where kint is the intraparticle diffusion constant (mg g−1 min−1/2), C is a constant (mg g−1) that is directly proportional to the boundary layer thickness. Both values of kint and C can be obtained by calculations from slope and intercept resulted from the linear curve of qt versus t1/2, respectively [24].

#### Kinetic Validation

To determine the most applicable model for describing the adsorption process, the normalized standard deviation Δqe (%) is used and calculated using Equation (11) [25].

$$
\Delta \mathbf{q}\_{\mathbf{e}} \left( \% \right) = 100 \times \sqrt{\frac{\left[ \left( \mathbf{q}\_{\text{t,exp}} - \mathbf{q}\_{\text{t,cal}} \right) / \mathbf{q}\_{\text{t,exp}} \right]^2}{N - 1}} \tag{11}
$$

where qt,exp is the experimental adsorption capacity (mg g−1), qt,cal is the calculated adsorption capacity for pseudo-first and pseudo-second models (mg g−1) and N is the number of data points.

#### 2.4.4. Adsorption Isotherm for CR Dye

Studying the adsorption isotherm is important for describing the ability of adsorbate molecules to distribute between the liquid and the solid phases at equilibrium state

of the adsorption process. For studying adsorption isotherm, 10 mL of dye solution (400–1000 mg L−1) was shaken with 10 mg of adsorbent in a shaking water bath (70 rpm) at 55 ◦C and pH 9. After reaching the equilibrium, the concentrations of un-adsorbed dye were determined by measuring absorbance of dye solution using UV-vis spectrophotometer at λmax = 497 nm. The adsorption isotherm was studied using four models; Langmuir, Freundlich, Temkin and Dubinin–Radushkevich models.

#### Langmuir Isotherm Model

Langmuir model supposes the formation of a monolayer coverage of adsorbate molecules onto a homogenous surface of adsorbent without interaction between the adsorbate molecules. This isotherm can be expressed by Equation (12).

$$\frac{\mathbf{C\_e}}{\mathbf{q\_e}} = \frac{1}{(\mathbf{q\_{max}} \cdot \mathbf{K\_L})} + \frac{\mathbf{C\_e}}{(\mathbf{q\_{max}})} \tag{12}$$

where qe is the amount of adsorbed dye at equilibrium (mg g−1), Ce is the equilibrium concentration of dye in solution (mg L−1), qmax is the maximum monolayer coverage adsorption capacity (mg g−1), and KL is a Langmuir coefficient that is concerned with adsorption energy (L mg−1). By plotting Ce/qe versus ce, results in a linear relationship, the values of both qmax and KL can be obtained from the slope (1/qmax) and the intercept (1/(qmax KL)), respectively. There is an important parameter for Langmuir model which is the separation factor (RL), Equation (13).

$$\mathcal{R}\_{\rm L} = \frac{1}{(1 + \mathcal{K}\_{\rm L}\mathcal{C}\_{\rm o})} \tag{13}$$

where KL is the Langmuir constant and *Co* is the initial concentration of the adsorbate in solution. This parameter indicates whether the process is irreversible, favorable, unfavorable and linear if RL = 0, 0 < RL < 1, RL > 1 and RL = 1, respectively.

#### Freundlich Isotherm Model

A Freundlich isotherm model is used for describing the adsorption onto heterogeneous surfaces. It possesses the formation of multilayer adsorption with interaction between the adsorbate molecules. Moreover, the adsorption active sites are non-identical for the occupation of adsorbate due to differences in adsorption energy. Its linear form is given by Equation (14).

$$
\ln \mathbf{q}\_{\mathbf{e}} = \ln \mathbf{K} \mathbf{p} + \frac{1}{\mathbf{n}} \ln \mathbf{C}\_{\mathbf{e}} \tag{14}
$$

where qe is the adsorbed dye amount at equilibrium (mg g−1) and Ce is the equilibrium concentration of dye in the solution at a constant temperature (mg L−1). KF is the adsorption capacity ((mg g−1) (L mg−1) 1/n), while 1/n shows an indication of the favorability and feasibility of the adsorption process and the surface heterogeneity, which is related to adsorption intensity. The value of 1/n indicates the type of isotherm; irreversible (1/n = 0); favorable (0 < 1/n < 1), and unfavorable (1/n > 1). By plotting ln qe against ln Ce, the constant *KF* and the indictor 1/n can be obtained.

#### Temkin Isotherm Model

The Temkin model assumes the decrease in heat uptake for dye molecules with surface saturation under the effect of continuous interaction between the studied adsorbent and dye molecules. The Temkin model can be expressed by Equation (15).

$$\mathbf{q\_{\{e\}}} = \mathbf{B} \text{ In } \mathbf{k\_{\Gamma}} + \mathbf{B} \text{ In } \mathbf{C\_{\S}} \tag{15}$$

where B is the Temkin constant that is controlled by the adjusted uptake temperature (J mol−1), KT is the binding constant of Temkin isotherm (L g−1), qe is the amount of dye adsorbed at equilibrium (mg g−1), and Ce is the dye equilibrium concentration at a constant temperature (mg L−1). By plotting qe against ln Ce, the constants B and KT can be obtained from the slope and the intercept, respectively.

#### Dubinin–Radushkevich (D–R) Isotherm

The Dubinin–Radushkevich model is used for estimating the free energy of adsorption and the characteristic porosity. This model can successfully help in determining the type of the adsorption process whether it is chemisorption or physisorption. The (D–R) isotherm model does not assume that the surface is homogenous; so, it is more general than the Langmuir isotherm. The D–R isotherm is given by Equation (16).

$$\ln \mathbf{q}\_{\mathbf{e}} = \ln \left( \mathbb{X}\_{\mathrm{m}} \right) - \left\| \boldsymbol{\varepsilon}^{2} \right\| \tag{16}$$

where qe is the amount of dye adsorbed at equilibrium (mg g−1), Xm is monolayer saturation capacity (mg g−1), β is the activity coefficient that is related to adsorption mean free energy (mol2 j −2) and ε is the Polanyi potential which is determined by Equation (17).

$$
\varepsilon = \text{RT} \, \ln \left( 1 + \frac{1}{\mathbb{C}\_{\text{e}}} \right) \tag{17}
$$

where R is universal gas constant (8.314 J mol−1K−1) and T is the absolute temperature (K).

By plotting ln qe versus ε2 using Equation (16), a straight line is obtained with a slope and intercept β and ln Xm, respectively. The mean free energy, E, of adsorption (kJ mol−1) can be determined by Equation (18):

$$\mathcal{E} = \frac{1}{(2\beta)^{0.5}}\tag{18}$$

This means that the free energy can determine the type of adsorption; if 8 < E < 16 kJ mol−1 the process is chemisorption, while if it is less than 8 kJ mol−1 the process is physisorption.

#### 2.4.5. Thermodynamic Studies

Thermodynamic parameters provide in-depth information on inherent energetic changes associated with adsorption; therefore, these parameters should be accurately evaluated. Studying thermodynamic parameters give a better understanding of the adsorptive behavior for the dye towards adsorbents. Gibbs free energy change, ΔG◦ (kJ mol−1), enthalpy change, ΔH◦ (kJ mol−1) and entropy change, ΔS◦ (J mol−1 K−1) were calculated for the adsorption of CR dye using four different temperatures (298, 308, 318 and 328 K) according to Equations (19) and (20).

$$
\Delta \mathbf{G}^{\odot} = -\mathbf{R} \mathbf{T} \ln \mathbf{K}\_{\odot} \tag{19}
$$

$$
\ln \mathbf{k}\_{\mathbf{c}} = \frac{\Delta \mathbf{s}^{\odot}}{\mathbf{R}} - \frac{\Delta \mathbf{H}^{\odot}}{\mathbf{R} \mathbf{T}} \tag{20}
$$

where R is the universal gas constant (8.314 J mol−1 K−1), T is the absolute temperature (K) and Kc is the distribution coefficient (qe/Ce). By plotting ln Kc versus 1/T, the thermodynamic parameters can be obtained.

#### 2.4.6. Activation Energy

The activation energy (Ea) is determined by applying Arrhenius equation, which refers to the minimum energy for proceeding the reaction as shown in Equation (21).

$$\ln\text{ln} = \ln\text{A} - \frac{\text{E}\_{\text{a}}}{\text{RT}}\tag{21}$$

where A is the Arrhenius factor, R is the universal gas constant (8.314 J mol−1 K−1), Ea is the activation energy (kJ mol−1) and T is the absolute temperature (K). By plotting ln k of the pseudo-second order constant versus inverse temperature (1/T), a straight line is obtained with slope of –Ea/R, where Ea can be determined.

#### 2.4.7. Desorption Study

The dye desorption from adsorbent was performed by its washing with distilled water to remove any un-adsorbed dye molecules. Afterwards, the adsorbent (0.01 g) was immersed in the desorption medium (10 mL of ethanol, methanol, acetone or aqueous NaOH solution (0.1 N)) at 25 ◦C for 24 h. The amount of the desorbed dye can be calculated using Equation (22).

$$\text{"\% Dye desorption = \text{\(q\)}\text{\(q\)}}\text{\(\text{q}\)}\text{\(10\)}\tag{22}$$

where qd is the amount of desorbed dye from the adsorbent surface (mg g−1) and qa is the amount of the dye adsorbed onto the adsorbent (mg g−1) [24].

#### **3. Results and Discussion**

#### *3.1. Synthesis of Novel CCs Adsorbent*

The CCs adsorbent (Scheme 1) was prepared via a four-step procedure since the primary amine groups in chitosan were firstly protected by reaction with benzaldehyde for achieving chitosan Schiff's base, in which the primary hydroxyl groups on C6 reacted with epichlorohydrin for generating epoxy chitosan Schiff's base, followed by reaction of the epoxy rings with cyanoguanidine for attaining cyanoguanidine chitosan Schiff's base, which was finally hydrolyzed in acidic medium to eliminate the benzaldehyde moieties and retrieve the amino groups to get CCs adsorbent. The amino and hydroxyl groups on chitosan in addition to the basic functional groups of cyanoguanidine incorporated into chitosan can potentially remove the acidic pollutants such as acidic dyes.

#### *3.2. Characterization of Novel CCs Adsorbent*

#### 3.2.1. FTIR Spectra of CCs Adsorbent

FTIR spectra of chitosan and its modified derivatives were demonstrated in Figure 2. In the spectrum of chitosan, the existence of the saccharide moieties was confirmed by the appearance of four absorption peaks at 1158, 1074, 1029, and 894 cm−1. A dense broad absorption peak at around 3700 to 3000 cm−1 appeared, relating to the stretching vibration of -OH groups overlapped with that for -NH2 and their hydrogen bonds. The symmetric absorption peak corresponded to -CH and -CH2 groups in the pyranose rings appeared at 2924 and 2864 cm−1, respectively. The high extent of deacetylation of chitosan was confirmed by the appearance of two weak absorption peaks at 1658 and 1593 cm−1 assigning to amide I and amide II, respectively. The overlapping between the amino groups deforming vibration at 1600 cm−1 and the stretching vibration peak of amide I at 1658 cm−1 resulted in an intensive peak [22,26].

The spectrum of chitosan Schiff's base displayed similar absorption peaks of chitosan in addition to some new peaks as follows: (1) at 3052 and 3027 cm−1 indicated to C-H groups in aromatic ring, (2) at 1691 cm−1 corresponded to C=N groups, (3) at 1600, 1579, 1493 and 1454 cm−1 related to C=C bond in aromatic rings, and (4) at 757 and 692 cm−1 (strong) due to mono-substituted benzene rings [16,21].

Epoxy chitosan Schiff's base spectrum, in addition to the afore-mentioned peaks, displayed a new peak at 1250 cm−1 due to the epoxide moieties [27]. The spectrum of the cyanoguanidine chitosan Schiff's base showed that the disappearance of the peak corresponded to the epoxide linkages at 1250 cm−1 and the appearance of two new peaks at 2207 and 2162 cm−1 related to C≡N group of the cyanoguanidine moiety [28], indicating occurrence of the interaction between the epoxide rings and the NH2 groups of cyanoguanidine. Moreover, the stretching vibration peak corresponded to C=N group of the cyanoguanidine moiety appeared at 1641 cm−1.

Removal of benzaldehyde moieties to obtain CCs adsorbent was confirmed by the disappearance of the absorption bands of mono-substituted benzene ring at 757 and 692 cm−1.

**Figure 2.** FTIR spectra of the modified chitosan derivatives.

#### 3.2.2. Powder X-ray Diffraction of CCs Adsorbent

X-ray diffractometry was used to explore the inner structures of the modified chitosan derivatives and their X-ray diffraction patterns were shown in Figure 3. There were two broad peaks in chitosan appeared near to 2θ = 10◦ and 20◦ which were attributable to its amorphous and crystalline regions, respectively [29]. This can be ascribed to the formation of the hydrogen bonds along its chains due to its possession of a great number of hydroxyl and amino groups. On the other hand, the modified chitosan derivatives were less crystalline than chitosan. This was illustrated by a disappearance of the peak at 2θ = 10◦ and a decrease of the intensity of the peak at 2θ = 20◦. After modification of chitosan, its functionality greatly changed with a reduction of the hydrogen bonds due to consumption of its polar –NH2 and/or –OH groups and incorporation of the modifiers' moieties that separated the chitosan chains away from each other. This led to an increase of the amorphous region and a decrease of the crystalline region.

#### 3.2.3. SEM Analysis of CCs Adsorbent

SEM micrographs of surface topography of the modified chitosan derivatives were shown in Figure 4. The original chitosan showed a smooth surface; however, its derivatives showed a raucous surface, containing lumps of diverse sizes because the incorporated substituent groups have assorted sizes. It can be noted that the lumps were homogeneously distributed over each derivative, suggesting that every stage for the chitosan modification process was successfully completed. The inserted substituent groups separated chitosan chains away from each other, reduced the formation of their hydrogen bonds and created a porous matrix with a large surface area.

**Figure 3.** X-ray diffraction patterns of the modified chitosan derivatives.

#### *3.3. Adsorption of CR Dye Using CCs Adsorbent*

CCs was used in the present work for the removal of CR dye from water. Its adsorption capacity came from the incorporation of functional chelating cyanoguanidine moieties, which possessed more binding centers for CR dye. Various adsorption factors were studied, in addition to studying kinetics, isotherm and thermodynamics for understanding the adsorption mechanism.

#### 3.3.1. Optimization of the Adsorption Conditions Effect of pH

One of the main parameters that has a major role on adsorption of dyes is the pH of the solution due to some mechanisms such as protonation processes between dyes and adsorbent surface.

The effect of pH on the adsorption of CR dye onto CCs adsorbent was studied at different pH values of 4, 7 and 9, and the results were presented in Figure 5. It can be noted that the adsorption of CR dye onto CCs decreased with increasing pH values. From the results, the percentages of CR dye removal were 92.55, 85.59 and 82.40% at pH 4, 7 and 9, respectively. This is in agreement with the study for the removal of CR dye using chitosan-coated quartz sand [30].

At low pH, an electrostatic attraction between the positively charged active sites on CCs adsorbent and dye anions takes place. The incorporation of cyanoguanidine moieties improved the adsorption performance of CCs, because they introduced more positive charges onto the CCs surface [31].

At a high pH, a strong competition between OH− groups and dye anions for the positively charged sites of CCs occurred. Since the OH− groups have a smaller size than the dye anions, binding of OH− groups with CCs predominated, decreasing the adsorption performance of CCs.

Interestingly, the cyanoguanidine moieties in CCs have an important role in enhancing and improving its adsorptive capacity. The % R.E. values of CR dye (adsorbent dose = 50 mg, dye solution (50 mL, 600 mg L−1) pH = 9 and temperature = 55 ◦C) by chitosan Schiff's base, epoxy chitosan Schiff's base and cyanoguanidine chitosan Schiff's base were 20.66, 23.70 and 41.48%, respectively, which were lower than that of CCs (82.40%).

**Figure 5.** Effect of pH on CR dye adsorption onto CCs (Experimental conditions: 50 mg of adsorbent, 50 mL of dye solution (600 mg L−1) and temperature 55 ◦C).

#### Effect of Temperature

Temperature is an important parameter which can change the adsorption capacity of adsorbent and highly affect the adsorption of dyes from aqueous solution. The adsorption of CR dye onto CCs was studied at four different temperatures (25, 35, 45 and 55 ◦C) and the results were illustrated in Figure 6. It was obvious that increasing the temperature was accompanied by increasing the adsorption capacity of the adsorbent towards CR dye molecules. The % R.E. values were 55.51, 72.22, 77.0 and 82.40% at temperatures 25, 35, 45 and 55 ◦C, respectively. This indicated that a high temperature can facilitate the adsorption of CR dye onto CCs. This might be attributed to the swelling effect of the internal surface of an adsorbent, which contributed for the penetration of dye molecules into the interlayer of the adsorbent. This indicated that the adsorption of CR dye onto CCs was endothermic in nature [32]. Additionally, the gradual increase of temperature was followed by an increase in dye diffusivity and an increase in the dimensions of adsorbent pores. This consequently led to a reduction in the contribution of intraparticle resistance and decreased the effect of the boundary-layer [30]. The maximum adsorption capacity was achieved at 55 ◦C, which was used as the optimum temperature for batch experiments.

**Figure 6.** Impact of temperature on adsorption of CR dye onto CCs (Experimental conditions: 50 mg of adsorbent, 50 mL of dye solution (600 mg L−1) and pH = 9).

Impact of Initial Concentrations of the Dye

The adsorbed dye amount is strongly dependent on the initial dye concentration. This finding comes from the relationship between the initial dye concentration and the available active sites onto the adsorbent surface. In the present study, the use of 150 mg L−1 dye solution resulted in a complete adsorption, which reached 100% of CR dye removal (Experimental condition: pH = 9, temperature = 55 ◦C and equilibrium time = 6 h).

The adsorption of CR dye using CCs was studied at different concentrations (400, 500, 600 and 1000 mg L−1) and the results were illustrated in Figure 7. It could be noted that increasing the initial concentration of the dye was accompanied by a decrease in the removal percentage of CR dye. It decreased gradually to 99.5, 94.9, 92.55 and 76% at concentrations 400, 500, 600 and 1000 mg L−1, respectively. This can be attributed to the limited obtainable reactive centers available for adsorption process at higher concentrations. While the adsorption capability increased with increasing initial dye concentration, this is attributed to the high driving force of the concentration gradient at higher initial CR dye concentrations [30]. The maximum removal rate of CR dye was achieved at initial dye concentration of 400 mg L−1, which was considered the optimum concentration.

**Figure 7.** Impact of concentration of CR dye on adsorption onto CCs (Experimental conditions: 50 mg of adsorbent, 50 mL of dye solution, temperature 55 ◦C and pH = 4).

#### 3.3.2. Adsorption Kinetics

Studying adsorption kinetics is essential for suggesting the mechanism of adsorption, determining the optimum parameters for reaching the maximum removal percentage. and explaining the rate of dye uptake. The effect of time on the adsorption of CR dye using CCs was studied at different pHs, temperatures and initial dye concentrations using

three kinetic models: pseudo-first-order, Equation (6), pseudo-second-order, Equation (8), and Elovich model, Equation (9), in order to understand the adsorption mechanism.

At a Different pH

The results of kinetic data of the adsorption of CR dye onto CCs are summarized in Table 1. The plots of the adsorption for the three kinetic models at pH 4, 7 and 9 were illustrated in Figure 8.

**Table 1.** Kinetic model constants and correlation coefficients for the adsorption of CR dye onto CCs at different pH.

The pseudo-second-order model showed the highest value of correlation coefficient (R2) compared with the other models. It ranged from 0.997 to 0.991 at pH values from 4 to 9. This indicated that the pseudo-second order model has a perfect fit to describe the CR dye adsorption by CCs surface. In addition, there is a good agreement between the values of experimental and calculated qe for the pseudo-second-order, which is equal to 555.33 and 555.56 mg g−1, respectively, at pH = 4. This reflected and confirmed the excellent fitting of the pseudo-second-order kinetic model for the adsorption process of CR dye onto CCs.

On the other hand, the values of R2 obtained the for pseudo-first order showed relatively lower values than that obtained by the pseudo-second order model, since they ranged from 0.989 to 0.958 at pH values from 4 to 9. Additionally, the theoretical equilibrium adsorption capacities, qe, for pseudo-first-order were significantly different from that of the experimental ones which equal to 555.33 and 310.38 mg g−1, respectively, suggesting the inadequacy of the pseudo-first-order model for describing the adsorption kinetics of CR dye onto CCs.

The normalized standard deviation, Δqe, (%) was applied to check the validity of the pseudo-first-order and pseudo-second-order models quantitatively. Based on the values of Δqe (%) shown in Table 1, it is clearly noted that the pseudo-second-order model showed the best fit since it yielded lower values of Δqe (%) than those obtained from pseudo-first-order at different pH values. These results confirmed that the adsorption kinetic studies of CR dye onto CCs would be more favorable when applied by pseudo-second-order. Comparable data were published in a previous study to remove CR dye using chitosan-coated quartz sand [30].

**Figure 8.** Linear regressions of kinetic plots at different pH for the adsorption of CR dye onto CCs.

The fitness of the pseudo-second order kinetic model for the adsorption of CR dye onto CCs indicated that the adsorption process was proceeded by various mechanisms, including chemical interaction and electrostatic attraction of anions of the dye with the binding reactive centers onto the surface of CCs [30].

On the other hand, it is noted that the values of rate constant, *k*2, decreased gradually with the increase in the pH from 4 to 9, suggesting that pH 4 was the optimum value for adsorption. At pH 4, an enhancement in the electrostatic attraction force between the protonated sites of CCs surface and the negatively charged sulfonate groups (–SO3 −) of CR dye molecules took place. This led to an increase in the adsorption process at a low pH.

The Elovich model is one of the most interesting kinetic models that has a wide applicability to describe the adsorption systems related to the chemical nature. This model occurs when the adsorption involves the chemisorption reaction onto adsorbent surface, and the adsorption speed decreases as time passes due to covering of the adsorbent surface with an adsorbate [33].

Figure 8 showed the linear plots of the Elovich kinetic model, which was obtained by applying Equation (9) at pH 4, 7 and 9. The values of α, β and the respective R2 were summarized in Table 1. The obtained correlation coefficients were relatively high, which confirmed that this model fitted fully to the experimental results of the adsorbent. The values of R2 ranged from 0.96 and 0.94 at pH from 4 to 9. The good fitness of the experimental data for Elovich confirmed that the adsorption of CR dye onto CCs was controlled by a chemisorption process.

On the other hand, it was obvious that the values of the initial adsorption rate, α, decreased with increasing pH values. They ranged from 325.64 to 148.17 mg g−1 min−1 at pH values from 4 to 9 for the adsorption of CR dye onto CCs. Whereas, the values of the extent of surface coverage, β, increased with increasing pH; they ranged from 0.019 to 0.020 g mg−1 at pH values from 4 to 9.

#### At Different Temperatures

Temperature is a crucial parameter which highly affects both adsorption performance and adsorbent behavior. The experimental results of kinetic data were summarized in Table 2 for the adsorption of CR dye onto CCs. The plots showing the three kinetic models for CR dye adsorption by CCs, at 25 to 55 ◦C were illustrated in Figure 9.

**Table 2.** Kinetic model constants and correlation coefficients for the adsorption of CR dye onto CCs at different temperatures.

Based on the high values of correlation coefficient (R2) and the low values of the normalized standard deviation, Δqe (%), it was clearly apparent that the pseudo-secondorder model showed fully fit for the obtained results compared with pseudo-first-order model. R2 values obtained from pseudo-second-order adsorption model ranged from 0.990 to 0.991 at 25 to 55 ◦C. In addition, the qe for pseudo-second-order model was similar to that of the experimental one, since the values of experimental and calculated qe for pseudo-second-order model at 55 ◦C equal to 494.44 and 500 mg g−1, respectively. On

the other hand, the correlation coefficient obtained from the pseudo-first order model showed lower values than those obtained from pseudo-second order, since they ranged from 0.957 and 0.958 at temperatures from 25 and 55 ◦C for the adsorption of CR dye onto CCs. In addition, the experimental and calculated capacities showed different values and equalled 494.44 and 322.18 mg g−1 at 55 ◦C, respectively. This indicated the unsuitability of the pseudo-first order model for describing the adsorption process.

**Figure 9.** Linear regressions of kinetic plots at different temperatures for the adsorption of CR dye onto CCs.

Additionally, the value of rate constant, k2, increased gradually by increasing the temperature, as it increased from 1.64 × 10−5 to 1.81 × 10−5 g mg−1 min−1 when the temperature increased from 25 to 55 ◦C. This can be attributed to the low viscosity of the dye solution, which subsequently entrapped more dye molecules onto the adsorbent surface [34]. Furthermore, this might be due to the increase in collision probability between the active sites of adsorbent and the adsorbate and the decrease in the boundary layer thickness of adsorbent at an elevated temperature [35]. In addition, this might be due to the effect of swelling of the adsorbent internal structure, resulting from increasing temperature. This led to a greater diffusion capability for dye molecules into the adsorbent surface, reaching to its internal structure, and finally increasing the rate of adsorption [32].

Figure 9 showed the linear form of the Elovich model (Equation (9)). The values of the initial adsorption rate constant (α) and the desorption constant (β) at different temperatures (25, 35, 45 and 55 ◦C) were summarized in Table 2. Obviously, the experimental data for the adsorption of CR dye onto CCs agreed with the Elovich model. This result was reflected by R2 values which ranged from 0.963 to 0.940 at temperatures from 25 to 55 ◦C. The Elovich model assumes that the adsorbent active sites were heterogenous, and exhibited various energies for chemical adsorption [33]. This model is used for describing the kinetics of ion exchange systems. Hence, it is concluded that the adsorption of CR dye onto CCs was chemisorption. This finding is in agreement with the above results observed throughout the present study, which concluded that pseudo-second order is the best fit model. The obtained results also suggested that the chemisorption is the rate-determining step for the adsorption of CR dye onto CCs.

On the other hand, based on the results in Table 2, the values of initial adsorption rate constant, α, showed an increase with the increasing temperature. It ranged from 36.24 to 148.17 mg g−1 min−1 when the temperature increased from 25 to 55 ◦C for the adsorption onto CCs, whereas the desorption constant, β, showed a decrease with increasing temperature; the values of β ranged from 0.027 to 0.020 g mg−1 when the temperature increased from 25 to 55 ◦C for the adsorption onto CCs. This is attributed to the low number of available binding sites for the adsorption at an elevated temperature [36].

#### At Different Initial Dye Concentrations

The experimental kinetic data were summarized in Table 3. Figure 10 showed the graphical linear forms of the studied three models of kinetics for adsorbing CR dye onto CCs at different initial dye concentrations (400, 500, 600 and 1000 mg L−1).

**Table 3.** Kinetic model constants and correlation coefficients for the adsorption of CR dye onto CCs at different dye concentrations.

**Figure 10.** Linear regressions of kinetic plots for the adsorption of CR dye onto CCs.

The adsorption of CR dye onto the adsorbent showed excellent compliance with the pseudo-second-order, better than the pseudo-first-order, according to their respective correlation coefficient (R2) values and the values of the normalized standard deviation, Δqe (%). The R2 values obtained for pseudo-second-order ranged from 0.998–0.990, which were higher than those obtained from the pseudo-first-order (0.972–0.879) at concentrations which ranged from 400–1000 mg L−1. On the other hand, the values of the normalized standard deviation, Δqe (%), showed lower values for the pseudo-second-order than the pseudo-first-order. In addition, the calculated equilibrium capacities were close to the experimental ones for the pseudo-second-order. These findings indicated that the adsorption kinetics of CR dye by CCs perfectly followed the pseudo-second-order model, implying that the overall rate of the adsorption process was controlled by chemisorption [35]. Similar results were obtained for adsorption of CR dye onto trimellitic anhydride isothiocyanate-cross-linked chitosan hydrogels [23].

Regarding the values of rate constants for pseudo-second-order model, k2, there was a significant decrease for k2 values with increasing dye concentrations, as it decreased from 3.97 × 10−5 to 0.88 × 10−5 g mg−1 min−1 when the dye concentration increased from 400 to 1000 mg L−1. The same results are in agreement with another study [36]. The obtained results confirmed that the optimum concentration value was 400 mg L−1. This might be ascribed to the decrease in the unoccupied available reactive centers with time, which led to a decrease in the adsorption rate [35].

The linear form of Elovich model, represented in Equation (9), was illustrated in Figure 10 at different dye concentrations (400–1000 mg L−1) for the adsorption onto CCs. The values of the initial adsorption rate, α, and the extent of surface coverage, β, were summarized in Table 3. It is noted that the experimental results agreed adequately with the Elovich model, since R2 ranged from 0.963 and 0.913 at a dye concentration from 400 and 1000 mg L−1. The good applicability of the Elovich model indicated that the process was governed by chemisorption. On the other hand, the values of the initial rate of adsorption increased with increasing concentrations, and it increased from 111.97 to 1874.60 mg g−1 min−1 by increasing concentrations from 400 to 1000 mg L−1. This result might be ascribed to the higher concentration gradient, whereas a decrease in the extent of coverage (β) with increasing concentration was observed and decreased from 0.23 to 0.017 g mg−1 with increasing concentrations from 400 to 1000 mg L−1. Since at a higher concentration the gradient forces the particles of CR dye to show more adsorption towards the surface of adsorbent, this might negatively affect the desorption process [37].

#### 3.3.3. Mechanism of CR Dye Adsorption onto CCs

Mainly, in the aqueous solution, the CR dye (CR-SO3Na) would dissolve and dissociate giving the dye anions (CR-SO3 −) as shown in Equation (23).

$$\text{CR-SO}\_3\text{Na} \rightarrow \text{CR-SO}\_3\text{"}^- + \text{Na}^+\tag{23}$$

CCs is characterized by polar functional groups (–NH2 and –OH) on its surface. At a low pH, the protonation of these groups took place, leading to the formation of NH3 + and -OH2 +, respectively. Thus, an electrostatic interaction occurred between these two positively charged functional groups with the negatively charged dye anions, as shown in Equations (24) and (25).

$$-\text{OH} \stackrel{\text{H}^{+}}{\rightarrow} -\text{OH}\_{2}^{+} \stackrel{\text{CR}-\text{SO}\_{3}^{-}}{\rightarrow} \text{CR}-\text{SO}\_{3}^{-}\text{OH}\_{2}^{+} - \tag{24}$$

$$\text{C}-\text{NH}\_2 \overset{\text{H}^+}{\rightarrow} -\text{NH}\_3^+ \overset{\text{CR}-\text{SO}\_3^-}{\rightarrow} \text{CR}-\text{SO}\_3^-\text{NH}\_3^+ - \tag{25}$$

Additionally, the surface of CCs contains several heteroatoms, comprising oxygen and nitrogen. Therefore, binding of CR dye on the CC surface via hydrogen bonding and Van der Waals forces cannot be ruled out [30]. Thus, the possible mechanism involves a combination of adsorbent–sorbate electrostatic interactions, between the positively charged protonated groups and the negatively charged CR dye ions, in addition to H-bonding interactions and other physical forces such as π–π stacking and van der Waals forces [38]. Whereas, the electrostatic interaction might be the main mechanism for the removal behavior (Scheme 2).

**Scheme 2.** Schematic illustration of the interaction of CR dye with CCs at a low pH.

To identify and explore the diffusion mechanism for the adsorption of the CR dye, the intraparticle diffusion model was applied. By studying the intraparticle diffusion, the adsorption process involves a multi-step process: (i) bulk diffusion of dye molecules, (ii) film diffusion; dye molecules can diffuse to the surface of adsorbent through the boundary layer, (iii) intraparticle diffusion; dye molecules are transported from the surface into adsorbent pores, and (iv) dyes are adsorbed onto the adsorbent active sites through chemical or physical reaction. Due to the continuous stirring of the batch system, ignoring diffusion by bulk could be suggested and the step that determines the rate would be the intraparticle diffusion in biosorption [39].

#### The Intraparticle Diffusion

The intraparticle diffusion model was studied for the determination of the rate determining step. When the plot of qt versus t1/2 (Equation (10)) gives a straight-line passing via the origin, this implies that the process of adsorption is governed by diffusion, and the particle diffusion would be the rate controlling step. However, if the plot shows two or more linear regions, this indicates that the adsorption process takes place by a multistage adsorption [36].

The intraparticle diffusion results for the adsorption of CR dye onto CCs were analyzed at different pH values (from 4 to 9), temperatures (from 25 to 55 ◦C) and initial dye concentrations (from 400 to 1000 mg L−1), and are illustrated in Figures 11–13, respectively. The obtained plots did not pass through the origin. This confirmed that the intraparticle diffusion model was not the only rate-limiting step, suggesting other kinetic factors controlling and contributing the rate and the mechanism of adsorption of CR dye onto CCs. The multilinearity of the curves confirmed that the adsorption of CR dye molecules into the adsorbent surface took place in three stages. Firstly, the dye molecule diffused into the external surface of adsorbent by film diffusion, then in a slower step it diffused into the internal pores of the adsorbent throughout the intraparticle diffusion. The second stage was more stable than the first stage, indicating that it was the step for determining the rate, and the final step corresponded to the stage for final equilibrium [40].

**Figure 11.** The intraparticle diffusion plots for adsorption of CR dye onto CCs using different pH values.

**Figure 13.** The intraparticle diffusion plots for adsorption of CR dye onto CCs using different initial concentrations.

The first slope (kint.1) represented the fast adsorption of dye molecules onto adsorbent surface by electrostatic interactions, while the second slope (kint.2) revealed to the diffusion of dye molecules into the internal pores of adsorbent. The third slope (kint.3) corresponded to the saturation of all active sites on adsorbent surface (equilibrium stage). These findings suggested that CR dye adsorbed onto the surface of adsorbent and in its internal structure [38].

Studying the intraparticle diffusion with the variation of pH was carried out using pH values of 4, 7 and 9. The values of intraparticle diffusion rate (kint) and correlation coefficients (R2) were summarized in Table 4 for the adsorption of CR dye onto CCs. It is noted that the values of kint.2, decreased with increasing pH values; they equal 5.50, 5.24 and 5.17 at pH 4, 7 and 9, respectively. A similar result was illustrated in a previous study for adsorption of CR dye onto pine bark [41].

**Table 4.** Parameters of intraparticle diffusion model using different pH for adsorption of CR dye onto CCs.

The intraparticle diffusion at 25, 35, 45 and 55 ◦C was studied, and kint and R2 values were summarized in Table 5. The results showed an increase in the diffusion rate constant with the increasing temperature. For the adsorption of CR dye onto CCs, the diffusion rate constant (kint.2) increased from 2.84 to 5.17 mg g−1 min−1 when the temperature increased from 25 to 55 ◦C. All values of R2 are higher than 0.9, suggesting that the adsorption process took place by a combination mechanism.

**Table 5.** Parameters of intraparticle diffusion model for adsorption of CR dye onto CCs at different temperatures.

The intraparticle diffusion was studied also at various initial dye concentrations (400, 500, 600 and 1000 mg L−1) and the values of intraparticle diffusion rate (kint) and correlation coefficients (R2) for adsorbing CR dye onto CCs were summarized in Table 6. It was found that the values of kint increased with the increase in the concentration, since the values of kint.2 for the second stage equal to 4.46 and 5.51 mg g−1 min−1 at concentrations of 400 and 1000 mg L−1, respectively. These findings indicated that the diffusion of CR dye molecules into the interior pores of adsorbent surface increased with increasing the initial dye concentration, and it is in agreement with some previous studies [40].

#### 3.3.4. Adsorption Isotherm

The adsorption isotherm is considered one of the main important fundamentals for describing the mechanism of dye adsorption onto the adsorbent surface. These isotherms efficiently show the way of interaction between the dye and the sorbent surface.

The adsorption isotherms for CR dye removal by CCs were represented in Figure 14 for Langmuir, Freundlich, Temkin and Dubinin–Radushkevich (D–R) isotherm models. Additionally, Table 7 listed the results of the four models and their corresponding fitting

correlation coefficients (R2). The suitability of the model would be determined by the value of R2 which is the closest to unity [36].

**Table 6.** Parameters of intraparticle diffusion model for adsorption onto CCs using different CR dye concentrations.

**Figure 14.** *Cont*.

**Figure 14.** Adsorption isotherms for CR dye onto CCs at 55 ◦C and pH 9.

**Table 7.** Adsorption isotherm constants of CR dye by CCs.

**Isotherm Model Parameters**

According to R2 values, the adsorptive systems were better modeled by Langmuir adsorption isotherms than Freundlich adsorption isotherms, over the concentration ranges studied, since the value of R2 for Langmuir was 0.980 and for Freundlich was 0.917.

The good applicability of Langmuir adsorption isotherm for the removal of CR dye by the adsorbent indicated that the adsorption process occurred in a monolayer coverage manner with no interaction between adsorbed dye molecules with each other. In addition, the adsorption active sites were homogenously distributed on the adsorbent surface and they were identical for all dye molecules. Thus, each active site binds only with one dye molecule [39]. The maximum monolayer coverage capacity, qmax, at 55 ◦C was 666.67 mg g−1 for the removal of CR dye using CCs. Previous studies also reported the Langmuir model as the best fit for the adsorption of dyes using chitosan [42]. The obtained RL values for adsorbing CR dyes onto CCs ranged from 0.217 to 0.111, which indicates that the adsorption of CR dye onto CCs was a favorable process.

The Temkin model took into consideration the effect of the indirect interaction between the molecules of adsorbate. Assuming the linear decrease of adsorption heat of all the molecules in the layer due to the interactions between adsorbent and adsorbate, the adsorption process is characterized with a uniform distribution of binding energies [40].

Temkin constants, BT and KT, were listed in Table 7. The obtained plot provided a curve which indicated that the adsorption process did not follow this model.

The Dubinin–Radushkevich (D-R) isotherm model is used to describe the nature of the adsorption process. It is similar to Langmuir, while it is more general because it rejects the homogenous surfaces [40]. The linear form of D-R isotherm (Equation (16)) was obtained by plotting lnqe versus ε2, as illustrated in Figure 14. The values of β and ln Xm were calculated from the slope and intercept, respectively, and were listed in Table 7. Based on the value of R2, the D-R isotherm cannot be used to describe the adsorption of CR dye onto CCs, as it equals 0.807 for the adsorption onto CCs.

From the analysis of the experimental results showed in Table 7, it is concluded that the Langmuir isotherm model was the best model can fit the experimental results due to its highest R2. Freudlich and Temkin models did not show high accuracy as their correlation coefficients were not as high as the Langmuir isotherm, while the D-R model could not describe the adsorption process using CCs due to its low value of R2. These findings reflected the homogenous nature of the adsorbent. Therefore, it is assumed that the adsorption of CR dye by CCs took place uniformly onto its active sites [40].

#### 3.3.5. Adsorption Thermodynamics

The adsorption process of CR dye onto the adsorbent was studied using various temperatures (298, 308, 318 and 328 K). The values of thermodynamic parameters ΔG◦ were calculated according to Equation (19), while ΔH◦ and ΔS◦ were calculated from the Van't Hoff linear plot (Equation (20)).

The results were illustrated in Figure 15 and Table 8 for the adsorption CR dye using CCs. The values of ΔH◦ was 34.49 kJ mol−1, this positive value suggested that the interaction of CR dye adsorbed onto CCs was endothermic in nature [23].

**Figure 15.** Adsorption thermodynamic of CR dye onto CCs.

The negative values of ΔG◦ at all selected temperatures indicated that the adsorption of CR dye onto CCs was a spontaneous process. It decreased with increasing temperature, and this indicated the feasibility and favorability of the process at elevated temperatures [32].

The positive value of ΔS◦ (118.48 J mol−1K−1) indicated that the randomness increased at the interface between adsorbent and adsorbate solution during CR dye adsorption by CCs [42]. It can be explained as follows: throughout the adsorption process, the dye molecule received more entropy due to the displacement of adsorbed solvent molecule which resulting finally in an increasing randomness [43]. This actually occurred due to increasing the number of free ions of adsorbate that were found in order form near adsorbent surface than the adsorbed ions before adsorption process. Hence, the distribution of rotational energy increased, which consequently increased randomness at the interface between solid and liquid, and finally resulted in an increasing adsorption [35]. These findings are in a good agreement with previous studies [42].

#### 3.3.6. Activation Energy

The type of adsorption process, whether it is physisorption or chemisorption, can be determined by the magnitude of activation energy. When the activation energy is found in the range 0–40 kJ mol−1, this indicates that the reaction is physisorption, while if it is in the range 40–800 kJ mol−1, the reaction is chemisorption [24]. The value of Ea for the adsorption of CR dye onto CCs was 2.47 kJ mol−1, as illustrated in Figure 16. This value indicated that the adsorption was a physisorption in nature and was ascribed to a low potential barrier [44]. This result agrees with the adsorption of CR dye onto activated Moringa oleifera seed [45] and onto guava leaf-based activated carbon [36].

**Figure 16.** Arrhenius plot for adsorption of CR dye onto CCs.

#### 3.3.7. Comparison between CCs and Other Adsorbents for CR Dye Removal

To evaluate the affinity and efficiency of the CCs adsorbent, it became of interest to compare its adsorption maximum capacity for removal of CR dye with other adsorbents in previous reported studies as shown in Table 9. It is clearly observed that CCs possesses the highest removal capacities for CR dye than other reported adsorbents. It could efficiently adsorb CR dye at 55 ◦C with qmax of 760.44 mg g−1. This proved that CCs is an efficient and a promising adsorbent for the removal of CR dye from aqueous solution.

**Table 9.** Comparison of adsorption capacity of CCs with other adsorbents for the removal of CR dye.

#### 3.3.8. Desorption Studies

The reusability is considered one of the most important aspects to minimize the cost of adsorption by regeneration of adsorbent [51].

The desorption of CR dye was applied using Equation (22). No desorption was obtained by using ethanol, methanol and acetone as desorption medium, while using 0.1 N NaOH solution showed that the desorption percentage reached 58% after 5 cycles. These findings confirmed that CCs can be efficiently reused for the adsorption of CR dye from the aqueous solution.

#### **4. Conclusions**

Modification of chitosan was successfully done via a four-step procedure using cyanoguanidine at the final step. The produced cyanoguanidine-modified chitosan (CCs) possessed free primary amino and hydroxyl active centers in its main chains, in addition to nitrogen-rich cyanoguanidine as crosslinking linkages. Its structure was characterized by different techniques; FTIR, XRD and SEM. Its active basic groups acted as efficiently binding sites for the anionic CR dye from its aqueous solution. The adsorption capacity for CR dye onto CCs increased with increasing temperature reaching its maximum at 55 ◦C. The adsorption capacity decreased with increasing pH achieving its optimum value at pH 4. The removal percentage of CR dye decreased with the increase in the initial concentration of the dye. The pseudo-second-order model showed a perfect fit to describe the adsorption process, indicating the chemisorption process. The values of rate constant, k2, decreased gradually with the increasing pH from 4 to 9 and initial concentration of dye solution from 400 to 1000 mg L−1, while it increased with increasing temperature from 25 ◦C to 55 ◦C. The relatively high correlation coefficient R2 obtained from the Elovich model confirmed that this model fitted fully to the experimental results for the adsorbent at the studied parameters. This indicated that the adsorption of CR dye onto CCs was controlled by a chemisorption process. According to intraparticle diffusion, the transfer of CR dye molecules into the adsorbent surface took place in three stages. The plots of intraparticle diffusion showed that it was not the only rate-limiting step. The good applicability of the Langmuir adsorption isotherm indicated that the adsorption process occurred in a monolayer coverage manner with no interaction between adsorbed dye molecules with each other. The maximum monolayer coverage capacity, qmax, at 55 ◦C was 666.67 mg g−1. The Freudlich and Temkin models did not show a high accuracy as their correlation coefficients were not as high as the Langmuir isotherm, whereas, the experimental data confirmed the unsuitability of the Dubinin–Radushkevich isotherm model. The positive value of ΔH◦ (34.49 kJ mol−1) suggests that the adsorption was endothermic in nature. The negative values of ΔG◦ at all selected temperatures reflected that the adsorption was a spontaneous process. The positive value of ΔS◦ indicated the randomness elevation at the interface between adsorbent and adsorbate solution. The value of activation energy Ea was 2.47 kJ mol−1, assigning that the adsorption was physisorption. Thus, it is concluded that the adsorption of CR dye CCs involved physisorption and chemisorption processes.

**Author Contributions:** Supervision, N.A.M., N.F.A.-H.; Conceptualization, N.A.M. and N.F.A.-H.; Methodology, N.A.M., N.F.A.-H.; Investigation, E.F.A.; Formal analysis, N.A.M. and N.F.A.-H.; Writing—original draft, E.F.A.; Review and editing, N.A.M., N.F.A.-H. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research received no external funding.

**Institutional Review Board Statement:** Not Applicable.

**Informed Consent Statement:** Not Applicable.

**Data Availability Statement:** The data presented in this study are available on request from the corresponding author.

**Conflicts of Interest:** The authors declare no conflict of interest.

#### **References**

## *Review* **Advanced Polymeric Nanocomposites for Water Treatment Applications: A Holistic Perspective**

**Adedapo Oluwasanu Adeola 1,2,3,\* and Philiswa Nosizo Nomngongo 2,3,\***

**Abstract:** Water pollution remains one of the greatest challenges in the modern era, and water treatment strategies have continually been improved to meet the increasing demand for safe water. In the last few decades, tremendous research has been carried out toward developing selective and efficient polymeric adsorbents and membranes. However, developing non-toxic, biocompatible, costeffective, and efficient polymeric nanocomposites is still being explored. In polymer nanocomposites, nanofillers and/or nanoparticles are dispersed in polymeric matrices such as dendrimer, cellulose, resins, etc., to improve their mechanical, thermophysical, and physicochemical properties. Several techniques can be used to develop polymer nanocomposites, and the most prevalent methods include mixing, melt-mixing, in-situ polymerization, electrospinning, and selective laser sintering techniques. Emerging technologies for polymer nanocomposite development include selective laser sintering and microwave-assisted techniques, proffering solutions to aggregation challenges and other morphological defects. Available and emerging techniques aim to produce efficient, durable, and cost-effective polymer nanocomposites with uniform dispersion and minimal defects. Polymer nanocomposites are utilized as filtering membranes and adsorbents to remove chemical contaminants from aqueous media. This study covers the synthesis and usage of various polymeric nanocomposites in water treatment, as well as the major criteria that influence their performance, and highlights challenges and considerations for future research.

**Keywords:** fabrication techniques; inorganic contaminants; organic pollutants; polymer nanocomposites; water treatment

#### **1. Introduction**

Water plays a major role in the evolution of human civilization and industrialization. The population of the world is predicted to increase to 9 billion people, which will result in an increase in demand for freshwater and necessitate wastewater treatment and reuse [1–3]. In recent decades, human activities have exacerbated major environmental and conservation challenges [4–6]. The environmental challenges we currently face that pose a threat to sustainable life on land (15th Sustainable Development Goal) include water and air pollution, poor waste management, fallen groundwater tables, loss of biodiversity, land/soil debasement, global warming/climate change, depletion of natural resources, etc. [7–9].

Nanomaterials could be a viable and effective technique for overcoming significant obstacles in the development of efficient remedial technologies and environmental protection [10,11]. However, within the last century, the volume of industrial chemicals produced has increased dramatically, from 1 to 400 million tons in the year 2000 [12]. Furthermore, between 2000 and 2017, the worldwide production capacity of industries increased from

**Citation:** Adeola, A.O.; Nomngongo, P.N. Advanced Polymeric Nanocomposites for Water Treatment Applications: A Holistic Perspective. *Polymers* **2022**, *14*, 2462. https:// doi.org/10.3390/polym14122462

Academic Editors: Irene S. Fahim, Hossam E. Emam and Ahmed K. Badawi

Received: 20 May 2022 Accepted: 14 June 2022 Published: 16 June 2022

**Publisher's Note:** MDPI stays neutral with regard to jurisdictional claims in published maps and institutional affiliations.

**Copyright:** © 2022 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https:// creativecommons.org/licenses/by/ 4.0/).

1.2 billion tons to 2.3 billion tons [13]. The advances in science and the phenomenal growth in production capacity in the 21st century have also resulted in the emergence of troubling realities, such as the exponential increase in the spectrum of different classes of emerging pollutants detected in our water systems. As a result, new materials are urgently needed to clean up the polluted environment.

Nanostructured sorbents have a high capacity for the treatment of polluted water and may be tailored to target specific pollutants [10,11]. There have been recent improvements in the development of polymer nanocomposites (PNC), which have enhanced their novel applications in pollution remediation. Many chemical pollutants, such as dyes, heavy metals, and hydrocarbons, have been removed from wastewater using polymer nanocomposites [14,15]. Natural and synthetic polymers are both accessible. Natural polymers are those that are found in nature and can be extracted for usage. Water-based natural polymers include silk, wool, DNA, cellulose, and proteins [16]. Synthetic polymers, on the other hand, include nylon, polyethylene, polyester, Teflon, and epoxy, etc. Polymer nanocomposites possess unique synergistic features that are not possible to achieve with individual components functioning alone [17]. Inorganic nanofillers such as nanoclays, metal-oxide nanoparticles, carbon nanomaterials, and metal nanoparticles, can be introduced into a polymer matrix to create a PNC with better properties for a specific application [18,19]. Polymer nanocomposites have gained scientific prominence due to their broad range of applications in environmental remediation and treating various environmental challenges [16,20].

The aim of this review is to comprehensively examine the current state of polymeric nanocomposites as water remediation materials around the world, discussing various classifications of polymer nanocomposites, state-of-the-art synthetic methods, applications, merits, limitations, and potentials. The search keywords for this review are polymeric material, polymer nanocomposite, adsorbent, photocatalyst, organic pollutant, heavy metal, photocatalytic degradation, adsorption, and water treatment, and the literature scope is based on SCOPUS and Web of Science published papers and books.

#### **2. Synthetic Methods and Remedial Application of Polymeric Nanocomposites**

The choice of design, precursors, and synthetic methods for polymer nanocomposite development are germane. It entails choosing monomers, fillers, and other composite materials, as well as synthesis methods from a vast variety in order to produce PNC with the required property [21]. The determination of optimal processing strategy, considering the application intended for the PNC, and lastly, fabricating the product yield of the composite. This emphasizes the significance of the design and synthesis steps in the manufacture of PNC [18]. The ultrasonication-assisted mixing, shear mixing, microwaveassisted synthesis, roll milling, in-situ polymerization, ball milling, selective laser sintering, double-screw extrusion, and 3D printing (additive manufacturing) are among the most often utilized synthesis procedures [18,22]. Generally, there are two methods for making polymer nanocomposites: direct compounding and in situ synthesis [23]. The choice of synthetic methods demonstrates researchers' trend toward simple, scalable, and ultimately commercially viable and reproducible PNC [24].

Atomic layer deposition and plasma-assisted mechanochemistry have been reported to solve nanoparticle aggregation challenges in melt-mixing techniques. In addition, highfrequency sonication plays a similar role in the disaggregation of nanoparticles in mixing procedures. In-situ polymerization allows for the creation of thermodynamically stable nanocomposites, while electrospinning is an efficient way of creating porous objects. Furthermore, fabricating nanocomposites using selective laser sintering offers obvious advantages in terms of overcoming the aggregation problem [25]. Due to the unique features and various suitable applications of nanofillers, they are increasingly being used in the fabrication of PNC. One-dimensional (1D) nanofillers are fillers with one dimension smaller than 100 nanometers [18]. They normally come in sheets ranging in thickness from a few nanometers to hundreds of thousands of nanometers in length. One-dimensional nanofillers include montmorillonite clay and graphene nano-platelets. Both academics and

industry are interested in polymer nanocomposites made from different nanofillers and polymers [2]. Due to its consistent volumetric heating and the huge increase in reaction rate, microwave-assisted synthesis is an emerging technology used in numerous domains of materials science and chemistry. Microwave-assisted fabrication has also been used to create functionalized polymer-based nanoadsorbents [26].

#### *2.1. Dendritic Polymers*

Dendrimers are branched structures with a specific form, size, and molecular weight [27,28]. Dendrimers, which are made up of monomers that radiate from a central core, are becoming a popular type of polymer. Different classes of dendritic polymers are depicted in Figure 1. Due to the large number of functional groups in the core, on the surface, and the periphery and pocket, dendrimers are a high-capacity nanoscaled multidentate chelating agent/ligand, suitable for ion separation technology. The microenvironment within the dendrimer scaffold, on the other hand, may be an appropriate host for diverse contaminants. Dendrimers' potential for environmental clean-up has also been explored. Dendrimers' value for removing inorganic contaminants as well as organic pollutants has been extensively demonstrated in this regard [29]. Dendrimers are also attractive chemicals for the preparation of electrochemical sensors [30].

**Figure 1.** Six subclasses of dendritic polymeric material. Adapted with permission from Ma et al. [28]. Copyright (2016) Ivyspring International Publisher.

Poly(amidoamine) (PAMAM) dendrimers are extensively utilized dendrimers for the adsorption of pollutants in aqueous media. They are made up of ethylenediamine core, repeating units, and terminal units that make up the three fundamental units [29]. Their synthesis is performed by serially repeating two reactions: amino group addition to the double bond of methyl acrylate, accompanied by amidation of the resulting methyl ester with ethylenediamine. Each reaction step results in the production of a new dendrimer

(Figure 2). The addition-amidation reaction increases the diameter of PAMAM dendrimers by increasing the repeating units. This is a 1 nm per generation increase [29,31,32].

**Figure 2.** PAMAM growth on magnetic chitosan nanoparticles. Adapted with minor modifications with permission from Wang et al. [32]. Copyright (2015) Elsevier.

Most polymers are synthesized via single step; however, dendrimers are made in a series of steps, giving them well-tuned structures and narrow polydispersity. Divergent or convergent techniques, or a combination of both, can be used to synthesize dendrimers [33,34]. The divergent strategy entails the preparation of the dendrimer from a multifunctional core that is stretched outward by repeated reaction sequence (Figure 3A). On the other hand, a convergent method involves a bottom-up approach where dendrimers are constructed with tiny molecules that build-up on dendrimer surface and get connected to a central core through a sequence of inward-oriented interactions (Figure 3B). To combine the advantages of divergent and convergent synthesis, a combined divergent/convergent strategy, also known as the double-stage convergent approach (Figure 3C), has been further developed and employed for dendrimer synthesis [31,35].

However, structural flaws are a major problem in the synthesis of high-generation dendrimers because they are often caused by incomplete reactions or side reactions that take place as steric hindrance increases. Furthermore, due to their comparable chemical compositions and physical properties, dendrimers with structural flaws are frequently difficult to distinguish from intact dendrimers [33]. The convergent technique gives better control of the preparation of dendritic polymers, thus limiting structural flaws and impurities can be gotten rid of easily, as they conspicuously differ in morphology from the synthesized dendrimers. However, steric congestion is a challenge in the convergent approach, thus they are only suitable for small-scale dendrimers [36–38]. In the combination approach, a divergent technique is used to synthesize building blocks, which is then followed by convergent dendrimer assembly. Furthermore, when compared to either divergent or convergent synthesis, the sequence of reactions necessary for dendrimer preparation and purification can be minimized, thus ensuring more efficient production of higher generation [31,39].

**Figure 3.** Covalent dendrimer synthesis using: (**A**) divergent synthetic route, (**B**) divergent synthetic route, (**C**) Combined approach (Adapted with permission from Lyu et al. [31]. Copyright (2019) Elsevier).

Environmental Remediation Using Dendritic Polymers

Researchers are interested in dendritic polymer-based nanocomposites because of their tunable architectures and characteristics. Using appropriate functional moieties, dendritic polymers can be tailored to the target contaminants (Table 1). Dendritic PNC has high specific surface areas and pore volumes for trapping chemical contaminants. They can be combined with a variety of supports and other materials. A challenge to field/large scale applications is the multi-step synthesis, which requires profound expertise in polymer development. This challenge is solvable by collaborative research between organic chemists and separation science professionals.

Dendrimers have shown good potential for remediation of inorganic and organic pollutants (Table 1) in aqueous matrices based on the following reasons:

Heavy metal decontamination of polluted water with aid of dendritic polymeric materials is influenced by pH due to its impacts on the moieties on the surface of dendritic polymer. Furthermore, solution pH plays a crucial role in the adsorption performance of dendritic polymeric materials by altering the speciation and shape of metal ions [29,40]. The mercury (Hg) adsorption onto silica gel-supported salicylaldehyde-modified PAMAM dendrimers was investigated using the density functional theory (DFT) approach in another study [41]. Hg ions may bind with dendritic polymers by chelation due to the presence of oxygen and nitrogen species. The results of the Dubinin–Radushkevich (D–R) isotherm model indicate that mercury was removed via chemisorption onto sorbent pores, while thermodynamic tests revealed the endothermic and spontaneous nature of the sorption process, similar to what was reported in a recent study [42].

There are hundreds of polycyclic aromatic hydrocarbons (PAHs), which are essentially hydrocarbons containing two or more fused aromatic rings. Partial combustion or burning of carbon-based materials generates hazardous PAHs [43]. They can cause endocrine disruption and are carcinogenic [44]. Arkas et al. used hyperbranched poly(ethylene imine) in combination with silicic acid to remove pyrene and phenanthrene from an aqueous solution [45]. Sol–gel processes were used to make the dendrimer–silica nanoparticle composite. It was discovered that adding dendrimer to the silica nanospheres increased the adsorption of PAHs. The development of complexes with transferable charges between the tertiary amino groups of the dendrimer and PAHs resulted in a greater water treatment performance of PEI–silica nanoparticles.

Textile dyes, particularly acid dyes, are frequently utilized and they are poisonous, causing nausea, sleepiness, diarrhea, blood clots, and breathing difficulties [37,46,47]. Textile production is critical to commerce and industry. Unexhausted dye is being thrown into natural streams without being adequately treated by some businesses due to noncompliance with environmental requirements and wastewater treatment before discharge [24]. Various colors have been removed from aqueous solutions using dendritic polymers (Table 1). Hydrogen bonding, Vander Waals forces, electrostatic attractions, and dye trapping in dendrimers facilitate dye adsorption on dendrimer-based hybrid complexes [48]. A dendritic polymer was used by Hayati et al. to remove Acid Blue 7 (AB7), Direct Red 23 (DR23), Acid Green 25 (AG25), and Direct Red 80 (DR80) aqueous solution [49]. The adsorption process was discovered to be affected by pH, dye concentration and adsorbent dosage. The Langmuir (monolayer) adsorption isotherm model was shown to be the best fit for all dye adsorption by PPI dendrimers. The dye removal rate decreases as the pH rise, with the highest adsorption capacity occurring at pH 2 as a result of substantial electrostatic interaction between the anionic dye and the positively charged dendrimer surface. Removal of dye anions reduces when pH rises as a result of the decline in the number of positively charged sites.

Dendrimers enhance the activity of the photocatalysts for dye degradation when used as supports. PAMAM dendrimer, for example, was used to modify the photocatalytic activity of a polyoxometalate (POM) cluster for methyl red (MR) degradation by blocking POM-MR aggregation. After 20 min of irradiation under the same reaction conditions, POMdendrimer showed an increased dye degradation of 83% compared to 11% for POM [48,50]. Dendritic polymers and composites have demonstrated extraordinary potential in the field of water remediation. Dendrimers can be utilized on their own or with other materials. In dynamic systems, they can be immobilized over membranes. Dendrimer production, solution pH, contact time, and other factors influence their performance [37,51]. They are less toxic and biocompatible, which bodes well for their future commercialization and applications in water treatment.

 water.

#### *2.2. Polymeric Aerogels and Hydrogels*

Polymeric aerogels are unique types of porous material with interesting physicochemical features, including ultra-low thermal conductivity, low density, high porosity, large specific surface area, and controllable surface chemistry [67]. Aerogels are polymeric nanoparticle networks that are expanded by a gas across their whole volume, while hydrogels are cross-linked polymeric networks that may contain water within the interstitial spaces between chains [68,69]. The feasibility of this porous material in various applications has been thoroughly studied, thanks to recent breakthroughs in the synthesis of several forms of aerogels [70]. Aerogels and hydrogels are used as adsorbents for removing a variety of pollutants that are hazardous to the environment and human health. Figure 4 describes the general approach for the synthesis of aerogels, which involves the mixing of starting materials, gelling and a crucial step of removing solvents from pores of wet gels without distortion of the structure or network of molecules [71]. The ability to regulate the construction of a tunable aerogel network is enabled by the flexibility of the process conditions (change of the synthesis parameters, composition, etc.) [72]. The drying stage is critical for retaining porosity and integrity. The supercritical drying approach is the best method so far for achieving well-defined structures among the many drying processes, although freeze-drying is considerably easier, less expensive, and ecofriendly [73–75].

**Figure 4.** A general method for the preparation of aerogels. Adapted from Nita et al. [71].

On the other hand, hydrogel materials are frequently made by polymerizing acrylic monomers using a free radical initiator [76]. The resins can be made in an aqueous medium with solution polymerization or in a hydrocarbon medium with well-dispersed monomers (Figures 5 and 6). Hydrogel has gained a lot of attention in recent years due to its exceptional mechanical properties, swellability, biocompatibility, etc., but there are still a few things that can be done to improve it. A simple approach was reportedly used to make a novel alginate/graphene double-network (GAD) hydrogel, and its mechanical characteristics, stability, and adsorption performance were compared with alginate/graphene singlenetwork hydrogel (GAS) [77]. It was discovered that GAD has a smaller swelling power than GAS, resulting in better gel stability in highly concentrated alkali/salt solution. The GAD beads have a considerably greater adsorption capacity for heavy metal ions and dye than GAS beads.

**Figure 5.** Preparation of macroporous MnO2-based aerogel crosslinked with cellulose nanofibers. (**a**) scheme for the synthesis of cellulose nanofibers/MnO2 hybrid aerogels; (**b**–**d**) the height, diameter and density of cellulose nanofibers/MnO2 hybrid aerogel; (**e**) the SEM image of cellulose nanofibers/MnO2 hybrid aerogel. Adapted with permission from Cao et al. [75]. Copyright (2021) Elsevier.

**Figure 6.** The primary steps for making hydrogels on a semi-pilot and industrial scale. Adapted from Ahmed et al. [76].

Apart from the adsorption of aqueous- and gas-phase pollutants, several aerogel composites are acknowledged as useful tools for photocatalytic remediation by harnessing the (photo)catalytic moieties present in their framework [78]. Although aerogels and hydrogels are good remedial tools, there is still the challenge of sophisticated drying technology requirements, structural fragility and instability, and high processing costs, which should be subject to advanced research [78,79]. To strengthen the structure of aerogels and hydrogels, for improved performance and wider applications, aerogel and hydrogel-based nanocomposites have been developed (Table 2).

A starch-graft-poly(acrylamide)/graphene oxide/hydroxyapatite nanocomposite hydrogel was prepared by free radical cross-linking copolymerization [80]. It is worth noting that the n-HAp nanoparticles served as cross-linkers, allowing for ionic interactions and hydrogen bond formation between phosphates and the polymer network's grafted acrylic groups. It also demonstrated that many hydrogen bonds and potential covalent connections developed between electrons of graphene oxide (GO) sheets and acrylamide (AM) monomers grafted on a starch framework (Figure 7). The bioadsorbent was used for malachite green dye adsorption. The results revealed that the adsorptive interaction was viable, spontaneous, and endothermic. The pseudo-second-order model was used to characterize the malachite green (MG) sorption rates. The sorption data were best-fit by the Langmuir model with a maximum adsorption capacity of 297 mg/g. The hydrogel-based adsorbent demonstrated good regeneration capacity for up to five cycles [80]. The polymeric nanocomposite could be an environmentally benign and promising adsorbent for water treatment applications.

Similarly, a starch/MnO2/cotton hydrogel nanocomposite has been synthesized [81]. 0.008 M potassium permanganate, 0.7 g starch, and 0.6 M sodium hydroxide were used to make the optimal starch hydrogel nanocomposite at 50–55 ◦C. Potassium permanganate was used as a strong and cheap oxidizing agent to prepare manganese dioxide nanoparticles and cross-link the starch molecular chains to cellulose molecular chains. Because of its simple one-step synthesis approach, in-situ preparation of nanoparticles, cost-effectiveness, and desirable activity such as photocatalysis, biocompatibility, antibacterial properties of 93 percent against *S. aureus*, the starch hydrogel nanocomposite is a good material for various applications such as agriculture, medical, textile engineering, and water treatment [81].

A selective adsorbent was developed for the remediation of mercury (Hg) contaminants in aqueous solution [82]. The functionalized silica-gelatin hybrid aerogel with 24 wt.% gelatin content is ideal for the high efficiency and selective adsorption of aqueous Hg(II). By exposing cultures of *Paramecium caudatum* to Hg(II) and observing the model cultures with time-lapse video microscopy, the remediation efficacy of this adsorbent was assessed under realistic aquatic settings. With Hg(II) concentration, the viability of *Paramecium* shows a definite exposure-response relationship. When the Hg(II) concentration is greater than 125 g/L, viability diminishes. Only at Hg(II) concentrations greater than 500 ppm do the cells lose viability in the presence of 0.1 mg/mL aerogel adsorbent. The importance of the provided quasi-realistic aquatic toxicity model system in meeting the needs of environmental and chemical engineering technology is highlighted by the need for actual testing during adsorbent development [82].

**Figure 7.** Hydrogel adsorbent with starch-graft-poly(acrylamide)/graphene oxide/hydroxyapatite nanocomposite as constituent. Adapted from Hosseinzadeh and Ramin [80].

**Table 2.** Selected hydrogel/aerogel nanocomposites for water treatment applications.

#### *2.3. Polymeric Membrane and Biopolymers*

The need for freshwater will continue to rise over time, necessitating the development of cost-effective, eco-friendly water treatment technology [90,91]. Membrane processes such as reverse osmosis, membrane distillation, pervaporation, and others are among the existing technologies for drinking water management [10,92]. The mechanical and thermal stability of the present polymer membranes employed in water treatment procedures has various drawbacks [93].

Simple and cost-effective water filtration technologies based on recyclable biobased natural polymers such as chitosan, cellulose, and carbohydrate polymer modified with nanoparticles are needed [94]. These novel chitosan-based nanomaterials have been shown to effectively remove a range of pollutants from wastewater to permissible levels [95]. Other biobased polymer and polymeric support includes porous resins, polyaniline, polyacrylamides, cellulose acetate, cellulose or carboxymethyl cellulose, chitosan, alginate, eggshells, nanofibers and cellulose nanofillers (CNFs) [19,96]. CNFs potentially hold an important role in the advancement of the development of polymer nanocomposites. However, the challenge of extracting, isolating, and refining them, as well as surface improvements, can be developed systematically. It is safe to assume that CNF exploits will first find use in high-end applications, since their utilization offers significant benefits in limited quantities.

Membrane-based separation technologies are regarded as the most advanced separation technology due to various merits: they can meet a wide range of separation requirements; change of phase is not required, which saves energy; they are highly selective; they can retrieve trace amounts of contaminant; and they are simple, adaptable, and versatile [91,97]. Cellulose nanofillers have found applications in membranes/filters, injury dressings, dental implants, sophisticated adhesives, and materials requiring high transparency and mechanical qualities [98]. Findings suggest that when the choice of the polymer matrix is properly done, this enhances mechanical properties and lead to a high degree of optical transparency in the finished product [99]. The greatest barrier to their industrial applications remains the efficiency of recovery and facile strategies for morphological modifications as integrated with a system. It is worth noting that some recent research suggests that CNFs have increased barrier qualities as a function of cellulose crystal content, implying that they could potentially be used in more environmentally friendly packaging materials [100,101].

Adsorption and rejection are the two strategies for extracting solutes from wastewater using adsorptive membranes. Molecular sieving and filtration rejects any solutes larger than the membrane's pore size when water-containing solutes come into contact with the active layer of the membrane (Figure 8). Smaller solutes will pass through the active layer and into the support layer, which acts as an adsorption microsphere [19,102]. These solutes then react/bind to form a tight internal spherical complex, resulting in filtered water permeating from the adsorptive membrane that meets the requisite standards [103]. The nanofiltration (NF) and reverse osmosis (RO) membranes are widely utilized due to their high-water permeability, low-pressure need, and low cost (Table 3). The improvement of the membrane's functional adsorption sites has been reported to enhance its performance [104]. Therefore, using hydrophilic nanoparticles to boost the clean treated water flux can be quite successful, as long as the number of nanoparticles introduced is low. Having less than 6% adsorbent in the membrane matrix is typically recommended to avoid clogging membranes and counterproductive effects [105,106].

**Figure 8.** Capabilities of different separation/adsorptive membranes using microfiltration (MF), ultrafiltration (UF), nanofiltration (NF) and reverse osmosis (RO) processes. Adapted with permission from Hmtshirazi et al. [19]. Copyright (2022) Elsevier.

Recently, a carbon molecular sieve (CMS) with different concentrations was used to make asymmetric polyethersulfone (PES-15 wt. percent) mixed-matrix membranes via the phase inversion method [107]. Loading 1 wt.% of CMS, the membranes improved in hydrophilicity, heterogeneity, porosity, net surface charge and mean pore diameter. Furthermore, 1 wt.% CMS addition to PES, led to increased pure water flux from 55.77 to 75.05 L/m2.hr. This polymeric nanocomposite showed improved dissolved ions-removal efficiency, compared to the unmodified PES membrane [107]. Table 3 presents various polymeric membranes that have been used to remove several types of pollutants from water, and the process parameters as well as the treatment technology adopted by various researchers.

Polymeric adsorptive membranes are generally considered potent pollution remediation technologies and have various applications in wastewater treatment plants and household water polishing. Cellulosic and other polymeric membranes have the capacity to remove various classes of persistent and emerging chemical pollutants from wastewater that are recalcitrant to conventional methods. The combined advantage of adsorption and filtration mechanisms, as well as the variety of forms and configurations, make these imprinted membranes (or membranes in general) and biopolymers appealing (Figure 9). Adsorptive membranes address fouling, minimize operational cost, adsorbent reusability, and enhances adsorption capacity, membrane permeability, rejection rates, and selectivity.

**Figure 9.** Removal of multiple contaminants selectively via an adsorptive membrane. Adapted from Salehi et al. [106].

polymeric membrane nanocomposites for water treatment applications.

**Table 3.** Selected

**Table 3.** *Cont.*

#### **3. Considerations for Future Research**

Pollutants in aqueous media have been remedied using a variety of polymeric nanocomposites adsorbents [128]. The toxicity of the adsorbent, when used for water filtration, is a major concern. Cellulose and other carbonaceous adsorbents (i.e., graphene, carbon nanotubes) have received significant attention for water purification due to their large surface area and affinity for inorganic and organic pollutants, as well as microbial contaminants [129,130]. Their toxicity, on the other hand, is relatively unknown and this requires more evaluation. Carbonaceous compounds and organic-based nanomaterials with reactive moieties may be hazardous to human health [131]. Therefore, only rational chemical design combined with a comprehensive understanding of potential biological interaction and risks can result in the fabrication of polymeric nanocomposites that is both safe and effective [18].

In recent research, polymers have been combined with materials such as carbon nanotubes, graphene oxide, and magnetic nanoparticles, in order to improve physicochemical and morphological properties [51,53]. Therefore, it makes sense to evaluate the toxicity of the entire composite system and not just the polymers. It is a prevalent fallacy that certain metals are only harmful in their pure form and functionalizing with a polymeric framework lessens toxicity. However, such statements should be backed up by scientific evidence.

Polymeric membrane technology has some concerns and challenges that need to be addressed by the scientific community. Examples include, but are not limited to:

#### **4. Conclusions**

Polymer nanocomposites' development and application in water and wastewater treatment holds immense potential due to the added benefits of the synergism between adsorption and membrane filtration approaches. Polymeric membranes have shown remarkable performance in removing both established and emerging chemical pollutants

from water. In the synthesis of polymer nanocomposites, water, atomic layer deposition, and plasma-assisted mechanochemistry have been identified to solve the problem of nanoparticle aggregation in melt-mixing techniques. Mixing techniques also benefit from sonication at high frequencies. In-situ polymerization allows for the creation of thermodynamically stable polymeric nanocomposites. Electrospinning is an efficient way for developing porous materials. In addition, fabricating nanocomposites with selective laser sintering offers obvious advantages in terms of avoiding aggregation.

Furthermore, the diversity in the types of materials, i.e., dendritic, biopolymeric, cellulosic and various polymeric composites, available for adsorptive membranes application is advantageous. However, the use of polymeric nanocomposites as adsorbents or as membranes (involving both filtration and adsorption) have various challenges associated with their preparation, human and environmental health, and as well as scalability and adaptability for industrial and field applications. As a result, future research should focus on overcoming these obstacles in order to effectively and sustainably use polymeric nanocomposite adsorbents and membranes in water treatment sectors.

**Author Contributions:** A.O.A.: conceptualization; investigation; data curation; formal analysis; writing—original draft preparation; writing—review and editing. P.N.N.: fund acquisition, project administration; supervision; validation; writing—review and editing. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded by Department of Science and Innovation-National Research Foundation South African Research Chairs Initiative (DSI-NRF SARChI) (grant no. 91230).

**Acknowledgments:** This work was supported by the Department of Science and Innovation-National Research Foundation South African Research Chairs Initiative (DSI-NRF SARChI) (grant no. 91230) and Global Excellence Stature (GES 4.0) funding received from the University of Johannesburg.

**Conflicts of Interest:** The authors declare no conflict of interest.

#### **References**

## *Review* **Functionalized Nanomembranes and Plasma Technologies for Produced Water Treatment: A Review**

**Anton Manakhov 1,\*, Maxim Orlov 1, Vyacheslav Grokhovsky 1, Fahd I. AlGhunaimi 2 and Subhash Ayirala 2**

**\*** Correspondence<EMAIL_ADDRESS>Tel.: +7-9158-494-059

**Abstract:** The treatment of produced water, associated with oil & gas production, is envisioned to gain more significant attention in the coming years due to increasing energy demand and growing interests to promote sustainable developments. This review presents innovative practical solutions for oil/water separation, desalination, and purification of polluted water sources using a combination of porous membranes and plasma treatment technologies. Both these technologies can be used to treat produced water separately, but their combination results in a significant synergistic impact. The membranes functionalized by plasma show a remarkable increase in their efficiency characterized by enhanced oil rejection capability and reusability, while plasma treatment of water combined with membranes and/or adsorbents could be used to soften water and achieve high purity.

**Keywords:** plasma; produced water; membranes; desalination; gliding arc; nanofiltration; nanofibers

#### **1. Introduction**

The sustainable development of modern society is increasing the energy demand, leading to a higher power being produced [1]. At the same time, the importance of all environmental concerns is growing substantially every year [2]. While attention to renewable sources is great for multiple reasons, the improvements and optimization of the technologies in the current format of the energy sector are also highly essential [3]. This necessitates the need to revise processes in the oil & gas industry towards more sustainable strategies in terms of energy efficiency, process-water reusability, and recycling of chemicals [4].

The water resources have a significant impact on the efficiency of the oil & gas industry, as the energy sector is 6th in the most water-intensive industries with 52,000,000,000 m3 of freshwater consumed annually for global energy production [5]. The reuse of wastewater streams became an important research topic with more than 130 000 publications in this field. The number of publications focused on the produced water treatment is also rapidly growing and will be even more significant in the coming years (Figure 1a).

In the oil & gas sector, the produced water is the most significant waste stream, especially for the wells with high maturity [6]. Production activities generate increasing loads of produced water up to thousands of tons each day or more, depending on the geographic area, formation depth, oil production techniques, and age of oil supply wells. The regulations related to the discharge of the produced water are stringent, although they might vary for different countries. For example, in Denmark, it must be below 30 ppm [7]. The Regional Organization for Protection of the Marine Environment (ROPME) issued several regional agreements controlling drilling operations in Arabian Gulf. According to the ROPME protocol concerning Marine Pollution resulting from Exploration and Exploitation of the Continental Shelf, the effluent discharge should contain less than 15 ppm of oil, and and no drilling waste above 40 ppm [8]. The water is produced from different

**Citation:** Manakhov, A.; Orlov, M.; Grokhovsky, V.; AlGhunaimi, F.I.; Ayirala, S. Functionalized Nanomembranes and Plasma Technologies for Produced Water Treatment: A Review. *Polymers* **2022**, *14*, 1785. https://doi.org/10.3390 /polym14091785

Academic Editors: Irene S. Fahim, Ahmed K. Badawi and Hossam E. Emam

Received: 7 April 2022 Accepted: 24 April 2022 Published: 27 April 2022

**Publisher's Note:** MDPI stays neutral with regard to jurisdictional claims in published maps and institutional affiliations.

**Copyright:** © 2022 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https:// creativecommons.org/licenses/by/ 4.0/).

oil fields containing different chemical compositions. Currently, produced water is also known as industrial wastewater containing heavy metals that are toxic to humans and the environment, requiring special processing so that they can be properly disposed [9]. The ever-evolving and increasingly stringent regulatory standards for discharging produced water pose major environmental and economic implications [6]. Hence, the treatment of produced water creates a significant challenge that new materials and technologies must be capable of solving.

**Figure 1.** The evolution of publications number (per annum) related to produced water treatment (**a**) and plasma wastewater treatment (**b**). The data was acquired from Scopus database.

The importance of this topic is also supported by the decisive engagement/increase of R&D funds, among international organizations and private companies, with a total number of awarded research grants of about 328 (determined according to the literature analysis using Scopus database). For example, the efforts to enhance the economic efficiency of sustainable produced water treatment processes were supported by different initiatives, including the U.S. Department of Energy (DOE). These initiatives actually aim to develop advanced technologies that can improve oil and gas production while protecting water resources, reducing water use [10].

The topic of produced water treatment was reviewed several times in the last decade, e.g., in [11–13]. The water treatment methods are generally based on chemicals, thermal processes, electrodialysis, membrane distillation, etc. However, chemical water treatment and thermal treatment have certain limitations, including economic efficiency and environmental impact (production of chemical, energy-intensive processes). In contrast, membrane technologies, plasma treatment, or combined technologies have excellent potential for oil/water separation, desalination, and heavy metals/toxic chemical removal. The use of plasma technologies for wastewater treatment also became rapidly growing topic with several hundred publications per annum, as shown in Figure 1b.

This review is focused on hybrid strategies and technologies based on nanomaterials and plasma processes with a significant focus on the sustainability of each proposed methodology. Indeed, as the plasma process is an environmentally friendly technology enabling new materials, it has excellent potential to enhance the water/oil separation process, induce new reactive cites to improve the desalination, and, finally, destroy toxic organic chemicals or convert heavy metals into sediments.

#### **2. Plasma Modified Membranes Used for Oil/Water Separation**

Plasma modification of different materials enables enhanced properties by adjusting surface chemistry and roughness without damaging the bulk properties [14]. The nonthermal plasma is generally used to modify the surface of materials as it will not damage the surface polymeric materials, including temperature-sensitive materials [15]. Nonthermal plasma can be ignited at different pressures by applying the high voltage to the electrodes separated by the gap (filled by noble gases, air or a mixture of precursors) [16]. To obtain homogenous plasma at atmospheric pressure, the surface of the electrodes could be covered by a dielectric layer (to prevent arcing). At the same time, the frequency of the applied voltage can be maintained in the range from a few kHz up to several MHz. It was often used to increase the superhydrophobicity or superhydrophilicity of nanomaterials to develop highly efficient nanofiltration membranes [17,18].

Nanofiltration (NF) is a perspective technology that provides a combined solution to reject both organic and inorganic pollutants, thereby making it more beneficial than ultrafiltration (UF). However, compared to reverse osmosis (RO), the salt and metal ion rejection percentages are typically lower (50–80% vs. 99%). The rejection of organic molecules is also somewhat lower. In comparison, ultrafiltration membranes allow passing almost all components with MW < 10,000 g/mol, while the pore size of nanofiltration membrane is much smaller, allowing many smaller organic molecules to be rejected. As a result, nanofiltration can be classified between reverse osmosis and ultrafiltration on the filtration spectrum [19].

Nanofiltration has a strong potential for oil/water separation, as it allows to obtain cleaner water from oily solutions, including the possibility for in-house recycling [20]. If the discharge of the produced water is preferred, nanofiltration permeate can also be clean enough to meet the stringent discharge requirements [21]. However, different challenges, including economic viability, robustness, and modest efficiency, hinder the largescale nanofiltration application. This section summarizes the most advanced eco-friendly oil/water separation solutions based on the modified nanofiltration membranes.

A novel environment-friendly method to prepare nanofiltration membranes for highly efficient separation of water-in-oil emulsions was reported by Yang et al. [22], where durable biphilic (oleophilic and hydrophilic) surface was prepared by atmospheric air plasma treatment of cellulose filter paper followed by various grafting procedures. The hydrophobic surface with a microscale hierarchical structure was constructed using long-chain alkyl silane grafting. The authors utilized silanization by acting aminopropyltrimethoxy silane

(APTMS), glycidyl propyl trimethoxy silane (GPTMS) and hexadecyl trimethoxy silane (HDTMS) onto plasma-activated SiO2 nanoparticles using surface chemistry approach as shown in Figure 2. The samples exhibited excellent separation performance for surfactantstabilized water-in-oil emulsions with the filtration efficiency of up to 99% and the oil recovery purity > 99.98 wt%. As shown in Figure 2, the water-amine attraction (for APTMS functionalized layers) and oil-alkyl attraction (for the GPTMS side) synergistically destabilize the emulsion and separate the two phases. The formatted coating was very durable against ultrasonication in acetone and can be stored for 12 months without degradation upon aging. The coating also exhibited excellent recyclability (over 20 cycles), although the multi-step process is too complicated and not cost-effective for large-scale application.

**Figure 2.** Scheme for the nanofiltration membranes preparation using atmospheric air plasma and modifications by dip-coating.

Various materials, including fluoropolymers, treated by plasma exhibited enhanced oil-water separation properties. For example, the traditional polytetrafluoroethylene (PTFE) membrane was pretreated using the atmospheric pressure glow discharge (APGD) plasma. Then, the surface of the plasma-treated PTFE membrane was modified by the photoinduced graft polymerization of 2-acrylamido-2-methyl-propyl-sulfoacid (AMPS). The obtained surface demonstrated a highly negative charge, ensuring substantial oleophobic property of the PTFE membrane and leading to improved anti-fouling performance compared to the unmodified sample [23].

A superhydrophilic PVDF membrane obtained by acrylic acid plasma polymerization followed by TiO2 nanoparticles self-assembly was employed for oily produced water treatment [24]. The TiO2 nanoparticles were immobilized onto the membrane surface via the coordination of Ti4+ with a carboxylic group from plasma layers. It is worth noting that the immobilization step did not affect the valence state of Ti4+ (Figure 3). The TiO2 decorated PVDF membranes exhibited high uniformity and dramatically improved surface hydrophilicity. After modification, the permeation flux was increased more than four times, and the oil rejection rate was higher than 92%.

**Figure 3.** Preparation of nanofiltration membrane by TiO2 self-assembling on a PVDF surface.

Lin and co-workers developed Janus membrane composed of the PVDF substrate functionalized with a special silicification layer inducing the asymmetric wetting selectivity [25]. The hydrophobicity of the membrane was facilitated by grafting the hydrophobic perfluorodecyltriethoxysilane, while unilateral plasma etching treatment of modified membrane allowed to induce the hydrophilicity. By tuning the surface chemistry of hierarchical nanosphere-like architecture of a robust membrane with superhydrophobicity/superoleophilicity (perfluorodecyltriethoxysilane modified side), and underwater superoleophobicity/superhydrophilicity (plasma-modified side) were obtained. The membranes showed an extremely opposing surface wettability with a high water/underwater oil contact angle (CA) difference up to 150◦ due to the exceptionally asymmetric wetting selectivity incorporated with the 15 nm ultrathin silicification layer. As a result, the wateroil separation efficiencies in surfactant-stabilized oil-in-water and water-in-oil emulsions reached 99.8% and 99.0%, respectively. In addition, the authors demonstrated nearly 100% recovery ratio of permeate flux after several cycles of oil-in-water and water-in-oil emulsion filtration tests.

It is worth noting that the multi-step functionalization of PVDF membranes has a particular drawback: this process requires several reactors and a relatively long treatment time. As a result, it increases the production cost. Therefore, many research works were focused on utilizing cheaper materials and improving the cost-efficiency of the modification to one-step processes.

Complimentary to the PVDF, polyethersulfone (PES) ultra- and nanofiltration membranes are attractive materials due to their high efficiency and stability [26]. Sadeghi and co-workers prepared efficient PES ultrafiltration membranes using a corona plasma treatment to reduce membrane fouling during the separation of oil/water emulsions [21]. These PES membranes were fabricated by the phase inversion method and thoroughly analyzed. The authors investigated the effects of solvent in casting solution and corona treatment conditions on membrane surface properties, morphology, and separation performance. The corona treatment induced the grafting of hydrophilic OH groups leading to the higher

water flux in all configurations of membranes without significant changes in oil rejection. Finally, the rejection rate up to 99.5% was observed at the oil/water permeability range from 34 to 70 (L m−2 h−1 bar−1). Similar approach was exploited by Adib et al. [27] to explore the potential of corona plasma-treated PES membranes for water-oil separation. By treating the membrane in corona discharge at 360 W for 6 min, the authors achieved their most promising sample with the rejection rate of 92%. Membranes exhibited mean pore size~30 nm at a porosity of 85%. To enhance the grafting of hydrophilic COOH groups and, additionally, to utilize CO2 to improve the sustainability of the approach further, carbon dioxide can be admixed to the plasma-forming gas mixture. CO2 plasmas were used to modify hydrophobic polysulfone ultrafiltration membranes to create hydrophilic surfaces throughout the membrane structure [28]. The water contact angle (WCA) of the upstream side of the membrane (facing the plasma) decreased to zero after treatment and remained at the same condition even after several months of aging.

Fedotova and co-workers employed radio-frequency (RF) glow discharge to treat polysulfonamide membranes (pore size of 0.01 μm) [29]. Low-pressure RF discharge in an argon and nitrogen atmosphere at an anode voltage of Ua = 1.5 kV and contact time τ = 1.5 min led to a decrease in the contact angle from 59.6◦ to 47.9◦, and grafting of oxygen groups. Experiments on the membrane separation of a 3% oil-in-water emulsion have been conducted. The results have shown that the use of plasma-treated polysulfonamide membranes leads to the intensification of the oil/water separation process.

In order to enhance the wettability or superhydrophobicity, many researchers prefer to use the materials with advanced nanoporous structures, e.g., electrospun nanofibers. Indeed, the production of nanofibrous membranes has approached the commercial scale processes with many products, including air-filters and masks, showing high cost efficiency [30–32]. Thus, the implementation of nanofibrous membranes for water treatment offered significant potential.

Nanofibrous membranes with superhydrophilic/superoleophobic surfaces were constructed by grafting acrylic acid onto low-pressure nitrogen plasma treated electrospun polystyrene/polyacrylonitrile (PS/PAN) nanofibrous foils as shown in Figure 4 [33]. The WCAs of the PS/PAN membranes decreased to 0◦ after grafting treatment of acrylic acid, thereby proving that the modification improved the surface hydrophilicity of the membranes due to the introduction of hydrophilic groups. Oil/water permeation tests confirmed a high oil/water separation potential of plasma-treated PS/PAN membranes. The results showed that these membranes effectively separated the layered oil/water mixture with permeate flux up to 57,509 L m−2 h−1, while high fluxes of 1390–6460 L m−2 h−1 were observed for the separation of different oil-in-water emulsions. The prepared membranes exhibited superhydrophilic and underwater superoleophobic surfaces, which could prevent oil droplets from adhering to the surface, thereby contributing to membrane anti-fouling. In the separation step, the resultant membranes, solely driven by gravity, had high separation efficiency as well as one to two orders of magnitude higher fluxes than that of traditional polymeric filtration membranes with similar permeation properties [33].

Notably, the membranes maintained high flux and efficiency even after several separation cycles to demonstrate excellent recyclability. Therefore, such flexibility in reusing these membranes is expected to provide potential cost-effective materials for oil/water treatment.

The strategy shown above for PS/PAN membrane modification by acrylic acid grafting using plasma activation required a wet chemical step that is preferably switched to a gasphase process or a single-step plasma modification for better cost-efficiency. Of course, the one-step atmospheric plasma method would be the most desirable, but it is pretty challenging at the same time [34,35]. Moreover, the wet chemical steps are often eliminated by low or atmospheric pressure plasmas to improve the economic viability.

**Figure 4.** PS/PAN membrane modification by acrylic acid grafting using plasma activation by grafting of radical Oxygen species (O\* ).

Several researchers demonstrated the capability of atmospheric plasma for modification of membranes to stimulate the oil/water separation. Chen et al. proposed the environment-friendly oil/water separation methodology based on atmospheric plasmatreated nylon meshes [36]. Nylon fibers were exposed to atmospheric pressure plasma for surface modification, leading to micro/nanostructures and oxygen-containing groups. Consequently, the functionalized meshes possess superhydrophilicity in air and thus superoleophobicity underwater. The obtained materials exhibited an efficiency above 97.5% for various oil/water mixtures. Moreover, the authors showed that the functionalized nylon meshes have excellent recyclability and durability in oil/water separation. Fedotova et al. optimized the plasma modification of polysulfonamide membranes (with a pore size of 0.01 μm) for the separation of model "oil-in-water" emulsions containing 3% of industrial oil [37]. The atmospheric pressure plasma torch was ignited for short time periods from 1 to 7 min at the anode voltage ranging from 1.5 to 7.5 kV. The highest performance was observed in the case of membranes treated with plasma for 7 min, leading to increased membrane efficiency from 90 to 99% owing to the surface hydrophilization. Indeed, the membrane exhibited a water contact angle 59.6◦ before treatment, and after a 4-min plasma treatment at 7.5 kV it decreased to 19.5◦. The authors also confirmed that the plasma treatment leads to a change in the surface structure of the membranes, namely, a decrease in their roughness. The membrane internal microstructure also undergoes changes that increase their crystallinity.

You et al. employed He/CH4/C4F8 atmospheric plasma deposition to modify the surface of stainless steel meshes [38] and applied them for oil/water separation. The finetuning of the plasma process parameters enabled the selective functionalization of each side of the membranes, leading to superhydrophobic-superoleophilic and superhydrophobicoleophobic sides. The authors demonstrated the successful separation in a simple test, where a 50 mL oil-water solution in 6 min was treated. According to the FTIR analysis, approximately 88% of the hydrocarbons in the oil-water mixture were separated in 6 min without any external forces, and no water was detected in the collected oil. The separation efficiency of up to 99% has been achieved. A similar concept of oil/water separation using oxygen plasma treatment of Cu coated meshes to induce hydrophobic and oleophilic behaviors is presented by Agarwal et al. [39]. This concept's advantage is the possibility of using cheaper and more robust steel grades for separating meshes.

Sometimes, produced water may contain special chemicals, e.g., hydrolyzed polyacrylamide (HPAM), which is added to the displacing fluid to enhance the mobility and extraction of the oil phase. Optimized ultrafiltration, using ceramic membranes with a surface pore size of 15 kDa, to effectively separate HPAM from produced water was demonstrated by Ricceri et al. [40]. The precipitation and ultrafiltration may be used in sequence as they complement each other in several ways. Moreover, ultrafiltration membranes with 15 kDa cut-off provided the best combination of productivity and HPAM removal (92%). However, the authors highlighted the problem of membrane fouling under severe conditions.

#### **3. Nanofiltration Membranes for Desalination**

The rejection of salts from the produced water is very challenging task. However, a sustainable nanofiltration approach using modified nanomaterials significantly progressed to solve this problem by using various advanced 2D nanomaterials, including graphene and boron nitride [41,42]. An exciting approach based on membrane crystallization was demonstrated by Ali et al. [43]. The authors employed membrane crystallization for desalination and salt recovery from produced water streams at a semi-pilot scale using convetional polypropylene or in-lab prepared PVDF hollow fiber membranes. The experiments were carried out in lab scale and semi-pilot scale. It is worth noting that the salts may also be in high demand, and the recovery of ions can be highly interesting if economically viable. Unlike conventional crystallizers, in membrane crystallization, well-controlled nucleation and crystal growth are achieved through a uniform evaporation rate through the membrane pores. The crystals recovered by membrane crystallization have higher purity and narrow size distribution. The recovered crystals were sodium chloride with high purity (>99.9%). It was demonstrated experimentally that at a recovery factor of 37%, 16.4 kg NaCl per cubic meter of produced water could be recovered.

Another method to extract salts from produced water is the airgap membrane distillation. Woo et al. modified the PVDF membrane by using electrospinning of PVDF, modified the surface of the nanofibrous foil, and evaluated the membrane's performance using real reverse osmosis brine with a produced water as a feed [44]. The results suggest the successful chemical modification of the membrane by plasma treatment without significantly altering the morphology and its physical properties, providing increased fluorination primarily through the formation of CF3 and CF2-CF2 bonds, with the treatment duration. The optimal 15 min plasma treatment induced the omniphobic property of the membrane with the highest hydrophobicity and high wetting resistance to low surface tension liquids such as methanol, mineral oil, ethylene glycol, and consequently high liquid entry pressure (187 kPa). These improved properties translate to high flux (15.28 L/m2h) and salt rejection (~100%) performances of membrane even with the addition of up to 0.7 mM sodium dodecyl sulfate in the reverse osmosis brine feed during air gap membrane distillation.

#### **4. Plasma-Aided Technologies for Produced Water Treatment**

The plasma-aided water treatment is generally achieved by activating highly efficient adsorbents or direct plasma treatment of water by atmospheric discharges. Each technique has its own advantages and drawbacks, and the most noticeable results are reviewed below.

#### *4.1. Plasma-Activated Adsorbents*

The economic aspects for the oil/water separation play a crucial role in the evolution of produced water treatment technology. Often cost-effective reusable materials with lower performance may have gained economic benefits against polymeric membranes based on PVDF, fluoropolymers, polysulfone, or other expensive materials. Hence cheaper feedstock material can play a crucial role in reducing costs and enhancing the business value for membrane oil/water separation. In this regard, the two-step surface modification of polyamide meshes and nonwoven fabrics for oil/water separation investigated by Zhao et al. attracted extreme attention from the scientific community [45]. Their methodology was based on pre-etching the polyamide surface using plasma treatment and coating of a pre-etched surface by eco-friendly polydopamine (PDA)/cellulose, as shown in Figure 5. The pre-etching increased the surface roughness, which further improved the underwater superoleophobicity of the coating. Therefore, the modified polyamide separated various oil/water mixtures and showed a higher intrusion pressure than the original sample and the samples, which were only etched or coated. The grooves on the surface that resulted from the pre-etching prevented the coating from peeling off. In durability tests, after six repeated uses, the modified nonwoven sample lost its underwater oleophobicity due to severe oil fouling, coming to a complete failure in oil/water separation. After 19 cycles, the modified mesh could still separate a certain amount of oil/water but showed reduced intrusion pressure because of slight oil contamination. Filters with different structures, like meshes with one layer of pores and nonwoven fabrics with complex three-dimensional pores, had different oil fouling levels that affected oil/water separation. The recoverability of filters from oil contamination should be considered for practical applications.

**Figure 5.** A fabrication of hydrophilic polyamide fabrics for produced water treatment.

Anupriyanka et al. presented the efficient and cost-effective sustainable oil/water separation by fabricating innovative oil-recovery materials (fluorine and silane-free) using an environment-friendly route [46]. Researchers produced superhydrophobic PET (polyethylene terephthalate) fabric by a DC glow discharge oxygen plasma treatment followed by the incorporation of the superhydrophobic agents. The WCA of this fabric was 163◦. The prepared fabric possesses excellent repellency to water while absorbing oil and hydrocarbons. Indeed, PET is a significantly cheaper material, and even disposed fabrics can be used as a feedstock to fabricate such adsorbents.

A superhydrophobic cotton nonwoven fabric aimed at oil-water separation was prepared by a two-step strategy based on (1) atmospheric-pressure N2/O2 plasma treatment and (2) graft polymerization of siloxane [47]. Different process conditions (plasma power, composition of the plasma gas) influenced the contact angle, stability, surface morphology of the hydrophobic coating, and the growth of silica nanoparticles. The water contact angle of treated nonwoven was up to 155◦. The superhydrophobic nonwovens treatment showed excellent stability toward severe acid and alkaline conditions and subsequently the resulting fabrics may be used under harsh environmental conditions. The modified cotton nonwoven exhibited an oil/water separation efficiency higher than 97%, and the result was repeated at least ten times for each sample.

#### *4.2. Direct Plasma Treatment of Water*

Atmospheric pressure discharges can directly decompose the organic pollutants. The decomposition of model organic pollutants: chloroform, benzene, and methanol through the gliding arc (GA) plasma technology was reviewed by Gong et al. [48]. The conventional GA reactor is depicted in Figure 6. It is equipped with an AC power supply and flat

electrodes contributing to the dispersion of heat and restriction of the gas flow. When a high voltage is applied, the gas flow initiates the GA plasma.

**Figure 6.** A typical gliding arc setup scheme.

GA plasma devices can be directly used to treat the organic waste gas derived from industrial emissions and be coupled with other technologies or devices, such as catalysts, washing towers, combustion chambers, and activated carbon. GA plasma has extensive adaptability for the types and concentrations of various pollutants, including linear hydrocarbons, aromatic hydrocarbons, halogenated compounds, ethers, and alcohols. Nevertheless, although the gliding arc requires less energy for organic waste gas treatment than other nonthermal technologies, it is still a little higher considering the practical engineering applications. It is worth noting that the plasma treatment can be used to treat produced water and emulsions of water in heavy oils, allowing the separation of oil from water and improving oil properties, as shown elsewhere [49].

Plasma discharges were also used to soften produced water, and the investigation of such a process was funded by US Department of Energy (US DOE) [50]. It was shown that plasma treatment enables the decrease of bicarbonate ions down to 100 ppm.

US DOE has also investigated novel processes for water softening [51]. Specifically, in the plasma discharge project, as mentioned above, the research was focused on reducing the "temporary hardness" (caused by the presence of calcium and magnesium bicarbonate). In this DOE-supported research by Drexel University, plasma arc discharges were used to dissociate and remove bicarbonate ions. In turn, this effect helps prevent fouling from the calcium carbonate scale providing a more prudent strategy for fouling prevention.

Later, Cho et al. investigated stretched arc plasma to increase the volume of produced water treated by plasma, increasing the practical efficiency of the process [52]. Stretching of an arc discharge in produced water was accomplished using a ground electrode and two high-voltage electrodes: one positioned close to the ground electrode, and the other placed farther away from the ground (Figure 7). The contact between the arc and water significantly increased, resulting in twice more efficient removal of bicarbonate ions from produced water when compared to plasma with and without stretching. The usage of plasma treatment for removing bicarbonate ions was further studied by Kim [53]. Although the results were quite promising, the energy cost for this method to treat the produced water was too expensive, and there were no recent publications after 2016.

**Figure 7.** The scheme of a stretch arc plasma applied for produced water treatment.

Choi et al. analyzed the decomposition of water-insoluble 1-decanol by a water plasma system operated at atmospheric pressure [54]. 1-decanol was dispersed in water by a surfactant generating an oil-in-water emulsion. The emulsion was used as the water plasma jet's feeding liquid and plasma-forming gas (Figure 8). A high decomposition rate of over 99.9999% was achieved by converting 1-decanol emulsion into H2, CO, CO2, CH4, condensed liquid, and solid-state carbon despite relatively low input power (<1 kW). The main organic substance in the treated liquid of 1-decanol emulsion was methanediol produced by the hydration of formaldehyde. The authors concluded that the gas conversion rate of carbon in 1-decanol and the removal rate of total organic carbon concentration were increased by increasing the arc current due to enhanced O radicals in the high temperature of the water plasma jet.

**Figure 8.** Atmospheric plasma jet setup for water purification.

The decomposition of organic pollutants (dyes as a model substance) in water under plasma micro-discharges was studied by Wright [55]. The low-temperature plasma contacting with aqueous solutions, including organic dyes, was analyzed. Using atmospheric plasma for cleaning water was demonstrated by relatively good energy efficiency: 1.3–12 g of removed pollutants per 1 kWh.

Millie et al. and Lemont et al. designed a plasma tool ELIPSE to decompose organic impurities in water using an atmospheric plasma torch [56,57]. The ELIPSE process is a novel technology of organic liquid destruction involving a thermal plasma working under a water column, ensuring the cooling, filtration, and scrubbing of the gases coming from the degradation (Figure 9). The ELIPSE demonstrated the ability to destroy the pure organic liquids and then eliminate the organic compounds remaining in the aqueous solution through plasma's thermal or radiative properties. Preliminary tests have shown how efficient the process is to destroy the organic liquids when they are directly fed in the plasma. By applying plasma for 60 min, the concentration of organic contaminants decreased from 1200 to 400 ppm.

**Figure 9.** Atmospheric plasma torch employed for the destruction of organic pollutions in water.

Yu et al. studied the decomposition of different polycyclic aromatic hydrocarbons (acenaphthene, fluorene, anthracene, and pyrene) by DC gliding arc discharge. The results indicated that the highest destruction rate was achieved with oxygen as a carrier gas and the external resistance of 50 kΩ independently of the type of hydrocarbons. Furthermore, experimental results suggest that the destruction energy efficiency of gliding arc plasma can be improved by treating higher concentration pollutants [58].

The efficient decomposition of toluene in a gliding arc was shown by Du et al. [59]. The toluene removal efficiency increased with the inlet gas temperature, while the presence of water vapors accelerated the toluene decomposition in the plasma. The energy efficiency was 29.46 g per 1 kWh at a relative humidity of 50% and a specific energy input of 0.26 kWh/m3, which is higher than other types of nonthermal plasmas. The primary gas-phase decomposition products were determined using FT-IR analysis of the gas components: CO, CO2, H2O, and NO2. Some small deposits of benzaldehyde, benzoic acid, quinine, and nitrophenol were found in the reactor. The authors concluded that these are the minor products from the reaction of toluene with radicals.

#### **5. Combined Techniques for Efficient Water Treatment**

The increase of the water quality purification related to the final use of the treated water might be extremely challenging to achieve in an economically viable scenario if a single technology is used (e.g., only sorption or filtration). However, the researchers recently implemented an exciting approach based on a combination of various techniques: plasma and sorption of membranes and photocatalysis. These combination approaches allow separating or adsorb parts of the pollutants and decomposing the rest of the contaminants. For example, Gushchin et al. developed a combined process involving the sorption of oil (C22H38) on a sorbent (diatomite) followed by regeneration of the sorbent by plasmaoxidative destruction in an atmospheric dielectric barrier discharge (DBD) [60]. The reactor depicted in Figure 10 was composed of a glass tube diameter of 22 mm, which served as a dielectric barrier; the outer electrode was made from the aluminum foil and located outside the glass tube, and the internal electrode, a 15 mm diameter aluminum rod. The sorbent was placed inside the reactor along the length of the discharge zone.

**Figure 10.** A scheme of combined DBD plasma-sorption technology used for water treatment.

The energy efficiency of the decomposition was 0.169 molecules of oil per 100 eV of input energy. It was shown that the complete deterioration of oil on the diatomite surface reached after 5 min DBD treatment, while the decomposition products were water-soluble and non-toxic carboxylic acids, aldehydes, and CO2. The complete removal of acids and aldehydes requires a time of about 40 min [60].

Another emerging sustainable approach for decontamination of water is a photocatalytic membrane treatment. Photocatalytic membrane reactor (PMR) is an emerging green technology for removing organic pollutants, photoreduction of heavy metals, photoinactivation of bacteria, and resource recovery.

The combination of photocatalysis and membrane separation improves the removal of contaminants and alleviates membrane fouling. However, the turbidity and color of produced water reduce the efficient light transmission. Efficient pretreatment improves the resiliency for produced water treatment and minerals recovery using PMR processes [61].

Recently, Butman and co-workers compared photocatalytic, plasma, and combined plasma–photocatalytic water treatment processes using a model pollutant, Rhodamine B dye, solutions with a concentration of 40 mg/L [62]. The TiO2-pillared montmorillonite was used as a photocatalyst, while plasma was ignited in dielectric barrier discharge (DBD) plasma. By using the combined DBD and photocatalysis process, a significant increase of degradation efficiency has been observed: combined approach (100%, 8 sec), plasmolysis (94%), and UV photolysis (92%, 100 min of UV irradiation). In contrast to photolysis, destructive processes are more profound and lead to the formation of simple organic compounds such as carboxylic acids. The plasma–catalytic method enhances by 20% the energetic efficiency of the destruction of Rhodamine B compared to simple DBD plasma. The efficiency of dye destruction with the plasma–catalytic method increases with specific surface area and total pore volume, and the size of the TiO2 crystallites. This approach may bring significant added value for produced water treatment if the methodology is further optimized.

Implementing the oxidation to the produced water treatment system may provide additional advantages in terms of continuous process reliability. For example, the advanced oxidation processes based on reactions with ozone and Fenton's reagent efficiently clean the polluted water streams, as revealed by Simões et al. in their bibliometric study [63]. They indicated the increase of publications dedicated to this approach in the last few years, although the first work appeared only in 1995. However, the selected works revealed the limitation of these processes due to the high costs with energy consumption.

Finally, the problem of low water flux attributed to commercial ceramic membranes applied in the treatment of produced water was also studied using complex combined technologies [64]. For minimizing this problem, the titanium dioxide (TiO2) nanocomposites, synthesized via a sol-gel method, were deposited on the active layer of the hydrolyzed bentonite membrane [65]. The grafting time of TiO2 nanocomposite improved the performance of the coated bentonite membranes. The pure waterpermeability performance showed an increment from 262.3 L h−1 m−2 bar−1 (pristine bentonite membrane) to 337.1 L h−1m−2bar−1 (bentonite membrane with TiO2 grafted for 30 min) and 438.3 L h−1m−2 bar−1 (TiO2 grafted for 60 min). The oil rejection performance also revealed an increase in the oil rejection performance from 95 to 99%. These findings can be an excellent example to further investigate and exploit the advantages of modified ceramic membranes in produced water treatment.

#### **6. Discussion**

The produced water treatment solutions based on plasma-aided technologies and nanomembranes show serious potential for industrial applications. However, it is obvious that not every kind of treatment would reach the stage of commercialization. To sum up the advantages and limitations of all of the reviewed technologies, the oil/water separation and desalination technologies are listed in Tables 1 and 2, respectively.

#### **Table 1.** Summary of the oil/water separation technologies.

**Table 2.** Summary of the desalination technologies.

Proposed technologies can provide the oil/water separation process with high efficiency. The applicability of the proposed methods would greatly depend on the pretreatment of an effluent and the remaining oil contaminants. In this review, the water/oil separation by most of the membranes allowed to achieve the oil rejection rate ranging from 92% to 99.8%. Hence, knowing the ROPME protocol standard value of 15 ppm (oil in water), it is quite easy to estimate the range of maximum oil concentrations passing

membranes. The very well-known formula for oil rejection rate available elsewhere [68] and is presented below:

$$\text{Rejection Rate} = \left(1 - \frac{C\_{\text{permeate}}}{C\_{fecd}}\right) \times 100\%$$

*Cpermeate* and *Cfeed* are the oil in water concentrations in the permeate and the feed, respectively. Hence, for the technologies exhibiting a rejection rate of 92%, the maximum concentration in the feed can be 187 ppm to ensure the oil concentration in the permeate is below 15 ppm. For the technologies exhibiting a rejection rate above 99%, the *Cfeed* may exceed 1500 ppm. Therefore, the majority of reviewed technologies may satisfy the requirements for the effluent discharge composition according to ROPME protocol.