Introduction

Water is the foundation of all life on Earth. Water pollution becomes more and more serious trouble due to the fast-growing global population (Mekonnen and Hoekstra 2016; Pedro-Monzonís et al. 2015; Yin and Deng 2015). Phenolic pollutants are classified as one of the most serious hazardous pollutants in wastewater since they are harmful to organisms even at low concentrations. Discharging of phenol into the surface water, without treatment, leads to serious health risks to animals, humans, and the aquatic life systems. Hence, there is a persistent need to develop cost-effective technologies to extend water resources and to solve the problem of water pollution with phenol.

Recently, various treatment processes were introduced for the removal of phenol from wastewater such as distillation (El-Ashtoukhy et al. 2013), adsorption (Mukherjee and De 2014; Park et al. 2015), extraction (Rahmanian et al. 2014; Shokry et al. 2020), chemical oxidation (Du et al. 2012), ion exchange (Elkady et al. 2016a), electrochemical oxidation (Tasic et al. 2014), and biological treatment processes (Jalayeri et al. 2013). Among these techniques, the adsorption technique is considered the most extensively used route for wastewater treatment (Amer et al. 2018; Ayad et al. 2018). Many adsorbent materials were utilized for the phenol removal from wastewater. For instance, Shokry Hassan et al. (2017a) monitored the phenol removal from aqueous solutions in the presence of magnetic ZnO cellulose acetate composite nanofibers. However, Salari et al. (2019) investigated the phenol adsorption onto magnetic chitosan using genetic algorithm and surface methodology. In our previous study, magnetic ZnO nanotubes were fabricated via microwave technique toward phenol adsorption from aqueous solutions by batch experiments (Elkady et al. 2017). The fabricated nanomagnetic ZnO performed a relatively high adsorption efficiency which reached 87% within 90 min in the presence of 2 g/L of adsorbent material. Moreover, Babuponnusami and Muthukumar (2012) utilized nanozero valent iron for phenol removal by heterogeneous photo-electro-Fenton-like process. Similarly, Hayat et al. (2011) investigated the efficiency of different nano-ZnO samples which fabricated via the modified sol-gel method for phenol removal from water by heterogeneous photocatalytic process. Accordingly, polymeric nanofibrous membranes are currently one of the widest membrane type used for water treatment due to its higher flexibility and relatively low costs compared with the other membrane matrices (Yin and Deng 2015). Consequently, electrospinning technology is considered one of the most promising, efficient, and useful methods for fabrication of ultrathin fibers (Frenot and Chronakis 2003; Kimmer et al. 2009; Ramakrishna 2005). Electrospinning is a simple, great, cost-efficient, and multilateral technique to produce continuous ultrafine fibers from a wide range of materials involving composites, polymers, and ceramics (Dersch et al. 2005; Jayaraman et al. 2004; Li and Xia 2004; Persano et al. 2013; Subbiah et al. 2005). This technique can be easily established in the laboratories as shown in Fig. 1 and also can be scaled upward to industrial processes (Ji et al. 2011). The nonwoven electrospun fibers have many advantages such as high surface area, small pore size, and high porosity as compared with regular fibers. These unique features make the nanofiber polymeric matrices to be good candidates for several applications including filtration, sensors, drug delivery, wound dressings, cosmetics, tissue engineering, energy conversion, and storage (Huang et al. 2003; Shokry Hassan et al. 2017b).

Fig. 1
figure 1

The electrospinning technique to produce polymeric composite nanofibers

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Many natural and synthetic polymer nanocomposites were employed for phenol decontamination from wastewater due to their suitable functional groups (Elkady and Shokry Hassan 2015; Yamasaki et al. 2006). Among the natural polymers, chitosan (CS) and alginate (Alg) are frequently used polysaccharides for bio-applications, where both these polymers can be produced from crustacean shell waste and brown seaweed (Anitha et al. 2014). CS is a linear cationic polysaccharide composed of β-(1,4)-linked-2-deoxy-2-amino-d-glucopyranose units and contains active amino and hydroxyl functional groups (Ayad et al. 2017a; b). Moreover, CS has many interesting properties such as biodegradability, biocompatibility, nontoxicity, hemostatic properties, antifungal, and antibacterial properties (Lee et al. 2009) and thus CS has been widely used in many applications such as biomedical and wastewater treatment. On the other hand, Alg is a natural hydrophilic anionic polymer, which has excellent biocompatibility, low toxicity, and capability to be used for pollutants removal from aqueous solutions (Sill and von Recum 2008). Furthermore, Alg polymer can improve the mechanical properties of the fabricated nanofibers, where it forms optically clear homogeneous blended nanofibers with PVA that enhances the tensile stress and Young modulus of the blended nanofibers compared with pure PVA (Islam and Karim 2010). Hence, there is a big challenge to form a uniform fiber without any beads in the presence of only CS and Alg. Thus, a high molecular weight polymeric material such as poly(vinyl alcohol) (PVA) is often added to increase the chain entanglement of polymers (Chang et al. 2012). The use of naturally based biopolymers such as CS as a cationic polysaccharide and Alg as an anionic polysaccharide attracted much attention due to their chemical stability, low price, high reactivity, and affinity for removal of various selected pollutants from wastewater (Barbusiński et al. 2016; Gokila et al. 2017; Nechita 2017; Ngo et al. 2015). For instance, Abdolmaleki et al. (2018) fabricated chitosan/PVA nanofibers by the electrospinning method and tested as an adsorbent for phenol decontamination. Moreover, Bahareh et al. (Alizadeh et al. 2018) used magnetic EDTA/CS/TiO2 cross-linked nanocomposite for phenol and cadmium (ӀӀ) removal from wastewater. To the best of our knowledge, the electrospinning of nanomagnetic ZnO immobilized with the three polymeric matrices (PVA, Alg, and CS) is lacking in the literature works. Moreover, the use of the formed composite nanofibers for phenol removal is missing. Therefore, the present work aims to perform a comparative study between the free PVA/Alg/CS composite nanofibers and nanomagnetic ZnO immobilized with PVA/Alg/CS composite nanofibers to get the maximum removal capacity toward phenolic compounds from aqueous solutions. Moreover, the influence of water treatment operation conditions on the adsorption process of phenol was investigated. Finally, the adsorption behavior and the reaction mechanism of phenol adsorption onto the optimum fabricated polymeric composite nanofibers were established.

Materials and methods

Materials

Poly(vinyl alcohol) 95.5–96.5%, with average molecular weight 85,000–124,000 g/mol, was purchased from Sigma-Aldrich (USA) and used without any further purification. Sodium alginate powder and aqueous glycerol were obtained from Universal Laboratories Pvt. Ltd. (Mumbai). CS of high molecular weight (∼ 310,000 to > 375,000 Da) was supplied from Sigma-Aldrich (USA). Glacial acetic acid 99% (ADWIC, Egypt), glutaraldehyde (25% aqueous solution, ACROS Organics, USA), zinc acetate (Zn(CH3COO)2·2H2O, Rankem, Gurgaon, India), iron (III) chloride (Fisher Scientific, UK), iron (II) sulfate (Fisher Scientific, UK), phenol (LOBA Chemie, India), calcium chloride (Across Organics, USA), sodium hydroxide (Sigma-Aldrich, Darmstadt, Germany), and absolute ethanol (Fisher Scientific, UK) were used as received.

Synthesis of zinc oxide nanotubes

Nanomagnetic zinc oxide was synthesized as described in our previous work (Elkady et al. 2017; Hafez et al. 2014). As a stabilizing agent, 25 mg of PVA was mixed with 14 mM of aqueous zinc acetate solution (150 mL), and sodium hydroxide was added dropwise for zinc salt reduction. The resulting solution was maintained in a microwave (THOMSON-COMBI1, Thomson Premier Lighting & Appliance, Logan, UT, USA) at 800 W for 1 h. The produced white powder was washed several times with absolute ethanol and distilled water to remove residual salts and then centrifuged at 4000 rpm. Finally, the resulting nano-ZnO was dried at 60 °C overnight.

Synthesis of nanomagnetic zinc oxide

For magnetite immobilization, 0.5 g of the synthesized zinc oxide nanotubes were suspended in 200 mL mixed solution of iron (III) chloride and iron (II) sulfate (with a molar ratio of 2:1) until a homogeneous suspension was obtained using a direct sonication probe ultrasonic homogenizer (Vibra-Cell VCX 500, SONICS, Newtown, CT, USA). Afterward, 5 M sodium hydroxide solution was added dropwise to the previous suspension at 70 °C and maintained for 30 min under constant stirring until black precipitate of magnetic zinc oxide appeared. Typically, the prepared black powder was washed several times with distilled water and absolute ethanol and then separated via centrifugation at 4000 rpm. Finally, the prepared material was dried at 70 °C overnight.

Fabrication of nanomagnetic zinc oxide/PVA/Alg/CS polymeric composites

Poly(vinyl alcohol) (10 wt%) was dissolved in distilled water at 85 °C and kept under continuous stirring overnight. 1 wt% of Alg was prepared in 50%wt aqueous glycerol under stirring at 25 °C. Afterward, 1.5 wt% aqueous CS solution was prepared by dissolving CS powder in 5 wt% of aqueous acetic acid at 25 °C under stirring. Subsequently, the sodium Alg solution was added to PVA solution under magnetic stirring, and then the CS solution was added to the mixture solution in a PVA/Alg/CS volume ratio of 8:1:1. Finally, six concentrations of nanomagnetic zinc oxide (0.0 wt%, 0.2 wt%, 0.4 wt%, 0.6 wt%, 0.8 wt%, and 1.0 wt%) were added to the previous prepared polymeric solution under continuous stirring at 25 °C.

Electrospinning processes

The electrospinning process was done at room temperature (25 °C). The polymeric solutions were fed into the electrospinning system (NanoNc, Model: ESR100D, Republic of Korea). In this study, the optimum applied voltage was 23 kV. The distance between the needle tip and the collector was 14 cm with applied flow rate was 1 ml/h. The six resulted composite nanofibers (M-1, M-2, M-3, M-4, M-5, and M-6) with the different concentrations of nanomagnetic zinc oxide (0.0 wt%, 0.2 wt%, 0.4 wt%, 0.6 wt%, 0.8 wt%, and 1.0 wt%, respectively) were collected and dried at room temperature for further use.

Cross-linking of the fabricated composite nanofibers

The cross-linking step is required for hydrophilic polymers (PVA, Alg, and CS) to be useful in wastewater treatment applications without dissolving in water. In order to cross-link the Alg polymer, the fabricated fibers were immersed in 1% CaCl2 ethanolic solution for 1 h and the resulting membranes were then treated with 10% glutaraldehyde vapor in a closed bottle for 24 h at 25 °C. After cross-linking, the membranes were washed out several times with pure methanol to eliminate any possible residual of glutaraldehyde and acids. Finally, the cross-linked composite nanofibers were soaked in distilled water for 24 h to confirm their water insolubility.

Characterization of the fabricated composite nanofibers

The fabricated M-6 composite nanofibers were characterized by Fourier transform infrared (FTIR) (Bruker, Bremen, Germany), X-ray diffractometer (XRD) (Schimadzu-7000, Shimadzu Corporation, Kyoto, Japan), vibrating sample magnetometer (VSM) (Dexing, Model: 250, Lake Zurich, IL, USA), scanning electron microscopy (SEM) (JEOL JSM 6360LA, JEOL, Japan), (JEOL, JEM-2100, Japan with an accelerating voltage of 80 kV), atomic force microscopy (AFM) (Schimadzu SPM-9700, Japan), and thermogravimetric analysis (TGA) (Shimadzu TGA-50 instrument). Moreover, the Brunauer-Emmett-Teller (BET) surface area was examined via surface area analyzer (Beckman Coulter SA3100, Brea, CA, USA) by measurement of the N2 adsorption-desorption isotherms at 77 K. All samples were degassed before measurements for 12 h under vacuum at 25 °C.

The tensile strength and elongation of the fabricated composite nanofibers were examined using a mechanical testing system (INSTRON-5500R, USA). To evaluate the surface hydrophilicity of the composite nanofibers, surface contact angles were measured using optical contact angle (OCA20, Data-physics Instruments GmbH, Germany).

Batch adsorption for phenol decontamination

The phenol adsorption efficiency of the prepared composite nanofibers (M-1, M-2, M-3, M-4, M-5, and M-6) with various immobilized magnetic ZnO ratios was investigated. Typically in the batch method, 20 mg of different composite nanofibers were mixed with 10 mL of constant phenol solution concentration of 10 mg/L of pH = 5 at room temperature using an orbital shaker (Yellow line, Germany) for 90 min. The residuals of the phenol were measured using UV-Vis spectrophotometer instrument (JASCO V-630) at wavelength λ = 510 nm. The influence of different processing parameters such as contact time (0–180 min), solution pH (1–11), composite nanofiber dose (0.5–10 g/L), initial phenol concentration (0–100 ppm), and solution temperature (25–85 °C) on the adsorption process was examined. The removal percentage of phenol via the fabricated composite nanofibers was calculated from the following equation (Elkady et al. 2016a):

$$ \mathrm{Removal}\%=\left(\left({C}_{\mathrm{o}}-{C}_{\mathrm{e}}\right)/{C}_{\mathrm{o}}\right)\times 100 $$
(1)

where Co is the initial phenol concentration (mg/L) and Ce is the phenol concentration at equilibrium in aqueous solution after phase separation (mg/L). The phenol adsorption capacity (mg/g) was determined from the following equation (Elkady et al. 2017):

$$ {q}_{\mathrm{e}}=V\ \left({C}_{\mathrm{o}}-{C}_{\mathrm{e}}\right)/m $$
(2)

where qe is the capacity of phenol adsorption (mg/g), V is the volume of phenol (L), and m is the mass of M-6 composite nanofibers (g). Moreover, the thermodynamic parameters were calculated, and the adsorption was evaluated using Langmuir, Freundlich, and Temkin equilibrium isotherm models. Finally, the kinetics of the phenol adsorption process were examined using pseudo-first-order, pseudo-second-order, Elovich, and intraparticle diffusion models.

Results and discussions

Characterization of the synthesized composite nanofibers

Fourier transform infrared

FTIR spectrum of the fabricated M-6 composite nanofibers was compared with the spectrum of the prepared M-1 composite nanofibers in Fig. 2. The FTIR spectrum of M-1 composite nanofibers shows a broad band around 3400–3500 cm−1 due to the combination of stretching vibration of hydroxyl groups of polymeric matrices, and N-H2 groups at CS (Ayad et al. 2014; Kumar et al. 2010; Siva Kumar et al. 2009). The band at 2925 cm−1 could be ascribed as the asymmetric stretching of CH2 group of CS. Moreover, CS displayed characteristic band of CH3 group and CH3-O at 1028 cm−1 (Ayad et al. 2014; Liu et al. 2003). On the other hand, the antisymmetric carbonyl stretch band of Alg appears at 1629 cm−1, while the symmetric carbonyl stretching vibration band is found at 1440 cm−1.

Fig. 2
figure 2

FTIR spectra of (a) M-1 and (b) M-6 composite nanofibers

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For the M-6 composite nanofibers, few bands were shifted to shorter wavenumbers that indicates the successful combination of the M-ZnO with polymeric matrices. For example, the bands at 3456, 2925, and 1629 cm−1 were shifted to 3447, 2923, and 1626 cm−1, respectively. The bands at 1626 and 1384 cm−1 were corresponded to carbonyl and C-N stretching vibration at the polymeric matrix (Anžlovar et al. 2012; Kumar and Rani 2013; Shokry Hassan et al. 2017a). Moreover, the absorption band at 873 cm−1 is due to C-H rocking vibration at the polymeric matrix. The vibrational bands at 538 and 397 cm−1 are characteristic to the vibration of both magnetite and ZnO (Cornell and Schwertmann 2003; Zheng et al. 2011). These results indicated the successful incorporation of nano-M-ZnO into CS/Alg/PVA composite nanofibers (Ahmed et al. 2018). The slight change in characteristic peaks of CS, Alg, and PVA spectrum could be attributed to the uniform incorporation of M-ZnO nanotubes with the polymeric matrices (Gutha et al. 2017).

X-ray diffraction

The crystallinity degree of both M-1 and M-6 composite nanofibers are investigated as shown in Fig. 3a, b. The XRD patterns of the optimized M-1 composite nanofibers show abroad peak around 2θ of 20° which is corresponding to PVA (Alakanandana et al. 2016; Jiao et al. 2017; Voronova et al. 2015), Alg (Jana et al. 2015), and CS (Ayad et al. 2014; Jiao et al. 2017). Meanwhile, these XRD patterns confirm the lower crystallinity of the prepared composite nanofibers due to the presence of PVA and Alg (Karim 2013) that possess amorphous character (Shokry Hassan et al. 2017b; Voronova et al. 2015). Moreover, the crystalline peak of CS exhibited at 20.14° is assigned to the intra-molecular and inter-molecular hydrogen bonding between amino and hydroxyl groups, which lead to the linear accumulation of the structure chains (Ayad et al. 2014). On the other hand, the XRD patterns of M-ZnO (Fig. 3b) demonstrated that the all characteristic peaks at 2θ of 31.73, 34.38, 36.21, 47.48, 56.53, 62.77, 66.30, 67.86, 69.00, 72.45, and 76.86 of the synthesized nanomagnetic ZnO can be indexed to the wurtzite hexagonal structure with a high degree of crystallinity without any impurities (Pascariu et al. 2018). These results are in good agreement compared with the obtained values of ZnO reference card (JCPDS, card number 01-089-1397) (Kumaraswamy et al. 2017; Roy et al. 2013; Shokry Hassan et al. 2017b; Yang et al. 2004).

Fig. 3
figure 3

XRD patterns of (a) M-1 and (b) M-6 composite nanofibers

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Vibrating sample magnetometer

The magnetic properties of the fabricated M-6 composite nanofibers were investigated at room temperature using a vibrating sample magnetometer. The hysteresis loop of the prepared M-6 composite nanofibers is shown in Fig. 4. It is observed that the fabricated nanofibers are a typical superparamagnetic material. The saturation moment per unit mass (Ms) of the synthesized M-6 composite was recorded 287.48 memu/g which is relatively lower than that of the magnetic nano-zinc oxide prepared at our pervious study (Elkady et al. 2017). The assigned data shows the low immobilization of magnetite nanoparticles into the polymeric membrane. Moreover, these results agree with previous studies showed that magnetite nanoparticles display superparamagnetic properties due to their particle size effect (Park et al. 2010; Shokry et al. 2020).

Fig. 4
figure 4

Magnetization curves of M-6 composite nanofibers

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Scanning electron microscope

Scanning electron microscope analysis was used to investigate the surface morphological structure of PVA/Alg/CS-based nanofibers before and after nanomagnetic ZnO immobilization. From Fig. 5, it is demonstrated that the fabricated nanofibers have a uniform, homogenous, and ordered shape with acceptable average diameters of 235 ± 52 nm and 295 ± 65 nm for M-1 and M-6 composite nanofibers, respectively. This result confirms the successful fabrication of uniform polymeric composite nanofibers (Shokry Hassan et al. 2017a).

Fig. 5
figure 5

SEM photographs of a M-1 and b M-6 composite nanofibers

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Atomic force microscope

AFM technique was used to demonstrate the three main properties of the surfaces of the fabricated fibers: surface morphology including pore size distribution, membrane adhesion (fouling), and surface electrical properties. The AFM images introduce the possibility of observing the surface topography. Figure S1 shows (a) 2D, (b) 3D, (c) phase AFM images of M-1 composite nanofibers, while Fig. S1 images (d), (e), and (f) characterize 2D, 3D, and phase of M-6 composite nanofibers. Comparing the AFM images indicated that the roughness increased from 0.24 μm for M-1 composite nanofibers to 1.35 μm for M-6 composite nanofibers. This increment of surface roughness may be returned to the increase in membrane hydrophobicity (Elkady et al. 2015).

BET surface area

The specific surface area of M-1 and M-6 composite nanofibers was 26.28 and 66.27 m2/g, respectively, as shown in Fig. S2 (a, b). These results of the surface area are relatively high for material adsorption (Shokry et al. 2020; Sun and Uyama 2013; Wang et al. 2015). In addition, the total pore volume of M-1 and M-6 composite nanofiber membranes were 0.24 and 0.25 cm3/g respectively. Furthermore, the mean pore diameters of the obtained composite nanofibers were 36.97 and 15.29 nm for the free and M-ZnO composite nanofibers, respectively. According to these results, the addition of M-ZnO has a positive impact on the surface area measurements. These data indicate that the fabricated composite nanofibers achieve significantly high surface area and porosity, which are very important for various potential applications (Elkady and Shokry Hassan 2015).

Thermogravimetric analysis

The thermograms of the M-1 and M-6 composite nanofibers are shown in Fig. 6. The TGA curves of the fabricated composite nanofibers indicate the presence of a weight loss stage of about 8% at 160 °C that can arises from the loss of water molecules (Shokry et al. 2020). The second weight loss stage positioned between 240 and 350 °C was about 40 wt% and this step can be assigned to the cleavage of side chain of the polymeric fiber (Liu et al. 2003). The third stage of weight loss ranged from 350 to 450 °C that attributed to the cleavage of nanofibers backbone (Elkady and Shokry Hassan 2015). Finally, the total weight loss step recorded about 99 wt% for M-1 membrane and about 89 wt% at 800 °C for M-6 composite nanofibers (Sargazi et al. 2019). From these results, both of the fabricated composite nanofibers were found to possess good thermal stability (Khan et al. 2018).

Fig. 6
figure 6

TGA curves of (a) M-1 and (b) M-6 composite nanofibers

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Tensile strength measurement

The mechanical tensile strength and elongation at break of the M-1 and M-6 composite nanofibers were carried out using the universal material testing machine. The extension rate and the clip distance were found to be 5 mm/min and 50 mm, respectively. As shown in Figs. S3 and S4, the tensile strength and elongation at break of the M-6 composite nanofibers were improved as compared with M-1 composite nanofibers as a result of nanomagnetic ZnO addition (Peña-Reyes et al. 2017). The highest tensile strength values were recorded at 11.08 and 9.43 N/mm2 for M-6 and M-1 composite nanofibers, respectively.

Contact angle measurements

The contact angle was employed to characterize the relative hydrophilicity or hydrophobicity of the fabricated membrane surface as given in Tables S1 and S2. From Figs. S5 and S6, it is obvious that the immobilization of nanomagnetic ZnO into PVA/Alg/CS nanofibers positively influenced the contact angle values. Accordingly, the mean value of theta for M-1 composite nanofibers was recorded as 67.02 compared with 85.48 recorded for M-6 composite nanofibers. These results confirm the hydrophobicity character of the fabricated composite nanofibers particularly M-6 composite nanofibers (Ma et al. 2010; Zhao et al. 2011).

Application of the fabricated composite nanofibers for phenol removal

The feasibility of the prepared composite nanofibers with different immobilization M-ZnO ratios was examined for the adsorption of phenol from aqueous solutions at room temperature using the batch technique (Elkady et al. 2017). Figure 7 shows that the M-1 nanocomposite sample possessed the lowest phenol removal percentage of 37.65%. However, the M-6 nanocomposite sample was the best phenol adsorbent with a removal percentage of 84.22%; so, this sample was selected for the rest of study.

Fig. 7
figure 7

Phenol removal by the fabricated composite nanofibers (contact time = 90 min, pH = 5, initial phenol concentration = 10 ppm, agitation speed = 440 rpm, composite nanofiber dosage = 2 g/L, and temperature = 25 °C)

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Influence of contact time on the phenol adsorption

The influence of contact time on the percentage of phenol adsorption onto sample M-6 nanocomposite was traced at different time intervals. As shown in Fig. 8, the phenol adsorption onto the fabricated composite nanofibers increased with increasing the contact time initially till reaching the equilibrium after 120 min. The high adsorption rate of phenol may be referred to the highly available surface area of the prepared composite nanofibers (El-Aassar et al. 2016; Shokry et al. 2020). The quantity of phenol adsorbed onto the composite nanofibers is at dynamic equilibrium with phenol ions desorbed from the polymeric composite nanofibers (Malayeri et al. 2014). The optimum contact time of the phenol adsorption process was recorded after 90 min, with a maximum phenol percentage removal of 84.22%. At the second stage of the adsorption process, the active sites of the M-6 composite nanofibers were saturated with the phenolic pollutants after the equilibrium (Mohy-Eldin et al. 2010) time.

Fig. 8
figure 8

Influence of contact time on phenol adsorption process onto M-6 composite nanofibers (pH = 5, initial phenol concentration = 10 ppm, agitation speed = 440 rpm, composite nanofiber dosage = 2 g/L, and temperature = 25 °C)

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Influence of M-6 composite nanofiber dosage on the phenol adsorption

The adsorbent dosage has a significant impact on the adsorption process because of its strong influence on the capacity of the adsorbent material at a given initial concentration of the pollutant (Elkady et al. 2018; Shokry et al. 2020). The adsorbent dosage effect of M-6 composite nanofibers was investigated after 90 min. Figure 9 shows that the percentage of phenol removal was enhanced from 53.52 to 95.44% as M-6 composite nanofiber dosage increased from 0.5 to 20 g/L. However, a decrease in the adsorption capacity appeared with increasing of the composite nanofiber dosage. The decline of adsorption capacity at high material concentration may be attributed to the residual of the unsaturated adsorption sites onto the M-6 composite nanofibers (Vijayalakshmi et al. 2010). Moreover, the increase in the M-6 composite nanofiber dosage leads to the availability of more active sites for the phenol adsorption. These results may be due to the large determined active surface area of the fabricated M-6 composite nanofibers (Elkady et al. 2016a). Accordingly, the optimum dosage of M-6 composite nanofibers was achieved at 2 g/L, which represented to the economical dosage of the adsorption process.

Fig. 9
figure 9

Influence of M-6 composite nanofiber dosage on both phenol percentage removal and phenol uptake capacity onto (contact time = 90 min, pH = 5, initial phenol concentration = 10 ppm, agitation speed = 440 rpm, and temperature = 25 °C)

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Influence of initial pH on the phenol adsorption

The initial solution’s pH represents an essential factor in the control of the phenol adsorption process onto the fabricated composite nanofibers. This parameter affects both the degree of ionization of phenol and the charge of the adsorbent material surface (Elkady et al. 2018). The effect was studied from pH 1 up to 11. It can be noticed from Fig. 10 that high adsorption capacity and removal percentage was achieved at low pH values, which demonstrates that M-6 composite nanofibers have high adsorption efficiency in the acidic medium. The adsorption percentage of phenol was found to be high till pH = 5 with 84.22% phenol removal and then the adsorption gradually declines with increasing of solution’s pH with the lowest adsorption of 56% phenol removal at pH 11. Therefore, pH = 5 was selected as optimum pH for the rest of the adsorption study onto the prepared M-6 composite nanofibers. The zero charge point (pzc) of the synthesized nanomagnetic ZnO was recorded at around 6 (Farrokhi et al. 2014). Accordingly, at pH ˂ 6, the surface of the nanomagnetic ZnO is positively charged, so there is an electrostatic attraction between the positively charged surface of nanomagnetic ZnO and the unionized species of phenol (Parida and Pradhan 2010). Furthermore, as the solution’s pH increases, the adsorbent surface became negatively charged, which results in reducing phenol adsorption due to the repulsive forces between the negatively charged magnetic ZnO surface and the phenolate ions. These results attributed to the pKa value of phenol around 9.8 (Han and Tao 2006).

Fig. 10
figure 10

Influence of solution pH on the percentage of phenol removal onto M-6 composite nanofibers (initial phenol concentration = 10 ppm, composite nanofiber dosage = 2 g/L, agitation speed = 440 rpm, contact time = 90 min, and temperature = 25 °C)

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Influence of initial phenol concentration on the adsorption process

The impact of phenol’s initial concentration on the phenol decontamination process at equilibrium was achieved in the concentration range of 5 to 100 mg/L using 2 g/L of M-6 composite nanofibers at pH 5 for 90 min. Figure 11 demonstrates the enhancement in the adsorption capacity as the phenol initial concentration increased from 5 to 100 mg/L, which agrees with other reported studies (Elkady et al. 2017; Elkady et al. 2016a; Shokry Hassan et al. 2017a). This behavior may be due to the saturation of the active sites of the prepared M-6 composite nanofibers as the phenol concentration increased. From the previous results, it is clear that the fabricated M-6 composite nanofibers are effectively capable of removing phenol from aqueous solutions at various phenol concentrations; this behavior is due to the characteristic surface area of composite nanofiber that was estimated previously as 66.27 m2/g.

Fig. 11
figure 11

Influence of initial phenol concentration on both the percentage of phenol removal and phenol uptake capacity onto M-6 composite nanofibers (pH = 5, composite nanofiber dosage = 2 g/L, agitation speed = 440 rpm, contact time = 90 min, and temperature = 25 °C)

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Influence of solution temperature on the phenol adsorption

Figure 12 investigated the influence of the temperature on the phenol removal onto the fabricated M-6 composite nanofibers. The rising of solution temperature was found to increase the phenol adsorption process. This behavior arises from enhancing the rate of phenol diffusion into the pores of the fabricated M-6 composite nanofibers at higher temperatures due to the creation of new adsorption sites onto the M-6 composite nanofibers at high temperatures (Abd El-Latif and Elkady 2011; Shokry Hassan et al. 2015). According to these results, the phenol decontamination onto the M-6 composite nanofibers is an endothermic process (Elkady et al. 2017).

Fig. 12
figure 12

Influence of phenol solution temperature on the percentage of phenol removal onto M-6 composite nanofibers (initial phenol concentration = 10 ppm, composite nanofiber dosage = 2 g/L, agitation speed = 440 rpm, contact time = 90 min, and pH = 5)

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Thermodynamics and equilibrium modeling

The various thermodynamic parameters that should be considered to investigate any adsorption process are the changes in enthalpy (ΔH°), standard free energy (ΔG°), and entropy (ΔS°). The value of the standard enthalpy and entropy can be calculated from the Van’t Hoff equation as follows:

$$ \ln\ {K}_c=\frac{\Delta S{}^{\circ}}{R}-\frac{\Delta H{}^{\circ}}{RT} $$
(3)

where the universal gas constant (R) = 8.314 J/mol K, T is the solution temperature in Kelvin while Kc = Fe/(1 − Fe), and Fe = (Co − Ce)/Co is the fraction adsorbed at equilibrium. Figure 13 indicated that the Van’t Hoff plot of lnKc against 1000/T gives a straight line with an acceptable value of correlation coefficient (R2) at different investigated phenol concentrations. The values of ΔS° and ΔH° can be respectively calculated from the intercept and the slope of the Van’t Hoff plot. Furthermore, the activation energy (Ea) can be determined from the following equation (Lu et al. 2016):

$$ {E}_{\mathrm{a}}=\Delta H{}^{\circ}+\mathrm{RT} $$
(4)
Fig. 13
figure 13

The Van’t Hoff plot of ln Kc against 1000/T for phenol adsorption onto M-6 composite nanofibers

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The values of ΔH°, ΔG°, ΔS°, and Ea at different temperatures are listed in Table 1. The negative values of free energy changes ΔG° indicate the spontaneous and thermodynamically nature of the phenol adsorption process onto the fabricated M-6 composite nanofibers. Moreover, the calculated positive value of enthalpy indicates the endothermic in nature of the adsorption process. On the other hand, the positive value of entropy illustrates an increase in disorder at the solid/liquid interface during the phenol decontamination process (Thinakaran et al. 2008).

Table 1 Thermodynamic parameters for phenol adsorption onto M-6 composite nanofibers
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Equilibrium isotherm analysis for the phenol adsorption

The Langmuir, Freundlich, and Temkin isotherm models were applied to investigate the adsorption behavior and the reaction mechanism of phenol adsorption onto the fabricated M-6 polymeric composite nanofibers. The linearized plot of Ce/qe versus Ce represented a straight line with a relative high correlation coefficient (R2 = 0.993) (Elkady et al. 2016b):

$$ \frac{C_{\mathrm{e}}}{q_{\mathrm{e}}}=\frac{1\ }{q_{\mathrm{m}}K}+\frac{C_{\mathrm{e}}\ }{q_{\mathrm{m}}} $$
(5)

where qe is the amount phenol adsorbed at equilibrium (mg/g), Ce is the equilibrium concentration of the adsorbate ions (mg/L), and qm and KL are Langmuir constants referred to the maximum monolayer adsorption capacity (mg/g) and adsorption energy (L/mg), respectively. According to the calculated parameters in Table 2 and Fig. S7, the Freundlich model is the best suitable model to describe the adsorption process of phenol onto the fabricated composite nanofibers. Moreover, the value of the separation factor (RL) is 0.049, which falls in the range 0 to 1 that proves the favorable adsorption process description via the Langmuir model (Elkady et al. 2017; Üner et al. 2016). On the contrary, the best fit of the equilibrium data with the Freundlich isotherm assumes multilayer adsorption on heterogeneous surfaces (Üner et al. 2016). However, the adsorption intensity (nF) calculated from the Freundlich model recorded 3.639 which is greater than unity, illustrating that the phenol adsorption process onto M-6 composite nanofibers is favorable (Üner et al. 2016). Therefore, the Freundlich isotherm is the best favorable model for representing the phenol adsorption process onto the prepared M-6 composite nanofibers, which supposes that the multilayer adsorption demonstrates the predictions about the physical adsorption of phenolic pollutants onto the fabricated composite nanofibers (Elkady et al. 2018; Shokry Hassan 2019).

Table 2 Isotherm parameters of Langmuir, Freundlich, and Temkin models for phenol removal onto M-6 composite nanofibers
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Comparison of adsorption capacity for prepared composite nanofibers with other adsorbent nanomaterials

To compare the adsorption performance of the prepared M-6 composite nanofibers with the other similar based nanomaterials, the monolayer adsorption capacities (qm) were compared. Table 3 presents the comparable adsorption capacity of fabricated 1.0 wt% M-ZnO/PVA/Alg/CS composite nanofibers with that of the other adsorbent nanomaterials. Consequently, the fabricated M-6 composite nanofibers showed promising and appropriate results for phenol adsorption from aqueous solutions.

Table 3 Comparison of calculated monolayer adsorption capacity for different adsorbent nanomaterials
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Kinetic models of phenol adsorption

To verify the adsorption mechanism of phenolic compounds from aqueous solution onto the fabricated M-6 composite nanofibers, the kinetics of the adsorption process were studied by applying the theoretical modeling of pseudo-first-order, pseudo-second-order, Elovich, and intraparticle diffusion models. The Lagergren first order is represented by the following equation (Alkaram et al. 2009):

$$ \ln\ \left({q}_{\mathrm{e}}-{q}_{\mathrm{t}}\right)=\ln\ {q}_{\mathrm{e}}-{k}_1t $$
(6)

where qe and qt are the amounts of phenol adsorbed ions (mg/g) at equilibrium and at time t (min), respectively. k1 (min) is the rate constant of the first-order reaction. Furthermore, the phenol adsorption kinetic data were investigated using the pseudo-second-order kinetic model that can be expressed by:

$$ t/{q}_{\mathrm{t}}=\left(1/{k}_2{q}^2\right)+t/q $$
(7)

where k2 is the equilibrium constant of the second-order reaction rate (g/mg min). The Elovich equation was demonstrated for the multilayer adsorption as follows:

$$ {q}_{\mathrm{t}}=\alpha +\beta\ \ln\ t $$
(8)

where α signifies the initial adsorption rate (mg/g min) and β is correlated to the surface coverage degree and physical adsorption activation energy (g/mg). The constants α and β can be calculated from the slope and intercept of the linear plot of qt versus ln t. Finally, the possibility of the intraparticle diffusion affecting the adsorption process was applied by Weber and Morris (Elkady et al. 2017; Shokry et al. 2020) using the intraparticle diffusion model from the following equation:

$$ {q}_{\mathrm{t}}={k}_{\mathrm{i}}{t}^{1/2}+C $$
(9)

where ki is the constant of the intraparticle diffusion rate.

The linearity of t/qt versus time with high correlation coefficient value (R2 = 0.999) illustrates that the adsorption process of phenol follows the second-order rate kinetic model as shown in Table 4 and Fig. S8 (Elkady and Shokry Hassan 2015). Moreover, the experimentally measured value (4.211) of qe of pseudo-second order is very close to the estimated value (4.593). However, the intraparticle diffusion model was given two separate regions. These two regions suggest that the adsorption process contains both surface adsorption and intraparticle diffusion. The intraparticle diffusion constant (ki) and the value of C were separately calculated from the slope and intercept of the two linear regions as listed in Table 4. Additionally, the values of C give prediction about the thickness of the boundary layer. It is clear from Table 4 that the thickness of the boundary layer of the second region, which referred to the intraparticle diffusion (C2), is larger than the thickness of the boundary layer of the first region, which corresponds to the film diffusion (C1). Consequently, the value of the intraparticle diffusion constant (ki2) is lower than the film diffusion constant (ki1). These results give prediction that the sorption process was controlled mainly by intraparticle diffusion (Namasivayam and Sureshkumar 2008). According to these results, the pseudo-second-order and intraparticle diffusion kinetic models are considered the most fitted for describing the phenol adsorption process onto the fabricated M-6 composite nanofibers that confirms the physical adsorption process (Abd El-Latif and Elkady 2011).

Table 4 Pseudo-first-order, pseudo-second-order, Elovich, and intraparticle diffusion kinetic parameters for phenol removal onto M-6 composite nanofibers
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Regeneration of M-ZnO/PVA/Alg/CS composite nanofibers

The regeneration test of the adsorbent material is considered the most significant factor because it touches on the overall cost in the real applications (Elzain et al. 2019). From all of the elution reagents, sodium hydroxide is considered the most favorable for phenol desorption (Ozkaya 2007). Consequently, to investigate the regeneration process, the final centrifuged form of M-6 composite nanofibers was washed several times with distilled water and then soaked into 0.2 N sodium hydroxide for 4 h under continuous stirring at room temperature. The polymeric composite nanofibers were washed again with distilled water and then dried at 50 °C overnight. The regenerated composite nanofibers were reused in the batch experiment to evaluate its efficiency after the regeneration process. The adsorption-desorption cycle was repeated five times. The data demonstrate the ability of M-6 composite nanofibers to reuse after three times where the removal efficiency decreased to 54.12% as shown in Fig. 14. This regeneration ability of M-6 composite nanofibers was due to its durability as a result of its high tensile strength.

Fig. 14
figure 14

Recyclability test of M-6 composite nanofibers toward phenol removal

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Conclusions

Nanomagnetic ZnO was successfully immobilized into PVA/Alg/CS polymeric composite nanofibers using the electrospinning technique. The fabricated composite nanofibers were characterized using FTIR, XRD, VSM, SEM, AFM, TGA, tensile, and contact angle. Both the composite nanofiber dosage and phenol solution temperature have positive impacts on the phenol adsorption process. The thermodynamic parameters confirmed the endothermic and spontaneous nature of the phenol decontamination process onto the polymeric composite nanofibers. Moreover, the positive value of ΔS° indicates the increment in the randomness of the solid/solution interface in the sorption process. Furthermore, the mathematical equilibrium description of the phenol adsorption onto M-6 composite nanofibers was fitted using the Freundlich model. These mathematical equilibrium models supposed that the multilayer adsorption of phenol onto the M-6 composite nanofibers contains a physical adsorption process. Finally, the adsorption of phenol onto the prepared M-6 composite nanofibers was well described by the pseudo-second-order and intraparticle diffusion kinetic models, which demonstrate the physical adsorption process. The fabricated M-6 composite nanofiber illustrates its validity for reuse after three cycles of regenerations.