (L/mg) and KF (mg/g) are Langmuir and Freundlich constants, respectively. Also, 1/n is a factor related to the heterogeneity of the adsorbent surface.

To obtain more detailed insight into the adsorption mechanism of TC on CuFe2O4/CuS MNC, the pseudo-first-order (PFO) and pseudo-second-order (PSO) were employed. The PFO model is commonly used to describe the adsorption kinetics in a liquid/solid system, while the PSO model assumes chemisorption is the predominant process controlling the overall adsorption rate. The correlation coefficient (R2) was applied as an indicator of the agreement between the experimental data and the models. The PFO model, also known as the Lagergren equation, is expressed as follows (Eq. (7)):

$$\text{Log}\left({Q}_{\text{e}}- {Q}_{t}\right)=\text{Log }{Q}_{\text{e}}-\frac{{k}_{\text{l}}}{2.303}t$$
(7)

where Qt and Qe denote the amount of substance adsorbed on CuFe2O4/CuS MNC at time t and equilibrium (mg/g), respectively. k1 represents the PSO rate constant (1/min). The values of k1 and R2 are determined by plotting Log (Qe − Qt) versus time (t).

The pseudo-second-order (PSO) is described as follows (Eq. (8)) (Zhang et al. 2011):

$$\frac{t}{{Q}_{t}}=\frac{1}{{k}_{2}{Q}_{\text{e}}^{2}}+\frac{1}{{Q}_{\text{e}}}$$
(8)

where k2 reflects the rate constant of pseudo-second-order (g/mg min). By plotting t/Qt against t, the velocity parameters can be obtained. The values of Qe and k2 are determined by calculating the slope and intercept from the origin (Gao et al. 2012; Keramatinia et al. 2022).

To scrutinize the stability of the synthesized samples, TC adsorption tests were conducted over five experimental cycles under the following conditions: TC concentration of 20 mg/L, adsorbent dosage of 1.5 g/L, pH 7.0, and temperature of 25 ± 2 °C. After each run, the nano-adsorbent was isolated from the treated water by employing an external magnet. Subsequently, it was subjected to ultrasonication with ethanol for 15 min to remove contaminants from the adsorbent pores. Eventually, the adsorbent was rinsed with DIW and vacuum-dried at 80 °C to reuse it in the subsequent cycle.

Results and discussion

Characterization results

XRD analysis

The XRD patterns of CuFe2O4 and CuFe2O4/CuS nanomaterials are depicted in Fig. 2. In the case of CuFe2O4, the characteristic peaks observed at 2θ values of 18.62°, 30.14°, 35.47°, 38.86°, 43.23°, 53.62°, 57.08°, 62.52°, 74.13°, and 75.22° can be related to (1 1 1), (2 2 0), (3 1 1), (2 2 2), (4 0 0), (4 2 2), (5 1 1), (4 4 0), (5 3 3), and (6 2 2) planes of cubic CuFe2O4 GMC (JCPDS No. 75–1517) and correspond to the Fd3m space group (Zhao et al. 2011; Younes et al. 2021). The strong intensity of (3 1 1) peak indicates that the spinel phase was likely the predominant crystalline component. As for CuFe2O4/CuS, a series of characteristic peaks at 2θ values of 28.87°, 29.37°, 31.87°, 32.87°, 48.12°, 52.87°, and 59.47° have been detected which can be indexed to (1 0 1), (1 0 2), (1 0 3), (0 0 6), (1 1 0), (1 0 8), and (1 1 6) crystal faces of CuS, respectively (JCPDS No. 06–0446) (Gao et al. 2012; Amir et al. 2019). Also, the two most intense peaks of CuFe2O4 ((3 1 1) and (4 0 0)) appeared in the XRD profile of CuFe2O4/CuS MNC, elucidating the presence of both particles in the fabricated nanocomposite and the successful synthesis of the CuFe2O4/CuS MNC. To determine the crystallite size of the as-mentioned nanocomposite, Scherer's relation was utilized (Eq. (9)) (Mandal et al. 2023; Heiba et al. 2021):

$$D=\frac{0.98\lambda }{\beta \text{cos}\theta }$$
(9)

where D implies the particle diameter, β reflects the peak width at half height, θ is the diffraction angle at the peak location, and λ is usually 0.1540 nm. In this method, it is assumed that the particles are stress-free; therefore, the crystallite size is determined from the most intense diffraction peak. Based on this equation, the crystallite size of the CuFe2O4/CuS MNC was estimated to be 38 nm.

Fig. 2
figure 2

XRD patterns of CuFe2O4 and CuFe2O4/CuS MNMs

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FTIR analysis

The FTIR spectra of CuFe2O4/CuS MNC at different pHs are provided in Fig. 3. In ferrite crystal structures, the metal ions can occupy two different types of interstitial sites within the lattice: tetrahedral (A) sites and octahedral (B) sites. Thus, the peak observed at 586 cm−1 corresponds to the stretching vibration of Fe–O in the tetrahedral sites, while the second peak at 380 cm−1 is associated with the stretching vibration of Fe–O in the octahedral sites of the CuFe2O4 spinel structure. The prominent sharp band at 615 cm−1 is related to the Cu–S bond. Besides, the peaks detected at 1000, 1069, and 1217 cm−1 can be indexed to the vibrational modes of SO42− ions. The absorption band at 1440.2 cm−1 corresponds to the vibration mode of the carbon-hydrogen (C–H) or (N–O) bond in the sample, which is more closely associated with the plant extract in the magnetic core. Additionally, the two peaks observed at 1618 and 3410 cm−1 can be attributed to the stretching vibration of O–H and bending vibration of O–H from absorbed water molecules, respectively (Mandal et al. 2023; Kamranifar et al. 2018). The vibrational characteristics of CuFe2O4/CuS MNC at different pHs (3.0, 5.0, 7.0, and 9.0) showed no significant change, indicating minimal release of metal ions. Eventually, the presence of these peaks in the FTIR spectra of the fabricated nanocomposite confirms the successful coating of CuS onto the CuFe2O4 structure.

Fig. 3
figure 3

FTIR spectrum of CuFe2O4/CuS MNC

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BET analysis

To comprehensively evaluate the surface area and the pores of the fabricated nanomaterials, the BET technique was utilized, and the obtained results are displayed in Fig. 4. Based on this figure, the specific surface areas (Sa) of CuFe2O4 GMC and CuFe2O4/CuS MNC synthesized using the A. pseudalhagi extract were quantified to be 8.23 and 14.57 m2/g, respectively. Also, the average pore diameters of CuFe2O4 and CuFe2O4/CuS magnetic nanomaterials (MNMs) were 38.24 and 1.66 nm, respectively. Based on the above-mentioned findings, coating the CuFe2O4 core with CuS particles causes an increase in the specific surface area (Sa) combined with a reduction in the mean pore diameter. The possible reasons behind these phenomena are: (i) deposition of CuS particles on the CuFe2O4 surface could partially clog some of the large surface pores, breaking them up into smaller pores, and thus reducing the mean pore diameter; and (ii) the CuS particles could also attach to the interior pore walls, constricting the openings and decreasing the pore diameters and consequently enhancing the total number of pores and the surface area accessible to the molecules. The enhanced Sa of the fabricated sample could increase the active sites on the adsorbent surface, thus improving the adsorption performance of the nanocomposite in the purification system (Naghikhani et al. 2018).

Fig. 4
figure 4

N2 sorption–desorption isotherms of A CuFe2O4 and B CuFe2O4/CuS

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FESEM and TEM analysis

The FESEM and TEM micrographs of CuFe2O4 and CuFe2O4/CuS MNMs are displayed in Fig. 5A–D. The multifaceted and irregular nanostructures can be observed in the FESEM micrograph of CuFe2O4. Furthermore, it is observed that both the CuFe2O4 and CuFe2O4/CuS materials exhibit a highly agglomerated morphology, which can be attributed to their magnetic properties. According to these images, the sizes of the agglomerated particles are found to be within the nano-range. Moreover, coating the CuFe2O4 magnetic core with CuS particles reduced the magnetism level of CuFe2O4, leading to the aggregation of the CuFe2O4/CuS particles with a smaller diameter (< 100 nm) compared to CuFe2O4 particles.

Fig. 5
figure 5

FESEM images of A CuFe2O4 and B CuFe2O4/CuS, and TEM images of C CuFe2O4 and D CuFe2O4/CuS

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To more comprehensively assess the structures of the fabricated magnetic materials, the TEM technique was employed (Fig. 5A–D). Based on the TEM images of CuFe2O4 and CuFe2O4/CuS MNMs, the particles had polyhedron structures that were irregularly formed. The TEM micrographs supported the FESEM results, demonstrating the smaller sizes of CuFe2O4/CuS particles (around 40 nm) compared to CuFe2O4 particles (Maleki et al. 2019).

Elemental mapping and EDX analyses

To scrutinize the dispersion of the elemental composition in the CuFe2O4/CuS MNC, elemental mapping, and EDX analyses were conducted (see Fig. 6). As depicted in this figure, Cu had the highest elemental composition ratio in the as-synthesized composite, owing to the existence of this element in both CuFe2O4 and CuS nanoparticles. Also, all the elements used in the fabrication process existed in the structure of CuFe2O4/CuS MNC, elucidating the successful coating of CuFe2O4 with CuS nanoparticles (Younes et al. 2021). Moreover, uniform dispersion of all elements indicates the close contact of nano-constitutes of CuFe2O4/CuS MNC.

Fig. 6
figure 6

Elemental mapping and EDX results for CuFe2O4/CuS MNC

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VSM analysis

VSM analysis was used to assess the role of CuS coating on the magnetic properties of CuFe2O4 (Fig. 7). As the results of this figure, the magnetic saturation values of CuFe2O4 and CuFe2O4/CuS were 43.2 and 13.6 emu/g, respectively. The reduction of saturation magnetization of nanocomposite can be attributed to the coating of the magnetic CuFe2O4 material with non-magnetic CuS. Therefore, the CuFe2O4/CuS nanocomposite is superparamagnetic, so this new nanocomposite separation from aquatic solution easily by an external magnet.

Fig. 7
figure 7

VSM analysis of CuFe2O4 and CuFe2O4/CuS MNMs

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In several studies, the amount of magnetization of the nanoabsorbent synthesized in them has been reported in the range of the magnetization of the nanocomposite synthesized in this study. For example, in several studies by Naseh et al. in 2024, 2023, 2021 and 2019, the amount of magnetization of synthesized nanocomposites has been obtained in the range between 50 and 70, which is approximately equal to CuFe2O4/CuS nanocomposite(Azqandi et al. 2024; Golrizkhatami et al. 2023; Arghavan et al. 2021; Nasseh et al. 2019).

The impact of operational parameters

Water pH

The initial pH of the solution is an essential factor affecting the adsorption potential of the fabricated nanomaterial. Thus, to illuminate the effect of solution pH on the adsorption efficiency of the contaminant, a series of adsorption experiments were performed within the pH range of 3.0–9.0 at the adsorbent dosage of 1.5 g/L, TC concentration of 20 mg/L, and ambient temperature. As shown in Fig. 8A, by raising the pH from 3.0 to 7.0, the adsorption efficiency significantly increased from 19 to 92.5%, and then in alkaline conditions (pH 9.0), it diminished to 66%. Therefore, the highest adsorption rate of TC over CuFe2O4/CuS MNC was observed at pH 7.0, similar to other works reported so far (Chitra et al. 2021; Moslehi et al. 2024b).

Fig. 8
figure 8

A The influence of pH changes on adsorption rate (TC concentration: 20 mg/L, adsorbent dose: 0.5 g/L, temp:23 ± 2), B The influence of adsorbent dosage on TC adsorption efficiency (TC concentration: 20 mg/L, pH 7.0, temp:23 ± 2), C The influence of TC concentration on adsorption efficiency (TC concentration: 20 mg/L, pH: 7.0, adsorbent dose: 1.5 g/L, temp:23 ± 2), D The impact of temperature on TC adsorption rate (TC concentration: 20 mg/L, pH: 7.0, adsorbent dose: 1.5 g/L, temp:23 ± 2)

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According to Fig. S1, the TC molecule exhibits three pKa values at pH 3.3, 7.7, and 9.7, revealing the pH levels at which protonation/deprotonation of the TC molecule occurs. This indicates that TC possesses different ionization states (cationic, zwitterionic, anionic) depending on the water pH. At very low pH (< 3.3), TC is cationic (TCH3+) due to protonation. At pH between 3.3 and 7.7, TC is zwitterion (TCH20), indicating the existence of both positive and negative charges. Eventually, at pH > 7.7, TC exists as an anion form (TCH─ or TC2─) due to deprotonation. Hence, the adsorption rate of TC due to pH changes depends on the predominance of one or more types of this antibiotic in different pHs.

On the other hand, based on the pHzpc plot (Fig. S2), the pHzpc value of CuFe2O4/CuS MNC was determined to be 3.0. This indicates that at pH values higher than 3.0, the surface of the CuFe2O4/CuS is negatively charged, whereas at pH lower than 3.0, the surface charge of the aforementioned adsorbent becomes positive. Consequently, the decline in TC adsorption efficiency under both acidic and alkaline conditions can be attributed to the electrostatic repulsion force between TC molecules and CuFe2O4/CuS nanoparticles (Teixidó et al. 2012; Ersan et al. 2016). Notably, a previous study by Mohammed et al. (2020) (Mohammed et al. 2020) has reported an optimum pH value of 7.0 for TC molecule adsorption.

CuFe 2 O 4 /CuS dosage

To evaluate the impact of the adsorbent dosage, a set of adsorption experiments was conducted using different quantities of CuFe2O4/CuS (0.025–2 g/L), and the outcomes are presented in Fig. 8B. According to the figure, by increasing the adsorbent dosage from 0.025 to 1.5 g/L, the adsorption efficiency boosted from 35.01 to 97.56% at the contact time of 200 min. This increase in efficiency can be caused by an increase in the available surface of the adsorbent and a higher probability of the effective contact of the pollutant (TC molecules) with the nanoparticles (CuFe2O4/CuS). However, no significant change in the purification efficiency was observed when the CuFe2O4/CuS quantity rose from 1.5 to 2 g/L. Hence, enhancing the adsorbent dosage up to some extent leads to increased adsorption because it achieves an equilibrium between the number of adsorbate molecules and the available active sites on the adsorbent surface (Mohammed et al. 2020; Wang et al. 2021). Eventually, the dosage of 1.5 g/L was chosen as the optimal adsorbent dose for the subsequent adsorption experiments.

Tetracycline concentration

The influence of initial adsorbate concentration on the purification efficiency was assessed in the presence of variable amounts of TC (5–100 mg/L) during 200 min of adsorption time (). As depicted in Fig. 8C, by increasing the initial concentration of TC, the adsorption efficiency decreases. This reduction can be attributed to the limited adsorption sites and rapid saturation of these sites due to the increase in the initial concentration of the TC pollutant. Thus, the antibiotic adsorption efficiency decreases with the rise of TC concentration (Ali 2019; Aliyu et al. 2022).

Solution temperature

In an attempt to comprehend the effect of operational temperature, the adsorption efficiency of TC using CuFe2O4/CuS MNC was tested over a temperature span of 5–50 °C at pH of 7.0, with a TC content of 20 mg/L, and adsorbent dose of 1.5 g/L (Fig. 8D). By raising the temperature from 5 to 50 °C, the removal efficiency improved. This enhancement could be related to the following reasons: (i) Higher temperatures provide more kinetic energy to the TC molecules, allowing them to more easily interact with active surface adsorption sites and resulting in an increased adsorption rate; (ii) by increasing the temperature, the molecular movements become faster [3]. This can boost the effective collisions between TC molecules and CuFe2O4/CuS MNC, thus enhancing the possibility of the adsorption of antibiotic molecules onto the adsorbent; and (iii) the boosted energy at higher temperatures could feasibly cause swelling and increased porosity in CuFe2O4/CuS adsorbent. The additional internal surface area and cavity space exposed could allow enhanced interaction with contaminant molecules (Eskandarinezhad et al. 2021).

Effect of contact time

According to the results presented in Fig. 8A–D and the percentage of adsorption, it is clear that the adsorption is fast in the first minutes and can be seen with a steep slope in the graphs, which is related to physical adsorption. In the middle minutes, the slope becomes a little gentler, which is also due to the resistance to pollutant penetration on the adsorbent surface, so that after some time when the tetracycline antibiotic occupied the empty sites, the repulsive forces between the liquid bulk and the molecules Adsorbent is created. The amount of growth of adsorption decreases, and finally, the adsorption reaches the saturation state, from this time onwards, the amount of adsorption does not increase with the increase in time, which is the time of equilibrium. The amount of adsorption is fixed after that. or decreases slightly, which may be due to desorption(Aliyu et al. 2022). Based on the graphs in Fig. 8, it is clear that approximately 120 min can be considered as the equilibrium time.

Thermodynamic study

To scrutinize the adsorption characteristics of TC molecules on CuFe2O4/CuS MNC, thermodynamic parameters (Gibbs free energy, enthalpy, and entropy change) were determined at different temperatures (5–50 °C). The results are presented in Table 1. The values of ∆G° for physisorption (physical adsorption) and chemisorption (chemical adsorption) are ‘− 20–0 kJ/mol’ and ‘− 80 to − 400 kJ/mol’, respectively. As evident in Table 1, the negative ∆G° values suggest that the adsorption of TC onto the CuFe2O4/CuS MNC occurred spontaneously and physical interactions are stronger than chemical ones to adsorb pollutants (Al-Ghouti et al. 2005). Additionally, increasing the temperature from 5 to 50 °C resulted in a decline of ∆G° values, further suggesting the physisorption nature of the process. Except for 5 °C, the adsorption of TC onto the CuFe2O4/CuS adsorbent was found to be nonspontaneous.

Table 1 Thermodynamic parameters for TC adsorption onto CuFe2O4/CuS MNC
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On the other hand, the positive values of both ∆H° and ∆S° demonstrate that the adsorption process was naturally endothermic, and the degree of dispersion enhanced with an increase in temperature (Foroutan et al. 2021). The endothermic behavior observed from enthalpy values further confirmed the trend of adsorption of TC on the adsorbents, which were found to increase as the temperature was increased.

Adsorption isotherm studies

To get an overview of the mechanism, behaviors, and affinity of CuFe2O4/CuS MNC toward TC contaminant, the equilibrium data obtained were fitted to different adsorption isotherm models, including Langmuir and Freundlich isotherms. In the case of the Langmuir isotherm model, it is assumed that TC molecules are adsorbed on a uniform surface of the adsorbent based on monolayer adsorption. On the other hand, the Freundlich isotherm model assumes that the adsorption occurs on a heterogeneous surface based on multilayer adsorption. Figure 9A, B illustrates the fitted plots of Langmuir and Freundlich models, respectively, and the obtained outcomes are represented in Table 2. Based on the findings, the Langmuir model's correlation coefficient (R2) was higher than the Freundlich model, elucidating that the Langmuir model best represents the adsorption behavior of TC on CuFe2O4/CuS MNC. The excellent fit to the Langmuir equation implies TC binds in a monolayer arrangement on equivalent surface sites of the adsorbent.

Fig. 9
figure 9

Comparison of A Langmuir and B Freundlich isotherm models

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Table 2 Isotherm parameters for TC adsorption onto CuFe2O4/CuS MNC
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Adsorption kinetics

The nature of the adsorption process, whether through physical interactions (physisorption) or chemical bonding (chemisorption), is influenced by both the characteristics of the adsorbent and the conditions prevailing in the surrounding environment. In this regard, both pseudo-first-order (PFO) and pseudo-second-order (PSO) models were employed to evaluate the mechanism of TC adsorption by CuFe2O4/CuS MNC. Figure 10A, B exhibited the plot of PSO and PFO models for the TC adsorption on CuFe2O4/CuS MNC, respectively, and the kinetic parameters calculated for the two models are illustrated in Table 3. The correlation coefficient (R2) derived from the PSO model showed a significantly closer proximity to unity compared to that of the PFO model. Additionally, the calculated adsorption capacity (Qe, cal) value determined from the PSO model slightly exceeded the experimental adsorption capacity (Qe, exp) value, while the calculated adsorption capacity (Qe, cal) value obtained for PFO was significantly higher than the experimental adsorption capacity (Qe, exp). Based on the findings, the adsorption of TC onto CuFe2O4/CuS MNC can be effectively explained by the PSO kinetic model.

Fig. 10
figure 10

TC adsorption on CuFe2O4/CuS MNC based on A PFO and B PSO kinetic models

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Table 3 The calculated parameters for kinetic modeling
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Reusability study

The reusability of adsorbents is a crucial factor in determining the cost-efficiency of water purification systems. The magnetic properties of the synthesized CuFe2O4/CuS composite play a significant role in this regard, as they allow for easy separation of the adsorbent from the aqueous solution using an external magnetic field. This magnetic separation process not only helps reduce costs but also significantly minimizes the loss of the composite during washing and recovery stages.

In this study, the adsorption potential of CuFe2O4/CuS MNC was evaluated over five successive cycles under optimal operational conditions (Fig. 11). Based on the obtained results, the TC adsorption efficiency decreased by 16% after five cycles of adsorbent usage. Although a decrease in adsorption efficiency was observed, this reduction can mainly be attributed to the loss of some CuFe2O4/CuS particles during the separation and washing processes, as well as the pore clogging of the adsorbent by TC molecules (Zhang et al. 2015). However, the use of the magnetic properties of this composite significantly minimized these losses compared to non-magnetic adsorbents, which typically experience greater loss during the separation and recovery processes.

Fig. 11
figure 11

Reusability of CuFe2O4/CuS MNC (TC concentration: 20 mg/L, pH: 7.0, adsorbent dose: 1.5 g/L, contact time: 200 min)

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These findings demonstrate that CuFe2O4/CuS MNC is a sustainable adsorbent option owing to its magnetic properties, enabling reusability without significant loss of removal efficiency. This magnetic feature not only simplifies the recovery process but also reduces operational costs by minimizing the amount of composite loss during multiple adsorption cycles.

Comparative study

The adsorption potential of CuFe2O4/CuS MNC was compared to that of previously reported adsorbents toward the TC adsorption. The reaction conditions and the adsorption rate are provided in Table 4. CuFe2O4/CuS MNC exhibits a significant potential regarding the adsorption of TC molecules compared to other adsorbents. This study achieved a TC adsorption rate of 90.98% within a removal time of 200 min, demonstrating the effectiveness of the fabricated composite.

Table 4 Comparison of the adsorption ability of CuFe2O4/CuS MNC and other adsorbents that used in the adsorption of TC
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Conclusions

This investigation aimed to explore the elimination of TC antibiotic from water solutions utilizing CuFe2O4/CuS MNC. According to the morphological analysis, the size of the fabricated nanocomposite was 38 nm. It was also determined by the VSM analysis that the CuFe2O4/CuS MNC has superparamagnetic features, which allowed separation and recovery from water environments utilizing external magnetic. The complete TC uptake efficiency (100%) was obtained at pH 7.0, the nanocomposite dose of 1.5 g/L, an operating time of 200 min, and TC content of 5 mg/L. Besides, the kinetic manner of the TC adsorption obeyed the pseudo-second-order model. Moreover, the thermodynamic investigation revealed that the TC adsorption process is spontaneous and endothermic. The TC adsorption onto CuFe2O4/CuS MNC obeyed the Langmuir model (Qmax: 31 mg/g). Additionally, the TC adsorption rate diminished slightly after five reuse runs (16%), demonstrating that CuFe2O4/CuS MNC has significant stability. The outcomes indicated that CuFe2O4/CuS MNC has remarkable potential in eliminating TC antibiotics from aqueous environments.