**Preparation of Biomass Activated Carbon Supported Nanoscale Zero-Valent Iron (Nzvi) and Its Application in Decolorization of Methyl Orange from Aqueous Solution**

#### **Bo Zhang \* and Daping Wang**

School of Metallurgical and Material Engineering, Hunan University of Technology, Taishan Road 88, Zhuzhou 412007, China

**\*** Correspondence: 13747@hut.edu.cn

Received: 12 July 2019; Accepted: 8 August 2019; Published: 12 August 2019

**Abstract:** The nanoscale zero-valent iron (nZVI) has great potential to degrade organic polluted wastewater. In this study, the nZVI particles were obtained by the pulse electrodeposition and were loaded on the biomass activated carbon (BC) for synthesizing the composite material of BC-nZVI. The composite material was characterized by SEM-EDS and XRD and was also used for the decolorization of methyl orange (MO) test. The results showed that the 97.94% removal percentage demonstrated its promise in the remediation of dye wastewater for 60 min. The rate of MO matched well with the pseudo-second-order model, and the rate-limiting step may be a chemical sorption between the MO and BC-nZVI. The removal percentage of MO can be effectively improved with higher temperature, larger BC-nZVI dosage, and lower initial concentration of MO at the pH of 7 condition.

**Keywords:** biomass activated carbon; methyl orange; pulse electrodeposition; zero valent iron nanoparticles

#### **1. Introduction**

Dye has become a widespread environmental pollution problem because of its wide application in industry such as paper, textiles, plastic, and leather tanning industries in the past several decades [1,2]. Over 100,000 commercial dyes are associated with an annual production rate of over 800,000 tons [3]. Large amounts of dye-containing effluents pose great challenges to the environment because of their strong color, complex structure, stability, and low biodegradability [4]. Therefore, it is of vital importance to remove dyes from wastewater to protect the aquatic life and alleviate the crisis of water pollution.

Many methods such as adsorption [5], reduction [6], advanced oxidation processes [7], coagulation [8], membrane separation [9], and biological methods [10] are used to remove dyes from wastewater. In recent years, nanoscale zero-valent iron (nZVI) particle has been extensively used as a new tool for the treatment of wastewater contaminated with various pollutants as a result of its small particle size, large specific surface area, and high reactivity [11]. Unfortunately, there are still some technical challenges in the use of nZVI. On the one hand, because of interparticle Van der Waals and magnetic interactions, nZVI particles are prone to agglomeration, resulting in the significant decrease of their dispersibility [12]. On the other hand, nanoparticles tend to be oxidized, and the formation of oxide layers easily block the serviceable active surface sites, which finally diminishes the reactivity [13]. To get rid of these shortcomings, supporting nZVI particles is an option. For the past few years, most of previous studies have loaded nZVI particles on some porous materials as a

carrier, such as kaolinite [14], bentonite [15], resin [16], activated carbon [17], mesoporous carbon [18], mesoporous silica [19], titanium oxide [20], pumice [1,21], and grapheme [22], through the liquid phase reduction method. Nevertheless, there still has some problems associated with method, such as low production efficiency, high production cost, and large amounts of hydrogen the generation during the preparation process [23,24]. All of these problems lead to the limitations of large-scale application.

In this study, the compound material was prepared with two steps. First, the nZVI particles were directly obtained from steel scrap using the pulse electrodeposition. Second, the nZVI particles were quickly supported to the biomass activated carbon (BC) under the mechanical agitation condition. The obtained BC-nZVI material was characterized and its decontamination abilities were tested by the removal of methyl orange (MO). In addition, the effects of dosage, initial pH, initial concentration, and temperature on the removal percentage of MO were investigated in conjunction with the analyses of mechanism and kinetics. It is hoped that this study can provide a new preparation technique for nZVI composites for wastewater treatment.

#### **2. Materials and Methods**

#### *2.1. Materials and Chemicals*

The biomass activated carbon was carbonized from coconut shell with particle size of 2~4 mm, filling density of 0.5~0.55 g/mL, PH value of 6.5~7.5, and iodine adsorption value of 850~1000 mg/g. This was supplied by Lu-yuan Co. Ltd., Mianyang, China. Ferrous sulfate septihydrate (FeSO4·7H2O, ≥98.5%), methyl orange (MO, 99%), sodium hydroxide (NaOH, 98%), hydrochloric acid (HCl, 36%), sodium dodecyl benzene sulfonate (DBS, 99%), thiourea (≥98%), and absolute ethanol (99%) were obtained from Sinopharm Chemical Reagent Co. Ltd., Shanghai, China. All the chemicals were analytical reagent grade and deionized water from a Liceng UPA-L system (18.2 MΩ/cm, 25 ◦C) was used for all experiments.

#### *2.2. Preparation of nZVI and Composite with Biomass Activated Carbon*

FeSO4·7H2O was dissolved in deionized water to prepare an electrolyte with Fe<sup>2</sup><sup>+</sup> of 30 g/L. The additive (thiourea, 0.5 g/L and DBS, 1.0 g/L) was dispersed and emulsified in deionized water in an ultrasonic cleaner (VGT SONIC-L30 300 w, 28 kHz) and added to the electrolyte. The electrolysis was carried out at 50~60 ◦C and the anode plate for electrolysis was a waste carbon steel (Q235) plate, which was used after surface treated. The cathode plate for electrolysis was a stainless-steel box and several ultrasonic vibrators were added in the box. The ultrasonic power was 28 KHz. The cathode and the anode were separated by a filter membrane and the structure of the electrolytic cell is shown in Figure 1. After 10 h of electrolysis, the electrolyte near the cathode plate was rapidly pumped out and filtration, and the leached residue was washed with deionized water and absolute ethanol for three times, respectively. Subsequently, DBS and absolute ethanol were intermixed with leached residue and dispersed in a mechanical disperser (FS400D, 2000 rpm, Qiwei instrument Co. Ltd., Hangzhou, China). Finally, an emulsion of nZVI was prepared after 12 h of dispersion under nitrogen protective.

The electrolytic process occurs at the cathode and anode as follows [25]:

Anode:

$$\mathbf{F}\mathbf{e} - \mathbf{2}\mathbf{e} = \mathbf{F}\mathbf{e}^{2+}\tag{1}$$

Cathode:

$$\text{Fe}^{2+} + 2\text{e} = \text{Fe} \tag{2}$$

The iron particles generated on the cathode plate were quickly stripped and emulsified in the electrolyte under the action of cavitation effect of ultrasonic oscillation and combined with dispersant. The polymer molecular chain of dispersant was negatively charged, which can form a brush surface layer to hinder the aggregation. Under the action of dispersant, iron particles were difficult to agglomerate and oxidize.

**Figure 1.** Electrolytic cell for preparation of nanoscale zero-valent iron (nZVI).

The emulsion of nZVI was transferred to a conical flask (1000 mL) for 100 mL, and 200 mL deionized water and 200 g biomass activated carbon were added into conical flask. Then, the conical flask was sealed with a sealing plug and shocked on a horizontal vibrator (HY-5A) with a frequency of 240 rpm. After 10 h of shocking, the mixture was cleaned three times by absolute ethanol to remove the excess emulsion. Finally, the freshly composite material BC-nZVI was dried at 30 ◦C in the vacuum drying oven for 12 h and kept in a nitrogen atmosphere prior to use.

#### *2.3. Characterization and Analytic Methods*

The surface morphology images of BC-nZVI were obtained with a scanning electron microscope (SEM) (SIGMA 300; Carl Zeiss AG, Oberkochen, Hallbergmoos, Germany) operating at 30 kV. The crystal structure and composition phase was analyzed by X-ray diffraction (XRD) (Philips-X'Pert Pro MPD, Malvern Panalytical, Almelo, Holland, The Netherlands). The sample of BC-nZVI was dissolved by HCl and the total iron ions in acid solution were analyzed by inductively Coupled plasma spectrometer (ICP-MS 8000; Perkinelmer, Waltham, MA, USA) to determine the actual load of nZVI on BC. The concentration of MO solution was measured using a UV-Spectrophotometer (760CRT, Shanghai precision scientific instrument co., LTD, Shanghai, China) at λmax = 464 nm [26].

#### *2.4. Batch Experiments*

The removal percentage of MO by BC-nZVI was evaluated by batch experiments. The degradation test for MO was carried out in a 250 mL conical flask and the volume of the experimental solution was 150 mL. The flask was placed in the horizontal vibrator and wobbled at 240 rpm after the BC-nZVI was added. The initial pH value of solution was adjusted by 0.01 mol/L HCl and 0.01 mol/L NaOH. The five major factors (MO initial concentration (20–200 mg/L), weight of BC-nZVI dosage (0.25–0.75 mg/L), reaction temperature (20◦C~80◦C), pH (3.0~10.0), and interaction time (60–120 min)) were considered for optimizing MO degradation by BC-nZVI. The samples (4 mL) were collected within a specified time and the concentration of MO was estimated using UV-Spectrophotometer after centrifuged. In order to ensure the quality of the data, all experiments were conducted in three copies and the average value was reported.

The MO removal percentage was calculated by Equation (3):

$$R(\%) = \frac{(\mathbb{C}\_0 - \mathbb{C}\_c) \times 100}{\mathbb{C}\_0} \tag{3}$$

The equilibrium removal capacity of MO (*q*<sup>e</sup> (mg/g)) was calculated by Equation (4):

$$q\_{\varepsilon} = \frac{(\mathbb{C}\_0 - \mathbb{C}\_{\mathbf{e}}) \times V}{W} \tag{4}$$

where *C*<sup>0</sup> (mg/L) is initial concentration of MO, *C*<sup>e</sup> (mg/L) is the equilibrium concentrations of MO, *V* (L) is the experimental solution volume, and W (g) is dry weight of nZVI in BC-nZVI used.

#### **3. Results and Discussions**

#### *3.1. Characterization of nZVI and BC-nZVI*

The morphologies of nZVI and BC-nZVI were determined using SEM and presented in Figure 2. Figure 2a shows that the nZVI nanoparticles were substantially nearly smooth and spherical with sizes ranging from 40 to 80 nm. The nanoparticles were slightly agglomerated, and this may have occurred during the detection process when the emulsion of nZVI was placed on the observation table and rapidly dried by washing ear ball. The surface of BC presents a porous structure, which provides a good platform for loading the nZVI as shown in Figure 2b. The nanoparticles were substantially evenly distributed on the biomass activated carbon surface and pores under mechanical loading condition.

**Figure 2.** SEM images of (**a**) nZVI, (**b**) biomass activated carbon (BC)-nZVI.

The XRD patterns of nZVI, BC, and BC-nZVI were obtained and presented in Figure 3. No characteristic diffraction peaks of Fe were observed because of its weak crystallization and the inclusion of DBS. The characteristic peaks have several small peaks at 2θ = 17.41◦, 19.24◦, and 3.84◦ on nZVI wave lines, which represent iron oxide [22], indicating a certain oxidation in the nanoparticles. In the characteristic peak of the BC-nZVI wave lines, there were obvious carbon peaks at 2θ = 26.60◦, 43.45◦, 54.79◦ [23] and the iron oxide peak disappears, indicating that the loading was beneficial to prevent the oxidation of the nZVI.

**Figure 3.** XRD pattern of nZVI, BC, and BC-nZVI.

#### *3.2. Removal E*ffi*ciencies of MO*

The removal percentage of MO was investigated using BC, nZVI, and BC-nZVI, respectively, as shown in Figure 4. Obviously, the total removal percentage of BC-nZVI and nZVI are significantly higher than that of BC as a result of the contribution of nZVI. But for nZVI and BC-nZVI, the different tendencies were presented. In the early stage (within 20 min), nZVI holds the higher removal percentage than that of BC-nZVI. The reason behind this fact is that the nZVI was in direct contact with the aqueous solution and the reaction was fast, while the removal percentage of BC-nZVI is lower because the pore structure of BC prevents the contact bewteen nZVI and MO. However, the opposite results can be found after 20 min, and the BC-nZVI shows the higher removal percentage than that of nZVI. This is due to the fact that the sole nZVI particle was easily oxidized than BC-nZVI composite material, resulting in the decreased reaction activity.

**Figure 4.** Comparative degradation of MO using different materials. The dosage of BC, nZVI, and BC-nZVI were 3 g/L, 0.5 g/L, and 0.5 g/L (the dry weight of nZVI in BC-nZVI used) with an initial MO concentration of 200 mg/L, temperature of 25 ◦C, and original pH.

Figure 5 showed the UV–vis absorption spectra of the MO before and after the addition of BC-nZVI. For the original MO, there were two main absorption bands at 464 and 270 nm, which were attributed to a conjugate structure formed by the azo band under the effects of the electronic-donation of dimethylamino group and the π–π \* transition of aromatic rings [27]. Obviously, the band intensity of 464 and 270 nm decreased with reaction time, and almost disappeared after reaction 50 min. It proves the degradation of MO through the cracking of azo bonds. The XRD pattern of BC-nZVI after interaction with MO was represented in Figure 6. The characteristic peaks of 27.87◦, 50.17◦ present the iron (III) oxide Fe2O3. With the increase of reaction time, the intensity of Fe2O3 characteristic peaks increased. The radicals of H were generated by the reaction nZVI nanoparticles and H2O or hydrogen ion [28,29], which caused the azo bond to open and consequently the absorption bands at 464 nm and 270 nm vanished.

**Figure 5.** UV–vis patterns of degradation of MO.

**Figure 6.** XRD pattern of BC-nZVI after reaction.

The maximum removal capacities of BC-nZVI in this study were compared with those of other absorbents prepared by liquid phase reduction method, as shown in Table 1. Clearly, the present adsorbent (BC-nZVI) shows the same level of removal capacity as other efficient absorbents, such as B-nZVI, HJ-NZVI, nZVI/BC, and Bio-nZVI. However, it should be emphasized that the BC-nZVI in this study was synthesized by pulse electrodeposition and mechanical agitation method, and this method is easier and more cost-effective than the liquid phase reduction method.

**Table 1.** Comparison of MO removal with different absorbents reported in literature.


B: synthesized bentonite; HJ: Hangjin clay; BC: biochar; Bio: Biogenic; B \*: Functional clay; Or is 6.17; NA: not available.

#### **4. Impact of Operational Parameters**

#### *4.1. E*ff*ect of Dosage*

As illustrated in Figure 7a, the decolorization kinetics of MO was dependent on the BC-nZVI dosage. The larger the BC-nZVI dosage, the higher the removal percentage. The concentration of MO decreases dramatically in the initial 20 min at the highest BC-nZVI dosage of 0.75 g/L, while the removal percentage at 0.25 g/L was sluggish. Moreover, for the dosages of 0.25, 0.5, and 0.75 g/L, the maximum removal percentages of MO were 68.96%, 94.41%, and 97.43%, and the removal capacities of BC-nZVI were 344.1 mg/g, 300.8 mg/g, and 181.2 mg/g, respectively. Due to the loading effect of BC, the oxidation of nZVI was slowed down, resulting in the highest removal percentage.

When the amount of BC-nZVI dosage was insufficient, the slower kinetics and lower removal percentage were observed for MO removal, but the removal capacity of BC-nZVI was higher [32]. At a sufficient amount of BC-nZVI, enough nZVI activity sites was provided to react with MO in the initial stage, so the reaction kinetics were faster, but the removal capacity of BC-nZVI was lower.

**Figure 7.** (**a**) Effect of dosage on degradation of MO, with an initial MO concentration of 200 mg/L, temperature of 25 ◦C, and original pH; (**b**) Effect of the pH values on degradation of MO, in the context of 0.5 g/L BC-nZVI with an initial MO concentration of 50 mg/L; (**c**) Effect of initial concentration on degradation of MO, at the original pH using 0.5 g/L BC-nZVI; (**d**) Effect of temperature on degradation of MO, in the context of 0.5 g/L BC-nZVI with an initial MO concentration of 50 mg/L, temperature of 25 ◦C, and the original pH.

#### *4.2. E*ff*ect of pH*

The effects of initial pH on MO removal were illustrated in Figure 7b. With the increase of pH from 3.0 to 10.0, the removal percentage of MO was declined from 97.75% to 89.13%, and decreased from 76.01% to 21.08% in the first 10 min. It suggests that the lower pH value was beneficial to MO reduction by BC-nZVI. The possible explanation is that the more H<sup>+</sup> was released at lower pH values, which could accelerate the corrosion of nZVI particles, and eliminate ferrous hydroxide on the surface of nZVI particles to generate fresh active sites. The removal capacities of BC-nZVI were 193.5 mg/g, 195.4 mg/g, and 180.1 mg/g as the pH were 3.0, 7.0, and 10.0, respectively. The alkaline pH reduced the removal capacity of BC-nZVI because OH− would significantly enhance the formation of the iron hydroxide, which formed a surface layer on the nZVI particles and inhibited further reactions [20].

#### *4.3. E*ff*ect of Initial Concentration*

The effects of initial MO concentrations ranging from 20 to 200 mg/L on removal percentage was shown in Figure 7c. The maximum decolorization efficiencies were 98.8%, 98.7%, 96.7%, and 93.3%, and the decolorization efficiencies in first 20 min were 77.6%, 82.3%, 75.6%, and 59.2% at initial concentrations of 20 mg/L, 50 mg/L, 100 mg/L, and 200 mg/L, respectively. The removal percentage with the initial concentration of 200 mg/L was significantly lower than that of the other lower initial concentration. It may be caused by that the highly initial MO concentration leads to a competition effect among the MO molecules and a decline in decolorization efficiency [33,34]. However, with the

increase of initial concentration, the removal capacity of nZVI was obviously improved. At a higher initial concentration, the removal capacity of nZVI was significant higher, at 327.7 mg/g.

#### *4.4. E*ff*ect of Temperature*

As shown in Figure 7d, the removal percentage of MO increased with increasing temperature. The final removal percentages of MO were 77.6%, 93.1%, and 96.2%, respectively, for 20, 40, and 60 ◦C. Similar results were reported previously [35–37]. There may be two possible explanations for this phenomenon: (1) The mobility of MO was increased by higher temperature, and (2) the activation energy of decolorizing reaction was increased as the temperature rose.

#### *4.5. Kinetics of Degradation of MO*

In order to investigate the mechanism of degradation, the pseudo-first-order kinetics model (PFO) (Equation (5)) and the pseudo-second-order model (PSO) (Equation (6)) were generally used to test the degradation of MO using BC-nZVI, which can be expressed as the following equation [13,29]:

$$\ln(q\_{\varepsilon} - q\_{t}) = \ln q\_{\varepsilon} - k\_{1}t \tag{5}$$

$$\mathbf{t}/q\_1 = \mathbf{1}/(k\_2q\_c^2) + \mathbf{t}/q\_6 \tag{6}$$

where *qe* and *q*t (mg/g) are the amount of MO adsorbed per gram of BC-nZVI at equilibrium and at time *<sup>t</sup>*, respectively, while *<sup>k</sup>*<sup>1</sup> (min<sup>−</sup>1) and *<sup>k</sup>*<sup>2</sup> (g/(mg·min)) are the PFO and PSO rate constants, respectively.

The effect of initial MO concentrations on the removal efficiency was calculated with these two kinetic models, and the fitting of the kinetics data was shown in Figure 8 and Table 2.

**Figure 8.** Kinetic modeling of MO adsorption using pseudo-first-order (PFO) (**a**) and pseudo-secondo-rder (PSO) model (**b**).


**Table 2.** Kinetic parameters for MO removal.

Significantly, the pseudo-second-order model matched better with the data (R<sup>2</sup> = 0.988) than the pseudo-first-order model (R2 = 0.823), suggesting that the chemical sorption may be the main rate-limiting step between the MO and BC-nZVI.

The rate constants for a pseudo second-order reaction were 0.0259 g/(mg·min) for 20 mg/L, 0.0048 g/(mg·min) for 100 mg/L, and 0.0025 g/(mg·min) for 200 mg/L, respectively. The result indicates that the degradation of MO occurs in the interface of BC-nZVI [20,22], hence the rate of degradation was closely linked to the initial concentration of MO and the active surface sites of BC-nZVI, as discussed in the previous section. It should be emphasized that the BC-nZVI materials in our study were prepared by the mechanical agitation method and most of nZVI particles are mainly distributed in the outer layer of BC. It means the MO could easily contact with the nZVI particles, so the chemical sorption is the main rate-limiting step compared to the mass transfer process. However, some other X-nZVI materials synthesized by liquid phase reduction method agreed with the pseudo-first-order model [14–18]. Here, X presents the carrier, such as clay [15], activated carbon [17], and mesoporous carbon [18]. This is due to that the nZVI particles can be uniform distributed in both the outer layer and internal space of carriers (X) using the liquid phase reduction method. Therefore, the mass transfer may be the main limiting step under this condition.

#### **5. Conclusions**

The present study demonstrated that BC-nZVI particles can be used to efficiently degrade MO in aqueous solution by cleaving the azo linkages. SEM, XRD, UV–vis, and batch experiments confirmed the characteristics of this composite: (1) The nZVI nanoparticles prepare dby means of pulse electrodeposition were substantially nearly smooth and spherical with sizes ranging from 40 to 80 nm, and successfully loaded on the surface and inside the pores of BC; (2) BC as a dispersant and stabilizer, which reduced the extent of aggregation of nZVI and therefore enhanced reactivity; (3) compared with other adsorbents prepared by liquid phase reduction method, BC-nZVI reflected with the same maximum removal rate; and (4) batch experiments show that various parameters such as dosage, pH, initial concentration of MO, and temperature did affect the removal of MO. In addition, the BC-nZVI in this study was synthesized by pulse electrodeposition and mechanical agitation method, and this method is easier and more cost-effective than the liquid phase reduction method, and the 97.94% removal percentage demonstrated its promise in the remediation of dye wastewater.

**Author Contributions:** B.Z. designed and carried out this research. D.W. prepared and analyzed the data. B.Z. wrote this paper. All authors have approved the manuscript.

**Funding:** This study was funded by the Specialized Research Fund for the National Natural Science Foundation of China (No. 51504090), and the Natural Science Foundation of Hunan Province, China (No. 2019JJ60062).

**Acknowledgments:** The authors are grateful to the three anonymous reviewers for their insightful and constructive comments, which greatly improved the quality of the paper.

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

#### **References**


© 2019 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 (http://creativecommons.org/licenses/by/4.0/).

## *Article* **Comparative Study of Four TiO2-Based Photocatalysts to Degrade 2,4-D in a Semi-Passive System**

**Gisoo Heydari 1,\*, Jordan Hollman 1, Gopal Achari <sup>1</sup> and Cooper H. Langford <sup>2</sup>**


Received: 14 February 2019; Accepted: 21 March 2019; Published: 26 March 2019

**Abstract:** In this study, the relative efficiency of four forms of supported titanium dioxide (TiO2) as a photocatalyst to degrade 2,4-dichlorophenoxyacetic acid (2,4-D) in Killex®, a commercially available herbicide was studied. Coated glass spheres, anodized plate, anodized mesh, and electro-photocatalysis using the anodized mesh were evaluated under an ultraviolet – light-emitting diode (UV-LED) light source at λ = 365 nm in a semi-passive mode. Energy consumption of the system was used to compare the efficiency of the photocatalysts. The results showed both photospheres and mesh consumed approximately 80 J/cm<sup>3</sup> energy followed by electro-photocatalysis (112.2 J/cm3), and the anodized plate (114.5 J/cm3). Although electro-photocatalysis showed the fastest degradation rate (K = 5.04 mg L−<sup>1</sup> h<sup>−</sup>1), its energy consumption was at the same level as the anodized plate with a lower degradation rate constant of 3.07 mg L−<sup>1</sup> h<sup>−</sup>1. The results demonstrated that three-dimensional nanotubes of TiO2 surrounding the mesh provide superior degradation compared to one-dimensional arrays on the planar surface of the anodized plate. With limited broad-scale comparative studies between varieties of different TiO2 supports, this study provides a comparative analysis of relative degradation efficiencies between the four photocatalytic configurations.

**Keywords:** photocatalysis; semi-passive; anodization; buoyant catalyst; 2,4-D; LED; mesh

#### **1. Introduction**

As a clean technology, photocatalysis holds a lot of promise. Research interest on photocatalysis has significantly increased since the initial publication on titanium dioxide (TiO2) photocatalysis in 1971 [1]. As an advanced oxidation process, it utilizes a semiconductor material as a reusable catalyst that is capable of mineralizing organic contaminants with light energy as the main input. This makes photocatalysis an appealing treatment option for both industrial and municipal wastewaters [2–7].

The most versatile semiconductor that has been used as a photocatalyst is titanium dioxide (TiO2), which owes its popularity to its low cost, non-toxicity, and photo-stability as well as its unique non-selectivity characteristic for oxidation reactions. It is also chemically and mechanically robust, that makes it an ideal photocatalyst candidate for various reaction media [8–11]. With a high oxidation potential (approximately +3.2 V), high energy ultraviolet (UV) radiation (λ ≤ 387 nm) is required to excite electrons in its valence band and move them to the conduction band to initiate the photocatalytic reaction [4,9,12]. Although other semiconductors such as zinc oxide (ZnO), cadmium sulfide (CdS) [13], tungsten oxides (WOx ≤ 3) [14,15] and other tungstate species [16] as well as their combinations [2,4,17] have been tested for photocatalysis, TiO2 is still a broadly studied photocatalyst and is utilized as a model photocatalyst in this study.

TiO2 can be used in various forms such as powder, being immobilized (on a surface such as a sphere or a plate), or one of many possible variations of nanoparticle materials. Each method of application presents unique advantages and challenges. While powdered TiO2 has been extensively researched, it has a significant disadvantage in environmental applications. As a photoactive material, the powdered TiO2 must be separated from water prior to its release that necessitates an additional treatment step [3,18,19]. Although immobilized TiO2 avoids the complications of a secondary separation step, its available surface for reactions is reduced and consequently has lower reactivity, leading to lower degradation rates in some cases [20–23].

Supported nanostructures are a viable option for the application of TiO2 in water treatment. They can provide higher reaction rates than simple immobilized TiO2 plates while having the advantage of being supported, making an extra separation step unnecessary. Many studies on TiO2 nanotubes boast high electron mobility, lower recombination rates, high surface area and high mechanical strength [24–29]. When nanotubes are produced on spaced materials such as a mesh, nanotubes can grow in all directions resulting in a three-dimensional structure. Several studies have been published demonstrating degradation of contaminants using these three-dimensional nanotube structures [26,30,31]. Research [26] has demonstrated that photocatalyst nanotubes with a 3D geometry were more efficient in absorbing light, minimizing photon loss in the liquid, and present a much higher photocatalytic activity per unit surface area compared to a plated one-dimensional array. The results showed that the photocurrent response in the mesh, as an indicator of the photocatalytic activity, was higher than the plate. This indicated a lower recombination of photo-generated electrons and holes, higher photoelectron transfer efficiency and higher light absorbance in the mesh. The increased light absorbance efficiency can be attributed to the different directionality of nanotubes with 3D geometry being more effective in capturing indirect light such as reflected or refracted photons. Beyond the improved light absorbance, better degradation rates in 3D structures have also been attributed to interstitial fissures between the nanotubes, which provides more access to the catalyst surface by the contaminant [26].

Amongst various fabrications and application methods, positioning a photocatalyst near the surface of water, and on a floating support have been investigated by several researchers as it allows higher oxygenation and illumination [17,32–36]. Although the initial floating supported photocatalysts were shown to be effective to clean oil slicks on water [17], the ingestion risk by fishes and animals in water and uncertain toxicity of the degradation intermediates limited their development. This led to various studies to improve their size scale and evaluate their efficacy to degrade various contaminants. Alternatively, TiO2 plates are a commonly considered photocatalyst. With a simple setup, they have been shown to degrade a variety of organic contaminants [8,37]. However, due to the tendency of charge-hole recombination in the semiconductor photocatalyst, the quantum yield is generally low in heterogeneous photocatalysis [19,20,38].

One of the methods to overcome the low quantum yield of heterogeneous photocatalysis on supported substrates is electro-photocatalysis. This refers to an anodic polarization being applied in an electrochemical cell, causing it to act as a photoanode. The applied voltage removes excited electrons from the surface of the photocatalyst, inhibiting the recombination of the electron-hole pairs [39,40]. Researches have demonstrated the enhanced rate of photocatalytic degradation when it is combined with an electrical bias. Research on an Azo dye showed that the degradation rate approximately doubled when an electrical bias of 1.5 V was applied on the nanotube arrays of an anodized Ti mesh electrode [41]. TiO2 coated electrode by the sol-gel method resulted in 65% enhancement in degradation efficiency of methyl tertiary butyl ether using 0.25 bias [42], and the electro-photocatalytic degradation rate of methyl orange was 1.7 times higher on the modified titanium nanotube electrode in comparison to the photocatalytic degradation [43].

There is extensive literature investigating various supports for photocatalysis. Most studies provide a comparison between different nanotube structures or various coating methods, without an experimental peer evaluation between the efficiency of different supports. Considering the advantages of buoyant photocatalysts and nanostructures, the intention of this research was to conduct an exploratory study to compare the efficacy between the conventional floating TiO2 coated on the glass spheres and nanostructured engineered supports, to be used as floating photocatalysts for water decontamination under ambient conditions.

In this research, an experimental evaluation of four types of supported TiO2 photocatalysts were conducted, allowing for a broad comparison between the different supports in a semi-passive mode. Floating TiO2 spheres, an anodized TiO2 plate with one-dimensional nanotube arrays, anodized TiO2 mesh with a three-dimensional nanotube structure, and electro-photocatalysis utilizing the anodized TiO2 mesh were investigated.

The herbicide Killex® was used as a model organic contaminant. It is a widely applied herbicide that is used in lawns and agricultural lands. It contains 2,4-dichlorophenoxyacetic acid (2,4-D), methylchlorophenoxy propionic acid (MCPP or Mecoprop-P) and 3,6-dichloro-2-methoxybenzoic acid (Dicamba). 2,4-D is a contaminant of priority due to its low biodegradability and runoff potential and high mobility [44,45]. Photocatalytic degradation of 2,4-D results in the production of 4-chloro pyrocatechol, 2-chlorophenol, 4-chlorophenol, 2,4-dichlorophenol [46–48]. During photocatalytic degradation of Killex®, degradation of 2,4-D is slower compared to its pure aqueous solution due to the competition between its three components [48,49].

The irradiation source was an ultraviolet – light-emitting diode (UV-LED) panel. UV-LEDs were selected for their energy efficiency and longer lifespan compared to low-pressure UV lamps (i.e., approximately four times) [6,49,50]. Additionally, LEDs contain no mercury, making them an environmentally friendly alternative to standard fluorescent UV bulbs.

The experiments were conducted in a semi-passive mode, where no mixing was involved during the irradiation. The objective was to investigate what can be expected of a semi-passive photocatalytic system if it was to be used to treat ponds containing contaminated water under an ambient environment. Thereby, the reaction rates can be representative of the rate of removal of organic contaminants in natural contaminated watersheds and industrial wastewater ponds that are exposed to sunlight.

#### **2. Materials and Methods**

#### *2.1. Materials*

The titanium mesh with 99.9% purity was supplied from Alfa Aesar. The titanium plate with 99.7% purity was obtained from Sigma-Aldrich, St. Louis, MO, USA. The 6061-grade aluminum was used as a cathode for anodization and was procured from New West Metals Inc., Winnipeg, MB, Canada. The hollow glass microspheres coated with anatase TiO2 (HGMT) were obtained from Cospheric Innovations in Microtechnology. They are referred to as photospheres in this paper and had a median diameter of 45 μm and density of 0.22 g/cm3.

A 99% pure-ethylene glycol was supplied by VWR and 99.9% pure acetone was obtained from EMD. A 98% pure ammonium fluoride, 99.8% ethyl acetate and 99.8% methanol were obtained from Sigma-Aldrich. The phosphoric acid was supplied by J. T. Baker. Commercially available Killex® containing 95 g/L of 2,4-dichlorophenoxyacetic acid (2,4-D), 52.5 g/L of Mecoprop-P, and 9 g/L of Dicamba was purchased from a local retailer.

An in-house UV-LED light source (λ = 365 nm) composed of 16 lamps (NSCU330B, Nichia Corporation, Anan, Japan) in a four by four array on a 10 cm by 10 cm circuit board was used as a light source. It was equipped with an air fan to cool down the generated heat from the lamps. A dual output power supply (Model TW 5005D) was used to drive the module with direct current, ranging from 10 mA to 500 mA [51]. A PANalytical Aeris X-ray diffractometer was used for X-ray diffraction (XRD) characterization of the photocatalysts. The HighScore Plus XRD analysis software was utilized for Qualitative XRD analysis and Rietveld Refinement. An SEM (JEOL JSM-IT300LV, Tokyo, Japan) by InTouchScope and a TEM (Tecnai F20, by Thermo Fischer Scientific) were used to characterize the morphologies of the anodized titanium mesh. A Hummer II sputter coater from Anatech, USA was used to coat the cathode (anodized titanium mesh) with platinum.

#### *2.2. Methods*

#### 2.2.1. Experimental Design and Setup

All experiments were conducted at room temperature, pressure, and under similar conditions. In order to be able to compare the photocatalyst supports (as a variable), all experimental parameters including light source, irradiation light intensity, lack of mixing, distance from the light source and the initial concentration of 2,4-D in the solution were identical in all experiments. The type of the photocatalyst support was different in each experiment, therefore the performance of the studied photocatalysts (i.e., TiO2 photospheres, anodized titanium mesh, plate and electro-photocatalysis using an anodized titanium mesh) were studied as a variable.

The output light intensity of the UV-LED module was linear in the range of 2.2 mW/cm2 to 17.3 mW/cm2. The light intensity was adjusted at 15 mW/cm2 [51]. The distance of the photocatalyst surface from the irradiation source was 5.5 cm in all experiments.

Wide mounted glass jars with a diameter of 4 cm were used as reaction vessels. In all experiments, 0.52 mL of Killex® was dissolved in 1 L of milli-Q water resulting in a solution containing 49.4 ppm 2,4-D, 27.3 ppm Mecoprop-P and 4.68 ppm Dicamba. In each experiment, a control sample without a photocatalyst was also monitored to evaluate the possible photolysis in the solutions. The length of time between sampling intervals created a requirement to correct the mass of the solution for evaporation. This was done by monitoring the mass of the solution during the experiments. The milli-Q water was added before each sampling to adjust the mass of the solution to its initial mass before each irradiation cycle.

In the experiments using an anodized titanium plate, the plate was placed at the bottom of the reaction jar. Dimensions of the plate were measured and used to calculate the surface area of the photocatalyst for energy calculations. When the titanium mesh was used, it was secured close to the surface of the reaction medium using a stainless steel wire mesh. The surface area of the mesh was calculated based on its dimensions and the percentage of the open area (64%) was used to calculate the effective surface area for energy calculations.

In the electro-photocatalysis experiments, a platinum coated anodized mesh was used as a cathode, and the anodized titanium mesh was used as a photoanode. Anode and cathode were connected to a battery (1.5 V) as a source of direct electric current. Current and voltage were measured with a voltmeter and ammeter during the experiment.

For the TiO2 photospheres experiments, the photospheres were mixed with water and those that were buoyant, were collected and dried overnight. The photospheres were then mixed with spiked water using a magnetic stirrer for 30 min before irradiation to allow adsorption to reach equilibrium. In order to identify an optimum photocatalyst loading for the purpose of comparison with the other studied photocatalysts, three different catalyst loadings were tested, 7.2 mg/cm<sup>2</sup> and 11.9 mg/cm<sup>2</sup> and 16.7 mg/cm2, equal to 1.8 g/L, 3.0 g/L, and 4.2 g/L, respectively. Since the photocatalyst was floating on the surface, the surface area of the reaction jar was considered as the irradiated surface by the light source and was used in energy calculations.

#### 2.2.2. Anodization of Titanium Mesh

Supported TiO2 nanotubes were prepared by anodization of a titanium mesh followed by a temperature controlled annealing to acquire anatase phase, which is the most desired crystalline morphology of TiO2 for photocatalytic applications [52]. The titanium mesh was first degreased by sonication in an acetone/methanol solution for 30 min. It was then rinsed with water and dried at room temperature. The degreased mesh was anodized in an aqueous solution of ethylene glycol, containing 2% water and 0.5% ammonium fluoride for 24 h. The anodized mesh was then rinsed with water and dried at room temperature followed by annealing in the oven for 3 h at 450 ◦C to form an anatase crystalline structure [24,26,45–47]. The mesh was cut into uniform pieces with a surface area of

2.9 ± 0.3 cm2, and each of them were attached to a reaction jar via a stainless steel wire and irradiated under the UV-LED lamp.

#### 2.2.3. Sample Analysis

Samples were filtered using 0.45 μm PTFE 25 mm syringe filters before analysis, and were analyzed using HPLC with a UV-visible detector, fitted with a Restek Pinnacle DB C18 column. Absorbance was recorded at a wavelength of 230 nm. Ingredients of Killex®, have all the peak wavelength at the range of 230 nm to 280 nm (i.e., 2,4-D: λmax = 280 nm, Mecoprop: λmax = 230 nm and Dicamba: λmax = 275 nm) [6,53,54]. The UV detection wavelength of 230 nm is selected so as to be able to detect all three compounds in the Killex® solution [55]. However, degradation of 2,4-D as a model contaminant is studied in this paper for the purpose of comparison of the photocatalysts. The eluent used was water:methanol (25:75) with ten mM phosphoric acid and a flow rate of 1 L/min.

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

#### *3.1. Photocatalyst Characterization*

Titanium dioxide exists in three different crystalline structures of rutile, anatase and brookite [56]. The anatase and rutile show different structural characteristics in terms of Ti–Ti and Ti–O distances. The Ti–Ti distance is larger in anatase than rutile but the Ti–O distance is shorter. These structural differences lead to different mass densities and different electronic structures that causes the difference in the mobility of the charge carriers under light excitation [57]. Amongst all crystal structures of TiO2, the anatase shows the highest photocatalytic activity [56]. The X-ray diffractometry (XRD) of the photocatalysts indicated that the primary crystal structure of all the photocatalysts was anatase. (The XRD graphs of all the photocatalyst samples are presented in the supplementary material).

The XRD quantitative results that show the weight percentage of titanium, TiO2, and its crystalline morphologies are presented in Table 1.


**Table 1.** XRD quantitative results of titanium dioxide (TiO2) photocatalysts.

**Figure 1.** Morphology of the anodized mesh: (**a**) Digital and SEM images before anodization; (**b**) TEM image of a nanotube detached from the surface; (**c**) digital and SEM images after anodization; (**d**,**e**) SEM image of titanium nanotube arrays on the surface of mesh.

Further analysis was performed on the mesh photocatalyst for characterizing the surface morphology, length, and thickness of the nanostructures by a scanning electron microscope (SEM) and a transmission electron microscope (TEM). Images are presented in Figure 1. As it can be seen on the images, the length of the nanotubes was about 300 nm with the opening diameter of 54 nm.

#### *3.2. TiO2 Photospheres*

A comparative study of the photosphere loadings was conducted. The results are presented in Figure 2. Samples were irradiated for nine h and 22 min with a total energy of 127.1 J/cm3. The 2,4-D has no absorbance at the peak wavelength of LED (λ = 365 nm). Therefore, no direct photolysis took place during its degradation [6]. The degradation showed a notable improvement when increased beyond a photosphere loading of 7.2 mg/cm2, while results of 11.9 mg/cm2 and 16.7 mg/cm<sup>2</sup> were similar. The degradation by photospheres in this study followed a zero-order reaction kinetics, with the kinetic rate constant (K value) of 4.12 mg L−<sup>1</sup> h−<sup>1</sup> and the half-life time of 6.05 h at a loading of 7.2 mg/cm2. K values were 4.55 mg L−<sup>1</sup> h−<sup>1</sup> and 4.36 mg L−<sup>1</sup> h<sup>−</sup>1, and the half-life times were 5.48 h and 5.72 h at the two higher loadings of 11.9 mg/cm<sup>2</sup> and 16.7 mg/cm2 respectively. No degradation took place in the control sample. Zero-order reaction kinetics was also reported in other studies, where TiO2 floating beads were used for the degradation of dyes [36]. The degradation efficiency was 80%, 86% and 89% at the loadings of 7.2 mg/cm2, 11.9 mg/cm2 and 16.7 mg/cm2 respectively. The results are summarized in Table 2.

**Figure 2.** Degradation of 2,4-dichlorophenoxyacetic acid (2,4-D) in Killex® solution at various loadings of TiO2 photospheres (mg/cm2).

**Table 2.** Summary of the experimental results using photospheres.


The results indicated that increasing the loading of photospheres in the solution increased the degradation rate up to an optimum value beyond which the degradation rate decreased as more photocatalyst was added. This phenomenon has been demonstrated by various researchers [58]. Since heterogeneous photocatalysis occurs inside the active sites on the surface of TiO2, the rate of degradation increases when the loading of the photocatalyst increases. This is due to the higher surface

area, hence the availability of the active reaction sites. However, an optimum TiO2 dosage exists after which degradation rates will decrease with higher catalyst loadings [59]. This phenomenon has been attributed to photon limitation or light scattering on the surface of the photocatalyst particles, decreased light penetration through active sites, and uneven competition between the nanoparticles for light adsorption at higher loadings [60]. Moreover, the high catalyst loading may deactivate the originally activated TiO2 through collision with ground state catalysts. The following reaction demonstrates the collision mechanism between TiO2 species. TiO<sup>∗</sup> <sup>2</sup> has the activated species adsorbed on its surface, TiO2 is the ground state and TiO# <sup>2</sup> is the deactivated state of TiO2 [61,62]. When the light activated TiO∗ <sup>2</sup> collides with the ground state TiO2, TiO<sup>∗</sup> <sup>2</sup> gets deactivated (TiO# 2), leaving the ground state behind. A high loading of the photocatalyst, increases the chance of collision between the species, resulting in the higher frequency in the deactivation process.

$$\text{TiO}\_2^\* + \text{TiO}\_2 \rightarrow \text{TiO}\_2^\# + \text{TiO}\_2,\tag{1}$$

A screening effect of UV irradiation is also reported to be responsible for lowering the rate of photocatalysis at the loading values beyond the threshold limit [59,63]. In addition, due to the uncontrolled catalyst agglomeration on the sidewalls of the reaction jar, increasing the catalyst loading does not enhance the available catalyst on the surface of the reaction media. Since most of the studies utilize photospheres in a reactor with mixing, this phenomenon is fairly reported in the literature [9,64].

The experiments resulted in similar degradation rates in two photocatalyst loadings. Therefore, it is possible that the optimum level of loading was reached. This allows for a comparison to other suspended photocatalytic methods as the measured level of treatment represents the highest level of attainable degradation by photospheres.

#### *3.3. Anodized Mesh*

An experimental study utilizing the anodized TiO2 mesh was performed to assess its photocatalytic activity in a semi-passive mode, and as well to compare it with the anodized plate and other methods studied here. The degradation of 2,4-D with time is presented in Figure 3. The zero-order kinetic rate constant was 3.45 mg L−<sup>1</sup> h−<sup>1</sup> with the half-life time (t1/2) of 7.09 h. Energy consumption to achieve 99.8% degradation was on average 160 J/cm3.

**Figure 3.** Degradation of 2,4-D using anodized mesh.

Experiments were performed in duplicate to account for considerable variances that can occur when growing TiO2 nanotubes on the non-uniform surface of the mesh. The degradation efficiency varied 2.2% between the replicate samples. Although the percentage was small, the difference between samples can be attributed to the variation on nanotube structure due to the variations in the electrical field alongside the wires of the mesh during the anodization process. This emphasizes the importance of the production of a uniform mesh.

Each TiO2 mesh was tested a second time to assess its durability in repeated applications. One of the samples showed a similar efficiency with degradation efficiency varied between 99.65 to 100%, but the degradation efficiency dropped 13% in the second replicate. The decreased efficiency has been reported for repeated runs on TiO2 nanotube meshes by Liao et al. [26], who found that after five runs the degradation efficiency decreased by 8%. The mechanism for loss of degradation efficiency is due to the imperfect mechanical stability of the nanotube layers as TiO2 nanotubes tend to flake off from the substrate during the experiment [65]. Previous research [66] showed that the mechanical stability of TiO2 nanotube arrays could be improved by surface modification methods such as incorporating carbon into microstructures of nanotube layers, which increases the hardness and mechanical strength of the layers. In another study, it was demonstrated that the presence of fluoride-rich layer improves the adhesion of the nanotube layer [65].

Variables affecting anodization are electrolytes types, voltage, and duration of the anodization, annealing temperature and duration, as well as the distance between the electrodes. They cause a difference in the crystal phase, length and width of the produced TiO2 nanotubes which is the reason behind the difference in their mechanical stability on the surface of the titanium mesh [25,26,28,67–70]. When the nanotubes are not mechanically stable, they can fall off from the surface in repetitive usage, resulting in a reduced photocatalytic activity.

#### *3.4. Anodized Plate*

The degradation efficiency was 57% after 10 h of irradiation, consuming 111.9 J/cm3 of energy. Degradation followed a zero-order degradation rate constant of 3.07 mg L−<sup>1</sup> h−<sup>1</sup> with the half-life time of 7.97 h. Since the photocatalyst plate was placed at the bottom of the reaction jar, an additional experiment was performed to investigate the effect of the depth of water column on the degradation efficiency. The plate was covered with contaminated water, which resulted in various water depths ranging from 1 cm to 4 cm on top of the photocatalyst plate. The concentration, distance from the irradiation source and the light intensity were identical in all samples. All samples irradiated for 10 h. The results are presented in Figure 4.

**Figure 4.** Degradation efficiency versus energy at various depths using the anodized plate.

It was observed that at a depth of 4 cm, energy consumption was 55.9 J/cm3 to achieve 38% degradation efficiency (defined as ΔM/M0 = (M0 − Mt)/M0), which was 27% lower compared to a 1 cm depth (76.3 J/cm3). The results showed that the increase in depth, did not decrease the energy efficiency of the system. Although the Beer Lambert equation (A = ε b c) states light absorbance in a solution (A) is directly proportional to its path length (b), molar concentration (c), and molar absorptivity (ε), the results showed that the effect of path length at the studied depth ranging from 1 cm to 4 cm was negligible.

This could be associated with the mass of the contaminant which is a significant parameter in the photocatalytic degradation reactions due to its effect on the adsorption of the reactant on the surface of the photocatalyst and mass transfer in the solution during the photocatalytic degradation reaction [9,71,72]. Since the mass was higher when the water column was higher, the effect of mass supersedes the effect of light absorbance between the studied depths of 1 to 4 cm. It was then concluded that when the anodized plate is placed at the bottom of the reaction jar, the effect of the depth of water column is not as important as the mass of the contaminant; hence it can be neglected. Therefore, the results are comparable with the floating photospheres and mesh, where the photocatalyst is brought near the surface.

#### *3.5. Electro-Photo Catalysis Using Anodized Titanium Mesh*

A study for comparative analysis of electro-photocatalysis was conducted on the anodized TiO2 mesh. Voltage, as the driving force in preventing electron-hole pair recombination [39], was kept constant at 1.5 ± 0.1 V during the 6 h irradiation. The photocurrent increased from 6 mA before irradiation to about 40 mA after UV irradiation. Voltage was variable at the first 20 min which was corrected by manually adjusting to the initial value of 1.5 V. The current was approximately 45 mA during the 6 h irradiation period. Results of the experiment are presented in Figure 5.

**Figure 5.** Concentration of 2,4-D with time during electro-photocatalysis.

Kinetics followed a zero-order model with a rate constant of 5.04 mg L−<sup>1</sup> h−<sup>1</sup> with the half-life time of 4.86 h during the first run. The enhanced degradation efficiency is due to the presence of the polarized anode, working as a photoanode in the photoelectrochemical cell. The applied voltage produced an electric field which is opposed to the attraction forces between excited electrons and the holes that are produced during photocatalysis. This results in excited electrons being removed from the surface of TiO2, minimizing the electron–hole recombination [39–41,73]. The same mesh was tested a second time, showing an efficiency drop of 14% compared to the initial run. Literature on the loss of degradation efficiency varies considerably. It was reported from no loss [41] to 7% [74] and 8% [26] efficiency reduction during the repeated cycles. The mechanism for the reduction in the photocatalytic activity after repeated use is the same as discussed for TiO2 mesh earlier.

#### *3.6. Comparison of Three Forms of the Photocatalyst and Electro-Photocatalysis*

The required energy to achieve 60% degradation was used to compare the efficiency of the four studied photocatalysts. In the experiments using the mesh photocatalyst, the average value of the energy consumption of the four experiments is used for the purpose of comparison.

Energy (Elamp in J/cm3) was calculated based on the LED output in mW/cm2 and the surface area of the photocatalyst. Therefore, the variation of the surface area of the different photocatalysts was taken into account. The degradation efficiency or the normalized degraded mass of 2,4-D (ΔM) was calculated based on the mass difference during the photocatalytic reaction using the following equation where M0 is the initial mass of 2,4-D in the solution, and Mt is its mass at various sampling intervals:

$$
\Delta \mathbf{M} = \frac{\mathbf{M}\_0 - \mathbf{M}\_t}{\mathbf{M}\_0},
\tag{2}
$$

In electro-photocatalysis, the applied excessive voltage was accounted in energy calculations. The following equations were used to calculate the total consumed energy per volume of the sample (Etotal in J/cm3):

$$\text{E}\_{\text{total}} = \text{E}\_{\text{lamp}} + \text{E}\_{\text{electrode}\text{---chemical cell}} \tag{3}$$

$$E\_{\text{electrode}-\text{chemical cell}} = \frac{\text{V} \times \text{I} \times \text{t}}{\text{V}\_{\text{s}}} \tag{4}$$

where V is the average potential of the electrochemical cell (in V), I is the average of measured current during irradiation (in A), t is time (in s) and Vs is the volume of the solution (in cm3) [75,76]. The results are summarized in Table 3, and illustrated in Figure 6.

**Table 3.** Experimental results of 2,4-D degradation in various photocatalyst substrates.


It was observed that energy consumption was at the same level in the anodized plate (i.e., 114.5 J/cm3) and electro-photocatalysis (i.e., 112.2 J/cm3). Energy consumption was also similar in the photospheres experiment at 80.3 J/cm3, with the average energy consumption of the four studied anodized mesh at 80.3 J/cm3. In comparison to the photospheres experiment and the mesh, energy consumption was 39% higher on average when the anodized plate and electro-photocatalysis were used.

The lower efficiency and higher energy consumption of the anodized plate is due to the 2D geometry of its planar surface compared to the 3D nanostructures on the mesh [26]. Although the dimensions of the plate and mesh were similar, void spaces in the mesh resulted in a significantly lower surface area available for light absorbance. In contrast, the three-dimensional surface morphology of TiO2 nanotubes grown on a mesh are more efficient in the absorbance of the scattered light compared to the 2D geometry of the planar surface of the titanium plate despite having a lower surface available for irradiation. Moreover, the interstitial fissures between the nanotubes allow contaminant molecules easier access to the photocatalyst surface area, which enhances the photocatalytic activity.

This phenomenon has been studied by other researchers [26] in the photocatalytic degradation of methyl orange under a high-pressure mercury lamp with the wavelength of 365 nm. It was demonstrated that under similar conditions and equal time (360 min), the photocatalytic degradation of methyl orange was 18% higher when an anodized mesh was compared with an anodized plate.

The outperformance of the photospheres in comparison to the plate can be attributed to the increased access to the catalyst surface area by photons and the reactants—a primary factor in determining the degradation rate in photocatalysis. Other studies have also demonstrated similar results [51].

**Figure 6.** Comparison between photocatalysts based on the mass difference versus energy.

It was observed that with a similar amount of energy, the degradation efficiency of the anodized mesh was comparable with photospheres. This is due to the improved photocatalytic activity of the supported TNTs and 3D structures of TiO2 surrounding the mesh wires, compared to anatase TiO2, mounted on the glass spheres. Moreover, it has significant advantages because a separation step after treatment is not required, and it can be removed easily when the treatment is completed.

Albeit electro-photocatalysis resulted in the lowest half-life time (i.e., 4.86 h) compared to the methods studied here, but its energy consumption was at the same level as the anodized plate. This outcome was also in-line with other published comparative studies [41,42,75–77]. The enhanced efficiency is attributed to the presence of the electrical bias between the electrodes which minimized the electron-hole recombination on the photocatalyst and increased the quantum yield of the photocatalytic degradation process. The primary disadvantage of this system is the required additional electricity to the photoanode as well as the increased complexity of the process, which leads to a higher capital and operational costs [75].

The studied semi-passive system using the anodized plate is also compared with a flow-through photoreactor, using the same anodized plate and the same light source to degrade 2,4-D in a Killex® solution [49]. During 2 h of irradiation, the photoreactor resulted in 78% higher degradation efficiency of ΔM = 0.25 compared to ΔM = 0.14 in the semi-passive system. Energy consumption of the semi-passive system was 44.8 J/cm<sup>3</sup> at ΔM = 0.25 which was 3.6 times higher than the photoreactor (12.5 J/cm3). This difference is due to the slower kinetics of the photocatalytic degradation in the semi-passive system due to the lack of mixing that minimizes mass transfer. The reported pseudo first-order kinetic rate constant was 648 h−<sup>1</sup> in the reactor, compared to the zero-order rate constant of 3.07 mg L−<sup>1</sup> h−<sup>1</sup> in the semi-passive photocatalysis. It can be concluded that the semi-passive photocatalysis consumes more energy and results in a lower degradation efficiency during a same period of time when compared to a photoreactor.

Although photospheres have recently been broadly and successfully studied for degradation of environmental contaminants [20,78–81], their usage is still constrained by problems with separation after treatment. Moreover, ensuring the even dispersion of the photospheres for a semi-passive treatment, with no mixing, may be a challenge in practical applications. Furthermore, a direct comparison between the publications in this field and the results presented in this paper is not feasible due to the different operational parameters and the variety of the studied contaminants. The feasibility of utilization of the robust anodized TiO2 based photocatalysts, designing a system for field applications at a larger scale, and studying the system to degrade various organic contaminants in water will be the direction for future research.

#### **4. Conclusions**


**Supplementary Materials:** The following are available online at http://www.mdpi.com/2073-4441/11/3/621/s1, Figure S1. XRD of titanium dioxide mesh, Figure S2. XRD of titanium dioxide plate, Figure S3. XRD of titanium dioxide photospheres.

**Author Contributions:** G.H. designed the study and conducted the experiments, analyzed data and wrote the original draft of the paper; J.H. edited the paper and supported data analysis; G.A. oversaw the research and contributed in experimental design, data analysis and reviewed the final draft; Late C.H.L. oversaw the research and experimental design. He could not review this manuscript.

**Funding:** This research was partially funded by CMC Research Institutes, MITACS and Natural Sciences and Engineering Research Council of Canada (NSERC).

**Acknowledgments:** The authors would like to thank Mitacs, and CMC Research Institutes and Natural Sciences and Engineering Research Council of Canada (NSERC) for providing partial funding for the project.

**Conflicts of Interest:** The authors declare no conflict of interest. The funders had no role in the design of the study; in the collection, analyses, or interpretation of data; in the writing of the manuscript, and in the decision to publish the results.

#### **References**

1. Fujishima, A.; Honda, K. Electrochemical evidence for the mechanism of the primary state of photosynthesis. *Bull. Chem. Soc. Jpn.* **1971**, *44*, 1148–1150. [CrossRef]


© 2019 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 (http://creativecommons.org/licenses/by/4.0/).

## *Article* **An Amphiphilic, Graphitic Buckypaper Capturing Enzyme Biomolecules from Water**

#### **Shahin Homaeigohar \* and Mady Elbahri \***

Nanochemistry and Nanoengineering, School of Chemical Engineering, Department of Chemistry and Materials Science, Aalto University, Kemistintie 1, 00076 Aalto, Finland

**\*** Correspondence: shahin.homaeigohar@aalto.fi (S.H.); mady.elbahri@aalto.fi (M.E.); Tel.: +358-50-431-9831

Received: 13 November 2018; Accepted: 18 December 2018; Published: 20 December 2018

**Abstract:** The development of carbon nanomaterials for adsorption based removal of organic pollutants from water is a progressive research subject. In this regard, carbon nanomaterials with bifunctionality towards polar and non-polar or even amphiphilic undesired materials are indeed attractive for further study and implementation. Here, we created carbon buckypaper adsorbents comprising amphiphilic (oxygenated amorphous carbon (a-COx)/graphite (G)) nanofilaments that can dynamically adsorb organic biomolecules (i.e., urease enzyme) and thus purify the wastewaters of relevant industries. Given the dynamic conditions of the test, the adsorbent was highly efficient in adsorption of the enzyme (88%) while being permeable to water (4750 L·h<sup>−</sup>1m−2bar−1); thus, it holds great promise for further development and upscaling. A subsequent citric acid functionalization declined selectivity of the membrane to urease, implying that the biomolecules adsorb mostly via graphitic domains rather than oxidized, polar amorphous carbon ones.

**Keywords:** carbon; nanofiber; membrane; urease; biomolecules; water treatment

#### **1. Introduction**

As a global challenge, water scarcity is expanding to major parts of the world, threatening human beings' lives. This crisis can have different origins but undoubtedly water pollution from industry and from urban communities is a main one. Amongst the variety of water pollutants, the organic ones such as proteins and biomolecules play a determining role. These substances even at a negligible amount, <1% of the entire contamination in a river, for instance, can deplete the oxygen present in water and cause the death of living creatures in that ecosystem [1]. Water recycling via purification can somewhat remediate this problem but necessitates the development of advanced water treatment systems. Micro-, ultra- and nanofiltration membranes are typically utilized for wastewater treatment. Their purification action mainly relies on sieving of the pollutants, and thus they require a porous structure whose pore size is less than the size of the solute to be separated. Other than the membranes, in a sustainable manner and using conventional and also emerging materials, functionalized adsorbents have shown applicability in the removal of even molecules and tiny pollutants based on physical/chemical interactions or biological functions [2–7]. Accordingly, there is no need for the construction of porous materials with small pore sizes that could impose high pressure differences. Moreover, a functionalized adsorbent with a surface decorated by particular functional groups can discriminate or entrap molecules in a selective manner [8].

Electrospun nanofibrous adsorbents have shown promising capabilities for selective water remediation. Their structure possesses a high interconnected porosity and huge surface area that in case of functionalization can efficiently separate functional pollutants, e.g., ions, dye molecules, organics, etc. While the high porosity realizes a significant permeability and with that, energy efficiency, the expansive surface area enables the notable functionalization necessary for highly selective

adsorbents. In this regard, several biofunctionalized nanofibrous membranes made of polyurethane, polysulfone, polyacrylonitrile, and cellulose have been tested for the separation of protein and enzyme (e.g., Immunoglobulin G (IgG), Bovine Serum Albumin (BSA), lipase, bromelain, etc.) [8–11]. In our studies [2,4,12], we also developed a biofunctionalized nanofibrous adsorbent composed of Bovine Serum Albumin and poly(acrylonitrile-co-glycidyl methacrylate) (PANGMA), as the functional agent and polymer nanofiber, respectively, that could offer a significant metal nanoparticle and biomolecule removal efficiency while being highly water permeable. This adsorbent was synthesized in a simple fashion versus the other previously developed systems [8,13]. The separation tests were performed under the most tricky conditions, i.e., dynamically and with a low protein amount (a few mg·L−<sup>1</sup> instead of mg·mL−<sup>1</sup> adopted by References [8,10,11,14]) and with a size scale of pollutants, potentially passing readily through a macroporous nanofibrous structure. Despite such circumstances, the adsorbent was successful in the removal of nanoparticles (97%) as well as proteins (88% BSA and 81% *Candida antarctica* Lipase B (Cal-B)). In another research, we developed a nanofibrous adsorbent comprising polyethersulfone (PES) nanofibers that were functionalized by the inclusion of vanadium oxide (V2O5) nanoparticles [6]. This adsorbent system was able to separate methylene blue (MB) dye from water with an efficiency of 85% under alkaline condition and high temperature.

Despite the various merits of the above-mentioned systems in the adsorption of diverse water pollutants, their synthesis and functionalization are not one pot. As a step forward to meet this need, recently, we developed carbon buckypapers based on amphiphilic carbon nanofilaments [15]. The nanofilaments are composed of oxygenated amorphous carbon (a-COx) and graphite (G), and thus are able to adsorb both polar (e.g., dye) and non-polar (e.g., oil) water pollutants efficiently. Already investigating the applicability of the amphiphilic graphitic buckypaper in discrimination of polar and non-polar contaminants, here, we challenge the buckypaper adsorbents with an amphiphilic water pollutant. For this sake, biomolecules (i.e., urease enzyme), one of the major organic pollutants that can adversely affect the water ecosystems, will be considered.

#### **2. Experimental**

*Materials*: polyacrylonitrile (PAN) (200,000 g·mol−1, purity 99.5%) and dimethylformamide (DMF) (purity 99%) were purchased from Dolan GmbH (Kelheim, Germany) and Merck (Darmstadt, Germany), respectively. Urease enzyme (impurity; ammonium < 4–10 <sup>μ</sup>mol·U−<sup>1</sup> enzyme) and citric acid (citric acid monohydrate, ACS reagent ≥ 99.0%) were obtained from Sigma-Aldrich (Saint Louis, MO, USA). All the materials were used as received.

*Synthesis*: The precursor PAN nanofibers were synthesized by electrospinning. To do so, employing a syringe pump (Harvard Apparatus, Holliston, MA, USA), a solution of PAN (8 wt % in DMF) was fed steadily (1 mL·h−1) into a needle (0.8 mm inner diameter with a circular opening). Upon electrifying the solution with a voltage of 20 kV (Heinzinger Electronic GmbH, Rosenheim, Germany), PAN was electrospun on an aluminum foil. The as-synthesized PAN nanofibers underwent oxidative stabilization and were heated in air at 250 ◦C for 2 h within a furnace with maximum operational temperature of 1250 ◦C (Linn Elektro Therm). In the next step, the oxidized nanofibers were carbonized under argon atmosphere at 1250 ◦C for half an hour with a heating rate of 5 ◦C·min−<sup>1</sup> and then cooled down to the room temperature at a same rate.

Due to the extreme brittleness of the graphitized nanofibers, challenging their handling as a freestanding membrane, they were suspended in distilled water (10 mL) and underwent an ultrasonication process for 2 min at a power of 20%. The a-COx/G nanofibers under the influence of ultrasonication are disintegrated as suspended nanofilaments that can be subsequently cast on a circular poly(phenylene sulfide) (PPS) technical nonwoven (3.5 cm in diameter). As a control group, a-COx/G nanofilaments were also functionalized by citric acid (CA). To do this, CA (30 mg·mL−1)) was added to the aqueous suspension to be ultrasonicated.

*Characterization*: The a-COx/G nanofilaments were characterized in terms of morphology by scanning electron microscopy (SEM) (LEO 1550VP Gemini from Carl ZEISS, Jena, Germany) and an

atomic force microscope (AFM) (MultiModeTM Atomic Force Microscope from Bruker AXS, Madison, WI, USA). The surface chemistry of the a-COx/G nanofilaments was analyzed by FTIR (ALPHA (ATR-Ge, ATR-Di) from BRUKER Optik GmbH, Ettlingen, Germany). The a-COx/G buckypaper's pore size distribution was determined by an automated capillary flow porometer (Porous Materials Inc. (PMI), Ithaca, NY, USA).

The urease retention efficiency of the buckypapers was assessed using the corresponding aqueous solutions in a dead-end manner and by employing a lab-built set-up [16]. The set-up's reservoir contained 200 mL urease solution (1 g·L−1) which permeated through the buckypapers under a 0.5 bar pressure. Based on a constructed standard urease calibration curve, the permeate's urease concentration was determined by UV-vis spectroscopy (HITACHI U3000, HITACHI, Tokyo, Japan). The urease retention efficiency (RE) was calculated via Equation (1):

$$\text{RE} = \left(1 - \frac{\text{C}\_{\text{P}}}{\text{C}\_{\text{f}}}\right) \times 100\% \tag{1}$$

where Cp and Cf represent the permeate's and feed's urease concentration, respectively. The permeation time was also recorded and the permeate permeance was calculated via Equation (2) [17]:

$$\mathbf{J} = \frac{\mathbf{Q}}{\mathbf{A} \times \Delta \mathbf{t} \times \Delta \mathbf{P}} \tag{2}$$

where J is the permeate permeance (L·h−1m−2bar−1), Q is the collected volume (L) of the permeate, A is the effective filtration area of the buckypapers (m2), Δt is the collecting time (h), and ΔP is the pressure difference (bar). The permeance measurements were done for three 50 mL permeates to ascertain the consistency of the buckypapers' performance. It is worthy to note that considering the hydrophobic, large microfibers of the PPS support layer, providing huge pore sizes, no significant contact and interaction with the urease molecules passing through the carbon layer can be envisioned. Thus, only the buckypaper is responsible for the reported removal efficiency and permeance.

The electrical conductivity of the buckypapers as non-functionalized and CA-functionalized before and after urease adsorption was measured by a four-point probe test. At least five measurements were done on different parts of the buckypapers, and the error bars were calculated. The thickness of the samples to be considered in the conductivity measurement was already measured by a digital micrometer (Deltascope®MP2C from Fischer, Windsor, CT, USA).

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

The developed buckypaper consists of the a-COx/G nanofilaments, randomly arranged but with no sign of clustering. SEM image, as shown in Figure 1a, clearly verifies this fact and the preservation of a porous structure that guarantees optimum water permeability. Moreover, as seen in Figure 1b, the nanofilaments' tips are exposed to the surrounding medium, and thus, this raises the interactivity of the material with the biomolecule pollutants. In fact, the nanofilaments are able to capture urease through adsorption not only on their body, but also on their cross-sections. AFM images, as shown in Figure 1c, provide insight into the dimensions and morphology of the nanofilaments individually.

**Figure 1.** Scanning electron microscopy (SEM) images show morphology of the nanofilaments at (**a**) a low and (**b**) high magnification. (**c**) Atomic force microscope (AFM) micrographs imply the nanofilaments' dimensions and morphology.

Pore size measurement via a bubble point test, as shown in Figure 2, implies that the pore size lies in the submicron range as small as 700 nm. This pore size distribution qualifies the structure as a microfiltration (MF) membrane [18,19] that could hardly stop the passage of tiny organic pollutants, particularly under hydrodynamic pressure. It is worthy to note that the adsorption tests reported in the literature are typically performed in a static, batch mode that maximizes the contact time of the adsorbent and solutes. To the contrary, here, we adopted an adsorption test in a dynamic mode under a hydrodynamic pressure that could challenge the adsorption efficiency of our samples. This test was performed in three successive steps, in each, 50 mL of the solution was passed through. Such a style can highlight the repeatability of the result and also stress robust bonding of the urease molecules with the nanofilaments in case they are not released in the next steps and if efficiency does not decline.

**Figure 2.** Pore size distribution of the a-COx/G buckypaper measured by a bubble point test.

Despite possessing pore sizes in the MF scale, the buckypaper showed a promising urease separation efficiency, as shown in Figure 3a. A removal efficiency of 88.5% was recorded after permeation of 150 mL urease aqueous solution. An ascending trend from 50 mL (75%) to 150 mL (88%) in urease removal efficiency is observed. As we previously proved [15], the nanofilaments have oxygen-based functional groups including carbonyl and hydroxyl that enable interaction, i.e., hydrogen bonding with amino acid units of urease. In addition to hydrogen bonding, the positively charged amine groups of urease and the negatively charged oxygen-based functional groups of a-COx segments can electrostatically interact. On the other hand, major graphitic regions allow for π-π interaction with non-polar domains of urease. For a similar DNA–CNT system, van der waals forces have been introduced as an adsorption driving factor with a larger impact than hydrophobic forces [20]. For the urease molecules, several intramolecular bondings between different functional groups could also be envisaged. Accordingly, some molecules interact through their less polar and non-polar zones with the nanofilaments. Thus, collectively, different parts of the nanofilaments are able to adsorb urease molecules via interaction with their corresponding regions. This feature can stabilize the enzyme on its substrate and prevent its conformational change that can lead to loss of enzyme activity, which is beneficial for further application as, e.g., a biocatalyst [21–23]. Moreover, the huge surface area of the buckypaper minimizes the diffusion pathway for the reaction products, thus enhancing the efficiency of the immobilized enzyme. ATR-FTIR spectra, as shown in Figure 4a, clearly verify the adsorption of urease onto the nanofilaments. Before the adsorption, the strong peak located at 1589 cm−<sup>1</sup> represents the unoxidized sp<sup>2</sup> C=C groups of the graphitic segments of the nanofilaments, which resulted from the aromatization process during the thermostabilization of PAN nanofibers [24,25]. The second evident groups at 1000–1300 (two bands) and 3800 cm−<sup>1</sup> imply a C–OH bond [15]. After the adsorption, the main chemical bonds related to urease emerge on the nanofilaments. These bonds represented by ATR-FTIR characteristic peaks are marked in Figure 4a.

**Figure 3.** (**a**) Urease removal efficiency; (**b**) permeate permeance of the buckypapers in two classes of non-functionalized and CA-functionalized.

**Figure 4.** ATR-FTIR spectra compare the surface chemistry of carbon nanofilaments before and after (**a**) urease adsorption and (**b**) CA functionalization.

The increasing trend of urease removal efficiency shows that the adsorption of urease is robust and further passage of the solution does not result in its release into water. This enhancement of removal efficiency can be attributed to a strong intermolecular interaction between the adsorbed urease molecules and solutes via peptide–peptide interactions [26]. Interestingly, adsorption of urease molecules enhances permeate permeance of the buckypaper after the first round, due to a hydrophilization effect, as shown in Figure 3b. It is worthy to note that the pure water permeance of non- and CA-functionalized were measured as ≈8670 and 14000 (L·h−1m−2bar−1), respectively, and adsorption of urease declines, most likely due to pore blockage and loss of porosity. In contrast to the non-functionalized samples, CA-functionalization and the emergence of various functional groups such as carboxyl and hydroxyl (Figure 4b) slightly lower the removal efficiency as far as the filtration is continued. While a high efficiency of 87% is seen at the onset of the experiment, it declines to 78% at 150 mL permeate volume. The reason could be found at less available binding sites for urease molecules or even the release of the previously adsorbed ones because of less graphitic regions that most likely have played a more important role in the stable adsorption of urease molecules rather than polar groups (hydrogen bonding or electrostatic interaction). However, still efficiency is as promising as 78%. It is worth nothing that permeate permeance for CA-buckypapers are significantly higher than that for the non-functionalized ones due to their hydrophilicity. The descending trend of permeate permeance in this class of adsorbents could be attributed to their declined hydrophilicity compared to the neat or fresh CA-functionalized samples due to the adsorption of the urease molecules. Slightly enhanced hydrophobicity along with the accumulation of the adsorbed molecules on the nanofilaments that lower pore size, cooperatively increase the resistance against water permeation.

The adsorption experiment performed here can be regarded as a proof of concept witnessing the applicability of the buckypaper adsorbent in the removal of urease molecules as a model for biomolecule pollutants from water. In this regard, taking into account the effect of environmental factors such as pH, temperature, ionic strength, adsorption time, and urease concentration, further experiments are in progress. The results will be later used in isotherm, thermodynamic and kinetic calculations. There is also a need for more strict tests considering a diverse range of pollutants, different applied stresses and environmental conditions that can affect the separation performance of such an adsorbent. In this regard, software-assisted design of experiments could be helpful. It can reduce the consumption and waste of chemicals, help with regards to the eco-friendliness of chemical processes and actually shorten the pathway to industrial applications [27,28].

As an extra bonus, the enzyme immobilization successfully performed here can be promising for further applications of the buckypaper with respect to biosensing, e.g., the buckypaper adsorbent can potentially act as a biosensor as well. Adsorption of urease can change the electrical conductivity of the nanofilaments, and thus, the entire buckypaper. To validate this proof of concept, electrical conductivity of the buckypaper before and after adsorption of urease was recorded. As shown in Figure 5, CA functionalization can lower the electrical conductivity of the buckypaper due to the inclusion of carboxyl groups that act as electron-withdrawing elements, raising electrical resistivity. In contrast, adsorption of urease enhances the electrical conductivity notably, but with a lower rate for CA-functionalized buckypapers. One reason for the enhancement of conductivity could be the formation of electron transfer bridges between the nanofilaments by urease molecules. This observation can be interpreted in another way, i.e., bridging between enzyme molecules (i.e., the biosensing element) by the carbon nanofilaments. This phenomenon, i.e., the direct electrical connection of redox enzymes and electrodes through carbon nanomaterials have been reported earlier. Patolski et al. [29] showed this behavior through the alignment of glucose oxidase enzymes on the SWCNTs' tips that were structured as an array on a conductive substrate. Exposure of the enzyme-immobilized buckypaper to the urea, often monitored in blood to track kidney diseases, can alter the electrical conductivity and be considered as the sensing mechanism for such an analyte. It is worthy to note that the immobilization of enzymes is indeed the simplest technique that can tackle the bottleneck of their high solubility [30]. Enzyme immobilization allows for the tailoring of the bioreactions' conditions, and thus enables a continuous process with minimum pollution by the reaction products, an extremely desirable characteristic in the food industry. Moreover, it guarantees an improved stability, lifespan and ease of removal of the enzyme from the reaction medium at the end of the process, enabling cost effectiveness and recycling of the enzyme. As mentioned earlier, immobilization can also lead to stabilization of biocatalysts, prevent their unfolding and immunize the polypeptide bonds against rupture.

**Figure 5.** Electrical conductivity of the various classes of buckypapers before and after the adsorption of urease measured via a four-probe test.

#### **4. Conclusion**

Taken together, we devised a buckypaper adsorbent based on amphiphilic carbon nanofilaments that could separate urease molecules from water effectively (as large as 88%). The separation tests were performed under dynamic conditions that could challenge the adsorbent more strictly. Desirable selectivity and permeance (over 4 kL·h−1m−2bar−1) of this novel adsorbent/membrane holds great promise for further development of the system for practical applications. Furthermore, firm immobilization of urease on conductive nanofilaments can assure the efficiency of a potential biosensing system. This proof of concept makes us optimistic with respect to the high potential of such nanomaterials for water treatment and biosensing in an industrial platform. However, first, we need to tackle some relevant bottlenecks for upscaling of their production. Electrospinning has shown to be a reliable method for large scale production of nanofibers, but post treatment (i.e., carbonization) of nanofibers must be performed in a controlled manner following a precise protocol that can govern a desirable chemistry for nanofibers. This step must be optimized and designed in a more economical way. For instance, the as-developed carbon nanofibers need to be stronger to exclude

the chopping step, while maintaining their uniformity, porosity and more importantly functionality. We are at the beginning of the development of this system for water treatment and biosensing, but the obtained results encourage and motivate us to start further working on our material either as is or coupled with extra reactive agents.

**Author Contributions:** S.H. conceived the idea, prepared samples, performed characterizations and drafted the manuscript. M.E. was involved in development of the idea and analysis of the results.

**Funding:** M.E. appreciates the financial support provided through Aalto University, Academy of Finland, and Helmholtz Association (Grant No. VH-NG-523).

**Acknowledgments:** The authors would like to acknowledge Kristian Bühr for the design of the water permeance measurement set-up, and Joachim Koll for the bubble point test.

**Conflicts of Interest:** The authors declare no conflict of interest. The founding sponsors had no role in the design of the study; in the collection, analyses, or interpretation of data; in the writing of the manuscript, and in the decision to publish the results.

#### **References**


© 2018 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 (http://creativecommons.org/licenses/by/4.0/).

## *Review* **An Overview of the Water Remediation Potential of Nanomaterials and Their Ecotoxicological Impacts**

#### **Mehrnoosh Ghadimi 1, Sasan Zangenehtabar <sup>1</sup> and Shahin Homaeigohar 2,\***


Received: 17 March 2020; Accepted: 15 April 2020; Published: 17 April 2020

**Abstract:** Nanomaterials, i.e., those materials which have at least one dimension in the 1–100 nm size range, have produced a new generation of technologies for water purification. This includes nanosized adsorbents, nanomembranes, photocatalysts, etc. On the other hand, their uncontrolled release can potentially endanger biota in various environmental domains such as soil and water systems. In this review, we point out the opportunities created by the use of nanomaterials for water remediation and also the adverse effects of such small potential pollutants on the environment. While there is still a large need to further identify the potential hazards of nanomaterials through extensive lab or even field studies, an overview on the current knowledge about the pros and cons of such systems should be helpful for their better implementation.

**Keywords:** nanomaterials; water treatment; environmental risks

#### **1. Introduction**

Water, a previously plentiful, free resource across the world, has become a rare, costly object over recent decades and currently, water shortage is going to be a challenge for sustainable development of the human community [1,2]. This crisis is dramatically expanding and is regarded a global systemic risk, mainly resulting from urban, agricultural, and industrial pollution. In these areas, water consumption has incremented up to 70% (agriculture), 22% (industry), and 8% (domestic) of the currently available fresh water and, accordingly, an enormous volume of wastewater containing a variety of pollutants has been produced [3]. No doubt, the release of wastewater from commercial and industrial sectors besides untreated domestic sewage and chemical pollutants into fresh water resources is horribly detrimental to both the human community and the ecosystem, including animals and plants. In this regard, the major water contaminants are heavy metal ions, organics (e.g., dyes), and oils that can disqualify any water stream for drinking.

To address the need for water remediation systems, during the past few decades, with the evolution of nanotechnology, a diverse range of new technologies based on nanomaterials has been developed. For instance, as adsorbent systems, nanomaterials offer an extremely large reactive surface area at a low mass, can be produced at a much less cost compared to activated carbon and they can remove pollutants efficiently [2]. In this regard, a plethora of nano-adsorbents in various forms and dimensionalities (D) like nanoparticles (0D), nanofibers and nanotubes (1D), nanosheets (2D), and nanoflowers (3D) has been investigated [2]. In terms of composition, the diversity is indeed extreme and many organic and inorganic nanomaterials have been synthesized that can help purify water streams. The separation mechanism can be based on chemical/physical affinity of the pollutant for the surface of the nanomaterial or through size exclusion of the pollutant by a porous nanomaterial system. In the latter case, nanomaterials act either as the main building block of the porous separator structure, as seen

for nanofibrous microfiltration or ultrafiltration membranes, or as an additive to a polymeric thin film membrane to improve its hydrophilicity and thermomechanical properties. The previously mentioned separation processes such as adsorption or filtration only gather the pollutant molecules in solid form but never entirely "eliminate" or "decompose" them. This issue could be problematic because disposal of the obtained sludge and fouling of the filtration system is challenging [4]. For such reasons, the separation process should be complemented by degradation processes such as photocatalysis, sonocatalysis, and reductive degradation that allow the decomposition of organic pollutants into non-toxic metabolites. Other than environmental remediation, nanomaterials can also be efficiently applied for environmental control and construction of sensors that can detect even trace amounts of water pollutants.

Despite all the beneficial potentials that nanomaterials offer for the sake of water remediation and control, their unwanted and uncontrolled release can harm the environment and health of human beings, animals and plants. While nanomaterials are produced in different dimensionalities and aspect ratios, nanoparticles, i.e., 0D nanomaterials, are indeed the most challenging ones in terms of environmental risks. Originating from anthropogenic and natural sources, nanoparticles have always been present in the environment. Nanoparticles suspended in air are normally regarded as ultrafine particles, while the ones existing in soil and water form colloids [5]. In urban areas, particularly metropolitan areas, the emissions of vehicles fueled with diesel and gasoline as well as stationary combustion sources, generate a large amount of particulate materials of different sizes. In this context, the amount of the manufactured nanoparticles exceeds 36% of the entire particulate number concentrations [6]. Apart from synthetic nanoparticles, naturally formed nanoparticles are also present in the atmosphere, though in a much lower concentration compared to the manufactured ones [5]. In aquatic media, the term "colloid" is generally ascribed to those particles whose size varies in the range of 1 nm to 1 μm. Aquatic colloids consist of macromolecular organics, including peptides and proteins, humic and fulvic acids, and also colloidal inorganic materials composed of hydrous iron and manganese oxides. The small size, extensive surface area, high surface energy, quantum confinement, and conformational behavior of the colloids enable them to bind to various organic and inorganic contaminants [5]. Lastly, with respect to soils, there is a variety of natural nanoparticles comprising organic matter, clays, iron oxides, and other minerals that are decisive in diverse bio-geochemical processes. The soil colloids encompassing nanoparticles and their impact on soil development (pedogenesis) and soil structure (dispersion and crusting) have been a demanding research topic for decades. In this regard, synthetic nanomaterials and the soil colloids made thereof could drive and ease the transfer process of contaminants in soils [5].

In this review, we introduce potential applications of nanomaterials for water remediation, and on the other hand, discuss the possible routes of their release into different environmental sectors like soils and water bodies and their harmful effects. Search in the "ISI Web of Knowledge" database on the main topics to be discussed in the current review, i.e., "Nanomaterials for Water Treatment" and "Environmental Impacts of Nanomaterials", provided 222 and 63 relevant articles, respectively, for the 10 year time period starting in 2010 (Figure 1). As seen in this graph, the number of studies on the former topic, that is nanomaterials for water treatment, has increased significantly over time during the past ten years. In contrast, the second topic, that is the environmental impacts of nanomaterials, has been less extensively investigated. This fact suggests the need to dedicate further time and cost investment for the purpose of uncovering the potential impacts and risks of nanomaterials on the environment that have been somewhat overlooked.

**Figure 1.** Annual number of publications in the last decade on the main keywords of the current review, according to ISI web of Knowledge (7 April 2020).

#### **2. Nanomaterials for Water Purification**

As mentioned earlier, nanomaterials offer several advantages for water treatment and control. This amazing potential stems from their large exposed surface area and functionality that can maximize their interactivity with water pollutants. In this section, we will take a glimpse on some important applications of nanomaterials at the service of water remediation.

#### *2.1. Nanomaterials for Adsorption and Photodecomposition*

In the water treatment field, the removal of dye pollutants due to their acute toxicities and carcinogenic nature is of paramount importance. Dyes have a history of thousands of years of application for textiles, paints, pigments, etc. Currently, almost 100,000 types of dyes are produced commercially. In terms of consumption volume, approximately 1.6 million tons of dyes are consumed annually. Thereof, 10–15% are wasted during use [7]. The dye pollutants released from industrial and agricultural wastes are refractory and potentially present carcinogenic effects. Therefore, they must be excluded from water streams through different kinds of traditional treatments, such as activated sludge, chemical coagulation, adsorption, and photocatalytic degradation [8]. Superior to the mentioned techniques, adsorption is relatively effective in the creation of a high quality effluent with no harmful byproducts in an energy/cost efficient manner [9–11]. This approach allows for exclusion of soluble and insoluble organic, inorganic, and biological water contaminants. The diverse merits of adsorption for dye removal are convincing enough to devise sustainable adsorbents that are manufactured on a large scale at low cost and enable fast and efficient dye removal. For this sake, within the course of the past few decades and with the evolution of nanotechnology, a variety of adsorbents of nanoscale size have been scrutinized. Nanomaterials provide an extensive reactive surface area at a low mass and versus activated carbon, i.e., the golden benchmark of adsorbents, they can be produced in a less expensive manner while removing dyestuffs and organic pollutants with a notably less amount [2]. Some examples of dye nano-adsorbents are as follow. Chitosan-coated magnetite (Fe3O4) nanoparticles showed a large adsorption capacity for crocein orange G (1883 mg/g) and acid green 25 (1471 mg/g). Interestingly, the adsorbent nanoparticles could be readily recovered by a magnetic field [12]. Based on such a concept, Fe3O4/activated carbon nanoparticles (6–16 nm) that can separate 138 and 166.6 mg/g methylene blue (MB) and brilliant green dyes, respectively, have been developed [13]. Dhananasekaran et al. [14] synthesized α-chitin nanoparticles (<50 nm) from Penaeus monodon shell waste and tested their dye (methylene blue (MB), bromophenol blue (BPB), and Coomassie brilliant blue (CBB)) adsorption

efficiency. The nanoparticle adsorbent showed an adsorption efficiency of 95.96–99%, depending on the adsorbent concentration and based on physical adsorption of the dyestuff to the nanoparticles. Other than nanoparticles, nanofibrous adsorbents have also found application in the removal of dye pollutants from water. Such adsorbents are typically made through electrospinning and are superior to nanoparticulate adsorbents due to their easy recovery. As an example for nanofibrous adsorbents, polyethersulfone (PES) electrospun nanofibers containing V2O5 nanoparticles have been employed for removal of MB dye pollutant from water [15]. The nanocomposite nanofibers show a low isoelectric point thus at elevated pHs they acquire an extensive highly hydroxylated surface area that facilitates adsorption of cationic MB molecules.

One drawback of adsorption is its inability to completely "eliminate" or "decompose" the dye pollutants. Instead, it solely collects the dye molecules by transferring them to other phases. This feature could be challenging due to the fact that discharge of the dye-related sludge is not straightforward and the adsorbent is rarely reusable [4]. Accordingly, there is a need to complementary degrading treatments such as photocatalysis, sonocatalysis, and reductive degradation that enable decomposition of dye to non-toxic metabolites. In this regard, a variety of advanced oxidation processes (AOPs), provoking release of hydroxyl radicals (OH•), have shown a promising potential for decolorization of textile effluents. With unpaired electrons, OH• is drastically reactive and oxidizes recalcitrant organic pollutants [16]. Due to the abundance of low cost, while operative photocatalysts, photocatalysis is indeed of the most researched AOP processes and is considered as a practical degradation process for organic dyes and pesticides. Various semiconductor metal oxide nanoparticles such as ZnO and TiO2 have shown notable efficiency in the photodecomposition of dye pollutants. For instance, Li et al. [17] developed an oil-in-water Pickering emulsion (PE) stabilized by the presence of TiO2 particles, wherein the dye containing wastewater and insoluble organic matter were regarded as the water and oil phases, respectively. The TiO2 particles could offer a large photoactivity effect and notably degrade the dye molecules (Figure 2a). In another relevant study, Kheirabadi et al. [18] synthesized a ternary nanostructure composed of Ag nanoparticle/ZnO nanorod/3-dimensional graphene (3DG) network via a coupled hydrothermal-photodeposition technique. While the 3DG can capture 300 mg/g MB dye by an adsorption process, Ag/ZnO component brings about the possibility of photodecomposition of the dye even under visible light irradiation. The dye removal mechanism of the synthesized adsorbent/photocatalytic system is illustrated in Figure 2b.

**Figure 2.** (**a**) Schematic illustration of an oil/water Pickering emulsion (PE) consisting superhydrophilic TiO2 particles enabling dye photodecomposition. Reproduced with permission from [17]. Copyright 2019, Elsevier. (**b**) The dye removal mechanism of an adsorbent/photocatalyst system comprising Ag nanoparticle/ZnO nanorod/3D-graphene hydrogel. Reproduced with permission from [18]. Copyright 2019, Elsevier.

Despite the high potential of semiconductor materials for photodecomposition of various organic pollutants, commercial visible light photocatalysts suffer from poor stability or inefficiency upon irradiation. Addressing such challenges, group II–VI semiconductors whose energy gaps cover the visible light spectral range have been proposed as superior, compatible alternatives [3]. Large aggregation tendency, challenging separation and recovery are the other bottlenecks that have

hampered the broader application of photocatalytic nanoparticles on a large, industrial scale [19,20]. One promising approach to overcome the abovementioned cons could be the nanocomposite strategy. By hybridizing the photocatalytic nanoparticles with polymeric nanofibers, not only is the large availability of the nanoparticles to the neighbouring water medium preserved, but also their intensive agglomeration is hampered and their recovery is facilitated. Nevertheless, due to the different polarity of the photocatalytic nanoparticles and polymeric nanofibers, the hybridization is not straightforward and can lead to clustering of the inorganic nanofillers in the polymer host [21]. In this regard, sol-gel treatment has shown promise in formation of tiny, isolated nanoparticles throughout the polymer nanofiber [22]. Even so, another important concern arises when recalling the possibility of photodecomposition of the encompassing polymer layer induced by the presence of photocatalytic nanoparticles [23]. Figure 3a–d show SEM images of the surface of TiO2/PVC composite films after UV irradiation for different durations. According to these images, photodegradation of the PVC matrix is initiated from the PVC–TiO2 interface and results in creation of cavities around TiO2 particle aggregates that grow and in fact coalesce over time during the irradiation [24]. This behavior is also seen in nanocomposite nanofiber systems. For instance, TiO2/PES nanofibers undergo photodegradation after exposure to UV irradiation, as reflected in their thermal and mechanical properties (Figure 3e,f, respectively [23]). Induced by UV irradiation, the electron/hole pairs formed in the conduction band (CB) and valence band (VB), respectively, react with O2 and thereby create various active oxygen species such as O2 <sup>−</sup>, 1O2, .O2H, and .OH [24]. In the next step, these active oxygen species start the degradation process and attack the neighbouring polymer chains in the surface and later in the polymer bulk and deeper regions. When the carbon-centered radicals diffuse into the polymer chain, their successive reactions end up with the chain scission with the oxygen incorporation and CO2 release [24].

**Figure 3.** SEM images show the surface morphology of the TiO2/PVC (1.5 wt.%) composite films after different irradiation times of (**a**) 0 h; (**b**) 25 h; (**c**) 50 h; (**d**) 100 h. Reproduced with permission from [24]. Copyright 2001, Elsevier. (**e**) Less notable increment of the glass transition temperature for the UV-irradiated TiO2/PES nanocomposite nanofibers versus the non-irradiated ones. (**f**) The dynamic thermomechanical (DMTA) analysis implies that storage modulus for the nanocomposite nanofibers drops upon UV-irradiation. Reproduced with permission from [23]. Copyright the authors 2019, assigned to MDPI under a Creative Commons Attribution (CC BY) license.

#### *2.2. Nanomaterials for Membrane-Based Water Treatment*

A membrane is a selective barrier located between two homogenous phases that splits a feed water stream into a retentate and a permeate fraction. The pressure difference between the feed and permeate sides acts as the driving force for the membrane's action and passes water through the membrane [25]. As a result, based on charge, size, and shape, solutes and particles are discriminated (Figure 4a). The new generation of membrane technologies employ nanomaterials for water treatment.

Nanocomposite membranes comprising a thin polymeric film surface decorated or incorporated with nanofillers are a distinguished class of membranes able to dynamically purify water. Composite materials possess a favorable package of properties that are derived from a combination of encompassing medium's and filler's properties [26–36]. These properties are not restricted to the classic and predefined ones, rather new properties and functionalities arise, especially, when the filler's dimensions lie in the nanoscale. Nanomaterials in different forms and dimensionalities can be used in construction of nanocomposite membranes. Nanoparticles, for instance, have been widely used as nanofillers for mechanical reinforcement or for hydrophilization of polymeric membranes. In this regard, Rodrigues et al. [37] incorporated clay nanoparticles into mixed matrix polysulfone ultrafiltration membranes to improve thermomechanical properties and water permeability of the membrane, while maintaining optimum rejection efficiency. Moreover, the membranes reinforced with clay nanoparticles showed a lower fouling tendency and higher flux recovery when tested with sodium alginate and natural water. One critical concern regarding ultrafiltration (UF) membranes is their biofouling and the presence of bacterial colonies on the surface and thereby clogging the pores and lowering the permeability of the membrane. In this regard, *extracellular polymeric substances* (EPS) are released upon bacterial cell lysis and are adsorbed on the UF membrane and thus reduce the longevity and permeability of the membrane [38,39]. The most promising solution to address the challenge of biofouling of the UF membranes is surface hydrophilization by incorporation of various antifouling agents [40]. In this regard, a diverse range of antifouling agents has been employed in membrane technology, including Ag, Au, Cu, graphene oxide (GO), Zn, and TiO2 nanoparticles [22,23], and also carbon nanotubes (CNTs) [38]. Despite the significance of industrial production of such nanocomposite membranes for water treatment, their toxicity that could originate from the release of the incorporated nanomaterials during the high pressure difference-driven filtration process should be carefully evaluated to minimize their adverse effects on human health and the environment [41]. The toxicity profile of the nanoparticles embedded in a polymeric matrix could be a function of their size, shape, charge and preparation conditions [38]. Among the nanofillers above mentioned, CNTs are resilient antibacterial agents whose toxic effect is derived from the ions and reactive oxygen species (ROSs) they release and thereby kill bacteria through oxidative stress stimuli [42]. Such a remarkable performance has led to wide application of CNTs in blended UF membranes, for the sake of improvement of filtration performance [43]. As reported in many studies, CNTs optimize water filtration and rejection of salts, nonpolar contaminants, micro- and macro-sized contaminants, and also waste chemical materials [44]. Ayyaru et al. [38] synthesized CNT- and sulphonated CNT (SCNT)-blended UF polyvinylidene fluoride (PVDF) membranes. For the latter group of the membranes, bovine serum albumin (BSA) rejection was 90%. As shown in Figure 4b, flux decline was less notable for the SCNT-PVDF membrane while permeating BSA solution through the membranes, thanks to its improved hydrophilicity. For the CNTand SCNT-PVDF membranes, the fouling recovery ratio (FRR) was 72.74 and 83.52%, respectively, Figure 4c, implying their optimum antifouling effect arisen from –SO3H and –OH groups found in SCNT and CNT, respectively. According to Figure 4d, the irreversible fouling value of the nanocomposite membranes, particularly that of SCNT-PVDF, is lower than that of the neat PVDF membrane. This again emphasizes the role of hydrophilicity induced by the presence of the nanofillers on lowering the fouling tendency of the membranes. Although CNTs are potentially versatile additives to membranes and also promising adsorbents for divalent metal ions, dyes, natural organic matters, etc., their relatively high unit cost is a limiting factor for their widespread practical use [3]. Moreover, the existence of metal catalysts in raw CNTs might induce a toxic effect. In contrast, chemically functionalized CNTs have not been shown yet to be toxic [45]. Accordingly, practical applicability of CNTs as adsorbents or inclusions in membranes for water treatment is tightly associated with finding cost effective production methods for CNTs and minimizing their toxicity effect by development of safer alternatives such as carbon nanocrystals (CNCs) [3].

**Figure 4.** (**a**) Schematic shows diverse membrane filtration processes including reverse osmosis, nanofiltration, ultrafiltration, microfiltration, and particle filtration. Other than the reverse osmosis membranes whose structure is almost dense and non-porous, the rest are different in terms of the average pore size (the image was obtained from www.muchmorewater.com). Antifouling properties of CNT/PVDF UF membranes represented in cycle filtration (**b**), water recovery flux (**c**), and fouling resistance (**d**). Reproduced with permission from [38]. Copyright 2019, Elsevier.

Other than film-like membranes widely used for ultrafiltration and nanofiltration, electrospun nanofibrous membranes have also been developed for size exclusion of contaminants. Such nanostructured membranes possess a large porosity and an interconnected porous structure with microscale pore sizes. While the high porosity assures notable permeability to gas and liquid streams, the interconnected pores enhance fouling resistance. Such features lead to low energy consumption for the membrane process. Moreover, extensive available surface area and flexibility in surface functionality optimize the adsorptive nature and selectivity of the nanofibrous membranes. The constituting nanofibers could be as neat, chemically functionalized, nanocomposite, and even biofunctionalized [46–50]. For instance, a PES electrospun nanofiber mat overlaid on a poly(ethylene terephthalate) (PET) non-woven was evaluated as a membrane for liquid filtration and removal of micro- and submicron sized polystyrene (PS) particles from water [51]. Despite a high initial flux, upon increase of the feed pressure, the nanofibrous membrane's porosity declines and thereby the water flux drops. Therefore, while the nanofibrous membrane shows a high potential for pre-treatment of water e.g., as a microfiltration (MF) membrane, it should be mechanically reinforced and hydrophilized to raise water permeability and flux. To meet such objectives, PES nanofibers were stabilized and hydrophilized by incorporation of ZrO2 [52] and TiO2 [22] nanoparticles. Nanofibrous membranes can also be used for UF, i.e., a particular liquid filtration process separating a variety of pollutants, such as viruses, emulsions, proteins and colloids that are as small as 1–100 nm. For this sake, a nanofibrous membrane needs to possess a surface pore size less than 0.1 μm that alongside a high surface area potentially render them prone to rapid fouling and notable flux decline. To address this concern, a nanofibrous membrane is coated with a thin film and makes up a thin film composite (TFC) membrane [53]. Such a concept has shown applicability for forward osmosis (FO) membranes, as well [54].

Self-sustained hydrophilic nanofiber supports have been investigated for construction of the TFC FO membranes [55]. With a particular scaffold-like structure, the nanofiber support optimally lowers the internal concentration polarization (ICP) and raises water flux. To address the challenge of biofouling, the nanofibers could be equipped with antimicrobial properties, as well. For this sake, antibacterial nanoparticles e.g., Ag nanoparticles can be incorporated within the nanofibers. This strategy has been previously applied for various applications with respect to wound dressings [56] and water filtration [57]. For FO water treatment, Pan et al. [55] synthesized a TFC FO membrane

based on an antibiofouling Ag nanoparticle-incorporated nanofibrous support layer, that could offer an improved water flux and reduce biofouling and ICP, Figure 5a–e. The as-formed FO membrane provides a remarkable bactericidal effect for *E. coli* (96%) and *S. aureus* (92%), thanks to release of Ag+-species into the solution.

**Figure 5.** (**a**) Schematic illustration of the structure of the antimicrobial nanofiber supported TFC FO membrane (Ag/PAN-thin film nanocomposite (TFN)) (left) and the antibacterial action of the Ag nanoparticles incorporated in the nanofibers that damage DNA and membranes of the bacteria (right). (**b**) Water flux of the Ag/PAN-TFN versus that of commercial FO membranes and (**c**) water flux trend of the Ag/PAN-TFN FO membranes over time (in the FO and PRO modes). (**d**) Reverse salt flux and (**e**) specific salt flux of the Ag/PAN-TFN membrane compared to those of commercial FO membranes (in the FO and PRO modes). Note that the experiments were performed using 0.5 M NaCl as draw solution and DI water as feed solution; In the FO and PRO mode, the active layer face feed and draw solution, respectively. Reproduced with permission from [55]. Copyright 2019, American Chemical Society.

As shown in Figure 5a, the Ag ions are able to (quasi)covalently bond with thiols, phosphates, and organic amines available in proteins, lipopolysaccharide, and phospholipid of the cell membrane, and cell wall, thereby damaging them. Moreover, some ions could pass through the cell wall and adversely influence ribosomal subunit proteins and enzymes [58] and disrupt DNA's structure, leading to cell death. Other than the bactericidal activity, the hydrophilic, porous nanofibrous support allows for a superior water flux compared to commercial FO membranes (HTI-CTA and HTI-TFC) in two modes of FO and pressure-retarded osmosis (PRO, Figure 5b). The water flux remains steady over time in the FO mode, whereas it declines in the PRO mod due to dilution of the draw solution, leading to a lower osmotic pressure difference (Figure 5c). Regardless of the operation mode, the nanocomposite membranes show an increased reverse salt flux versus the commercial FO membranes (Figure 5d). In contrast, as shown in Figure 5e, the nanocomposite membranes exhibit a much less specific reverse salt flux, implying their higher selectivity and efficiency for a FO process.

Graphene, i.e., a 2D, 1-atom-thick planar sheet of *sp*<sup>2</sup> bonded carbon atoms, possesses remarkable physical, mechanical, thermal and optical properties [59]. In relation to water remediation, graphene's atomic thickness can potentially guarantee a high fluid permeability (that is significantly larger than that of typical nanofiltration (NF) membranes) and therefore lower energy consumption and inexpensive operation. Moreover, the 2D nanochannels forming between stacked graphene sheets or the nanopores available in a single graphene layer enable size-selective transport and purification of water streams (Figure 6a) [60]. The graphene that has been employed for development of water desalination membranes can be in different forms, such as pristine graphene, graphene oxide (GO) and reduced GO (rGO). Also, structurally graphene membranes can be constructed either as single layers or as stacked, multilayer forms. Monolayer graphene membranes with intrinsic pores have been studied for NF purpose experimentally as well as theoretically via simulation. For instance, O'Hern et al. [61] mounted a monolayer of chemical vapor deposition (CVD) graphene onto a porous polycarbonate substrate and thereby fabricated a graphene composite membrane with an active filtration area of 25 mm2. The graphene monolayer contained intrinsic nanopores as small as 1–15 nm that could contribute to the size-selective passage of molecules such as KCl, tetramethylammonium chloride, Allura Red AC (496 Da dye) and tetramethylrhodamine dextran (70 kDa), through the membrane made thereof. While KCl and tetramethylammonium chloride permeated through the graphene membrane, the diffusion of tetramethylrhodamine dextran was hampered. Despite the feasibility of selective molecular transport through the monolayer graphene membrane, selectivity is not controllable thanks to arbitrary sizes and locations of the intrinsic pores. Conclusively, formation of graphene layers with a large number of nanopores with adjusted, near monodisperse sizes and chemistries is a sophisticated objective that must be targeted in the new generation of graphene membranes.

**Figure 6.** (**a**) Schematic illustration of the water purification mechanism for a single layer graphene membrane containing nanopores of tailored size and a multilayer graphene membrane comprising stacked GO sheets. Reproduced with permission [62]. Copyright 2018, Elsevier. (**b**) Schematic depicts the nanochannels forming between adjacent GO sheets that comprise hydrophobic and hydrophilic zones facilitating water flow and removal of tiny water pollutants. Reproduced with permission from [63]. Copyright 2016, American Chemical Society.

Despite various advantages of the monolayer graphene membranes for nanofiltration and even water desalination, large scale production of nanoporous graphene is indeed challenging and a single layer graphene is not robust enough to withstand the usual filtration pressures. In contrast, multi-layered GO membranes can be produced in a scalable manner and survive under large applied pressures. The nanochannels formed between the stacked GO nanosheets decorated with various polar functional groups allow water permeate through the membrane [64]. Induced by the notable slip length of water within the interlayer channels, the stacked GO inhibits passage of the solute particles. Figure 6b schematically shows how water molecules get into the hydrophilic zones and slip through the hydrophobic nanochannels. With respect to selectivity, because of hydration, the interlayer spacing of the GO nanosheets rises to 0.9 nm upon their immersion in ionic solutions. This structural change enables permeation of K<sup>+</sup> and Na<sup>+</sup> ions, and disqualifies the membrane for desalination purposes [65].

Despite the promising potential of graphene membranes for water purification, their release into water and thereby the environment is a concern that should be taken into account. In this regard, there are several reviews in the literature that widely discuss the fate, transformation, and toxicological impacts of such nanomaterials in the environment [66–68]. However, there is still a need to realistic, long term determination of the environmental implications of graphene nanomaterials. These precise ecotoxicological and life-cycle analyses enable us to better judge pros and cons of graphene nanomaterials and to find out how we can employ the safest ones with the least health and environmental concerns.

As highlighted so far, there are many promising achievements in relation to employment of nanomaterials for water treatment. Some of the recent developments (as of 2019) in this field are tabulated in Table 1. Many nanoadsorbents and nanomembranes are currently in development stage for the sake of large-scale production and industrialization. Given the fact that the water recycling is regarded an important aspect of sustainable development in the human communities, particularly when recalling the expanding water shortage crisis across the world, significance of realizing advanced water treatment technologies using nanomaterials is further stressed. However, this tiny functional

building blocks could be also problematic and harmful in terms of sustainability, if released into the environment in an uncontrolled manner. In the next section, we review how they would be scattered in different media of water and soil and how they influence the biota living in such systems.


**Table 1.** Some examples of recent studies (as of 2019) on nanomaterials for water treatment.

NP: nanoparticle; CNT: carbon nanotube; PVDF: polyvinylidene fluoride; Co: cobalt; MO: methyl orange; AC: activated carbon; RhB: Rhodamine B; MOF: metal-organic framework; CR: congo red; PANI: polyaniline; BG: brilliant green; AB: acid blue; AY: acid dye; PES: polyether sulfone; KGM: Konjac glucomannan; PVA: polyvinylalcohol; EPVC: emulsion polyvinyl chloride; BSA: bovine serum albumin; MB: methylene blue; PEI: poly(ether imide); PD: polydopamine; HA: humic acid; CV: crystal violet; PA: polyamide; PAN: polyacrylonitrile; PS: polysulfone; PEG: polyethylene glycol; CS: chitosan.

#### **3. Ecotoxicology of Nanomaterials**

As an inevitable result, with extensive use in water treatment, nanomaterials can be released into soils and water bodies, thereby threatening the quality of life of humans, animals, and plants (Figure 7a). In the case of uncontrolled release, they are accumulated and as suspended solids contaminate food and drinking water. Their final destiny strongly depends on their properties as well as the characteristics of the environment they are released into. Consequently, a variety of adverse ecotoxicological impacts can take place on microorganisms, plants, invertebrates, and fish. In the case of human, no significant risk from such nanomaterials has been reported, though the relevant studies performed so far are insufficient [87]. It is extremely necessary to research on the potentially hazardous effects of the nanomaterials on ecosystem and human health. This will encompass quantitative and qualitative assessment of such nanomaterials in different segments of the environment and determination of the likely consequences on the health of various species living in that segment. Accordingly, the research must be directed towards identification and testing of the environmental fate and transport, and ecotoxicology and toxicology of the nanomaterials.

**Figure 7.** (**a**) The cycle of release of nanomaterials into environment from water to soil and vice versa. (**b**) The aquatic colloid comprising hydrous iron oxide aggregates and organic fibrils (scale bar = 1 μm). (**c**) The aquatic colloid made form silver stain on organic fibrils along with the polysaccharide fibrils possibly decorated with hydrous iron oxides (scale bar = 200 nm). (**d**) The likely cytotoxicity mechanisms of nanomaterials exerted on bacteria. CYP = cytochrome P. Reproduced with permission from [5]. Copyright 2009, John Wiley and Sons.

Among nanomaterials, carbon and metal oxide ones (e.g., TiO2) are hardly biodegradable and persist in the environment [88,89]. It is anticipated that many of such insoluble nanomaterials would aggregate (as homo- or heteroaggregation) in the ecosystem and eventually settle out [90,91] and be subjected to various species as sediments. With respect to the soluble nanomaterials, the dissolution rate and solubility are determining in their fate. The release of toxic ions and generation of reactive oxygen species (ROS) are considered as the main toxicological pathways for metal/metal oxide nanomaterials, thereby threatening the life quality of aquatic and terrestrial creatures [5,92,93].

#### *3.1. Nanomaterials in Aquatic Systems*

As a general rule, the nanomaterials released into the environment interact with different substances available in that medium and thereby experience aggregation, dissolution, sedimentation, and transformation [94]. In 2010 it was reported that from 8 to 28% of the released nanomaterials are accommodated in soil, while between 4 and 7% in water [95]. These statistics are associated with the higher potential of soil (i.e., its constituents) and its interaction level with the liberated nanomaterials.

The nanomaterials are released into different ecosystems including the aquatic ones either intentionally or unintentionally. While the nanomaterial can be on purpose added to address the available contamination in the groundwater [96], they can also originate from the atmospheric emissions and solid/liquid waste streams delivered from factories, for instance. Furthermore, the nanoparticles present in paints, textile, and health care products, e.g., sunscreens and cosmetics, can potentially be released into the environment. As an example, thanks to desirable bactericidal ability, Ag nanoparticles are commonly employed in the consumer products (e.g., to inactivate the odor-causing bacteria in socks). However, their release into water streams can kill also useful bacteria that exclude ammonia from wastewater treatment systems [97]. Metal oxide nanoparticles, such as TiO2, ZnO and CeO2, are also extensively utilized in diverse products (e.g., sunscreens, paints, coatings, catalysts) and can reach water systems in different ways. The released nanoparticles are settled on soil and surface water systems, and in case of proper surface treatment they remain un-aggregated and float on the water. The nanoparticles that are deposited on the land, contaminate soil, transit through the surface, reach the ground water and in the course of this travel affect the existing biota. Wind and rainwater stream also displace the nanoparticles available in the solid wastes, wastewater effluents, direct or uncontrolled emissions into water bodies. Though accidental release of the nanomaterials can be strictly controlled within the production units, likely spillage during transportation to the consuming units is indeed challenging [5].

The majority of released nanoparticles aggregate as soon as hydrated, thereby being sediment in different rates. The extent of aggregation depends on the nanoparticle's surface charge and charge magnitude, and likely coverage of the nanoparticle's surface by mono- and divalent cations, by natural organic matter (NOM) or other organic molecules. In fact, such features determine the prevail of attractive and repulsive forces, thereby governing the nanoparticles' aggregation or their sticking to other surfaces [89] e.g., aquatic colloids.

Aquatic colloids, according to the definition established by the International Union of Pure and Applied Chemistry (IUPAC), are the materials whose dimension(s) are below 1 μm. Accordingly, aquatic colloids encompass (natural) nanoparticles, i.e., the materials with a minimum of two dimensions larger than 1 and smaller than 100 nm. In terms of composition, aquatic colloids are composed of organic (mainly humic substances and protein and polysaccharide fibrils) and inorganic fractions (e.g., metal oxides of Fe, Mn, and Al, and silicon oxide) and also microorganisms including viruses and bacteria (Figure 7b,c). It is very likely that nanoparticles (nanomaterials) interact with the aquatic colloids and thereby aggregate and sediment. This means that fate of the nanoparticle can be dominated by characteristics and concentration of aquatic colloids. For instance, it is reported that in estuarine and marine waters, density of aquatic colloids is notably low, on the other hand, concentration of nanoparticle is also low due to the high aggregation tendency and thus increased sedimentation rate at the aqueous systems with high ionic strengths [5]. It is worthy to note that the concentration of the nanoparticles commonly found in natural waters including those made of Ag and oxides of Ti, Ce, and Zn could range from 1 to 10 μg/L, and cumulatively, this concentration can reach 100 μg/L [5]. With respect to the interaction between the aquatic colloid and nanoparticles, humic substances coat the surface of the nanoparticle [98] and stabilize their surface charge, thereby minimizing the chance of aggregation [99], as shown for CNTs, for instance [100]. On the other hand, the fibrils raise this possibility via bridging mechanisms [101]. The main factors or properties of the nanoparticles that significantly influence their behavior in natural water systems are: chemical composition, mass, particle density, surface area, size distribution, surface charge, surface contamination (the likely shell and capping agents), and stability and solubility of the nanoparticle [5].

Given that industrial discharges are mostly exposed to marine environments, and the freshwater streams and coastal runoff end up to seas, this medium and its contamination is of utmost importance. The sea environment has a high ionic strength, is typically alkaline, and contains a diverse range of NOMs as well as colloids whose type and concentration depends on the location (coastal zone versus oceanic one). This medium can potentially be contaminated by the nanoparticles that are released via atmospheric deposition and/or coastal runoffs. The physicochemical properties of water such as temperature, salinity, and type of NOM varies by depth and affects the aggregation and colloid formation [102]. Similar to freshwater systems, the formed aggregates of the nanoparticulate contaminants precipitate slowly down to the ocean floor. In this route, they may stop at the interface of cold and warm streams or even be recycled by the existing biota. In either cases of complete sedimentation on the ocean floor or getting stuck in the mentioned interface, the living species corresponding to the zones might be affected. Additionally, the nanoparticles could be suspended and trapped in the surface microlayer of oceans and thereby impose risks to the birds, mammals and the species living in the microlayer [103].

The uptake of nanoparticles by living organisms can induce toxicity effects through different mechanisms. Nanoparticles can find a way into the cells by penetration through cell membranes, endocytosis as well as adhesion [104–106]. As soon as the nanoparticle is accommodated in the cell, it starts to damage the natural functions in different ways such as destruction of membrane structure or potential, proteins oxidation, genotoxicity, blockage of energy transduction, and generation of ROS and toxic substances [5]. Such mechanisms are illustrated schematically in Figure 7d. For instance, graphene is toxic to bacteria and damages the membrane and raises the oxidative stress level. In the graphene family, GO shows the most notable antibacterial activity, followed by rGO and graphite [107].

#### *3.2. Nanomaterials in Terrestrial Systems*

Given the large availability of a reactive sink, that can lead to overestimation of the exerted dose to the biota relative to the real one, soil is notably distinct from fresh and marine waters. In a similar manner, soil can be contaminated with nanomaterials intentionally (for the purpose of remediation, fertilizing, etc. e.g.) or unintentionally (by uncontrolled spills in the course of production and transport, e.g.) [5,108]. Regardless of the application aim, they can affect the biota present or dealing with soil.

Upon contacting soil, nanomaterials are physicochemically adsorbed to the soil particles' surface through hydrophobic interactions, hydrogen bonding, electrostatic interactions, etc. Induced by the presence of organics, they could experience chemical transformations. Additionally, nanomaterials can penetrate into the pores of macroparticles and stay there for a long time [109]. The behavior of nanomaterials i.e., their retention or mobility, in soil is dependent on several factors including soil texture, pH, humic acid, and chemistry of soil, surface coating and nanomaterial size, and pore water velocity. Particularly, pH and humic acid of soil determine the aggregation and colloidal stability of nanomaterials in soil [110].

Soil is a medium that accommodates diverse species such as microorganisms, plants and nematodes. These soil inhabitants are crucial in the cycle of nutrients, decomposition of materials, and nitrogen fixation [110]. Therefore, any damage to such important ecosystem elements can have non-compensable consequences on the life quality and nature.

Microorganisms are vital for performance and health of soil, taking into account their role in modulation of the organics' turnover and the cycle of mineral nutrients. They also play a crucial role in the physical characteristics of soil, thereby influencing water maintaining potential and the tendency of compaction or erosion. Accordingly, any undesired effect on microorganism community from the released nanomaterials could indirectly affect soil's quality and function. A diverse range of nanomaterials including fullerenes (C60 and nC60) [111], 3-aminopropyl/silica, palladium, dodecanethiol/gold and copper nanoparticles [112] have been challenged with respect to their effect on soil's microbial community. In this regard, specifically, the factors such as soil respiration, microbial biomass, phospholipid fatty acid quantity, methyl ester of fatty acids, enzymatic activities, colony forming unit and DNA profile of bacterial community have been analyzed. While the above cited reports imply no toxicity of the mentioned nanomaterials, there are other studies such as that of Johansen et al. [113] on nC60 that explicitly demonstrate a notable decline in bacteria density of the soils exposed to this kind of nanomaterial. Nogueira et al. [108] also showed that the bacterial communities in soil samples treated with gold nanorods, TiO2 nanoparticles, the polymeric nanomaterials composed of carboxylmethyl-cellulose (CMC), the hydrophobically modified CMC (HM-CMC), the hydrophobically modified polyethylglycol (HM-PEG), and the vesicles of sodium dodecyl sulphate/didodecyl dimethylammonium bromide (SDS/DDAB) are notably influenced by their toxicity effects. According to Rodrigues et al.'s study [114], bacteria can survive when exposed to SWCNTs, while fungal microbiota are unable to recover after the exposure. In another study, it was shown that FeO and Ag nanoparticles can decrease fungal biomass [115]. In contrast, quantum dots and super paramagnetic nanoparticles impose no notable toxicity to *Fusarium oxysporum* [116]. In general, the nitrifying bacteria are more vulnerable to nanomaterials' toxic effects compared to the bacteria with nitrogen fixing and denitrifying ability [110]. In the case of the ammonia oxidizing bacteria, several nanomaterials such as TiO2 nanoparticles have shown a toxicity effect [117].

Plants are the other main species that are influenced by the presence of nanomaterials in the soil. Several kinds of nanomaterials have shown the ability to diffuse into plants via their roots, translocation, biotransformation, and spread in different forms across their structure, thereby impacting on photosynthesis, growth and regeneration abilities [118,119]. The resulting perturbation of the physiological functions impacts on seed germination, seedling growth, higher ROS production, damage to cell walls, and changes in proteins, carbohydrates, lipids, pigments, and hormones [120]. In this regard, metal oxide nanomaterials have been widely used as fertilizers for agriculture [121]. These nanoparticles account for a major number of the nanoparticle contaminants in soil including: Al2O3, TiO2, CeO2, ZnO, CuO and ZrO2. The other important nanocontaminants are carbon fiber, SiO2 and Ag nanoparticles, as well as carbon black [122]. Of the aforementioned metal oxide nanoparticles, ZnO and CeO2 are especially detrimental for the plants [121]. As shown by Priester et al. [123] soybean (*Glycine max*) is vulnerable to high amounts of CeO2 nanoparticles in soil, as witnessed by large ROS production, lipid peroxidation, visible stains (Figure 8a,b), and also declined total chlorophyll amounts. Compared to the leaves exposed to ZnO nanoparticles, the ones treated with CeO2 nanoparticles show a higher percentage of visible damages, Figure 8c,d.

Feizi et al. [124] examined the effect of TiO2 nanoparticles in different concentrations of 0, 5, 20, 40, 60 and 80 mg.L−1, on the fennel seed germination. According to their results, after 2 weeks of seed incubation with TiO2 nanoparticles (60 mg.L−1), there was an increase in the total germination percentage. Other reports also show a rise in seed germination for tomato and rice when exposed to SiO2 nanoparticles [125] and CNTs [126], respectively. It is worthy to note that concentration of the nanomaterial is an influential factor on the plant reaction and various plants might react differently to a single specific nanomaterial. It is imaginable that nanomaterial undergoes transformation within the plant and the resulting product can potentially impose a further risk or even be beneficial for the plant growth [110]. Other than the impacts on the growth process of plants, nanomaterials can affect and raise the generation of ROS in a plant. The high concentration of ROS can engender protein oxidation, DNA and cell membrane destruction, lipid peroxidation, electrolyte leakage, etc., thereby inducing oxidative damage and eventually cell death [127].

**Figure 8.** Camera images of soybean leaves without (**a**) and with (**b**) exposure to CeO2 nanoparticles. The percentage of the leaves whose damage is recognized visually after exposure to (**c**) CeO2 and (**d**) ZnO nanoparticles. Reproduced with permission from [123]. Copyright 2017, Elsevier.

One important point that must be kept in mind is that nanomaterials are normally transformed from origin i.e., the fabrication facility to destination. Thus, the results obtained in the studies based on pure nanomaterials could be not applicable to the realistic situations. Also, environmental factors such as pH, surface charge, ionic strength, surface coating, (UV) light irradiation, humic acids, inorganic ligands, mono- and divalent cations, affect the toxicity of nanomaterials in their destination [110]. With respect to the inclusion of nanoparticles into the terrestrial systems and their consequences, there are several reviews [110,128–131].

#### **4. Conclusions and Future Outlook**

Nanomaterials are favorable candidate materials for water remediation and control and an exciting prospect for their integration into point-of-use systems, and also in absolute removal of the current and emerging inorganic and organic pollutants from water is foreseen. They can be used in construction of nanoadsorbents that effectively capture polar and non-polar pollutants from water depending on their surface functionality. In this regard, the extensive surface area offered by nanomaterials is decisive and maximizes the adsorption efficiency. In case, the used nanomaterial is a photocatalyst, adsorption is extended to photodecomposition. Accordingly, the stuck organic pollutant is degraded into harmless byproducts and the adsorbent's surface becomes ready for a next round of adsorption/photodecomposition process. As another opportunity originated from nanomaterials, membrane nanostructures can be mentioned. Nanomaterials can be exploited as building blocks of a porous separator, as seen in electrospun nanofibrous membranes or single/few layer graphene membranes. Additionally, they can be used as additives to conventional thin film polymeric membranes for ultrafiltration to reduce fouling tendency and to raise thermomechanical

properties. Despite such merits of nanomaterials in filtration and water treatment industry, there are a variety of bottlenecks that must be properly addressed:


and pressure. The former nano-adsorbent needs UV irradiation to photodecompose the dye pollutants that adds to the expenses of the treatment. In fact, it is highly necessary to produce large amounts of such nanomaterials at justifiable costs for water treatments, specific to different categories of wastewaters.

(7) *Environmental hazards*: This concern will persist in the future. This stems from the reality that many environmental and biological consequences of nanomaterials should be identified in the long term. Short term studies have shown that several nanomaterials are safe to human being, plants and animals. But, there is no certainty about their long term safety. For this reason, establishment of nanomaterial based water treatment systems should be followed with sufficient precautions. Technologically, it is also vital to secure such systems so that the release of nanomaterials into environment would be miminized.

**Author Contributions:** M.G. and S.Z. contributed to idea development and to drafting of the manuscript. S.H. conceived and developed the idea and drafted the manuscript. All authors have read and agreed to the published version of the manuscript.

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

**Acknowledgments:** S.H. would like to acknowledge the financial support received from the European Union's Horizon 2020 research and innovation program under the Marie Sklodowska-Curie grant agreement No. 839165. Mady Elbahri is acknowledged for his support during the research.

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

#### **References**


© 2020 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 (http://creativecommons.org/licenses/by/4.0/).

## *Review* **Role of Nanomaterials in the Treatment of Wastewater: A Review**

#### **Asim Ali Yaqoob 1, Tabassum Parveen 2, Khalid Umar 1,\* and Mohamad Nasir Mohamad Ibrahim 1,\***


Received: 27 December 2019; Accepted: 8 February 2020; Published: 12 February 2020

**Abstract:** Water is an essential part of life and its availability is important for all living creatures. On the other side, the world is suffering from a major problem of drinking water. There are several gases, microorganisms and other toxins (chemicals and heavy metals) added into water during rain, flowing water, etc. which is responsible for water pollution. This review article describes various applications of nanomaterial in removing different types of impurities from polluted water. There are various kinds of nanomaterials, which carried huge potential to treat polluted water (containing metal toxin substance, different organic and inorganic impurities) very effectively due to their unique properties like greater surface area, able to work at low concentration, etc. The nanostructured catalytic membranes, nanosorbents and nanophotocatalyst based approaches to remove pollutants from wastewater are eco-friendly and efficient, but they require more energy, more investment in order to purify the wastewater. There are many challenges and issues of wastewater treatment. Some precautions are also required to keep away from ecological and health issues. New modern equipment for wastewater treatment should be flexible, low cost and efficient for the commercialization purpose.

**Keywords:** nanocatalysts; nanomembranes; nanosorbents; nanomaterial applications; waste water treatment; nanomaterial challenges

#### **1. Introduction**

Water is a natural source on the earth and its availability in pure state is very essential for human beings as well as for other living creatures because the concept of life is unbelievable without water. Water is also called universal solvent due to its potential properties like solubility power etc. Currently, major problem of whole world is water contamination, due to several reasons like inadequate sewage treatment, industrial wastes, marine dumping issues, radioactive waste material, some agricultural perspectives etc. [1,2]. Water pollution has an adverse effect on environment, and it can also responsible for air pollution that reflects very dangerous results on human health. Water pollution also carries an adverse impact on economic growth and socials perspectives of the concern societies/countries. Recently, a UN report stated that purified and freshwater availability is a global issue and become a challenge in 21st century for whole world because the survival of living creatures is not safe with contaminated water [3,4]. Contaminated water means that unwanted materials come into water bodies or reservoirs and make it unsuitable for drinking and other purposes. To overcome this emerging problem, there are many chemical, physical and mechanical methods. Moreover, researchers are still working by exploring different new technologies to improve water purification methods with low cost [5,6]. Newly, emerging field nanotechnology provides a potential offer to purify water with a low expense, high working efficiency in removing pollutants and reusable ability [7]. In past era, nanomaterials are successfully applied to several places like in a field of medical science, catalysis, etc. Recently, when the world facing serious issues of drinking water, experts found that nanomaterial is better option to treat wastewater because it has some unique properties like nano size, large surface area, highly reactive, strong solution mobility [8], strong mechanical property, porosity characters, hydrophilicity, dispersibility and hydrophobicity [9–11]. Some heavy metal like Pb, Mo, etc. organic and inorganic pollutants and various harmful microbes are reported to be successfully removed by using different nanomaterials [12–16]. Currently, WHO (World Health Organization) reported that almost 1.7 million people died due to water pollution and four billion cases of different health issues were reported annually due to waterborne diseases [17]. Table 1 indicates different kinds of water pollutants with sources and their adverse effects.


**Table 1.** Indicates different water pollutants with their sources and adverse effects.

Recently, there are more advance developments occurred in nanomaterials such as nanophotocatalysts, nanomotors, nanomembranes and nanosorbents (Nanosorbents contain high sorption capacity which carries many applications for water treatment methods) and some imprinted polymers are effectively useable for treatment process of contaminated water. In short, the study of nanomaterial applications in water purification is considered to assess positive perspectives [16]. Therefore, we have summarized the role of nanomaterials for wastewater treatment to overcome the water crises in this review article. Nanoengineered materials provide a potential and significant water treatment approaches which can be easily adaptable, but some imperfections still need for further attention which are specially summarized in this article. Moreover, we also addressed the limitations, advantages, disadvantages and future perspectives related to these nanomaterials. Furthermore, the toxicity of nanomaterials and their several applications in wastewater treatment are briefly discussed which might be useful for researcher to plan new strategies.

#### **2. Water Treatment Methodologies**

#### *2.1. Nanophotocatalysts*

The word "photocatalysis" is composed of two Greek words "photo" and "catalysis" which mean decomposition of compounds in the presence of light. Usually, in scientific world there is no consensus definition for photocatalysis [28]. However, this term can be employed to define a process to activate or stimulate the substance by using light (UV/Visible/Sunlight). Photocatalyst which changes the rate of reaction without any involvement by itself during the chemical transformation process. Furthermore, the key difference between traditional thermal catalyst and photocatalyst is that the prior is activated through heat while the final is activated through photons of light energy [29]. Nanophotocatalysts are commonly used for wastewater purification, as they help to enhance the reactivity of catalyst due to having a greater surface ratio and shape dependent features [30].

The nano size-based materials show different response as compared to bulk materials due to their distinct quantum effects and surface properties. It assists to increase their electric, mechanical, magnetic chemical reactivity and optical properties too [31]. It has been showed that the nanophotocatalysts can expand the oxidation ability due to effective production of oxidizing species at surface of material which helps in degradation of pollutants from the polluted water effectively [32].

Nanoparticles like zero-valence based metal, semiconductor and some bimetallic type etc. are mostly used for treatment of environmental pollutants e.g., azo dyes, Chlorpyrifos [33–35], organochlorine pesticides, nitroaromatics, etc. [36]. There are also several reports in literature which illustrated that TiO2 based nanotubes can effectively be used in removal of pollutants (organic pollutants such as azo dyes, Congo red, phenol aromatic base pollutants, toluene, dichlorophenol trichlorobenzene, chlorinated ethene, etc.) from waste water [37–41]. However, the most common and significant metal oxide nanophotocatalyst are SiO2, ZnO, TiO2, Al2O3, etc. [42–44]. Among them, Titanium dioxide (TiO2) is one of excellent photocatalyst from all existing material due to its several reasons such as low cost, toxic free property, chemical stability, and its easy availability on earth. Moreover, TiO2 exist in three natural states, anatase, rutile, and brookite. So far, anatase considered as good nanophotocatalyst material [45]. The bandgap of this state is 3.2 eV and it can absorb ultraviolet light (below 387 nm) [46]. However, other photocatalysts like ZnO have also been produced to eliminate contaminants in wastewater and presents reusable ability effectively [47–51]. In case of composite nanomaterial, the degradation of reference substance (dimethyl sulfoxide) for evaluating the photocatalytic performance of water treatment by using CdS/TiO2 composite as catalyst under the irradiation of visible light also investigated [52]. The iron doped nanomaterials have ferromagnetism ability which helps to recycle and reuse it easily [53–56]. Similarly, due to some characteristics like high photocatalytic reactivity Pd incorporated ZnO nanomaterial have been used for the removal of *Escherichia coli* from wastewater [57]. Although, new efforts have been targeted on metal oxides in order to increase the photocatalytic performance under visible light irradiation through modifying them with other elements like metals or metal ions [6,33] carbonaceous-based materials, dye sensitizers [58] and many others but still there is a need for further modifications in nanophotocatalysts.

Furthermore, nanophotocatalysis process may occur in two states heterogeneously or homogeneously. The most intensively studied state is heterogeneous nanophotocatalysis in modern era, due to its wide scope in water decontamination and environmental related applications. In case of heterogeneous photocatalysis, it implies the prior development of an interface between fluid (both reactants and products of reaction) and solid photocatalyst (such as metal or semiconductor) [59,60]. Generally, the word "heterogeneous photocatalysis" is mostly employed where a light-based semiconductor photocatalysts are used, which is in interaction with gaseous or liquid phase [61]. The heterogeneous photocatalysis based applications are strongly depend on the scaled-up reactor based on advance developed designs with improved efficiency [62]. The major task in reactor designing is effective illumination of nanocatalyst and mass transfer optimization, particularly in case of liquid phase. The transfer of photon can be improved by using light-emitting diodes and optical fibres,

but productive revolutions in this field are still lacking. Moreover, an extensive effort has been focused toward the progress of solar photoreactors [63–65]. According to literature, the positive role of nanophotocatalyst has been demonstrated in research laboratory for air cleaning and water treatment. At the commercial level it is still not a perfect way to minimize the problem. Moreover, the present lack of extensive commercial applications is due to absence of effective photoreactor configurations and lower photocatalytic competence of photocatalysts. Despite all, heterogeneous nanophotocatalyst recommend fascinating advantages i.e., inexpensive usage of chemicals, additive free, work even at lower concentration, chemical stability (e.g., TiO2 stable in aqueous medium) [66]. Therefore, recently heterogeneous photocatalysis is achieving the pre-industrial scale.

#### 2.1.1. Advantage and Disadvantages of Nanophotocatalyst

Nanophotocatalysis has reflected a vital role for the mineralization of dangerous organic substances at 25 ◦C and proved very effective and efficient method for detoxification of water with help of nanophotocatalysts [67]. Furthermore, mostly nanophotocatalysts show some advantages such as they are less toxic, having low-cost, chemically stable, easily accessible and excellent photoactive properties with nano size i.e., 1–100 nm range [68]. Generally, nanophotocatalysts present various advantages such as stability (chemical/physical), low cost and eco-friendliness. Among them, TiO2 having good photostability but many nanophotocatalysts, such as zinc oxide, metal sulfide materials, copper-based materials, and so on exhibit relatively low chemical stability due to photocorrosion [69]. As light shine on them, they oxidized or reduced depend on materials and their oxidation states changes by generating holes or electrons, which leads to decomposition of photocatalysts and decreases the efficiency of photocalatalysts. Therefore, there is a strong need to synthesize a nano composite in order to achieve stable photocatalysts for long-time performance.

The major advantage of nano-sized is related to quantum-size effect, which enhance the energy bandgap and reduce the particle size [70]. Moreover, as a process photodegradation also have several advantages such as low cost, reusable, and generally complete degradation. In addition to all of the developments, still nanophotocatalyst facing some issues i.e., toxicity and recovery of catalysts from mixture. These type of issues limits the applications and scope of nanophotocatalyst at higher level [71]. So, the scientific community is now focusing others nanocomposite of different material which can reduce the toxicity while using in water treatment process. So, it is suggestible for the scholar community to synthesize new photocatalysts which can work in visible ranges for sustainable result and promote the doped photocatalyst with different material such as graphene and its derivatives to reduce the toxicity effect. To overcome the catalyst recovery drawback, one significant approach could be used i.e., magnetic nanophotocatalysts in wastewater treatment. When magnetic nanophotocatalysts are used, the recovery of catalyst can be achieved through external magnetic fields, therefore, permitting the many recycling of nanocatalyst and achieved more effective and naturally responsive water decontamination processes. Furthermore, the approach regarding nanophotocatalysis for removal of pollutants from water has been described as a very effective approach and some unique applications of nanophotocatalysts are shown in Figure 1.

**Figure 1.** Potential applications of nanophotocatalysts.

#### 2.1.2. Future Perspectives of Nanophotocatalyst

A wide research on nanomaterials is ongoing in field of nanophotocatalysis which led to few progress in reactor designing and developments regarding the modifications in nanophotocatalysts. Although many developments in nanophotocatalytic materials occurred, still some important inquiries required related to characteristics of nanophotocatalytic materials. The major challenges in intensification process is mass transfer limitations and higher consumption of photons [72]. The concept of nanocomposites is ideal in solving the electron pair recombination problem which can be prolonged by combining the nanocomposites with nanophotocatalytic reactor structures. The new modern reactors are known as microfluidic reactors which open a new door for intense characteristics study in reaction phase and synthesis phase. Microfluidic reactors are those reactors which are working on micro level with reactants [73,74]. The key features of micro-reactors are large surface-to-volume ratio, improved diffusion effect and great mass transfer coefficient factor, highly stable hydrodynamics, less Reynold's flow, and informal handling which makes them more ideal material than conventional reactors. However, still it is difficult for photocatalysis to apply on large scale in actual wastewater.

Moreover, the synthesis of significant structures such as nanorod, nanosphere, nanoflowers, nanoflakes and nanocones with enhanced functional and structural properties could be opened an extensive area of study. Several structures of nanomaterials with potential properties could be produced through the measured synthesis approaches. However, the future research, should be explored by producing new photocatalysts. The synthesis of novel nanophotocatalysts with excellent efficiency, inexpensive, eco-friendly and high stability is crucially needed. Moreover, to exploit pollutant treatment effectively, several approaches should be joint with sensible match, such as electrocatalysis, photocatalysis, adsorption and several thermodynamics processes. The preparation of nanocomposites in the presence of ZnO, TiO2 was well explored in early decades for the treatment of water pollutants. The nanocomposites preparation by using carboneous material, polymer and ceramic materials are still in initial stage. It can generate ideal nanocomposites with improved properties. The heterogeneous photocatalytic for wastewater remediation is inhibited by some main technical problems that need to be study effectively. Finally, a significant photocatalytic treatment with better solar-driven, excellent efficacy and less site area requirements can be comprehended in future with fast assessment.

#### 2.1.3. Photocatalytic Degradation and Mineralization Pathway

Nanomaterials also got much attention in degradation as well as mineralization of toxic organic pollutants due to having remarkable physiochemical properties. In photocatalysis, there are two types of processes that occur, namely mineralization and degradation of organic pollutants [5,75]. In the process of degradation, the organic pollutants are splitting or decomposed into several products while in the case of mineralization, the complete destruction of organic pollutant took place into water, carbon dioxide and some inorganic ions. The possible pollutant degradation mechanism in the presence of light is shown in Figure 2 [75]. Briefly, a semiconductor like TiO2 absorbs the light which is greater or equal to TiO2 band gap width, it carries to electron–hole pairs (e−–h+). If the separation of charge is continued, the electron–hole may travel to the surface of catalyst where they contribute with sorbed species in redox reactions. Particularly, h+vb react with water (surface-bound) to generate the hydroxyl radicals and simultaneously e−cb selected by oxygen to produce the radical anion (superoxide radicals) as designated below in equations.

$$\text{TiO}\_2 + \text{hv} \rightarrow \text{e}^-\_{cb} + \text{h}^+\_{vb} \tag{1}$$

$$\text{O}\_2 + \text{e}^-\_{cb} \to \text{O}\_2^- \tag{2}$$

$$\text{H}\_2\text{O} + \text{h}^+\text{}\_{v\Phi} \rightarrow \text{OH} + \text{H}^+\tag{3}$$

**Figure 2.** General mechanism of toxic organic compound degradation through nanophotocatalysts.

After pollutant degradation by using nanomaterials in the presence of light, HPLC-MS or GC-MS was employed to analysis the produced degraded products. Here, it is necessary to confirm regarding the obtained degradation products, whether these products are more toxic or less toxic as compared to parent compound using toxicity test.

As discussed above, mineralization concept is actually synonymous of complete photodegradation. It defines the degradation of a compound into CO2 and H2O. Sometimes others minerals also released during mineralization of compounds such as sulphate, ammonia, sulphite, fluoride, sulphide, chloride, phosphate, nitrite, etc. [76]. Generally, the rate of mineralization is less as compared to degradation, probably due to generation of stable intermediate during process. Therefore, it is supposed that a long irradiation is compulsory for entire removal of total organic carbon (TOC). Moreover, the mineralization concept is avoiding the generation of undesirable products is the excellent mode to degrade the organic compounds. The TOC is defined as the total quantity of bounded carbons in any organic compound and measured by TOC analyser. The possible mechanism of pollutant mineralization is shown in Figure 2 coupled with photodegradation mechanism.

For example, different degradation products as well as the mineralization products of Lignocaine are shown in Figure 3 as follows [77];

**Figure 3.** Different degradation and mineralization products of Lignocaine.

#### *2.2. Nano- and Micromotors*

Nanotechnology, an area of research that has progressive at such a quick pace in early decades and offered many approaches for water treatment. In modern era, nano/micromotors have been considered that can convert the energy from several resources into machine-driven force, empowering them to achieve special goals through various mechanisms. These innovative motors are motorized in both cases by using fuel or without fuel sources (acoustics, magnetic field, electric field) have several significant exciting applications [78]. They show more speed, high power, specific control movement, self-mix ability, etc. The removal of contaminant from polluted resources is a significant focus for environmental stability and sustainability [79]. The trend of water purification and its treatment has grown quickly throughout the world because of high demands for pure water resources and novel water superiority regulations. A large variety of approaches are used earlier for removal of pollutants from polluted groundwater, fresh water, sediments wastewater, etc. Different mechanisms employed by nano/micromotors for treatment of water pollutants are graphically shown in Figure 4 and some environmental applications of nano/micromotors are shown in Table 2 with their mechanism.

Traditional treatment Processes are inadequate through diffusion and, hence, entail outward agitation to stimulate the wastewater treatment process. Though, nano/micromotors could possibly overwhelmed the diffusion boundary by energetic mixing due to their self-propulsion competences. These self-propelled nano/micromotors expressively stimulate the water treatment efficiency through water decontamination efficiency, merging with materials nano/microstructure i.e., greater surface area and working activities [80–83]. Moreover, nano/micromotors can go in nano/microscale detentions in the presence of a magnetic field, serving as programmed cleaners, where conventional approaches are not actively working. However, most positive concept for nano/micromotors in term of wastewater decontamination depends on fuels as shown in Table 3. There is problem, this condition reduced the working potential of nanomotors in biological applications. Although some photocatalytic, biocatalytic and magnetically driven nano/micromotors have been industrialized, still there are some challenges in order to apply nano/micromotors for water treatment in future.


**Table 2.** List of nano/micromotors and their applications.

**Figure 4.** Different pollutant removal mechanisms used by nano/micromotors (adapted with permission from Royal Chemical Society) [84].


*Water* **2020** , *12*, 495

#### 2.2.1. Advantage and Disadvantages of Nano- and Micromotors

Recently, nano/micromotors are used to overcome the environmental issues such as water pollutant treatment and environmental sensing/monitoring. The nano/micromotor are considered as excellent option, because the reactive nano-based materials have potential properties which make it more efficient to convert toxic pollutants into toxic free. Nanomachines offer different advantages as compare to traditional remediation agents. The small machines improve a dimension which based on decontamination approaches, lead to in-situ and ex-situ nano-remediation rules, and have ability to decrease the clear-out time and entire cost [112]. Specifically, the constant movement of nanoscale substances can be utilized for transferring reactive nanomaterials for water purification through polluted samples, for discharging remediation agents to long distances, and for communication of important mixing throughout decontamination processes [113]. Existing technologies are insufficient for fulfilling the demand in term of scaling up, as stated in the treatment of polluted water, therefore more efforts are required. Some issues require to be explained prior to move toward applied application at commercial level. For example, the lifecycle of multi-functional nano/micromotors is restricted to the residual materials in its physique that were consumed in locomotion or oxidation reactions. Another disadvantage is poisoning of Pt layer because the compounds present in wastewater can bond chemically to other surface-active sites of the catalyst, or great viscosity-based wastewater which hinder the movement of micromotors [114]. The introduction of novel development in nano/micromotors will offer countless environmental treatment possibilities to achieve more multifaceted and challenging operations.

#### 2.2.2. Future Perspectives of Nano- and Micromotors

Nano/micromotor are still considered as immature techniques because it still on initial stages but this topic open a new door of research for researcher. As Compared with other traditional wastewater treatment approaches, this subject is still novel and but has a lot of restrictions in extensive and real-world applications. Despite all, current approaches are insufficient to fulfil the current demand in scaling up the technique described in water pollutant treatment, therefore considerable work is still required [115]. Novel materials should link with nano/micro-motors, such as graphene to treat wastewater. These kinds of materials show better performance in wastewater treatment as compared with other materials. Moreover, some new mechanism should functionalize in nanomotor design for better results. Soon, we believe that excellent nanomotor will be planned by associating with other approaches. These predictable new, novel micro/nanomotors would ultimately develop the ecological monitoring and wastewater treatment technologies, from a struggle to advance the class of life.

#### *2.3. Nanomembranes*

Nanomembranes are a unique type of membrane formed with different nanofibers which are employed to eliminate unwanted nanoparticles present in aqueous phase. Using this technique, the process takes places at a very high elimination speed with condensed fouling propensity and it also served as a pre-treatment process which is used for reverse osmosis [116]. There are many reported studies on membrane nanotechnology in order to produce multifunction membrane by using different nanomaterial substance in different polymers-based membranes. Water porous membrane for water treatment did nanofiltration, ultrafiltration, reverse osmosis, etc. The membrane contains a porous support with composite layer. Typically, the considerable composite layer is carbon-based material (graphene oxide/CNT) dispersed into polymer matrix for significant practice. This provide promising and significant progresses in fouling resistance and aqueous transport. For example, CNTs hold anti-microbial properties that can minimize fouling, biofilm formation and it can also reduce the chance of mechanical failures [117]. The doping process of nanomaterial (zeolite, alumina, TiO2, etc.) into polymer ultra-filtration membranes show the formation of amplified membrane on surface of hydrophilicity and fouling resistance [118–120]. The antimicrobial material like silver metal particles are

also doped with a polymer to produce polymeric membrane to prevent attachment of bacteria and inhibit biofilm production on surface of membrane [121,122]. It served to inactivate viruses and it has also ability to prevent the biofouling of membrane [123]. The production and growth of nanomaterial thin film are incorporated into active thin film composite through doping in surface modification. Usually, there are a few major challenges in nanomembrane i.e., membrane clogging and membrane fouling. Therefore, it is a need to overcome this problem. By the addition of super hydrophilic nanoparticles in making a thin film nanocomposite membrane, the prevention from clogging and fouling could be checked. Furthermore, metal oxide nanoparticles (Al2O3, TiO2), antimicrobial nanoparticles (nano silver, CNTs) and aquaporin-based membranes are very useful material to overcome the membrane clogging and fouling issue due to having high hydrophilicity, high porosity, better fouling resistance and a better homogenous nanopore. Generally, nanoparticles affect the selectivity and permeability of a membrane which depends upon the sort, quantity, dimension, etc. of nanoparticles. Moreover, there are many biological membranes present with highly permeable and selective ability [124,125] Nanomembranes are also used for wastewater treatment due to having several properties that make this material more prolific, these are high uniformity, homogeneity ability, optimization, short time required, easily handled and contain much order of reaction [126]. There are some nano photocatalysts which can be introduced in the nanocomposite membrane to make fit it for the degradation of organic pollutants. For example, TiO2 incorporated nanomembranes and films are effectively used to deactivate different microorganisms and degrade the organic pollutants [127]. The developments and progressive growth of nanotechnology especially in nanomaterial produced several nanostructured catalytic membranes which contain novel properties like improved selectivity, high decomposition rate, and larger foul resistance [128,129]. In order to synthesize these types of nanomaterials, there were many approaches used to produce it with multi-functionality features [130]. The incorporation of nanostructured catalytic materials like iron-catalysed based free radical and enzymatic catalysis within pore membrane showed a constructive progress in this technology. Therefore, it possible to carry out oxidative reactions for removal of pollutants and detoxification of water without use of any toxic chemicals. To prove the efficiency of nanostructured catalytic membrane in industrial/commercial applications, immobilized membrane nanoparticle (ferrihydrite/iron oxide) reaction was carried out with hydrogen peroxide to make free radicals' ions for removal of chlorinated based organic pollutant in real groundwater. The development of nanostructured materials is still very useful in many other environmental applications [131]. There were several reported studies of metallic nanoparticles immobilization on membrane (e.g., chitosan, cellulose acetate, polysulfones, etc.) for dichlorination, degradation of a toxic substance which contains novel properties such as, hindrance of nanoparticles, high reactivity, absence of agglomeration and surface reduction [132,133]. Palladium acetate and polyetherimide both are used to prepare nanocomposite films and a novel type interaction is present among hydrogen and Pd nanoparticles in order to improve water treatment efficiency [134]. In situ and ex-situ methods are also used to produce nanomaterials within the matrix with precursor film under many conditions [135–137]. This offers many opportunities to produce nanomaterial with tunable properties. Moreover, the developments of nanotechnology are also responsible to produce many significant catalytic membranes with high selectivity, better permeability and high resistance fouling. Modern methods contain hybrid and bottom-up approaches for empowering its multi-functionality characteristics in the field of wastewater treatment [138].

#### 2.3.1. Advantage and Disadvantages of Nanomembrane

The main objective of using membranes in case of drinking water is to separate the toxic particles from water resources. The nanomembrane filters were also used to measure the water safety level [139]. The advantages of nano membrane as compared to traditional approaches for filtration is that; in traditional approaches of filtration, throughout the entire process, calcium and magnesium required another ion to compensate, therefore mostly Na<sup>+</sup> ions are served as an exchanger but in case of nanomembrane, there will be no need [139,140]. Nanomembranes limitations are usually reducing

its efficiency compared to other conventional approaches. The first one is fouling of nanomembrane which comes after using the membrane few times. It is the major issue, which make this approach more expensive and inefficient. Fouling problems occur in the nanomembrane because it entirely depends on working conditions, sometime the working conditions are not appropriate such as over temperature, pressure, and optimization is also considered as responsible for membrane fouling [141]. Second major limitation is membrane stability. The nanomembrane could not keep the stability for long period. After sometimes, the stability start reducing its efficiency and it does not give excellent outcomes as earlier. So, there will be a need for changing the nanomembrane to get excellent results, but this will cause several other issues such as high cost, impurity chances during changing process, etc. [142]. The stability is depending on the essential chemical resistance which apply for material cleansing. The disadvantages are less reliability, slow operation process, less selectivity, high maintenance cost and working efficiency reduce with passage of time. There are some common disadvantages that is why it is not extensively explored [143]. Some nanomembranes types are summarized in Table 4 along with their advantage and disadvantage and applications.


**Table 4.** Advantages/disadvantages and applications of nanomembranes.

#### 2.3.2. Future Perspectives of Nanomembranes

Nanomembranes separate the inorganic ions, organic, nanoparticles viruses-based pollutants from water resources through solution using diffusion and size exclusion. Though these described nanomembranes are proved effectively at the research laboratory level, upscaling to lower cost, but making them ideal for industrial scale is still emerging as a challenge. To cross this barrier toward productive upscaling, the commercialization objective needs a joint collaboration struggle through research institutions and manufacturing companies. The selectivity of nanomembrane should be improved, secondly upgrade the nanomembrane resistivity to avoid membrane fouling. Surface grafting-based polymers like zwitterionic may be a significant candidate for the development of new generation membranes. Surface grafting, though, would not report fouling resistance to inner walls pore [149]. Thereafter, the prime of other appropriate anti-fouling convertors fixed into the membrane matrix that will be vital. It is also very significant to improve the high sensitivity of polyamide membranes to many types of oxidants like ozone and chlorine. Furthermore, the multi-functional membranes fabrication requires a significant attention for better innovation at industrial scale. However, still there are several challenges to be solve for manufacturing production of lower cost and effective nanomembranes for water treatment.

#### *2.4. Nanosorbents*

Nanosorbents hold wide properties like high sorption capability that make the nanosorbents more appropriate and powerful for water treatment [150]. These nanosorbents are very rare in commercial form but researcher and experts doing a lot of work on it to produce nanosorbents in larger quantity/at commercial level [151]. The most commonly reported nanosorbents are based on carbon (e.g., carbon black, graphite, graphene oxide). Furthermore, metal/metal oxide and polymeric nanosorbents were also exist [152]. The composite of different material like Ag/polyaniline, Ag/carbon, C/TiO2, etc. carrying huge significant importance in order to reduce the effect of toxicity in the wastewater treatment process. The carboneous material such as CNTs with a cylindrical form nano structure may present as single-walled and multiwalled nanotubes depending on the method of synthesizes. CNTs hold measurable adsorption sites and due to high surface area, they hold sustainable surfaces. It must be stabilized to prevent from aggregation that decreases the surface-active sites because CNTs has hydrophobic surface properties. So, it is an efficient material for the pollutants by adsorbing process. Similarly, the polymeric nanoadsorbents like dendrimers are functional for eliminating organic pollutants and heavy metals from wastewater [153]. For example, copper ions were reduced with the help of dendrimer-ultrafiltration system [154]. They simply regenerated by shifting of pH effect and show biocompatibility, biodegradability, and toxic free environment. Furthermore, the removal percent of dyes or others organic pollutants is almost 99% [155]. Another important nanosorbent is zeolites, which have an absorbent structure in which several nanoparticles like copper, silver ions can be implantable [156]. The advantage of zeolites is to control the amount of metals and it also served as anti-microbial agent [157]. Moreover, the magnetic nanosorbents have play vital role in water treatment and a unique tool to remove different organic pollutants from water [158]. Some organic containments are also removed by using magnetic filtration. Magnetic separation nanosorbents are synthesized by ligands coating with magnetic nanoparticles at specific affinity [159]. There were many methods like ion exchange, cleaning agents, magnetic forces, etc. reported to regenerate these nanosorbents. These regenerated nanosorbents have the ability of being cost-effective and more promotions of commercialization. Despite all, the carbon has some health risks. The reported studies demonstrated that toxic effect is depends on morphology of nanoadsorbents, chemical stabilizers and surface modifications [160,161]. So, there is a need to give an attention to synthesize more stable morphology (size and shape) to overcome the toxicity issues as well as health risks. Furthermore, the bioadsorbents have the properties of high biodegradable, good biocompatible and nontoxicity which could be replaceable with chemically synthesized nanosorbents. The graphene oxide is suggestable to scientific community because it is very emerging nanomaterial to use as nanosorbents to remove pollutant and it can give better result than others due to its superior properties. There are some reported nanosorbents and their functions as shown in Table 5.


**Table 5.** Most commonly used nanosorbents and their functions.

Advantages, Disadvantages and Future Perspectives of Nanosorbents

The role of nanomaterials as sorbents in explaining ecological issues such as decontamination of wastewater received a great interest due to having remarkable physiochemical properties. These properties distinguish them in several fields as compare to conventional traditional sorbents. For an ideal sorbent to treat the pollutant very effectively in a short time, it should hold large surface area, excellent rate of adsorption, short time adsorption and equilibrium times. Similar to nanosorbents, nanomaterials got interest because of having nano-size which can hold excellent rate of adsorption with short time. Furthermore, nanosorbents that can be used as a separation medium in water decontamination to eliminate the organic, inorganic-based pollutants from polluted water are nanoparticles [170]. Literature review shows that a lot of efforts are already done for wastewater treatment and used nanomaterials as sorbent for efficient results. Some challenges are required to be addressed entirely for commercialisation purpose of the nano-size sorbents for water decontamination such as production scalability, as well as excellent adsorption measurements, selectivity, stability of material, operational duration of material, etc. However, there is an enormous need for an active approach to treat the wastewater and fabricate some new nanosorbents which could be applied to control toxic ions and compounds from wastewater [171]. The future prospects looking admirable, as scientific community working on advances knowledge and improving the adsorption mechanisms. In the modern world, researchers, consultants and officials are concerned about wastewater pollutant which holds health risks and they all are dedicated to find an effective explanation as concerning the nanomaterials should be used at industrial scale.

#### **3. Self-Toxicity of Nanomaterials**

Nowadays, nanomaterials have become most attractive and widely used material for different applications in various fields such as electronics, medicine, agriculture, wastewater treatment plants, energy generations and other sciences. These nanomaterials are doped metal oxides, doped carbon nanotubes (single and double walled), nanosorbents, etc. There were many reports available in literature which show that these nanomaterials served as a leading role in order to remove different pollutants from waste waters [172]. In addition to their importance, utilities and applications in different fields, especially in water treatment process, it is very important to know about their self-toxicity effects. There are studies which show the toxicity effect of few nanomaterials, as shown in Table 6. Moreover, few metal oxides show some toxic character at high and some, even at low concentrations [173,174]. Furthermore, the toxicity of CNTs is depending on different properties like length, surface area, distribution ratio, aggregation degree, initial concentration of material [175]. Furthermore, single walled CNTs are less toxic as compared to double walled [176]. They are associated for pulmonary inflammation, oxidative stress, granuloma in lungs, basic inflammation, apoptosis and fibrosis [177]. Similarly, in case of TiO2, the toxicities of different composition ratio of TiO2 depend upon initial concentration and time. Generally, it is considered as non-toxic even at a higher concentration for 24 h [178]. Currently, the studies by scientific communities on these nanomaterials are ongoing to see the toxic and side effects and try to find out the mechanisms with a focus on explaining the outlines of nanomaterials transport, degradation, elimination, accumulation, etc. Furthermore, nanomaterials can harm body through various sources. Therefore, there is a critical need to find out the effects of nanoparticles on health.


**Table 6.** Observation of self-toxic effect of some common nanomaterials.

#### **4. Applications of Nanomaterials**

There were many opportunities to use engineered/modified nanomaterial for water treatment and many other industrial fields. Currently, nanomaterial attracts more attention of researcher in the field of wastewater treatment. The study trend in the field of nanomaterials is increasing day by day as we can observe through previously reported data. The number of publications in nanomaterial was less than 5% in 2005 but in 2019 number of reported works is more than 80% as shown in Figure 5 [184].

**Figure 5.** Publication trends in field of nanomaterial for wastewater treatment.

With emerging several aspects of nanomaterials, the wider environmental water resources effects are also considered. Such attentions might contain models to measure the significant advantages of reduction or inhibition of contaminants from engineering sources. Nanoscience expertise holds excellent potential for constant improvement for environmental protection. The present article has presented more indication to this matter and it has a response to what are the significant environmental impacts of this nanotechnology. For a quick assessment, the summary of nanomaterial applications for treatment of pollutants from water resources are briefly summarized in Table 7 [185].


**Table 7.** Summary of water treatments by using nanomaterial by different mechanisms [185].


materials. The Ag nanoparticles give remarkable results in order to remove different pathogens from drinking waters [190].


**Figure 6.** Schematic illustrating of TiO2 photocatalytic process (Adapted with permission from Jurnal Teknologi) [197].

The summary/mechanism of pollutant removal by using TiO2 nanoparticles as nanophotocatalyst is shown in Table 8.


**Table 8.** Observation of different pollutant removal by TiO2 nanophotocatalyst.

(9). Different types of pollutants such as organic pesticides, organic dyes, pharmaceutical drugs, etc. are photodegraded by many researchers under several conditions such as choice of UV or Visible light, doped nanoparticles or undoped, metal/non-metal doping, etc. According to literature, as shown in Table 9 indicate clearly that modified or doped form of TiO2 can give better results especially in photodegradation of pollutants. Table also demonstrates that the efficiency of doped-TiO2 in visible light showed better results as compared to UV light.

**Table 9.** Observation of different pollutant removal by Doped-TiO2 nanophotocatalyst.


#### **5. Nanomaterial Challenges for Water Treatment**

Currently, the emerging nanomaterial possesses some challenges in the field of wastewater treatment [206,207]. Nanomaterial provides several possibilities of treatment of wastewater and they contain different kinds of substances which are distinct on the basis of the particles morphology. The developments regarding the commercial applications of nanomaterial is too quick and production of nanomaterial is increasing at a global level [58,208]. Nanomaterials are used for purification of polluted water through different methods such as photocatalytic, adsorption, and nanosorbents [209,210]. These methods need some modifications to work more effectively. Moreover, understanding of risk pretended by nanotechnology has not improved as quickly as research has giving possibilities to different applications of nanomaterials. The major challenge is the lack of information about the nanomaterial and how nanomaterials are released into environment, how they travel into water, how they start to exist in water [196]. Another challenge is related to human health because these types of material have some adverse effects/toxicological effect. The reported studies show that nanomaterials may cause some health issues but still research is going on. It is not easy to give a conclusion and ongoing research requires some conducting experiments at appropriate concentrations and more about toxicological studies. The membrane process is also effectively used wastewater treatment but there were very few reported studied about this process. The major challenge is the use of the membrane for fouling process and water treatment because after performing work/filtration, their pores may block, and the efficiency starts to decrease. In case of nanosorbents, the reusability is the main drawback, so there is need to synthesize such types of nanomaterials which can efficiently remove the pollutants from wastewater and after that the recovery process of these nanosorbents should be very easy. The United State Environmental Protection Agency (USEPA) found some basic challenges in order to remove nanoparticles using a process of water treatment [210–213]. (a) What is the mechanism of removing nanoparticles from wastewater? (b) What is the effect of nanoparticles on other waste substances during wastewater treatment? (c) How coagulation, carbon adsorption, etc. methods are effective in working. Moreover, water contains some pollutants and after treatment by the above mentioned methods, they degrade or are decomposed. There is a basic need to design an experiment for finding the intermediate products during these processes. Furthermore, constant development in methodology is needed to evaluate nanomaterials with low cost and it should be appropriate for complex nanomaterials.

#### **6. Conclusive Remarks and Future Perspectives**

Water distinguishes and makes our planet superior as compared to other planets. Though the worldwide available supply of pure water is high to meet all existing and predictable water demands. There are several areas where the drinking water resources are insufficient to fulfil the basic, economic and domestic developmental needs. In such areas, the insufficient fresh water to fulfil human water need and sanitation requirements is certainly a limit on human health and for other living creatures. Scientific community/research institutes must find a path to eliminate these limitations. Moreover, the world is facing several challenges in doing that, especially given a fluctuating and undefined future environment, a fast-rising population that is driving enlarged community and financial growth, urbanization and globalization. How superlatively overcomes on these challenges which entails exploration in all features of water management. The trend of nanomaterial for water pollutant treatment is rapidly increasing in this modern era due to very horrible conditions of water and demand for fresh water in the whole world. There is an important need for innovative progressive water treatment approaches, in specific to certify a high class of water for drinking purpose, remove micro/macro pollutants and increase industrial production developments through flexibly modifiable water treatment approaches. Nanotechnology has proved great achievement for controlling water purification challenges and makes some future advancement. Nanomaterial approaches like nanosorbents, nanostructured catalytic membranes, etc. are very efficient, less time required, less energy and eco-friendly techniques but all these methods are not cheap, and they are not used yet for commercial purpose to purify the wastewater at a large scale.

Nanomaterials show high efficiency due to having high rate of reaction. However, there are still some weaknesses that must be negotiated. Up to now, no operational digital monitoring techniques exist that offer consistent real-time measurement facts on the superiority of nanoparticles which are existing in small amounts in H2O [212]. Furthermore, to reduce the health risk, some research institutes and international research communities should prepare proper guidelines to overcome this issue. Another, further mechanical restriction of nano-engineered water approach is that they are infrequently flexible to mass developments, and at present-day, in several cases are not modest with conservative treatment approaches [213]. However, nano-engineered materials provide excessive potential for water revolutions, in specific for decentralized water treatment technologies, point-of-use strategies, and seriously degradable pollutants. Furthermore, there is a great need to synthesize some modified nanomaterials which should be effective, having high efficiency, easy to handle and eco-friendly. It is also necessary to take the cost challenges and commercialization of these technologies for wastewater treatment. The different applications of nanomaterial can provide a tremendous offer in order to supply drinking water to whole world.

**Author Contributions:** A.A.Y., T.P. designed and wrote the first draft of review and drew the figures. K.U. participated in technical check and edited the full manuscript. M.N.M.I. was responsible for conceptualization of draft, technical check, and funding source for the manuscript. All authors discussed the results and commented on the manuscript. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded by Universiti Sains Malaysia, (11800 Penang, Malaysia). The grant number is 1001/ PKIMIA/8011070, and APC was funded by Universiti Sains Malaysia.

**Acknowledgments:** This research article was financially supported by Universiti Sains Malaysia, 11800 Penang Malaysia under the Research University Grant; 1001/ PKIMIA/8011070). The author (Khalid Umar) gratefully acknowledged the post-doctoral financial support. (USM/PPSK/FPD(BW)2/(2019).

**Conflicts of Interest:** The authors declared that they have no conflicts of interest.

#### **References**


*Water* **2020**, *12*, 495


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