*Article* **A New Polyvinylidene Fluoride Membrane Synthesized by Integrating of Powdered Activated Carbon for Treatment of Stabilized Leachate**

**Salahaldin M. A. Abuabdou 1, Zeeshan Haider Jaffari 2, Choon-Aun Ng 1, Yeek-Chia Ho 3,\* and Mohammed J. K. Bashir 1,\***


**Abstract:** Stabilized landfill leachate contains a wide variety of highly concentrated non-biodegradable organics, which are extremely toxic to the environment. Though numerous techniques have been developed for leachate treatment, advanced membrane filtration is one of the most environmentally friendly methods to purify wastewater effectively. In the current study, a novel polymeric membrane was produced by integrating powdered activated carbon (PAC) on polyvinylidene fluoride (PVDF) to synthesize a thin membrane using the phase inversion method. The membrane design was optimized using response surface methodology (RSM). The fabricated membrane was effectively applied for the filtration of stabilized leachate using a cross-flow ring (CFR) test. The findings suggested that the filtration properties of fabricated membrane were effectively enhanced through the incorporation of PAC. The optimum removal efficiencies by the fabricated membrane (14.9 wt.% PVDF, 1.0 wt.% PAC) were 35.34, 48.71, and 22.00% for COD, colour and NH3-N, respectively. Water flux and transmembrane pressure were also enhanced by the incorporated PAC and recorded 61.0 L/m2·<sup>h</sup> and 0.67 bar, respectively, under the conditions of the optimum removal efficiency. Moreover, the performance of fabricated membranes in terms of pollutant removal, pure water permeation, and different morphological characteristics were systematically analyzed. Despite the limited achievement, which might be improved by the addition of a hydrophilic additive, the study offers an efficient way to fabricate PVDF-PAC membrane and to optimize its treatability through the RSM tool.

**Keywords:** stabilized leachate; membrane fabrication; filtration technology; phase inversion technique; powdered activated carbon (PAC)

### **1. Introduction**

Sanitary landfills are the widely applied technique to tackle municipal solid waste (MSW). Inappropriately, the majority of these landfills do not fulfill the normal discharged limits [1]. In developing countries such as Malaysia, more than 80% of the MSW produced was received by open duping and landfill sites [2]. This resulted in the generation of highly contaminated leachate, which is the liquid generated due to the precipitation above these solid litters and could be toxic to the surrounding environment. This leachate could contaminate the sources of fresh water if not carefully treated before discharging to the environment [3]. Stabilized leachate, which is more than ten years old, has lower BOD5/COD ratio. Thus, it is almost impossible to treat this kind of leachate using some

M.J.K. A New Polyvinylidene Fluoride Membrane Synthesized by Integrating of Powdered Activated Carbon for Treatment of Stabilized Leachate. *Water* **2021**, *13*, 2282. https://doi.org/10.3390/w13162282

**Citation:** Abuabdou, S.M.A.; Jaffari, Z.H.; Ng, C.-A.; Ho, Y.-C.; Bashir,

Academic Editor: Giovanni Esposito

Received: 11 June 2021 Accepted: 28 July 2021 Published: 20 August 2021

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**Copyright:** © 2021 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https:// creativecommons.org/licenses/by/ 4.0/).

biological treatment technique [4]. To date, various purification techniques such as adsorption [5], coagulation [6], advanced oxidation [7], electro-Fenton [8], and combinations of these processes [9,10] have been successfully introduced to eliminate the organic contaminates from stabilized leachate. Among these techniques, membrane filtration could be one of the most suitable purification process [11]. The membranes acted as a selective barrier to achieve the objective of separation and purification. Nonetheless, there are still some shortcomings in membrane technology such as membrane fouling upon the higher contaminant concentration [12]. Fouling could affect the separation efficiency as well as permeability of membrane, which are the vital factors in the membrane filtration [13]. Several strategies, including pre-treatment of feed [14], optimization of operating parameters [15], selection and modification of membrane [16], hydraulic flushing [17], and applied field enhancement [18], have been performed to alleviate membrane fouling and water flux rate. Under different circumstances, the workability of membrane can be improved through the membrane characteristics and performance of treatment process. Hence, investigation of membrane characterization can be separated into four groups: membrane activity (permeability, surface wettability, average pore size, and porosity); morphological characterization (surface chemistry and roughness, and external and internal membrane texture); treatment efficiency (separation performance); and antifouling evaluation (pore size decrease and cake formation) [19].

Synthetic polymers such as polypropylene (PP), polyvinylidene fluoride (PVDF), and polysulfone (PS) are commonly applied in the membrane fabrication due to their higher flux, antifouling ability, and separation efficiency [20]. Among all these synthetic polymers, PVDF polymer proved to be an ideal membrane fabrication material due its durability [21], good thermal stability and higher chemical resistance [22]. Additionally, the PVDF polymer can also help to extend the membrane life, as well as reduce the damage caused by the concentrated pollutants [23]. However, the PVDF membranes antifouling capability could be enhanced due to its hydrophobic nature [24]. Many researchers have successfully applied dry–wet phase inversion technique to boost their membrane performance [25]. For instance, Zhou et al. [26] developed an ultrafiltration PVDF membrane using nanoparticles of titanium dioxide (TiO2) and polyvinylpyrrolidone (PVP) as blended additives to increase the fouling resistance and water permeability. The addition of PVP-TiO2 increases the average pore size and porosity of membrane, leading to the higher flux and hydrophilicity of membrane with more than 91.4% removal performance against sulfonamide antibiotics water. Moreover, polyethylene glycol and poly(acrylic acid) were also applied in the fabrication of membrane through chemical reaction with a key focus of enhancing hydrophilicity. Their batch filtration experiments clearly exhibit an increase in critical flux and a declined fouling rate. Similarly, various reports have presented effective ways to boost the antifouling abilities of PVA-based membranes due to their hydrophilic properties [27,28].

Recently, the incorporation of activated carbon (AC) on the surface of the membrane has proven to be an effective way to boost the membrane rejection performance [29]. The utilization of AC in membrane is a relatively new technology for the elimination of organic contaminates for wastewater, which not only enhances the adsorption capacity of AC, but also improves the particle removal capabilities of membrane [30].

To date, there are quite a number of studies which clearly demonstrate that the usage of PAC can significantly improve the filterability of membranes [13]. However, evaluation of PAC addition into PVDF flat sheet membranes with different concentrations, in terms of their treatment efficiency and productivity, has not been investigated. Therefore, the current study was performed to observe the potential of incorporating PAC, for the first time, into the PVDF polymeric membrane for stabilized landfill leachate purification. Furthermore, fabricated membrane was optimized using RSM technique, and the membrane properties and morphologies were systematically characterized.

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

#### *2.1. Collection of Leachate*

Leachate sample was taken form Sahom landfill site located in Perak, Malaysia, which is an operative landfill site with a daily production of 100 tonnes of MSW in average [31]. After collection of leachate sample, it was stored in a refrigerator at 4 ◦C. Initial leachate characterization was performed using standardized methods of water and wastewater [32]. All measurements, including dissolved oxygen (DO), colour, chemical oxygen demand (COD), 5-day biochemical oxygen demand (BOD5), and ammoniacal nitrogen (NH3-N), were undertaken in triplicate.

#### *2.2. Materials*

The PVDF polymer (Kynar®740) was purchased from Afza Maju trading (Terengganu, Malaysia), and utilized after drying for 24 h at 70 ◦C. 1-Methyl-2-pyrrolidone (NMP, 99.5%) was purchased from Sigma–Aldrich. Methanol, (99.8%) was supplied by Chem Soln. Ultrapure distilled water (DI) was utilized throughout the experiments. PAC was purchased from R&M Chemicals. The AC was charcoal-based, and consists of sulfide, chloride, calcium, sulphate, iron, lead, zinc, and copper. This PAC density was 1.8–2.1 kg/m<sup>3</sup> with pH (4–7). Particle size analysis (PSA) and field emission scanning electron microscopy (FESEM) tests were used to investigate the distribution and the size of PAC particles, respectively. All these chemical materials were of analytical grade, and used without additional treatment.

#### *2.3. Experiment's Design and Optimization Process*

Central Composite Design (CCD) is the design method used in response surface methodology (RSM) for the membrane fabrication's experimental design [33]. Both CCD and RSM were run by version 8 from the Design Expert. For membrane dope solution design, two factors, the polymer (PVDF) weightage and the additive (PAC) weightage, were set into the CCD. Based on preliminary experiments and the extensive literature [34,35], the total mass of fabricated membrane dope was fixed at 100 g, which represents 100% of the dope weight, thus each 1 g of the dope element is equivalent to 1% weightage. The dosage of the PVDF was set within the range of 10 g to 18 g, and the amount of PAC was set within the range of 0 g to 2 g. Regarding the CCD, the alpha value was selected to be 1.0, and thus the centre points were 14.0 and 1.0 wt.% for the polymer content and additive content, respectively. The rest of the dope weight (to complete 100 g) is the NMP solvent. The total concentration of PVDF/PAC was kept at 20% (as maximum) and 10% (as minimum), as concentration higher than 20% resulted in solutions of extremely high viscosity, and was difficult to be casted on the glass plate, while clumsy, non-thick membrane was the result of using concentration less than 10%. Five responses, which are the removing efficiencies of COD, colour, and NH3-N, as well as maximum transmembrane pressure (max. TMP), and pure water flux, were also set into the CCD to have the full design of experiments. The influence of various parameters was optimized by RSM using a combination of statistical and numerical techniques. In the current work, nine experiments were reinforced with four replications to assess the pure error [36]. The 13 different membranes were applied in double repetition and have their effluent collected. The quadratic model for every response was investigated by analysis of variances (ANOVA) to identify the results significancy, and to find the represented quadratic model after eliminating irrelevant terms. The frontal sign of each model term signifies to either antagonistic or synergistic effect on the response when it is positive or negative, respectively [4]. In RSM, it does mention that Prob > F less than 0.050 indicates model terms are significant, and Prob > F with the values greater than 0.10 indicates model term is not significant. "Not significant", in the description of lack of fit, is regarded a decent model, as it means the experimental reading is fitting the model [37]. Additionally, a good experimentally fitted data will have a higher coefficient (R2) value. The higher the R2 value, the closer the experimental data towards the predicted graph model by the RSM [38,39]. Selection of the best membrane takes into consideration

the membrane purification performance. Desirability value closer to 1.0 used to be selected as the ideal design for the data.

#### *2.4. PVDF-PAC Membrane Fabrication*

2.4.1. Dope Preparation

To produce the polymeric membrane, PVDF and NMP were applied as polymer and solvent, respectively. Figure 1 presents the process used for the dope preparation. Initially, the polymeric PVDF was entirely dissolved in the NMP solvent at a temperature ranged between 60 and 70 ◦C using a heating mantle (Figure 1a). In order to achieve a better permeate flux of the synthesized membrane, the heating mantle temperature should always be maintained within the above stated range [21]. The dope solution containing dissolved PVDF polymer in the NMP solvent was then infused into a clean Schott bottle. After that, the required amount of PAC was inserted into the dope solution to generate the dope for hybrid membrane. Lastly, the Schott bottle containing the dope solution was placed into a sonicator bath (Cole-Parmer, Vernon Hills, IL, USA) for eight hours to confirm the homogeneous mixing of the additives without any air bubbles raised in the prepared dope [40].

**Figure 1.** PVDF-PAC membrane dope preparation process.

#### 2.4.2. Membrane Casting

A semi- automated membrane casting machine (TECH INC, Chennai, India) was applied to synthesize a flat sheet membrane using the dry–wet phase process, as illustrated in Figure 2a. The membrane was produced at temperature 27 ◦C to 30 ◦C with an approximate thickness of 60 μm based on literature reports [41,42]. After 60 s of membrane casting above the glass board, it was submerged into a distilled water (DW) basin for 180 s (Figure 2b). As a result, a thin layered polymeric film was generated, which separated from the glass plate. Later, the newly produced membrane was transferred into a DW coagulation bath and remained there for 24 h. Afterwards, a methanol bath was used for 8 h, as shown in Figure 2c, to perform a post-treatment to ensure the excess solvent in the membrane can be removed completely [43]. Finally, the membrane was dried 24 h at the ambient temperature with 60% humidity, as shown in Figure 2d, to be ready to use in the filtration process [13].

**Figure 2.** The casting process of flat sheet PVDF-PAC membrane.

#### *2.5. Membrane Performance and Characterization*

The produced membranes have been characterized to investigate their treatment efficiencies, fouling, and permeability properties and surface morphologies. To ensure the accuracy of the findings, all of the tests have been duplicated. Each time, a fresh membrane has been utilized to investigate their characteristics and performance.

#### 2.5.1. Treatment Efficiency

The membrane filtration performance was investigated using laboratory scale crossflow filtration setup with a 3.34 cm disc diameter, as exhibited in Figure 3. The membrane rejection capabilities were studied against the treatment of landfill leachate. Before each experiment, initial characterization of leachate was measured to eliminate the small errors which occurred due to the minor changes in organics concentration with time. The steady flux for all individual membranes was acquired by a constant (200 mL/min) flow for 120 min. The volume of permeate, along with the recorded transmembrane pressure, were noted down under the flow of 200 mL/min for different intervals of time (0.5, 1, 2, 5, 10, 20, 40, 90, and 120 min).

**Figure 3.** CFR test configuration (filtration treatment set).

Final leachate characterizations were evaluated in terms of removing efficiencies for the COD, colour, and NH3-N pollutants using Equation (1):

$$\text{Removal efficiency }\%= \frac{(\text{C}\_{\text{F}} - \text{C}\_{\text{P}})}{(\text{C}\_{\text{F}})} \times 100 (\%) \tag{1}$$

where CF is the contaminant concentration at the feed (mg/L) and CP is the contaminants concentrations in the permeated solution (mg/L). All contaminants' concentrations were checked using the UV-V spectrophotometer (Hach DR6000, Loveland, CO, USA) in prior and post of filtration practice.

#### 2.5.2. Productivity of Membrane

Pure flux plays a dynamic role in the membrane productivity evaluation. Permeability of membrane was investigated through the pure water flux, which was measured via a deadend filtration apparatus, as illustrated in Figure 4. A metallic ring having 5 mm average pore size and 8.76 cm2 effective permeate area was applied to support the membrane. Initially, the impurities present in the membrane were removed by submerging the membrane in DW for 30 min. Then, a stable flux was achieved by pre-compacting the membrane with N2 gas at a pressure of 30 KPa for 2 min. After 30 min, the permeated water volume was noted at a similar pressure of 30 KPa. The pure water flux can be calculated using the Equation (2):

$$\mathbf{J} = \frac{\mathbf{V}}{\mathbf{A} \times \mathbf{t}} \left( \mathbf{L} / \mathbf{m}^2 \cdot \mathbf{h} \right) \tag{2}$$

where V is the permeated pure water volume (L), A is the membrane effective surface area (m2), and t is the time of permeation (h).

**Figure 4.** Dead-end test (pure water permeation set-up).

#### 2.5.3. Antifouling Valuation

Throughout the membrane filtration process, the overall decrease in flux, alongside the improvement of transmembrane pressure, were mainly caused by either membrane fouling, concentration polarization, or a combination of both [44]. Both of these components can be attained from the experimental data using both of the leachate permeate flux and maximum transmembrane pressure (Max. TMP) values which are measured by the cross-flow ring test. Max. TMP was applied to indicate the antifouling ability of fabricated membranes [45].

#### 2.5.4. Morphological Characteristics

It is a well-known fact that the membrane properties and performance are highly dependent on its morphology (pore size, surface texture, and microstructure). Therefore, investigation of membrane morphologies is considered a significant factor in the effectiveness evaluation of the produced membranes.

Fourier transform infrared spectroscopy (FTIR, Perkin Elmer Lambda 35, Waltham, MA, USA) was applied to investigate the membrane surfaces chemical compositions. The FTIR spectra ranged between 4000–400 cm<sup>−</sup>1.

EDX is a chemical microanalysis method used for quantitative, qualitative, and elemental mapping examination. Octane Silicon Drift Detector (SDD, EADX Inc., Mahwah, NJ, USA) was used at high voltage of 15 kV, using Mn Kα as source of energy. The fabricated PVDF-PAC membranes with different compositions were measured by INCA Energy 400 software (Firmware INCA, Version V1.09R13), along with the image taken by the Quanta FEG 450 instrument.

FESEM (Quanta FEG 450, FEI, Hillsboro, OR, USA) was applied to record the crosssectional and surface morphologies of the fabricated membrane. The cross-sectional morphologies were investigated by fracturing the membranes in liquid nitrogen and immediately cutting them after air drying. FESEM measurement starts by placing the sample on carbon tape, which was attached with the sample stub. The sample was also coated with the platinum nanoparticles in auto fine coater (JFC-1600, SUTD-MIT International Design Centre, Singapore) before performing the analysis.

An atomic force microscopy (AFM, Dimension 5000, Bruker AXS, Santa Barbara, CA, USA) was also applied to study the surface morphologies and roughness of the synthesized membranes. Herein, membranes were cut into small square pieces (1 × 1 cm) and pasted on a glass slide. Sample scanning were performed using a probe-optical microscope on tapping mode and images of 10 μm × 10 μm were taken by AFM. The root-mean-square

roughness (Rq) and average roughness (Ra) was applied to measure the surface roughness for each membrane.

Porosity of membrane could be easily defined as the pore's volume divided by the membrane total volume. Wet membranes were weighed after carefully wiping the surface (Ww). Afterwards, these membranes were dried in an oven at 50 ◦C for 24 h and weighed again (Wd). The porosity of membrane ε (%), was measured by gravimetric method using Equation (3) [25]

$$
\varepsilon = \frac{(\text{Ww} - \text{Wd}) / \rho \text{w}}{\frac{\text{Ww} - \text{Wd}}{\rho \text{w}} + \text{Wd} / \rho \text{p}} \times 100\% \tag{3}
$$

where, Ww is the weight of wet membrane (kg), Wd represents the weight of dry membrane (kg), ρw is the density of water (1000 kg/m3), and ρp, the polymer density (1770 kg/m<sup>3</sup> for PVDF).

Based on the measured distilled water flux, the average pore size (d) of the membrane was calculated by the Guerout–Elford–Ferry equation, Equation (4) [46].

$$\mathbf{d} = \sqrt{\frac{(2.9 - 1.75\varepsilon)8\delta \text{IV}}{\varepsilon \text{ A } \Delta \text{P t}}} \tag{4}$$

Herein, <sup>ε</sup> is membrane porosity (%), <sup>δ</sup>, the water viscosity (8.9 × <sup>10</sup>−<sup>4</sup> Pa s), l represents membrane thickness (60 × <sup>10</sup>−<sup>6</sup> m), V is the volume of the distilled water penetrating through the membrane (m3), t is the experimental time interval (s), A, the effective membrane surface area (m2), and ΔP is the working pressure (30 kPa).

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

#### *3.1. Landfill Leachate Characteristics*

Table 1 displays the key characteristics of the raw leachate sample of more than 10 years in age. The lower BOD5 to COD ratio (0.074) was another strong indication of highly stabilized leachate sample [3]. The other quality parameters of leachate, such as COD, BOD5, NH3-N, colour, and pH values, were around 1188 mg/L, 89 mg/L, 313 mg/L, 1360 PtCo/L, and 8.33, respectively. These obtained values were also compared with the standard discharged limits set by the Malaysian Environmental Quality was conducted (Table 1) [47]. As shown in Table 1, the COD, colour, and NH3-N concentrations were found to be far greater than the standard discharged limits.



#### *3.2. PAC Characterization*

Analysis test of the particle size was conducted to investigate the particle size distribution of fine samples in terms of volume. The particle size distribution of PAC sample is shown in Figure 5a. It can be seen from Figure 5 that PAC has small particle sizes which

varied between (0.02–50 μm) in diameter. The average particle diameter of the PAC is 25 μm. It is evident from Figure 5a that the distribution curve of PAC particles could be counted a uniform-distribution curve. The percentage of adsorption is higher for those adsorbents have smaller particle size due to the availability of more surface area [48]. The surface morphology of PAC was visualized via FESEM, with a magnification of 10,000×, as shown in Figure 5b. FESEM micrographs of PAC, shows uniform size particles, which confirmed the results obtained from the particle size analysis. To some extent, the PAC surface having small cavities, pores, and more rough surfaces indicates the presence of an interconnected porous network. Increasing the particles' number of an adsorbent material by decreasing its particles size resulted in increasing the adsorption surface area, and thus the material adsorption characteristics [49].

**Figure 5.** PAC characterization: (**a**) Particle size distribution; (**b**) FESEM image at 10,000× magnification.

#### *3.3. Membrane Filtration and Experimental Results*

Herein, the relationship among the independent factors (PVFD and PAC dosage in membrane) and responses (COD, NH3-N, colour removal, max. TEM, and pure water flux) were thoroughly investigated. There were 13 different experiments performed on the PVDF and PAC composition based on the central RSM composite design, as shown in Table 2. CFR test was performed to investigate the pollutants removal efficiency together with the max. TEM, while dead-end test was executed to measure the pure water flux.

**Table 2.** Experimental results for the PVDF-PAC membranes (RSM design).


\* Estimated by Equation (1). \*\* Estimated by Equation (2).

The COD, colour, and NH3-N removal efficiencies were found to be around 14.8–37.2, 14.6–56.3, and 7.5–23.8%, respectively, while the pure flux and max. TMP were ranged between 26.2–127.7 L/m2·h and 0.42–1.00 bar, respectively. ANOVA analysis was performed for the further investigation on the obtained experimental results.

It is observed from Table 2 that an increase in both PVDF and PAC concentrations on the membrane leads, to some extent, to an increase in the contaminants removal. When PVDF and PAC concentration are higher than 14 wt.% and 1.0 wt.%, respectively, the removal efficiency starts to decrease with increasing the amount of PVDF and PAC. This behaviour was attributed to the combination effect between polymer and additive in dope. This leads to the creation of large volume voids with increasing polymer dosage, and allows the small particles of contaminants to pass through the membrane [50].

#### 3.3.1. Removal Efficiency of Contaminants

Table 3 depicts the empirical model using the data obtained from COD, colour, and NH3-N removals. F-values of the model, together with the low probability values (P > F > 0.05), clearly suggest that the models were significant for all responses.

**Table 3.** ANOVA results and quadratic models of PVDF-PAC membranes for COD, colour, and NH3-N removal efficiencies.


<sup>a</sup> Significant. <sup>b</sup> Not significant.

The significant model terms for COD removals in the ANOVA analysis were sorted in descending order depending upon the influential terms (AB, B2, A, B, and A2). It was clearly seen that the PVDF and PAC (AB) had the highest impact on the COD removal with an F-value of around 32.25, followed by the quadratic term of PAC concentration (B2), PVDF concentration (A), PAC concentration (B), and finally the quadratic term of PVDF concentration (A2) with an F-value of 12.18, 4.34, 4.19, and 0.42, respectively. The quadratic terms of PVDF concentration together with the linear terms of PAC and PVDF contents caused a positive effect on the COD removal. Nonetheless, interaction and quadratic terms of PAC exhibited negative effects. In fact, an increase in the COD removal was recorded upon the change in the liner terms of PVDF and PAC concentrations, and PVDF concentration with quadratic term from lower to higher level. Hence, this change is complemented by the outstanding COD removal using PVDF-PAC membrane. On the other hand, a decline in COD removal was recorded when the interaction term and quadratic term of PAC was in the higher level.

The quadratic term of PVDF contents (A2) has the most significant effect towards the colour removal rate. This is due to the highest F-value (97.03), where other terms had the values of 21.10, 5.69, 3.89, respectively. The PAC content (B) had a progressive influence on the colour removing. However, the quadratic term of PVDF, PVDF concentration, and interaction among the PVDF and PAC displayed a negative effect. Thus, the removal of colour was enhanced with the enhancement of the PVDF contents in membrane fabrication until the optimum amount (>14 wt.% PVDF).

Additionally, in case of NH3-N removal, the A, B, B2, A2, and AB were sorted in descending order of their effecting strength. The highest F-value of 55.81 was recorded for the linear term of PVDF concentration (A), and thus it had the huge effect in NH3-N removal. On the other hand, the lowest F-value of 0.032 was recorded for interaction term, which regarded to have a negligible effect on the model. The PVDF linear term only offered a strong influence on removing of NH3-N, while the remaining terms were found to be the negligible influencers. Hence, the NH3-N removal was increased upon enhancing the PVDF contents in membranes. However, for PAC concentration after the point (PAC = 1.0 wt.%); when either the quadratic term of PVDF or PAC, or the interaction term is in the significant level, the NH3-N removal starts to decrease.

The lack of fit F-statistic was statistically not significant, as the values of (P) were higher than 0.05. A significant lack of fit suggests that there may be some systematic variation unaccounted for the proposed models. This may be due to the exact replicate values of the independent variables in the models that provide an estimate of pure error [15]. The correlation coefficient value (R2) resulted in the present study for COD removal (0.9482), colour removal (0.9411), and NH3–N removal (0.9456), indicating that only 5.18, 21.09, 5.89, and 5.44% of the total dissimilarity might not be explained by the empirical models. Zielinska et al. [10] stated that the correlation coefficient should be more than 0.80 for a good fit of a model. Moreover, the C.V.% of the obtained models for COD, colour, and NH3- N removals were 6.42%, 10.26%, and 7.52%, respectively, which designates an adequate model [51].

In the current study, all insignificant model terms which have limited effects were eliminated from the study to improve the model. Based on the findings, the response surface models for COD, colour, and NH3-N removal efficiency were constructed to predict responses, which were considered reasonable. The final regression models, in terms of their coded factors, are expressed by the second-order polynomial equations, and are presented in Table 3.

Typically, it is vital to study the effect of the operational factors on the different responses. The effect of PVDF and PAC concentration on the responses of COD, colour, and NH3-N removals over PVDF-PAC membranes could be evaluated using perturbation and three-dimensional (3D) response surface plots (Figure 6). Perturbation plots show the comparative effects of independent variables on the responses. For instance, in Figure 6, the different sharp curvatures in PVDF concentration (A) and PAC concentration (B) show that the three responses (COD, colour, and NH3-N removal efficiency) were very sensitive to the fabrication variables, but with different behaviours. In other words, PVDF and PAC contents have a major function in the treatment process under the experimental conditions. This is another confirmation of the important effects of the independent variables (PVDF and PAC concentrations) on the treatment removal efficiency. Therefore, the 3D surface response and contour plots of the quadratic models were utilized to assess the interactive relationships between independent variables and responses. The 3D response surface was introduced as a function of PVDF and PAC concentrations. Figure 6a,c shows a symmetrical 3D surface response for both COD and NH3-N removals. In the meantime, the removal of colour presents a different 3D surface (Figure 6b), which indicates that colour removal was influenced differently by experimental factors than the other responses.

**Figure 6.** Perturbation plots (**left**) and 3D response surface (**right**) of PVDF-PAC fabricated membrane for the removing efficiency of (**a**) COD, (**b**) colour, and (**c**) NH3-N.

Figure 6a,c indicated that the responses for COD and NH3-N removal rate was sufficiently enhanced upon the increase in PVDF contents in applied membranes. On the other hand, the increase in PAC contents in membrane fabrication led, to the removal of COD and NH3-N to some extent. It was seen that, when the PAC contents in membrane were higher than 1.0 wt.%, the removal rate for COD and NH3-N began to decline. According to Figure 6c, for the removal of NH3-N, the effect of interaction between PVDF and PAC concentrations have a noteworthy influence on removal percent. The NH3-N removal were

gradually increased with the increasing of PAC concentration to some extent, which means that the incorporated PAC has enhanced the membrane performance in terms of NH3-N removal, in addition to the main separation action gained by the membrane texture itself. This good result might be ascribed to the high adsorption characteristics of the used PAC, which significantly improved the fabricated membrane efficiency [5,52].

However, PVDF concentration has limited effect on COD removal efficiency compared with the PAC content. Where 35.5 and 38.5% of COD were removed at minimum and maximum PVDF concentration (10.0 and 18.0 wt.%), respectively, 18.5 and 35.5% of COD removal were removed at minimum and medium PAC concentration (0.0 and 1.0 wt.%), respectively. Likewise, the minimum NH3-N removal was found to be 7.5% at membrane concentration of 10.0 wt.% PVDF and 2.0 wt.% PAC, while the maximum NH3-N removal (24.5%) was observed at the PVDF and PAC concentration of 18.0 and 1.0 wt.%, respectively.

On the other hand, the 3D response surface in Figure 6b displays a different effect of interaction between the experimental factors on the colour removal rates. It was observed that an increase in the concentration of PVDF in the membrane leads to an improvement in the colour removal to some degree. When the concentration of PVDF was higher than (14 wt.%), the colour removal performance starts to decrease. This behaviour was credited to the combined effects of additive and polymer in the dope. This leads to creating large volume voids with increasing polymer dosage, and lets the fine particles from contaminants to permeate through the membrane [50]. Meanwhile, the enhancement of PAC concentration in a membrane drove a steady increase in the colour removal efficiency. As witnessed in Figure 6b, the predicted minimum and maximum efficiencies of colour removal were 15.0 and 56.5% present at fabrication concentrations of (18.0 wt.% PVDF, 0.0 wt.% PAC), and (14.0 wt.% PVDF, 1.0 wt.% PAC), respectively. This also confirms the effectiveness of PAC content in enhancing the removal performance of the filtration process using PVDF fabricated membrane.

Despite the incorporation of PAC into membrane enhancing the COD, colour, and NH3-N removal, the filtered leachate still did not meet the Malaysian Discharge Standard (Table 1). This is due to the highly concentrated pollutants of leachate that resulted in a reduction in membrane efficiency owing to the clogging caused by influent SS component. Therefore, a pre-treatment process such as PAC adsorption is suggested to be used before the membrane treatment [33].

#### 3.3.2. Pure Flux and Transmembrane Pressure Studies

By applying the factorial regression analysis on the experimental data related to PVDF-PAC membranes, both max. TMP and pure water flux responses were well agreed to a linear model of the second degree, as shown in the ANOVA analysis presented in Table 4.

In a general linear model or a multiple regression model: Y = *<sup>ß</sup>*<sup>0</sup> <sup>+</sup> *<sup>k</sup>* ∑ *ssi Xi* + *ε*, where: Y is

*i*=1 the response, *Xi* is the independent factor, k is the number of variables, *ß*<sup>0</sup> is the constant term, *ßi* represents the coefficient of the linear, and *ε* is the random error or noise [53]. The final linear models obtained for each response has been expressed by the first order polynomial equation, as presented in the last raw of Table 4.

The fitted model for the pure water flux suggests a large F-value (53.56), suggesting that the model is significant. As the value of Prob > F of all terms is less than 0.050, this suggests that all the model terms are significant. Based on their F-values, the PVDF concentration term (A) has the highest influence on the model, followed by PAC concentration term, and lastly the combination term. The term of PAC concentration presents a positive effect on pure flux, while the other two terms have been found to be negative influencers. Hence, the pure water flux was raised only with enhancing PAC contents in the membrane while, in contrast, it is decreased with the increasing of the PVDF content of a membrane.


**Table 4.** ANOVA results and quadratic models of PVDF-PAC membranes for pure flux and max. TMP.

<sup>a</sup> Significant. <sup>b</sup> Not significant.

On the other hand, the suggested model of max. TMP was significant with a high F-value (49.62), as can be seen from Table 4. Based on its effect on the model from the highest to the lowest, the model terms can be arranged as follows: PVDF content, PAC content, and the combination of both, with F-values of 131.07, 13.01, and 4.78, respectively. However, the PVDF concentration is the only factor which showed a positive influence on the max. TMP, due to the positive sign of its term; this indicates a worse impact on the max. TMP, as it could be increased with the increasing of PVDF content on the fabricated membrane. On the other hand, PAC concentration exhibited a better effect on the max. TMP, which showed a reduction in max. TMP occurred due to the increasing of the PAC content.

Additionally, both of the models display a non-significant lack of fit F-value, which indicates that well fitted models have been selected to present the experimental results with minor pure errors [15].

The R<sup>2</sup> values obtained in the present study for pure flux and max. TMP were 0.9470 and 0.9430, respectively. The high value of R<sup>2</sup> represents good agreement between the observed and the calculated results within the experimental ranges [37]. Moreover, C.V. % for the water flux and TMP were 10.50% and 6.56%, respectively. Where these small values indicate good fitness of the models [51].

Based on these findings, the resulted response surface models in the current work for predicting the two responses (pure flux and max. TMP) were considered reasonable.

The influence of integrated PAC and the interaction of content's concentrations on the max. TMP can be explored by the plots of perturbation and 3D response surface, as shown in Figure 7. From perturbation plots at Figure 7, it is easy to notice that pure flux and max. TMP responses are very sensitive to the experimental factors, and to conclude that both have a different (inversed) behaviour regarding the PVDF and PAC concentration values. As can be seen from Figure 7a, increasing of PVDF concentration (A) resulted in a linear decrease in pure water flux and increase in max. TMP, which attributed to the reduction in membrane porosity due to the increase in polymer concentration, which is well recognized for the system of a single polymer casting solution [50]. However, PAC concentration (B) showed a different effect, as any increase in its value causes a linear increment on the pure water flux, but a decrease in max. TMP.

**Figure 7.** Perturbation plots (**left**) and 3D response surface (**right**) of PVDF-PAC fabricated membrane for (**a**) pure water flux and (**b**) max. TMP.

Minimum and maximum predicted pure fluxes (26.0 and 128.5 L/m2·h) were found at the membrane compositions of 18.0 wt.% PVDF with 0.0 wt.% PAC, and 10.0 wt.% PVDF with 2.0 wt.% PAC, respectively. On the other hand, lowest and highest max. TMP according to the suggested model were found to be 0.38 and 0.98 bar at membranes of compositions (10.0 wt.%) PVDF with (2.0 wt.%) PAC, and 18.0 wt.% PVDF with 0.0 wt.% PAC, respectively. From the findings, membranes with lower PVDF concentration and high PAC concentration (10.0 wt.% PVDF and 2.0 wt.% PAC) exhibited the best water permeation and antifouling properties. Nonetheless, this membrane still falls short to produce the highest removing rates of COD, colour, and NH3-N based on the previous discussion.

#### *3.4. Fabricated Membrane Characterization*

The morphology of produced membrane can explain the effect of dope composition on membrane performance. A collection of membranes composed from different concentrations of PVDF and PAC (wt.%) were chosen from the fabricated membranes to represent the different membrane compositions, and consequently to be investigated by the morphological studies. These membranes were: FM1 with the content of (10.0 wt.% PVDF-0.0 wt.% PAC) to represent minimum PVDF concentration with no PAC; (10.0 wt.% PVDF-2.0 wt.% PAC) to represent minimum PVDF with high PAC, denoted as FM2; (14.0 wt.% PVDF-1.0 wt.% PAC) to represent intermediate composition of both PVDF and

PAC, named FM3; and finally FM4 with 18.0 wt.% PVDF and 0.0 wt.% PAC to represent maximum concentration of PVDF without PAC.

The FTIR spectrum of PVDF-PAC fabricated membranes with the various compositions is illustrated in Figure 8. It is clearly observed from Figure 8 that membranes displayed semi-typical distinctive spectra along the range of 4000 and 400 cm−1. Characteristic chemical groups are witnessed in the band of all membranes at waves with lengths 3020, 2990, 2370, 1400, 1070, 875, 590, and 490 cm−<sup>1</sup> with altered vibrations of strength depends on the different membrane compositions. The spectrum shows bands at 2990 and 3020 cm−<sup>1</sup> which are attributed to the symmetric and asymmetric stretching vibrations of C-H coming from ketones and carboxylic acids [54], where vibrations at 1070 and 1400 cm−<sup>1</sup> presented the deformation peaks of C-F related to PVDF.

**Figure 8.** FTIR spectra for PVDF-PAC membranes with different concentrations.

The notable peaks of the various membranes at 2370, 875, and 590 cm−1, assigned to CO2, CO3 −2, and C-O- groups, respectively, were the features distinctive of neutralization methanol, used after membrane casting [55]. Moreover, the OH group detected at 490 cm−<sup>1</sup> is attributed to the DW used for membrane solidification during the casting process [56]. Figure 8 also confirmed that the recorded wave numbers in the spectrum of both membranes without PAC (FM1 and FM4) have higher frequencies in comparison with the spectrums of the other two membranes with PAC content (FM2 and FM3).

Furthermore, it could be observed that the peaks of the membrane with higher content of PAC (FM2) have lower vibrations compared with the membranes with lower PAC content (FM3). Evidently, the peaks become narrow with less strength at the increasing of PAC weight, indicating that the hydrogen bonds were constructed well between PVDF polymer chains and the hydroxyl groups from PAC, which reduces the PVDF hydrophobic tendency [57]. These outcomes confirmed that PAC was well integrated to PVDF membranes, and partially relocated on the membrane surface, which leads to membrane treatment efficiency enhancement.

To investigate the elemental composition present in the fabricated PVDF-PAC membranes with different compositions, EDX analysis was recorded in the binding energy region from 0 to 15 keV as exhibited in Figure 9. The PVDF characterized elements C

and F were clearly observed in the spectra of the pure PVDF membranes (without PAC), while the AL element, which characterizes the presence of PAC, appeared only at the PVDF membranes incorporated with PAC [33]. Figure 9b,d shows the EDX analysis of 2.0 and 1.0 wt.% PAC, respectively. It is clearly witnessed that the presence of PAC was presented well.

**Figure 9.** EDX analysis for the selected PVDF-PAC fabricated membranes with different compositions: (**a**–**d**) for (FM1–FM4).

Table 5 shows the atomic percentages of the different elemental compositions of the selected membranes (FM1–FM4). From the EDX findings, the weight percentages of elemental AL on the FM2 and FM3 were determined as 1.04 and 0.79, respectively, which confirmed the presence PAC with representative weights on the integrated membranes.

**Table 5.** Elemental compositions of selected PVDF-PAC fabricated membranes based on EDX mapping.


Figure 10 presents the FESEM images for produced membranes with different compositions, which show the top surface morphology of membranes, along with its cross section. As can be seen from Figure 9a–d, there were many small pores available on the surface of FM1 membrane which contains the lowest PVDF polymer content (10.0 wt.%). Furthermore, the number and size of these pores start to be decreased, first on membrane FM3, with PVDF content 14.0 wt.% and PAC content 1.0 wt.%, followed by FM2 membrane with the highest PAC content (2.0 wt.%), while the membrane FM4 has a semi-impermeable surface due to its high PVDF polymer content (18.0 wt.%) with no PAC content. This was in agreement with the findings earlier discovered by Kunst and Sourirajan [58].

**Figure 10.** FESEM morphologies of PVDF-PAC membranes with different compositions (FM1 to FM4): (**a**–**d**) cross-sections and (**A**–**D**) top surfaces.

Referring to membrane cross sections on Figure 10a–d, all membranes display the formation of macrovoid with loosely packed structures. Typically, the membrane consists of two layers, which are a spongy porous support layer and a dense top finger-like layer. The establishment of these configurations can be attributed to the instantaneous demixing of polymer and solvent during the process of phase inversion.

FM1 membrane, with only PVDF and the weight of 10.0 wt.%, displayed an unimproved finger-like formation and a sponge-like support layer containing large, unconnected pores, delimited by polymer walls (see Figure 10a). The finger-like voids turn become flat, bigger, and even strained to the bottommost of the fabricated membranes with an increase in PAC concentration (i.e., in FM2 and FM3), and the spherical voids of the spongelike structures connect more closely with themselves (Figure 10b,c). However, the FM4 membrane, containing the highest concentration of PVDF, gives thin, smaller, and nonstretched figure-like pores, with less connection to the little sponge-like pores located on the cross-section's bottom. This produces low membrane flux due to the greater amount of polymer contributing a higher membrane viscosity, which lead to a decease in the membrane porosity and pore size. The overall FESEM micrographs have proved the significant effect of the PAC presence in improving the fabricated membrane characteristic in terms of membrane rejection, and therefore removal rate of contaminants.

Furthermore, an AFM test was carried out to investigate the membrane top surface, along with its roughness, as shown in Figure 11. The FM2 membrane might contain some extra PAC particles which made its top surface rougher compared to others (Figure 11b). Having less depth of facial peaks and valleys, the FM4 membrane surface (Figure 11d) is relatively smooth due to containing only PVDF polymer which received a homogeneous mixing at the preparation phase of dope solution [59]. However, the peaks and valleys of FM1 and FM3 membranes reduced gradually compared to FM2, where FM3 has the smoothest surface compared with other membranes (see Figure 11a–d). To confirm all above observations, the values of membrane surface roughness (Rq and Ra) given in Figure 11 can be considered.

**Figure 11.** AFM top surface images with average membrane roughness values (nm) for different compositions of selected PVDF-PAC membranes: (**a**–**d**) for (FM1 to FM4).

For membrane permeability analysis, the impact of PAC addition to membrane permeability, in terms of porosity and average pore size, were evaluated for the produced PVDF membranes. As presented in Table 6, the porosity and average pore size of the fabricated PVDF membranes incorporated with PAC were higher than the other membranes without PAC. Based on Table 6, the resulted fabricated membranes were "micro-filtration", and the highest mean values of porosity and average pore size were achieved by FM2 membrane at the values 77.48% and 24.43 μm, respectively. On the other hand, the lowest values of the same corresponding permeability parameters were found using membrane FM4 at 48.38% and 12.15 μm, respectively. These findings agreed with the above morphological results.


**Table 6.** Permeability measurements for selected fabricated PVDF-PAC membranes.

<sup>a</sup> Each parameter is expressed as average value ± standard deviation.

#### *3.5. Membrane Treatment Optimization*

The best synthesized membrane has been selected using the RSM tool, where the membrane efficiencies of COD, colour, and NH3-N removal were optimized during this study.

Based on the DoE software, the operational conditions (PVDF weight and PAC weight) were targeted to be within the range. While the dependents of treatment performance (COD, colour, and NH3-N removal) were chosen as ''maximum" to achieve the ultimate filtration treatment. The other responses were remained "within the range". Accordingly, the optimization tool assimilates the singular desirability into a particular number, and then aims to optimize the function.

Consequently, the composition of the optimum membrane, together with respective rates of removal efficiency, were obtained. The optimum removals and their corresponding water flux and max. TMP are presented in Table 7.

**Table 7.** Predicted and experimental removal efficiencies of the optimum PVDF-PAC membrane with the corresponding operating condition.


\* Optimum value.

The membrane with 14.9 wt.% of PVDF and 1.0 wt.% of PAC was found to be the optimum, and thus selected as the best membrane design, having optimum removal efficiency according to its highest desirability (0.870) [60].

As shown in Table 7, 35.34, 48.71, and 22.00% removal of COD, colour and NH3- N, respectively, was predicted by the software under optimized operational conditions. The corresponding (non-optimized) water flux and max. TMP were found at the values 61.00 L/m2·<sup>h</sup> and 0.67 bar, respectively. An additional experimentation was then performed to confirm the optimum findings.

As illustrated in Table 7, the error column indicates the differences between the predicted and laboratory values, which shows that the lab experiments agree well with the response values estimated by the software. However, less agreement between the predicted and the laboratory result was obtained in case of NH3-N removal (8.36% error).

#### *3.6. Membrane Performance Comparison with Other Reported Studies*

The performance of the optimum fabricated membrane with other reported PVDF produced membranes is shown in Table 8. It can be noticed from Table 8 that the current study offered the smoother surface among the existing works based on the average roughness (Ra = 36.39 nm), which accordingly improves the removing performance and antifouling properties of the created membrane [61]. There exists few values of pure flux that are higher than the achieved in the current work, such as the flux of 143.24 L/m2·h produced by Penboon et al. [62]. The low value of pure water flux of the current work (61.00 L/m2·h) could be ascribed to the differences in the experimental characteristics such as the type of wastewater or the rates of feed flow. In addition, the rejection efficiency in the current work is lower than previously reported, which could be solved through further enhancement of the produced membranes using hydrophilic additives such as PVA or PVP [25,57]. Based on previous studies, after saturation, membrane corroborated PAC can be washed back and reused [13,63].


**Table 8.** Comparison of performance with other modified PVDF membranes in wastewater treatment process.

#### **4. Conclusions**

The adsorbent material PAC was used to fabricate a novel PVDF membrane for the treatment of stabilized landfill leachate. The fabricated PVDF flat sheet membranes integrated with PAC showed better performance when compared with PVDF membrane (without PAC). The addition of PAC effectively enhanced the removal rate and the fouling control parameters of produced membranes. Increasing PAC content to a certain value has a positive influence on the removal efficiency of COD, colour, and NH3-N, as well as on membrane characteristics. Operational optimization was performed using RSM to select the optimum membrane design in terms of the removal efficiency. The best membrane composition was found at (14.9 wt.%) PVDF and (1.0 wt.%) PAC, which removed 36.63% of COD, 49.50% of colour, and 23.84% of NH3-N. This was in agreement with the predicted removals. The corresponding experimental values of water flux and max. TMP also agreed with the prediction, with the values of 61.10 L/m2·h and 0.64 bar, respectively. The performance and structure of fabricated membranes were investigated by filtration tests, FTIR, FESEM, and AFM spectroscopy. In general, this work shows the potential of treatment and hydrophilic improvement of hydrophobic PVDF polymer membranes using

PAC. For further removal efficiency, membrane properties or practice could be improved by either adding a hydrophilic material, or applying pre-treatment process such as adsorption via PAC.

**Author Contributions:** S.M.A.A.: experimental work, writing original draft, preparation, and revisions. Z.H.J.: visualization, investigation, and language reviewing. C.-A.N.: supervision. Y.-C.H.: funding and technical support. M.J.K.B.: supervision, conceptualization, methodology, software, and revision. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded by Higher Education Ministry for their fund (FRGS/1/2019/TK10/ UTAR/02/3 and PETRONAS through YUTP grant (015LC0-169).

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

**Informed Consent Statement:** Not applicable.

**Data Availability Statement:** Not applicable.

**Acknowledgments:** The authors are thankful to the Higher Education Ministry for their fund (FRGS/1/2019/TK10/UTAR/02/3). This research was funded by PETRONAS through YUTP grant (015LC0-169).

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

#### **References**


### *Review* **Treatment of Poultry Slaughterhouse Wastewater with Membrane Technologies: A Review**

**Faryal Fatima 1, Hongbo Du <sup>1</sup> and Raghava R. Kommalapati 2,\***


**Abstract:** Poultry slaughterhouses produce a large amount of wastewater, which is usually treated by conventional methods. The traditional techniques face some challenges, especially the incapability of recovering valuable nutrients and reusing the treated water. Therefore, membrane technology has been widely adopted by researchers due to its enormous advantages over conventional methods. Pressure-driven membranes, such as microfiltration (MF), ultrafiltration (UF), nanofiltration (NF), and reverse osmosis (RO), have been studied to purify poultry slaughterhouse wastewater (PSWW) as a standalone process or an integrated process with other procedures. Membrane technology showed excellent performance by providing high efficiency for pollutant removal and the recovery of water and valuable products. It may remove approximately all the pollutants from PSWW and purify the water to the required level for discharge to the environment and even reuse for industrial poultry processing purposes while being economically efficient. This article comprehensively reviews the treatment and reuse of PSWW with MF, UF, NF, and RO. Most valuable nutrients can be recovered by UF, and high-quality water for reuse in poultry processing can be produced by RO from PSWW. The incredible performance of membrane technology indicates that membrane technology is an alternative approach for treating PSWW.

**Keywords:** poultry slaughterhouse wastewater; microfiltration; ultrafiltration; nanofiltration; reverse osmosis

#### **1. Introduction**

Water is essential for all lives and a natural resource at the core of sustainable development. It is critical for socio-economic prosperity, healthy ecosystems, and human survival. Unfortunately, water is a finite and irreplaceable resource in time and space. The increase in water consumption has made water management a priority. On the other hand, improper wastewater treatment in some regions has intensified the inadequate discharge of wastewater into the environment and augmented natural water resource pollution. As a result, progressively stricter standards for effluent discharge worldwide have changed the target from wastewater disposal to water reuse and recycling, leading to advanced wastewater treatment technologies, which can recycle and reuse wastewater [1].

Food industries such as dairy, beverage, vegetable, fruit, oilseed, seafood, poultry, and other types of meat consume a high volume of freshwater. Among them, the poultry industry is at the top [2]. From 2018 to 2019, the world poultry market increased by 6% due to an increase in the per capita poultry consumption, which corresponds to 58 kg per person in the U.S., 57 kg per person in Brazil, and 48 kg per person in Peru. The high demand for poultry meat correspondingly increases freshwater consumption by poultry processing plants [3].

Poultry slaughterhouses discharge massive amounts of wastewater into the environment because of their high freshwater usage for the continuous operations of cutting

**Citation:** Fatima, F.; Du, H.; Kommalapati, R.R. Treatment of Poultry Slaughterhouse Wastewater with Membrane Technologies: A Review. *Water* **2021**, *13*, 1905. https://doi.org/10.3390/w13141905

Academic Editors: Amin Mojiri and Mohammed J. K. Bashir

Received: 1 June 2021 Accepted: 7 July 2021 Published: 9 July 2021

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

**Copyright:** © 2021 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https:// creativecommons.org/licenses/by/ 4.0/).

up, rinsing, and packaging meat. Other operations in poultry slaughterhouses such as scalding, de-feathering, evisceration, and bird wash are also water-intensive and generate a significant amount of wastewater. The eviscerating step and bird wash generate enormous wastewater, at 7.57 L/bird and 4.35 L/bird, respectively, as shown in Figure 1. On average, a 2.3 kg bird consumes 26.5 L of water [4,5]. The wastewater is highly contaminated with organic matter quantified as biochemical oxygen demand (BOD) and chemical oxygen demand (COD). It also contains high nitrogen and phosphorous constituents, including blood, fats, oil, grease, and proteins [6]. Thus, discharging improperly treated poultry slaughterhouse wastewater (PSWW) has a high risk of polluting freshwater sources. It can also cause serious environmental and health concerns such as deoxygenation of rivers, groundwater contamination, eutrophication, and the spread of water-borne diseases [7,8].

**Figure 1.** Water consumption during poultry processing in the poultry slaughterhouse.

Generally, PSWW is treated by physical, chemical, and biological methods. These conventional techniques are only responsible for discharging the treated water into the environment without recycling it. Besides, they face some challenges, such as lack of nutrient recovery, frequent use of chemical cleaning agents, and the degradation of valuable compounds in wastewater. Therefore, unconventional methods, e.g., pressure-driven membrane technologies, are being explored for PSWW treatment. These membrane filtration technologies can include microfiltration (MF), ultrafiltration (UF), nanofiltration (NF), and reverse osmosis (RO) [9,10]. They can overcome some limits of the conventional methods by removing colloids and suspended and macromolecular matter, and eliminating mineral substances and low-molecular organic compounds. Membrane filtration is a physical process that provides great separation efficiency and improves final product quality. Most importantly, membrane technology can produce water clean enough for the reuse of the treated water in industrial poultry processing. Furthermore, it can recover a fair amount of valuable nutrients, e.g., proteins, which could be utilized as animal feed, thus supplementing the global protein demand for animals [9,11,12].

This review identifies membrane technologies for advanced PSWW treatment. It evaluates the quality of treated water based on pollutant removal efficiency and the degree of permeate produced to meet the environmental regulation of discharging and recycling standards. In addition, we highlight the potency of membrane technologies to recover valuable nutrients from PSWW, which is not achieved by traditional methods.

#### **2. Characteristics of Poultry Slaughterhouse Wastewater**

Before any wastewater treatment, it is critical to characterize the wastewater to show pollutant levels in using some instruments and test reagents [7]. The parameters commonly used to describe PSWW are pH, COD, BOD, total organic carbon (TOC), total suspended solids (TSS), total nitrogen (TN), total phosphorus (TP), and pathogens [13,14]. COD indicates the amount of organic compounds in wastewater; a high concentration of COD suggests a large amount of oxidizable organic substances in the wastewater. Similarly, BOD indicates the biological oxidation of organic compounds, and a high BOD level also signifies large quantities of organic pollutants in wastewater. Nutrients in wastewater are TN and TP; nitrogen in wastewater is available in the organic form, primarily present in proteins, and the inorganic form, which includes nitrite (NO2 −) and nitrate (NO3 −). The most stable type of nitrogen in water is nitrate that originates from a natural decaying process of biological matter. Excessive nitrates in wastewater can lead to harmful algae bloom, oxygen depletion, fish poison, and putrid odors. Moreover, in wastewater, orthophosphate (PO4 <sup>3</sup>−) is the most common type of phosphorus, originating from disinfectants and cleaning agents; high phosphorus constituents in wastewater may prompt eutrophication. To reduce phosphorus in wastewater, a practical and straightforward technique is chemical precipitation [15]. These parameters vary from one slaughterhouse to another due to many factors, such as system type, operation method, and processing capacity [9]. The characteristics of PSWW, effluent discharge standards, and water reuse applications are tabulated in Table 1. It is required to treat PSWW to or below the standard limits because these measured parameters of raw PSWW are much higher than the acceptable standards by the World Health Organization (WHO) and other regulations [16].

*Water* **2021**, *13*, 1905


**Table1.**PSWWcharacteristics,discharginglimits,andreclaimingwastewatertreatmentgoals[6,15,17–20].

 Not reported; CFU = colony-forming unit.


#### **3. Conventional Treatment of Poultry Slaughterhouse Wastewater as Pretreatment Prior to Membrane Separation**

The conventional treatment for PSWW is similar to municipal wastewater treatment, consisting of preliminary, primary, and secondary treatments. There are various combined treatment methods after preliminary treatment, and the most common combination is physicochemical treatment as primary and biological treatment as secondary, as described below [18]. Prior to PSWW membrane filtration, some conventional treatment is necessary as pretreatment to alleviate membrane fouling and improve overall membrane performance. Without proper pretreatment, membranes will suffer severe membrane fouling, hindering the membrane performance, and some heavy fouling could even cause membrane failure. For example, Meiramkulova et al. [21] investigated the performance of an integrated membrane process with electrochemical pretreatment on PSWW purification, and their findings showed that the electrochemical pretreatment was highly efficient at reducing turbidity, color, TSS, COD, and BOD by 71–85%. In addition, the pretreatment resulted in a low rate of cake formation on the membrane [21].

#### *3.1. Preliminary Treatment*

The preliminary treatment removes suspended solids from PSWW; the most common preliminary treatment uses screeners, sieves, and strainers. A typical wire mesh screen retains solid fraction with a size of 10–30 mm. Rotary screeners extract solids greater than 0.5 mm diameters; they also protect the equipment from fouling, clogging, and jamming. In preliminary treatment, 60% of suspended solids and 30% of BOD are removed from PSWW. Other preliminary methods are catch basins, flotation, equalization, and settlers [1].

#### *3.2. Primary Treatment*

After the preliminary treatment, PSWW goes through primary treatment in which the physiochemical process eliminates BOD, COD, oil, grease, fats, and residual TSS. The typical primary methods are dissolved air flotation (DAF), electrocoagulation, coagulationflocculation, and sedimentation [18].

#### 3.2.1. Dissolved Air Flotation

DAF is a separation of solid from liquid in which air is introduced into PSWW from the bottom of the tank, resulting in moving the light solids, grease, and fats on the surface, creating a sludge blanket. The efficiency of a DAF system can be improved by adding polymers and other flocculants. Generally, DAF's efficiency at removing BOD and COD is 30–90% and 70–80%, respectively [1]. Additionally, DAF removes suspended solids from PSWW in the range of 38–70%, and it eliminates fats in the range of 63–95%. The drawbacks of DAF are regular malfunctioning and poor TSS separation [2].

#### 3.2.2. Coagulation-Flocculation and Sedimentation

In the coagulation process, the colloidal particles present in the PSWW are grouped with large particles to form flocs. Those colloidal particles are nearly negatively charged, so they can be destabilized by adding positively charged coagulants to rescind the formation of flocs and ease the sedimentation process [18]. Previous research showed that this process could remove oil, grease, and TSS by up to 85%, and the removal efficiency was reported as 62–78.8% of BOD and 74.6–79.5% of COD [22]. However, this process results in toxicity and health hazards, inefficient removal of heavy metals and emerging contaminants, and an increase in effluent color [23].

#### 3.2.3. Electrocoagulation

Electrocoagulation (EC) is an advanced method for removing large amounts of pollutants from wastewater, such as organics, heavy metals, and pathogens, using electric current. The EC process generates M3+, Fe3+, or Al3+ ions using different electrode materials; the most commonly used electrodes are Al and Fe [18]. EC is a three-step process. In the first step, electrolytic oxidation forms metal hydroxides and oxyhydroxides at the sacrificial electrode. Then, the produced coagulants destabilize and adsorb the pollutants. Finally, flocs formed by aggregation of the destabilized phase are removed by a downstream sedimentation and filtration process [24]. The EC sets up sacrificial anodes that need to be changed regularly, and chlorinated toxic compounds can form if chlorine is present. In some regions where electricity is expensive, the cost of operating EC is high [25].

#### *3.3. Secondary Treatment*

The pollutants present in wastewater that are not removed by primary treatment are further treated by secondary treatment. The main goal of secondary treatment is the removal of organic compounds to reduce the BOD level. In the secondary treatment, the biological process, aerobic and anaerobic digesters are used for treating PSWW [1]. In both treatments, organic matter is degraded into simple compounds with the help of decomposers, where the efficiency of the decomposers depends on the quality of wastewater [2].

#### 3.3.1. Anaerobic Digestion

In biological treatment, anaerobic digestion of organic waste, sludge, and highstrength wastewater is a widespread technique [26]. The anaerobic system's primary goal is to reduce high-level BOD [27]. Anaerobic digestion consists of hydrolysis, acidogenesis, acetogenesis, and methanogenesis steps. With the help of a diverse group of microorganisms (bacteria and archaea), complex organic compounds are degraded in the absence of oxygen. The degradation rate relies primarily on various bacterial activity rates [26]. In the anaerobic treatment, organic compounds are broken down into methane, water, and carbon dioxide by anaerobic bacteria in an anaerobic environment [28]. However, PSWW usually has high organic strength, which can negatively affect the anaerobic process's performance. Therefore, an anaerobic system for PSWW is often followed by additional treatment to remove TP, TN, and pathogenic microorganisms [2].

#### 3.3.2. Aerobic Digestion

The main goal of aerobic digestion is nitrification [27]. Aerobic digestion uses oxygen to break down organic matter and other pollutants; it degrades ammonia or other organic matter into less harmful compounds like carbon dioxide, water, and nitrate. The oxygen and time needed for this treatment depend on the organic strength of PSWW. Aerobic digestion is usually applied as the last nutrient removal when using anaerobic techniques for the decontamination of sludge water. Some drawbacks of aerobic digestion are daily maintenance, excess biomass production, and increased demand for oxygen and electricity [26].

#### **4. Membrane Technology for Poultry Slaughterhouse Wastewater**

Pressure-driven membrane processes such as MF, UF, RO, and NF are widely studied for wastewater treatment throughout the world. Pressure-driven membranes rely on hydraulic pressure to achieve separation [29]. Membrane filtration is one of the most emerging technologies to produce high-quality water because it utilizes zero chemical constituents and offers enormous advantages over conventional methods. Several research groups have reported the use of membrane technology for PSWW treatment [12]. Jason et al. [30] first used membrane technology for PSWW treatment with recovering nutritional by-products in the 1980s; they experimented on a laboratory scale with a commercial tubular UF membrane with a molecular weight cut-off (MWCO) of 50 kDa. The study reported that the membrane technology produced permeate with a significant reduction of 85% of TS and 95% of COD; the permeate was believed to be safe for discharge and potential reuse. In addition, it recovered 24–45% of fat and 30–35% of protein as byproducts. Since then, some scientists have experimented with membrane technology for PSWW treatment used as a standalone or an integrated process. The available literature on membrane applications in PSWW treatment is mainly based on UF membranes; in a few

other research activities, such as MF, NF, RO, and membrane distillation, experiments have also been carried out. All the studies conducted on PSWW using membrane technology showed excellent performance in separation efficiency and compliance with environmental regulations.

#### *4.1. Characteristics of Pressure-Driven Membrane*

A pressure-driven membrane separates the feed into concentrate and permeate using the pressure difference as a driving force to transport the liquid or gas. Pressure-driven membranes such as MF, UF, NF, and RO membranes differentiate based on their characteristics [31]. The most important characteristics of pressure-driven membranes are pore size, structure, and operating pressure, as presented in Table 2. In reference to structure, membranes can be divided as symmetrical or asymmetrical; symmetrical membranes show uniform pore sizes in their cross-section, whereas asymmetrical membranes' pore size gets larger farther from the filter surface [11]. All the pressure-driven membranes are asymmetrical except for some MF membranes [32]. As shown in Figure 2, an MF membrane has the largest pore, highest permeability, and it can reject large suspended particles. A UF membrane has a smaller pore size and lower permeability than MF membranes and can separate small suspended particles and macromolecules. An NF membrane has the properties of the second smallest pore size, the high rejection of multivalent ions, but the low rejection of monovalent ions. In contrast, an RO membrane has a very high rejection of monovalent ions. It can be seen that as the pore size becomes smaller, the operating pressure increases [31]. For PSWW treatment, the membranes are chosen according to the pollutant levels in PSWW. A UF membrane is used to remove a substantial amount of suspended solids present in PSWW, and an RO membrane can eliminate all the pollutants, such as BOD, COD, TSS, salts, etc.

**Table 2.** Characteristics of pressure-driven membranes [29,32–34].


Most membranes for MF, UF, RO, and NF are made from synthetic organic polymers. MF and UF membranes are usually produced from the same materials under different membrane formation conditions to achieve different pore sizes. The commonly used polymers for MF and UF membranes are PVDF, PS, poly (acrylonitrile), poly (ether sulfone), and copolymers of poly (acrylonitrile) and PVDF. The materials used for MF membranes also include nylons, poly (tetrafluoroethylene), polypropylene, polyethylene, and blends of CA and cellulose nitrate. NF membranes are made from CA blends or PA composites, whereas RO membranes are produced from CA or PS coated with aromatic PA. Moreover, ceramic and metals are used to create inorganic membranes. Ceramic membranes are microporous, thermally stable, chemically resistant, and mostly used for MF, UF, and NF [31,33]. Metallic membranes are usually fabricated from stainless steel and can be very finely porous. The most common configurations of membrane modules are plate-andframe, spiral wound, tubular, and hollow fiber.

**Figure 2.** Schematic representation of pressure-driven membranes.

#### *4.2. Microfiltration*

The pore size of MF membranes is in the range of 0.05–10 μm, and the operating pressure of MF is between 0.1 and 2 bar for separating colloids and particles, reducing the effluent's turbidity and COD [34,35]. MF membranes have been used widely in industries such as liquid clarification and wastewater treatment, especially as an initial filtration stage for wastewater treatment [36]. Goswami and Pugazhenthi [12] evaluated the performance of fly ash tubular MF membranes with a pore size of 0.133 μm and porosity of 40.17% for PSWW treatment. The study revealed that the MF membrane produced a filtrate with zero turbidity and almost 100% removal of COD and TSS, thus satisfying the COD and TSS norms for discharging and reusing, which can help attenuate the water shortage crisis. A study was led by Marchesi et al. [37] to recycle the pre-chiller wastewater generated from the poultry carcass chilling process by using MF membranes. Both hollow fiber PA membrane with a pore size of 0.20 μm on a bench scale and spiral-wound membrane with a 0.1 μm pore size on a pilot scale provided complete retention of turbidity, apparent color, fat, and microorganisms. The rejection efficiency was up to 92.5% of COD, up to 89.1% of TOC, and 100% of microorganisms; it was stated that the use of membranes was a promising approach for the recycling and reuse of poultry pre-chiller wastewater. Abboah-afari and Kiepper [38] investigated the effects of pore size on the performance of MF membranes for the treatment of pre-DAF poultry processing wastewater. The results indicated that the 0.3 μm PVDF membrane was the most effective among the three tested membranes (0.3 μm PVDF, 0.1 μm PS, and 100,000 MWCO Ultrafilic). In detail, the 0.3 μm PVDF achieved a maximum mean permeate flux of 115 L/m2/h, and the removal efficiency was 88% of COD and 34% of TS. Moreover, in their other research work, they identified the 0.3 μm PVDF membrane as an alternative to DAF in poultry processing wastewater [39].

#### *4.3. Ultrafiltration*

The pore size of UF membranes is in the range of 0.002–0.05 μm, and the operating pressure of UF is between 1 and 10 bar for separating macromolecules and suspended solids [34]. The UF process has been extensively explored for PSWW treatment due to its significant advantages, such as low pressure, high permeate flux, cost-effectiveness, and the capability to eliminate pathogens that are very harmful to PSWW recycling [40]. The UF's transport properties are influenced by concentration polarization, fouling, and interactions between the feed stream and the membrane [9]. The UF process is considered an economical and environmentally friendly substituent for conventional wastewater treatments by some researchers [9]. Coskun et al. [41] studied the PSWW treatment using laboratory-scale membrane processes. Their study reported that UF as pretreatment improved the removal efficiencies for NF and RO processes; NF reduced almost 90% of COD, RO removed 97.4% of COD, and the UF pretreatment resulted in higher final fluxes 8.1 and 5.7 times more for NF and RO, respectively, than for those without UF. Yordanov et al. [9] examined the efficiency of a UF 25-PAN membrane for PSWW treatment. The results indicated that UF exhibited excellent performance by removing 97% of BOD and 94% of COD and that it also reduced 99% of TSS and 98% of fats. Rinquest et al. [13] treated PSWW using a UF membrane system for the removal of organic matter and suspended solids coupled with aerobic single-stage nitrification-denitrification (SSND) and an anaerobic static granular bed reactor, as shown in Figure 3. The experimental efforts showed that the UF system further reduced COD and TSS by an average of 65% and 54%, respectively, after SSND. All the measured parameters of final effluent excluding PO4 <sup>3</sup><sup>−</sup> and NH4 +-N satisfied industrial wastewater discharge requirements. Moreover, water flux higher than 200 L/m2/h was obtained by identifying the optimal condition. Mannapperuma and Santos [42] assessed UF for reconditioning poultry chiller overflow. The UF operation reduced microbial counts by more than 5.4 log cycles and achieved rejection efficiency of over 73% for COD and above 99.2% for turbidity. Their results verified that UF produced water acceptable for reuse in the chiller to replace freshwater makeup. Meiramkulova et al. [20] evaluated the performance of an integrated PSWW treatment with electrolysis, UF, and ultraviolet radiation in terms of microbial inactivation from PSWW. The results showed that the integrated system achieved an overall microbial removal efficiency of 99.86–100%.

**Figure 3.** Static granular bed reactor coupled with single-stage nitrification-denitrification and ultrafiltration for PSWW treatment (reproduced with permission from [13]: Rinquest, Z.; Basitere, M.; Ntwampe, S. K. O.; Njoya, M. Poultry Slaughterhouse Wastewater Treatment Using a Static Granular Bed Reactor Coupled with Single-Stage Nitrification-Denitrification and Ultrafiltration Systems. *J. Water Process Eng.* **2019**, *29*, 100778. https://doi.org/10.1016/j.jwpe.2019.02.018. Accessed on 14 March 2021. Copyright (2019), Elsevier).

#### *4.4. Nanofiltration*

The pore size of NF membranes is in the range of 0.001–0.002 μm, and the operating pressure of NF is between 5 and 20 bar for separating low molecular weight particles [34]. The NF process is a great separation tool due to its versatile properties, which fall between UF and RO. It removes a large amount of multivalent inorganic salts and small organic molecules while operating at a moderate pressure; the moderate operating pressure makes the separation process consume little energy and is cost-effective [43]. Therefore, it has been applied to various industrial sectors for wastewater treatment, e.g., water recycling during fishmeal, lupin bean, and textile processing [44–46]. A few researchers explored NF for PSWW treatment by using standalone NF or combining NF with UF. Zhang et al. [10] evaluated some membrane filtration processes for poultry abattoir wastewater treatment to recycle the wastewater stream to meet the Canadian poultry wastewater reuse criteria. Their results showed that both NF membranes (DS: desal thin composite membrane and NF 45: thin-film composite membrane) produced permeate with less than 100 mg/L of TOC and gave a reasonable flux of 46 to 66 L/m2/h. The TOC level of permeate produced by NF satisfied the Agriculture and Agri-Food Canada criteria for recycled water. However, the tested UF membranes did not meet the TOC criterion, although UF removed all bacteria and significantly reduced other organic species.

#### *4.5. Reverse Osmosis*

The operating pressure of RO is between 10 and 100 bar, and it can remove microparticles smaller than 0.001 μm, carbohydrates, amino acids, or even monovalent ions, including NH4 +, from water [34]. The RO process has widely been used for seawater desalination since it is the most effective on a large scale [47]. It is shown that combining biological treatment with RO produces a permeate with a quality superior to the WHO standards for drinking water [48,49]. Bohdziewicz et al. [50] investigated the application of membrane processes such as UF and RO to treat the wastewater of the meat industry. The study revealed that the hybrid system consisting of RO was permissible in the removal potency of 100% phosphorus and 98.8% nitrogen compounds. The removal efficiency of COD and BOD both exceed 99%, and only the permeate produced by RO was satisfied for the reuse in the production cycle. Meiramkulova et al. [51] evaluated the performance of an integrated process for PSWW treatment on both laboratory and industrial scales; the RO step was designed to reduce the total salinity of the water. Their results showed that the removal efficiency was up to 100% of turbidity, color, and TSS, and 99.6% of BOD and COD for laboratory and industrial testing. Almost all the physical and chemical parameters of the produced water were within the recommended standards set by legislation. The water purified on an industrial scale was certified as excellent quality in terms of Kazakhstan's drinking water quality standards.

#### *4.6. Membrane Bioreactor*

A membrane bioreactor (MBR) is an integrated system with membrane filtration for the biological degradation of waste present in wastewater. Generally, it is composed of a biological unit and a membrane module, which separates water from the aerobically digested water and returns activated sludge to the biological unit, as shown in Figure 4. The MBRs can remove organic and inorganic contaminants and biological entities from wastewater [52]. They are widely used for recycling water in buildings, wastewater treatment for small communities, and industrial wastewater treatment by producing an effluent free of bacteria and pathogens. Williams [53] treated PSWW using MBR coupled with a single-stage nitrification-denitrification reactor and an expanded granular sludge bed reactor (EGSB). The study reported that the overall removal efficiency of the EGSB-SSND-MBR system was 99% of turbidity, 92% of TSS, and 99% of COD. Fuchs et al. [54] used a cross-flow MBR for PSWW treatment. The MBR produced effluent with a removal efficiency of over 90% of COD. Gürel and Büyükgüngör [55] investigated MBR to extract organic substances and nutrients from wastewater in the slaughterhouse plant. Hollow-

fiber UF membranes with a pore size of 0.03 μm were used in the bioreactor. The removal efficiency of MBR was reported as 97% of COD, 96% of TOC, 65% of TP, and 44% of TN. The COD and TOC levels of permeate were 16 and 9 mg/L, respectively, which complied with the discharge limits of the slaughterhouse plant. Meyo et al. [56] treated PSWW using a pretreatment stage, an EGSB, and an MBR, and their results showed that MBR as a final stage treatment further reduced over 95% of TSS and COD.

**Figure 4.** Schematic representation of an MBR process (reproduced with permission from [57]: Poerio, T.; Piacentini, E.; Mazzei, R. Membrane Processes for Microplastic Removal. *Molecules* 2019, 24, 4148. https://doi.org/10.3390/molecules24224148. Accessed on 15 March 2021. Copyright (2019), MDPI).

#### *4.7. Vacuum Membrane Distillation*

Membrane distillation (MD) is a thermal-based membrane separation process that was introduced in 1963. The driving force in MD is the vapor pressure difference across the hydrophobic membrane instead of the applied absolute pressure difference [58]. The MD process can be used for water desalination, removal of organic matter in drinking water production, water and wastewater treatment, recovery of valuable components, and treatment of radioactive wastes [59]. Vacuum membrane distillation (VMD) is one configuration of MD in which the permeate side is vapor or water under reduced pressure [60]. Bialas et al. [61] conducted protein and water recovery from poultry processing wastewater using an integrating process of MF, UF, and VMD. A hydrophilic PVDF MF membrane was used in the pretreatment to extract suspended solids from the processing water, and a cellulose UF membrane was used to isolate soluble protein. During UF, the COD concentration of the permeate consistently surpassed the maximum permissible level. As a result, at the final stage, VMD was used. The membrane was a flat-sheet hydrophobic membrane made of polypropylene with a pore size of 0.2 μm. The study showed that the increase in temperature and the decrease in downstream pressure led to a considerable increase in the permeate flux, as shown in Figure 5. The removal efficiency of COD, TSS, TN, and total organic matter extractable by petroleum ether (TOEM) were very high, exceeding 99%. Moreover, VMD retained 93.3% of protein; the 6.7% loss of protein could have been due to adsorption of proteins to the membrane surface and denaturation of the proteins due to the high temperature. In terms of consistency, the permeate obtained through VMD was comparable to the RO permeate. The integrated process comprising MF, UF, and VMD made it possible to recover 70% of the water. The performance of VMD along with other pressure-driven membrane technologies used in the PSWW treatment is summarized in Table 3.

*Water* **2021**, *13*, 1905



reported;TPispresentedbyphosphates.


**Figure 5.** Response surface for permeate flux, J, as function of downstream pressure and feed temperature (reproduced with permission from [61]: Białas, W.; Stangierski, J.; Konieczny, P. Protein and Water Recovery from Poultry Processing Wastewater Integrating Microfiltration, Ultrafiltration and Vacuum Membrane Distillation. *Int. J. Environ. Sci. Technol.* **2015**, *12*, 1875–1888. https: //doi.org/10.1007/s13762-014-0557-4. Accessed on 19 March 2021. Copyright (2015), Springer).

Summarized in Table 3, MF, UF, NF, and RO membranes have widely been used to treat PSWW, from lab tests to case studies on a large scale. Both polymeric and ceramic UF membranes can effectively remove organic matter at a low energy cost since MF membranes possess the largest pore size among these membranes. More impressive, a fly ash-based ceramic membrane recently developed by Goswami et al. [12] removed 100% of COD, TSS, and turbidity. Compared to other membranes, UF membranes have been used more frequently due to their capacity for removing suspended solids, proteins, and pathogens with high water flux and great energy efficiency. Although a study demonstrated that the TOC level of poultry abattoir wastewater treated by NF met the Canadian poultry wastewater reuse criteria, the quality of water produced by NF is not compatible with RO. RO membranes, which can even block monovalent ions, are ideal for producing qualified water for reuse in poultry processing. As an alternative to RO, it was confirmed that thermal-driven VMD equipped with a hydrophobic MF membrane could produce a similar quality of water as RO during PSWW treatment [41].

#### *4.8. Nutrient Recovery from PSWW by Membrane Separation*

In the stage of conventional treatment, valuable nutrients such as proteins could be recovered using coagulation or flocculation from PSWW. Unfortunately, the protein concentrate obtained by the traditional methods cannot be used as animal food because coagulants and flocculants introduce some harmful compounds and change protein properties. Pressure-driven membrane processes are good at protein recovery while keeping protein unchanged because membrane separation is a physical process. For example, Hart et al. [68], in their preliminary studies of MF for the reuse of food-processing water, concluded that MF is a suitable method for reconditioning processing water for reuse, leading to substantial energy savings and reducing disposal costs, and recovering by-products such as protein and fats. Lo et al. [64] investigated protein recovery from poultry processing wastewater using a PS UF membrane with MWCO at 30,000 Da. Their findings revealed that almost all crude proteins with a concentration of 390 ppm were retained, and it reduced 58.86% of COD in poultry processing wastewater. Bialas et al. [61] demonstrated a recovery of 84% of total protein using the integrated process of MF, UF, and VMD.

#### *4.9. Membrane Fouling and Cleaning Methods*

One of the biggest challenges of membrane technology applied to wastewater treatment is membrane fouling, which is caused by the deposition of molecules or particulates on the membrane surface or into membrane pores [69]. Fouling in the membrane is caused by various contaminants in water such as colloidal or particulate matter, dissolved organics, chemical reactants, and microorganisms and microbial products [70]. There is no unified statement regarding the mechanisms of membrane fouling. However, from the analysis of the causes of membrane fouling, four main reasons have been confirmed, including the blocking of membrane pores, adsorption of solute by the membrane, deposition of the activated sludge on the membrane surface, and compaction of the filter cake layer on the membrane surface [71]. Membrane fouling is mainly categorized into two types: reversible and irreversible. Reversible fouling occurs when there is no permanent permeate flux loss, whereas irreversible fouling is caused by permanent permeate flux loss. Other types of membrane fouling include organic fouling, scaling fouling, colloidal fouling, and biofouling. The main drawbacks of membrane fouling are that it could drastically reduce membrane lifetime, productivity, and permeate quality [72,73]. To control membrane fouling, several fouling control strategies have been explored. The commonly used technologies are chemical methods, including coagulation, chemical cleaning, and membrane surface modification; hydrodynamic methods such as a vibrating membrane, high shear, and rotating disk; and physical processes such as ultrasound and physical cleaning techniques [74]. Lo et al. [64] stated that for the treatment of PSWW, membrane fouling was inevitable after processing, and that flushing the UF membrane with a cleaning reagent containing 200 ppm sodium hypochlorite was found to be capable of effectively restoring 90% of membrane performance. Hart et al. [62] reported that the flux rate of MF membranes for PSWW treatment was restored by 15 min in-line cleaning with Micro brand detergent. Moreover, Marchesi et al. [37] recovered 95% of water flux by cleaning the MF and UF membranes sequentially with sodium hypochlorite, citric acid, sodium hydroxide, and ultrapure water. Some pretreatment approaches can be adopted to alleviate membrane fouling before the membrane filtration process. For example, Sardari et al. [40] reported that for the PSWW treatment, the EC as a pretreatment for UF significantly mitigated membrane fouling. Alternatively, Racar et al. [75], who worked on the treatment of rendering plant wastewater, confirmed that sand filtration was an effective pretreatment for later UF, decreasing fouling. Sand filtration primarily eliminated soluble microbial product from secondary effluents, thus improving the UF performance. Although current fouling control approaches are practical, further research shall be carried out to develop cost-effective pretreatments, advance membrane configuration, identify optimal membrane operating conditions, and design an effective hybrid physical/chemical cleaning process [74].

#### **5. Economic Assessment**

Achieving excellent performance with membrane technologies in PSWW treatment and making it cost-competitive with conventional methods are critical for the poultry processing industry. Białas et al. [61] reported that recycling 70% of water by integrating MF, UF, and VMD would yield a savings of EUR 10,850 per month. In addition, the use of clean water would decrease by 7000 m3/month with a savings of EUR 6166 per month, and an 84% protein recovery would generate a product value of EUR 33,000 per month. Jason et al. [30] reported that protein recovery could lead to USD 424 income per day where 100,000 chickens are processed each day. This income could be used as a partial operating cost, thus making the membrane process economic and competitive. Houston et al. [76] analyzed the economic feasibility of incorporating a UF chiller water recycling unit in the pilot poultry processing plant and indicated positive impacts by attaining a profit of more than USD 60,000 a year. Their work reported that recycling the water would reduce the cost by USD 219,465 annually, which came from USD 84,600.75 in water savings, USD 90,695.00 in sewage cost savings, and USD 44,169.84 in energy savings. After detecting the cost of recycling, depreciation, labor, filter cleaning, and miscellaneous, a

net income of USD 68,756 per year with a return rate of 45.6% was achieved. Coskun et al.'s [40] economic analysis for poultry processing wastewater treatment stated that raw water consumption with the conventional method was large as 1220 m3/day, which was decreased to 396 m3/day when using the UF+RO process, as the recycled water was used for various poultry processing purposes. The unit costs of the conventional treatment were reported as USD 738,600/year, the sum raw consumption costs were USD 4,402,500/year, and the treatment costs were USD 336,000/year. Coskun et al. [40] declared that the most economical process was UF+RO with a total cost of USD 295,700/year. As the processing water was recycled and reused, the raw water consumption and treatment costs were decreased to USD 130,800/year and USD 164,900/year, respectively. The experiments by Mannapperuma and Santos [42] verified that UF could treat the 480 L/min chiller overflow to produce 380 L/min reconditioned water at about 80% recovery. It would replace 346 L/min freshwater with the chiller based on guidelines. The use of this system resulted in total savings of USD 165,800/year, which includes savings in freshwater, disposal costs, and energy. This assessment indicates a 2.4-year payback period.

#### **6. Future Perspective and Recommendation**

Membrane technology is promising for PSWW treatment mainly due to its advantages for producing high-quality water for reuse, nutrient recovery, and the operational perspective of compactness and modularity. This review identifies that the research on pressure-driven membrane filtration and membrane distillation for PSWW treatment has mostly been done on a lab or pilot scale, making it unclear for application on a larger scale. Therefore, it is necessary to assess its properties and efficiency with an analysis of energy and operating costs on the industrial level to implement the membrane technologies in poultry slaughterhouse plants. Correspondingly, most research has been conducted only on UF for the PSWW treatment; thus, for a broad perspective, NF and RO should be examined often. In most research articles, the quality of the product water is certified by quantifying parameters such as BOD, COD, TSS, TOC, TN, and TP; therefore, to provide more authenticity to membrane technology, a wide range of physical, chemical, and biological parameters should also be considered.

As an alternative to the traditional membrane approaches, dynamic membrane technology (DMT) is an attractive method for municipal and industrial wastewater treatment and surface water treatment [77–79]. The concept is that when a cake layer forms on a support, such as a mesh or woven filter cloth instead of a conventional membrane, the cake layer acts as a dynamic membrane by properly controlling its thickness. There are some examples of DMT integration with anaerobic BMR for treating high-strength wastewater [80,81]. Recently, Mahat et al. [82] evaluated the 90-day performance of dynamic anaerobic MBR by utilizing low-cost non-woven filter cloth as the support material and producing biomethane as in situ renewable energy while treating high-strength food processing wastewater. Their success indicates that dynamic anaerobic MBR has great application potential for the treatment of high-strength wastewater, including PSWW, at a low operating cost. On the other hand, due to ceramic membranes having a super chemical/thermal stability, low fouling propensity, and long lifespan, the applications of ceramic membrane technology in water and wastewater are rapidly growing, even on a full-size/industrial scale [83,84]. The lower lifecycle cost of ceramic membranes than PVDF membranes in water treatment [85] also implies that the use of ceramic membranes is an excellent option for PSWW treatment in the near future. It is also possible to integrate PSWW treatment with electricity generation. Recently, Roshanravan et al. [86] conducted some tests of polymer-electrolyte membrane microbial fuel cells by feeding meat poultry wastewater, and they found that the cell equipped with SPSU20/MIL7 composite membrane could generate electric power at a power density of 27.50 mW/m2 and Coulombic efficiency of 31.01% with a COD removal rate of 57.65%. In addition, we highly recommend conducting more research to purify PSWW for recycling water than discharging it into the environment by using RO membranes on a pilot or industrial scale. RO is a very

suitable solution to alleviate water scarcity, especially for slaughterhouses located in arid or semiarid regions.

#### **7. Conclusions**

It is concluded that the poultry slaughterhouses produce a large amount of wastewater, which is generally treated by conventional methods. The traditional technologies' inability to recover water and nutrients has given great attention to membrane technology for PSWW treatment. Membrane technology is energy efficient, with a reduction in the number of processing steps, and it provides greater separation efficiency and improved final product quality. It produces water clean enough for reuse and recycling for industrial processing when appropriate two-stage or hybrid membrane separation, e.g., UF+RO, is used. The UF membranes are used as the pretreatment prior to RO by removing suspended solids and macromolecules and recover a good amount of valuable nutrients. Therefore, membrane separation is a promising approach for PSWW treatment. It exhibits excellent performance as a standalone or integrated process by providing high efficiency for pollutant removal and the recovery of valuable products. It removes almost all the pollutants and purifies the water as required to discharge into the environment and reuse for industrial poultry purposes. The summarized economic assessment shows that membrane technology is an economical alternative for the treatment of PSWW. In the near future, robust ceramic UF membranes and DMT using a low-cost mesh or woven filter cloth will have great potential for PSWW pretreatment. The integration of membrane separation with power generation for PSWW treatment is worth further exploration.

**Author Contributions:** Conceptualization, H.D. and R.R.K.; methodology, F.F.; resources, R.R.K.; data curation, F.F.; writing—original draft preparation, F.F.; writing—review and editing, H.D. and R.R.K.; supervision, R.R.K.; project administration, R.R.K.; funding acquisition, R.R.K. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded by the USDA-National Institute of Food and Agriculture, grant number 2020-38821-31091, and National Science Foundation CREST Center for Energy & Environmental Sustainability, NSF grant number 1914692.

**Acknowledgments:** This work was primarily funded through a grant from the USDA-National Institute of Food and Agriculture, Award #2020-38821-31091. Partial support was also received from National Science Foundation CREST Center for Energy & Environmental Sustainability, NSF grant #1914692.

**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, or in the decision to publish the results.

#### **References**


### *Review* **Multi-Integrated Systems for Treatment of Abattoir Wastewater: A Review**

**Larryngeai Gutu 1, Moses Basitere 2,\*, Theo Harding 3, David Ikumi 3, Mahomet Njoya <sup>1</sup> and Chris Gaszynski <sup>3</sup>**


**Abstract:** Biological wastewater treatment processes such as activated sludge and anaerobic digestion remain the most favorable when compared to processes such as chemical precipitation and ion exchange due to their cost-effectiveness, eco-friendliness, ease of operation, and low maintenance. Since Abattoir Wastewater (AWW) is characterized as having high organic content, anaerobic digestion is slow and inadequate for complete removal of all nutrients and organic matter when required to produce a high-quality effluent that satisfies discharge standards. Multi-integrated systems can be designed in which additional stages are added before the anaerobic digester (pre-treatment), as well as after the digester (post-treatment) for nutrient recovery and pathogen removal. This can aid the water treatment plant effluent to meet the discharge regulations imposed by the legislator and allow the possibility for reuse on-site. This review aims to provide information on the principles of anaerobic digestion, aeration pre-treatment technology using enzymes and a hybrid membrane bioreactor, describing their various roles in AWW treatment. Simultaneous nitrification and denitrification are essential to add after anaerobic digestion for nutrient recovery utilizing a single step process. Nutrient recovery has become more favorable than nutrient removal in wastewater treatment because it consumes less energy, making the process cost-effective. In addition, recovered nutrients can be used to make nutrient-based fertilizers, reducing the effects of eutrophication and land degradation. The downflow expanded granular bed reactor is also compared to other high-rate anaerobic reactors, such as the up-flow anaerobic sludge blanket (UASB) and the expanded granular sludge bed reactor (EGSB).

**Keywords:** bio-membrane; multi-integrated system; expanded granular bed reactor; anaerobic digestion; activated sludge; membrane bioreactor

### **1. Introduction**

The continuous influx and increase in urbanization and industrialization have led to an increase in the consumption of goods and services. Relative to other commodities such as winery and car manufacturing, the abattoir industries have also increased and doubled in production in the past decade, increasing water consumption. This increase in water consumption inevitably poses a threat to the environment due to added pollution and increasing water scarcity such that by 2050 global water demand is projected to be 20–30% higher than current levels given both population growth and socio-economic development [1].This is caused by the presence of organic matter such as chemical oxygen demand (COD), which poses a threat to the environment by accelerating the deoxygenation of rivers and contamination of ground water [2]. Abattoir industries consume about

163

**Citation:** Gutu, L.; Basitere, M.; Harding, T.; Ikumi, D.; Njoya, M.; Gaszynski, C. Multi-Integrated Systems for Treatment of Abattoir Wastewater: A Review. *Water* **2021**, *13*, 2462. https://doi.org/10.3390/ w13182462

Academic Editors: Amin Mojiri and Mohammed J.K. Bashir

Received: 21 July 2021 Accepted: 30 August 2021 Published: 7 September 2021

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

**Copyright:** © 2021 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https:// creativecommons.org/licenses/by/ 4.0/).

26 L of potable water per bird to clean the blood off of slaughtered animals, clean off the slaughtering surfaces, cleaning of by-products, steam generation, and for chilling [3]. The slaughtering process and the periodic washing of residue particles in the slaughterhouse result in large quantities of water containing high amounts of biodegradable organic matter [4,5]. The contribution of organic load to these effluents usually comes from different materials such as undigested food, blood, fats, oil, and grease (FOG) and lard, loose meat, paunch, colloidal particles, soluble proteins, manure, grit, and suspended materials [4]. Farzadkia, Vanani [6] stated that the characterization of abattoir wastewater contaminants is influenced by the type of treated water, the kind of animals that have been slaughtered for the particular time frame leading up to water collection, the sampling techniques of the individuals involved, as well as the cleaning and sanitizing procedures of a specific abattoir. These wastewater contaminants can be further characterized into three categories, as shown in Figure 1. Biological oxygen demand (BOD), chemical oxygen demand (COD), and total organic carbon (TOC) are the most widely used parameters for testing effluent quality before discharge according to discharge standards, as shown in Table 1.

**Figure 1.** Characteristics of abattoir wastewater [Abbreviations: BOD biological oxygen demand; COD chemical oxygen demand; TOC total organic carbon; TN total nitrogen, P phosphorus; DO dissolved oxygen] [7].

**Table 1.** Maximum limits permitted by the City of Cape Town: Wastewater and Industrial Effluent By-law 2013 and the characteristics of different abattoir wastewater.


The discharge of untreated water not only poses a severe threat to public health but also causes the death of aquatic species and eutrophication, leading to the depletion of dissolved oxygen (DO) and possible emanation of harmful gases [9,11]. Blood and fat are a major problem in contaminated AWW. Blood has a COD of 375 000 mg/L which is considered very high and on the other hand, fats cause physical problems in treatment plants such as blockages, clogging, scum formation and possible shut downs [2]. Governments have imposed strict regulations on the discharge of water to mitigate the expenses of pollution, for which non-compliance results in heavy penalties. Each municipality in South Africa has regulation standards for water discharge, whether it is into the sewers, land applications or for onsite reuse.

Due to the high costs associated with the efforts to reduce and handle waste, abattoirs are aiming to treat the wastewater onsite with the possibility of reusing and recycling to reduce plant running costs, have a smaller footprint, as well as upgrading to newer cost-effective technologies. The increase in onsite treatment and waste eradication requires advanced refuse-handling equipment and methods to produce organic-rich and less biotoxic waste [12]. The wastewater can be treated using biological and chemical treatment. Recently, chemical treatment has become less popular as the use of chemicals increases the cost of treatment, leaves the difficult task of disposing of the chemical sludge and is environmentally unfriendly, making this option uneconomical and unfavorable [9]. As a result, aerobic and anaerobic treatment systems have become dominant and favorable options [9,13]. AWW contains high concentrations of organic contaminants and is rich in proteins and lipids, making it ideal for biogas production [4], as well as being a good candidate for the highly attractive anaerobic digestion [5]. According to Ozdemir and Yetilmezsoy [14], analysis confirmed that the bio-diesel produced from the waste fats, oils and grease (FOG) obtained from slaughterhouse waste showed excellent fuel properties when compared to biodiesel produced from other common crop-based feedstocks. This is because AWW is protein and lipid rich and has great potential to produce high methane yields at different concentrations of volatile solids.

Anaerobic treatment is advantageous as it has excellent eco-friendly organic matter removal, less sludge production, lower energy consumption, execution of higher organic loading rates (OLR), fewer nutrients and chemical requirements, high COD and BOD removal efficiency, and requires a smaller footprint as well as the considerable production of renewable energy in the form of biogas [15,16]. However, anaerobic digestion poses some limitations, such as having longer start-up and running periods, sensitivity to higher temperature conditions and the inability to effectively remove nutrients such as nitrogen and phosphates, which results in low to moderate effluent quality [17]. Additionally, the process often faces operational challenges due to the difficulties related to the treatment of suspended solids, fats, oils and grease (FOGs) accumulating in the reactors, leading to reduced methanogenic activity, as well as sludge and biomass washout [4,18,19]. These challenges result in process failure, hence the need to incorporate pre-treatment for FOG removal, initiate hydrolysis, and remove solid particles and feathers.

Mondal, Jana [20] stated that aerobic treatment is superior to anaerobic treatment for treating water with a high organic content because it is quicker and more effective for degrading contaminants. However, aerobic digestion also has its flaws, such as high energy requirements for aeration compared to anaerobic, which adds to running costs. Hence a combination of both anaerobic and aerobic processes must be employed to tackle this predicament and effectively remove the nutrients and organic matter [4,9]. The fraction of lipids presents in AWW poses a threat due to their slow hydrolysis rate [21]. Typically, induced and dissolved air flotation is used to remove the oils and grease before aerobic– anaerobic digestion. However, the costs of the air and reagents used, if chemically assisted, tend to make this process uneconomical and expensive. Additionally, the removal efficiency is low and sometimes produces difficult sludges to treat [15]. Other methods tested include alkaline, thermal [22,23] and ultrasonic [24] pre-treatment; however, these all fall short in one way or another. Enzymatic pre-treatment is a good option to satisfy the concerns of

improving methane production, reducing the number of suspended solids before anaerobic digestion and is environmentally friendly [19]. Enzymes hydrolyse the triglycerides to fatty acids and glycerol, which improves the efficiency of biodegradation by microorganisms and eases operation during biological treatment [15]. A study done by Zhang, Zou [19] compared the stability of anaerobic digestion by feeding enzyme pre-treated water vs non-pre-treated water. The reactor containing the enzyme pre-treated feed showed higher stability during operation, even at higher organic loading rates (OLR).

Although it may be a great option, it is not economically feasible to use commercial enzymes practically in engineering practice, as most enzymes have to be significantly monitored as they are sensitive to temperature and pH, and some cannot digest all the organic matter present [19]. An economic and feasibility study done by [15], without considering the ability of methane production to offset costs, revealed that using enzymes to pre-treat wastewater with high fat content has lower installation and operational costs than the traditional technologies. Therefore, it is still a better and cheaper alternative with great potential, despite its complex operation if methane generation is considered as an income generating byproduct. Alternatively, the application of biosurfactants produced by micro-organisms has recently been reported in studies as a cheaper alternative to commercial enzymes [25]. The biosurfactants enhance biodegradation by dissolving FOGs and can be incorporated simultaneously into the biological aeration process, reducing the number of stages for pre-treatment. Other advantages include lower capital and operation costs, reduction in operational problems, as well as an increase in methane production through anaerobic digestion [15,26].

This review highlights the importance of using biological processes in wastewater treatment. The use of a bioremediation agent known as the eco-flush, a product developed by Mavu Biotechnologies (Pty) limited during aerobic treatment, is a novel method that has not been extensively researched. Still, it can pose as an economical and more preferable approach when compared to pure commercial enzymes. Since biological processes are generally slow and not adequate, a multi-integrated system approach can be used, where each stage focuses solely on removing a particular nutrient or pathogen.

#### **2. Analytical Methods for Testing Water Quality**

Measurements need to be performed to check if the treated water complies with municipal discharge regulations. The analytical methods are all outlined in the Environmental protection agency (EPA) handbook, and each analysis is specifically coded. Analysis can be tested on: pH (EPA 9040C), temperature, total dissolved solids (TDS) (EPA 160.1), salinity (EPA 320), turbidity (EPA 180.1), total suspended solids (TSS) (EPA 160.2), volatile suspended solids (VSS) (EPA 1684), COD (EPA 410.4), ammonium (EPA 350.1), nitrates concentration (EPA 353.4), biological oxygen demand (BOD) (EPA 405.1), volatile fatty acids (VFAs) (EPA 8260D) and fats, oils and grease (FOGs) (EPA 1664A). Monitoring the efficiency of a wastewater treatment plant is essential. One of the widely used methods for presenting water quality data is the water quality index (WQI) approach. A summary of different water quality parameters is calculated to a single number, which helps define the general quality status of water and its suitability for various purposes like drinking, irrigation, fishing etc., [27].

#### **3. Aerobic Treatment**

Aerobic treatments involve the treatment of sludge with air in the presence of aerobic or facultative anaerobic microbes before anaerobic digestion [9]. Oxygen is injected into the treatment system, which accelerates the hydrolysis rate of the organics by enhancing the activity of the micro-organisms [28]. Aerobic treatment prior to anaerobic digestion improves the hydrolysis stage, the sludge solubilization, accelerates hydrolytic activities, increases the methane yield by 20–50%, and decreases VS by 21–64% [28]. This suggests that aerobic pre-treatment does not decrease the methanogenic activity of methanogens within the anaerobic digester and can be a great addition to a multi-stage system [28]. Besides

being used in the pre-treatment stages, aerobic processes can be employed after anaerobic digestion to enhance nutrient removal. The required oxygen and treatment time correlate with the strength of the AWW being treated [29]. Due to the expenses incurred during the pumping of artificial oxygen and maintenance, using aerobic treatment for extended periods becomes uneconomical and produces large volumes of biomass. Furthermore, due to the benefits of aerobic treatment, it can be incorporated for shorter processes such as before anaerobic digestion and for nutrient removal after the digester. This will ensure maximum organic matter removal and lower costs, as the processes are relatively short. Despite the higher running costs compared to anaerobic digestion, aerobic treatment has some advantages, such as low odor production and a fast-biological growth rate [30].

#### **4. Aeration Pretreatment Using Enzymes**

Enzymes are used to accelerate the hydrolysis of macromolecules to enhance anaerobic digestion [21]. Pre-treatment is included to ensure complete degradation during anaerobic digestion at shorter hydraulic retention times (HRT). Enzymes breakdown the bonds between the triglycerides and hydrolyze them to basic components of fatty acids and glycerol, thereby giving the aerobic micro-organisms a higher chance to biodegrade the FOGs [15]. An eco-flush is an Ergofito's commercially manufactured bioremediation agent supplied in South Africa by Mavu Biotechnologies. An eco-flush is a mixture of natural ingredients and bacteria with the ability to remain dormant until a rich organic source, which acts as a substrate (such as AWW), is applied to activate them, primarily producing enzymes for hydrolysing FOGs [31,32]. Its natural ingredients are derived from glaucids and essential amino acids, which form powerful decomposing agents that stimulate the natural predisposition of certain bacteria to produce enzymes. These enzymes are capable of breaking down the hydrocarbon chains in FOGs and also compete with the bacteria that are responsible for producing Ammonia (NH3) and Hydrogen Sulphide (H2S), which results in no to less odor during pre-treatment [31]. The eco-flush can be added to raw AWW at the desired ratio as shown in Figure 2, a systematic diagram representing the pretreatment stage before anaerobic digestion. Artificial aeration is required to facilitate the bacteria to produce enzymes to degrade the FOGs by providing oxygen as an electron acceptor. For successful enzymatic pre-treatment, several parameters such as temperature, pH, substrate quantity and enzymes stability have to be assessed and optimized [28].

**Figure 2.** AWW pre-treatment stage using an eco-flush.

Generally, the oxidation of 1 kg of COD requires 1 kWh of aeration energy when the aerobic treatment is selected for wastewater treatment [33]. Oxygen is slightly soluble in water and has to be transferred from the gas phase to the liquid phase, which is called absorption, driven by the concentration gradient between the atmosphere and the bulk liquid [33,34]. The aeration requirement results in the need for a large surface for efficient oxidation of the organic matter, which increases the running costs [33].

Although an aerated pre-treatment stage improves the anaerobic digestion as mentioned previously, the presence of dissolved oxygen in the treated wastewater can also inhibit the methanogenic activity in the anaerobic digestion stage. One critical parameter for a good performance of anaerobic treatment is the lack of oxygen. This is usually determined through the redox potential that should remain <−50 mV for anaerobic digestion and <−300 mV for a good methanogenic activity [35]. For a hermetically closed digester, there is usually no need to attempt to remove the oxygen present, as the BOD in the wastewater consumes the oxygen present rapidly since aerobes and facultative aerobes normally use 100 mg/L of dissolved oxygen to degrade 100 mg/L of BOD [33]. Furthermore, for lab studies and industrial scales, oxygen removal must be implemented through nitrogen purging, which includes three main methods [35], namely: Displacement purging, Pressurizing purging, and Dilution purging. Purging consists of the replacement of one gas by another in an enclosed chamber or space, e.g., removal of oxygen and replacing it with nitrogen gas in anaerobic digestion [34]. Therefore, before the pre-treated water is fed into the anaerobic digester, the Dissolved Oxygen must be monitored.

A study done by [36] a pre-treatment using an Ecoflush bioremediation agent was implemented and resulted in FOG removal of 80% and the TSS and COD removal which reached 38% and 56%, respectively, before feeding the slaughter wastewater into the anaerobic digester. Meyo, Njoya [32] also did a similar study on the pre-treatment of Poultry Slaughter Wastewater (PSW), and the removal percentages varied between 20 and 50% for total suspended solids (TSS), 20 and 70% for chemical oxygen demand (COD), and 50 and 83% for fats, oil, and grease (FOG) before anaerobic treatment using an EGSB reactor. These studies are among the few that reported the use of an Ecoflush reagent. The removal efficiencies do suggest there is potential in bioremediation technology as a pre-treatment stage for high fat content wastewater.

#### **5. Anaerobic Digestion**

Anaerobic digestion is a degradation process that occurs in the absence of oxygen to produce methane and carbon dioxide. It consists of four stages: hydrolysis, acidogenesis, acetogenesis, and methanogenesis, as shown in Figure 3. The hydrolysis stage reduces insoluble organic matter and high molecular weight compounds such as polysaccharides, proteins, and lipids into monosaccharides and amino and fatty acids [37]. During acidogenesis, acidogenic bacteria produce volatile fatty acids, carbon dioxide, hydrogen sulphide, ammonia and other by-products, using the components formed during hydrolysis [29]. Acetogenesis is the third stage in which acetic acid, carbon dioxide and hydrogen are produced from the digestion of higher alcohols and organic acids. Methanogenesis is the last and final step in which methane gas is produced by methanogenic bacteria [37]. The production and accumulation of volatile fatty acids (VFAs) can cause a drop in pH, which can affect methane production. Consequently, the VFA: alkalinity ratio is a critical factor in determining reactor performance and should in no case exceed 0.3 [9,38]. Besides a pH range of 6.8–7.2, the organic matter loading/substrate ratio largely affects biogas production, where either too little or too much can cause a slow digestion process and should in no case be >0.3 [9].

**Figure 3.** Anaerobic digestion stages in an anaerobic reactor.

#### **6. High-Rate Anaerobic Reactors (HRABS)**

High-rate anaerobic digesters have been a subject of increasing interest, due to their high loading capacity and low sludge production. The commonly used high-rate anaerobic digesters include: anaerobic filters, up flow anaerobic sludge blanket (UASB) reactors, anaerobic baffled, fluidized beds, expanded granular sludge beds (EGSB), sequencing batch reactors, anaerobic hybrid/hybrid up-flow anaerobic sludge blanket reactors and the downflow expanded granular bed reactors (DEGBR) which is a hybrid of the EGSB and static granular bed reactor (SGBR) which is shown in Figure 4 [39].

**Figure 4.** High-rate anaerobic reactor (Downflow Expanded Granular Bed Reactor).

Biological processes heavily rely on the growth and bio-preservation of the required microorganisms through controlling essential operational parameters such as temperature, pH, organic loading rate, carbon to nitrogen ratio, inoculation and start-up of the biodigester, mixing, and inhibition factors [34]. The stability of the HRABS is usually reliant on the maintenance of the mentioned operational parameters within a specific prescribed range for growth of microorganisms [33,35]. Table 2 below describes some of the inhibition parameters for anaerobic digestion and how they affect methanogenic activity.

**Table 2.** Inhibition factors in anaerobic digestion.


Good methanogenic activity in HRABS results in the production of biogas and biogas production can be used as a direct measure of biodegradability efficiency. However, there were instances where a good removal of the substrate from the influent, which usually translates to a good COD or BOD5 removal percentage, didn't align with consequent production of biogas [8,41]. This may have been due to biogas entrapment within the anaerobic granular bed as a result of loss in kinetic energy due to friction losses, a weak connected porosity of the anaerobic granular bed or high surface tensions weakening the emergence of biogas bubbles [8,42].

Numerous studies have been carried out to develop high-rate bioreactors; however, most studies show various drawbacks, ranging from large space requirements, a massive volume of sludge generation, intensive use of energy, and the high overall cost of maintenance [11]. For instance, in the expanded granular sludge bed reactor (EGSB) and the up-flow anaerobic sludge blanket (UASB), the liquid up-flow velocity causes low and inadequate removal of nutrients, pathogens and suspended solids, which results in the requirement of post-treatment for compliance with environmental regulations [43]. Unlike the EGSB and UASB, the downflow expanded granular bed reactor (DEGBR) as shown in Figure 4, takes advantage of gravity as a supplementary force through the granular bed, hence using less energy, as there are no gravitational forces or upward frictional forces to compensate for [42]. The DEGBR consists of a recycle stream, which aids in wastewater distribution of the influent to the anaerobic biomass, and also develops a counter-current flow inside the bioreactor for enhanced mixing of its contents [42,43]. Furthermore, the downflow configuration results in the effluent being collected at the bottom and the gas naturally rising to the top, which eliminates the need for a three-phase separator to separate

the gas and biomass compared to the UASB and EGSB [42]. Moreover, the DEGBR also has several advantages like design simplicity, low anaerobic granular sludge (AGS) production, high treatment efficiency, and low operating costs, all of which have turned this bioreactor into a sustainable alternative to mitigate the crisis of water pollution [16].

#### **7. Multi-Integrated Systems**

Anaerobic treatment does not produce discharge compliant effluent on its own. The complete degradation of the organic matter is difficult due to the high organic content levels in AWW, the long hydraulic retention times (HRTs) required to remove all the organics as well as the anaerobic process being slow as compared to aerobic processes. An additional treatment stage(s) is/are recommended to remove the organic matter, nutrients, and pathogens that remain after anaerobic treatment [30]. The integration of multi-stage systems can be used to remove pollutants such as heavy metals, grease and oils, color, BOD, TSS, COD and can be handled within one system with multiple stages [18]. Several studies have been done to incorporate additional stages after anaerobic digestion, as shown in Table 3. Comparing single systems and multi-stage systems shows that the latter provides higher removal efficiencies. The data from Table 3 was used to plot a graph, as shown in Figure 5.

**Table 3.** Comparison between effluent qualities of single systems vs. multi-integrated systems.


**Figure 5.** Comparison between single systems and MIS in removing BOD & TN [Abbreviations: An—anaerobic process; Ae—aerobic process; AO—advanced oxidation process, CC—chemical coagulation].

Treatment technologies such as (i) membrane separation using reverse osmosis, (ii) anaerobic, (iii) aerobic, (iv) anaerobic–aerobic–UV, (v) anaerobic–aerobic advanced oxidation, and (vi) anaerobic–aerobic chemical coagulation were compared graphically to show the effect of introducing multiple stages. Figure 5 shows that all single-stage processes have a BOD removal efficiency below 50%, whilst in multi-integrated systems, the values are above 90%. The TN removal follows the same trend, with reverse osmosis having the highest efficiency despite being a single-stage process. This further supports why membranes are necessary for nutrient recovery after anaerobic digestion as a separation process. Although multi-integrated systems offer many benefits, the type of water, cost, and effluent quality will determine the number of stages and processes to be used.

The use of multi integrated systems provide a significant impact on the effluent quality. Dyosile, Mdladla [36] had a higher overall removal efficiency when an integrated system of using enzymatic pre-treatment–DEGBR–MBR was analyzed as compared to anaerobically digesting the poultry slaughterhouse wastewater (PSW) with the DEGBR with no prior or post treatment. The pre-treatment had FOG removal of 80% and the TSS and COD removal reached 38% and 56%, respectively. The removal results on the DEGBR, at an OLR of 18–45 g COD/L.d, was 87%, 93%, and 90% for COD, TSS, and FOG, respectively. The total removal efficiency across the pre-treatment–DEGBR–MBR units was 99% for COD, TSS, and FOG which is much higher than the single stages. Their effluent quality also met requirements for effluent discharge after post treatment using a membrane bioreactor (MBR).

A similar setup of incorporating pre-treatment–EGSB digester–MBR system was used by [32] to reduce the concentration of organic matter in PSW. The pre-treatment stage resulted in a 50% for TSS removal, 80% for COD removal, and 82% for FOG removal. The EGSB effluent had removal percentages of 90% for TSS, >70% for COD, and >90% for FOG. Further removal was also observed using the MBR with the removal performance being >95% for both TSS and COD and 80% for FOG. Their effluent after the MBR process met the discharge standards. These studies add to the fact that single stages alone do not possess the ability to treat AWW to the required discharge standards. Pre and post treatment is required with any anaerobic processes.

Figure 6 shows a proposed process flow diagram of a multi-integrated system to treat AWW. The raw wastewater is first aerobically pre-treated to remove suspended solids and FOGs and enhance anaerobic digestion. Oxygen is artificially added using an adjustable pump. A stainless-steel sieve is used to filter out any suspended solids remaining from pre-treatment. The pre-treated wastewater is added to a holding tank, which feeds into the DEGBR at the desired organic loading rate. The DEGBR operates anaerobically to biodegrade the nutrients, and biogas is produced as a by-product. The effluent from the DEGBR does not meet the required discharge standards as mentioned previously. The effluent becomes the feed to the membrane bioreactor (MBR) where nitrification and denitrification takes place. The micropollutants that pass through membranes can be disinfected using the ultraviolet system (UV). Bustillo-Lecompte, Mehrvar [13] and Bustillo-Lecompte and Mehrvar [30] demonstrate an evaluation on treating AWW using combined advanced oxidation processes. The evaluation factored in treatment capability and overall costs for treatment technologies, including ABR, AS and UV. It was proven that the combined process of the ABR–AS–UV system was the most cost-effective solution compared to single processes for TOC removal under optimal conditions. However, as this may be a guide, different wastewaters have different characteristics, and analysis must be done to find the best method possible.

Ultraviolet (UV) light is frequently used for pathogen inactivation in wastewater treatment [48–51]. UV light effectively inactivates viruses, bacteria, and cysts by penetrating cell walls and damaging DNA or RNA without chemical addition. Traditional UV lamps are low-cost and accessible in developing economies, but also contain toxic mercury vapor. On the other hand, UV LEDs are more expensive but mercury-free [1,52].

**Figure 6.** Process flow diagram of a proposed multi-stage integrated system to treat AWW.

The study by Beck, Suwan [1] evaluated a cost-effective, user-friendly, and relatively fast treatment process involving a woven-fiber microfiltration (WFMF) membrane to filter domestic wastewater followed by UV disinfection to disinfect the permeate. With an effective pore size of 1–3 μm [53] the membrane was capable of removing *Ascaris lumbricoides* eggs (50 mm) and *Giardia* cysts (10 μm), whereas bacteria (1–2 μm), viruses, and *Cryptosporidum* oocysts (3 μm), which are small enough to pass through the filter pores, were inactivated by exposure to UV light. The bacteria (total coliform and *Escherichia coli*) and viruses (MS2 bacteriophage) passing through the membrane were disinfected by flow-through UV reactors containing either a low-pressure mercury lamp or light-emitting diodes (LEDs) emitting an average peak wavelength of 276 nm. For domestic wastewater from a university campus that they used in their study, the membrane reduced TSS (by 79.8%), turbidity by 76.5%, COD by 38.5%, BOD by 47.8%, and NO3 by 41.4%. UVT at 254 nm improved by 19.4%, and UVT at 280 nm by 12.4% [1]. Following UV disinfection, wastewater quality met the WHO standards for unrestricted irrigation. UV lambs can succumb to fouling and scaling after extensive use and it is reversible through citric acid circulation [54].

#### **8. Hybrid Membrane Bioreactor**

A membrane bioreactor (MBR) is an integrated system with membrane filtration for the biological degradation of waste present in wastewater. Generally, it is composed of a biological unit and a membrane module, which separates water from the aerobically digested water and returns activated sludge to the biological unit [55].

Anaerobic membrane bioreactors (AnMBRs) retain solids selectively through microfiltration membranes which offer an alternative to lagoons and granule based high-rate anaerobic treatments [56]. They produce an excellent effluent quality, have a high tolerance to OLR variations, as well as the ability to produce a solid free effluent for the purposes of reuse [57]. The hybrid membrane bioreactor consists of (i) an anoxic stage and (ii) an aerobic membrane filtration stage. Since the DEGBR operates anaerobically, it has two significant drawbacks, (i) it is ineffective in removing nitrates and phosphorous, and (ii) it reduces the organic nitrogen and sulphur to ammonia and hydrogen sulphide, which are toxic, hence the need for incorporating a membrane bioreactor stage as post-treatment. The advantages of MBR compared with a conventional activated sludge process include high effluent

quality, decreased reactor volume, elevated solid retention time (SRT) and high mixed liquor suspended solids (MLSS), low sludge yield, and easier operation [58,59]. However there are still some drawbacks associated with MBRs such as; membrane fouling, high energy consumption and low removal efficiency of poorly biodegradable micropollutants like diclofenac, atrazine, and carbamazepine [58].

MBR technology has been widely used recently for nutrient recovery. Coagulation or flocculation can be used to recover valuable nutrients in the conventional process. Unfortunately, the protein concentrate obtained by the traditional methods cannot be used as animal food because coagulants and flocculants can introduce some harmful compounds and change protein properties. Pressure-driven membrane processes are good at protein recovery while keeping protein unchanged since membrane separation is a physical process [55].

Recovering nutrients from wastewater reduces the environmental effects of wastewater treatment, and subsequently, the recovered nutrients can be used to produce fertilizers. Phosphorus and nitrogen are essential for organism growth and result in eutrophication in surface water sources, leading to the death of aquatic life [56]. If ammonium and phosphate ions were to be removed from wastewater using processes such as chemical precipitation, it would consume a large amount of electricity and cost about 4% extra compared to nutrient recovery [60]. Besides the extra costs involved, nutrient recovery is better than complete removal because i) nutrient-based fertilizers can be produced for agricultural purposes, (ii) the environmental impact from wastewater discharged is reduced, hence less eutrophication occurs, and iii) N recovery can reduce the consumption of natural resources and save costs on nitrogen fixation [56].

The hybrid membrane bioreactor is required to provide an anoxic–aerobic stage where oxygen is utilized by bacteria to oxidize the ammonia and hydrogen sulphide to less harmful substances. Nitrification occurs due to two specific autotrophic bacteria, the ammonia oxidising organisms (ANOs) and the nitrite oxidising organisms (NNOs), and occurs in two steps. The ANOs convert free and saline ammonia to nitrite. In the second step, the NNOs convert nitrite to nitrate. Ammonia and nitrite are used for catabolism [33]. Nitrification is a prerequisite for denitrification, and without it, biological N removal is not possible. Denitrification becomes possible once nitrification takes place by incorporating anoxic zones in the reactor. The denitrification occurs anoxically via facultative heterotrophic biomass [33]. During nitrification, the N remains in the liquid phase because it is transformed from ammonia to nitrate. In the denitrification step, the N is transferred from the liquid to the gas phase and escapes to the atmosphere.

The proposed study referred to in Figure 7 employs the modified Ludzack–Ettinger system (MLE), which separates the anoxic and aerobic reactors by putting them in series, as shown in Figure 7 below. It also consists of a recycle for the underflow feeding back to the first anoxic reactor as well as a mixed liquor recycle from the aerobic to the primary anoxic reactor. The influent is discharged to the first or primary anoxic reactor, which is maintained in an anoxic state by mixing without aeration and provides conducive conditions for denitrification. The second reactor is aerated and is where nitrification takes place. However, the MLE system has one major drawback: complete nitrate removal cannot be achieved because a part of the total flow from the aerobic reactor is not recycled to the anoxic reactor but instead exits the system with the effluent [33].

**Figure 7.** MLE system for nitrification and denitrification.

Phosphorus can be removed biologically through enhanced biological phosphorus removal, exploiting the ability of polyphosphate-accumulating organisms (PAOs) to take up P in excess of metabolic requirements and accumulate it intracellularly as polyphosphate [61]. This metabolic phenotype is facilitated by a continuing cycle of provision of organic carbon, mainly in the form of volatile fatty acids (VFAs) to the microorganisms, and then exposure of the organisms to first anaerobic and then aerobic conditions. Organic carbon is often the limiting substrate for both denitrification and P removal, and many wastewater treatment plants add extra carbon for denitrification to balance the processes. A combination of denitrification and enhanced biological phosphorus removal in one process could offer substantial savings on carbon for the overall nutrient removal process, which makes this approach highly attractive [62].

The performance of the membrane is mainly characterized by the permeate flux and retention properties [5]. Membrane separation has one particular advantage over other separation processes such as distillation, crystallization and adsorption because it relies on physical separation without phase change and usually no addition of chemicals. Therefore, energy consumption is usually much lower compared to distillation and crystallization [63]. Two main MBR configurations exist: side stream and submerged, as shown in Figure 8. A recirculation pump provides cross-flow velocity in the side stream configuration to reduce blockage by suspended solids on the membrane surface. The side stream MBR is widely used in industrial wastewater treatment but has a higher energy demand. On the other hand, the submerged MBR operates at lower flux and offers higher permeability. They are often used in municipal wastewater treatment. Coarse aeration is provided to the system to reduce fouling as well as provide oxygen to the biomass [64].

**Figure 8.** AnMBR configurations: (**a**) side stream configuration, (**b**) submerged configuration.

#### **9. Applications of Membranes in Wastewater Treatment**

Pressure-driven membrane processes such as microfiltration (MF), ultrafiltration (UF), reverse osmosis (RO), and nanofiltration (NF) are widely studied for wastewater treatment and they rely on hydraulic pressure to achieve separation [65]. Membrane filtration is one of the most emerging technologies to produce high-quality water because it utilizes zero chemical constituents and offers enormous advantages over conventional methods [55]. Mostly used membrane filtration in wastewater treatments are RO, UF and MF. Although reverse osmosis (RO) is a well-established technology for water reuse and desalination [66,67] it is still limited by its high energy consumption and operating costs as the flow is against the pressure gradient.

An alternative is the use of low-pressure RO (LPRO) membranes which have been developed to reduce the RO operation pressure when maintaining high rejections to small soluble organic molecules and ionic species [68–70]. The operation pressure is an important operation parameter of LPRO, which affects the filtration productivity (flux), membrane fouling, and energy consumption. The performance of RO in the treatment of secondary effluent of SWW was reported by [45] to remove organic matter and the removal efficiencies of BOD, COD, TN, and TP were found to be 50.0, 85.8, 90.0, and 97.5%, respectively. It was concluded that LPRO was a suitable technique for the post-treatment of AWW effluent.

A study done by [71] on the performance of the UF membrane treating AWW showed BOD and COD removal efficiencies of around 97.8–97.89% to 94.52–94.74%, whereas TSS and FOG removal were 98% and 99%, respectively [72]. Pressure driven membrane processes have proven to be successful in the separation of valuable organic and inorganic compounds in black liquor as well as being energy-efficient in several studies [73–75]. In recent studies, separation processes are being coupled to improve effluent quality. For example, UF/NF combinations have been reported to be a promising solution in wastewater with high amounts of organic material such as black liquor. In these cases, UF is used as a pre-treatment for NF [73,76].

The ultrafiltration (UF) pre-treatment and the control of the operation pressure were found to be essential for mitigating LPRO membrane fouling. Water quality analyses showed that an integrated process of the UASB + UF + LPRO could achieve an effluent quality characterized by concentrations of 10.4–12.5 mg/L of chemical oxygen demand (COD), 1.8–2.1 mg/L of total nitrogen (TN), 1.3–1.8 mg/L of ammonia nitrogen (NH3-N), and 0.8–1.2 mg/L of total phosphorus (TP) [70].

Coskun, Debik [77] studied the PSW treatment using laboratory-scale membrane processes. Their study reported that UF as pretreatment improved the removal efficiencies for NF and RO processes; NF reduced almost 90% of COD, RO removed 97.4% of COD, and the UF pretreatment resulted in higher final fluxes 8.1 and 5.7 times more for NF and RO, respectively, than for those without UF.

Ionic species can be removed to meet the reuse requirements of brewery wastewater effluent discharge by the inclusion of reverse osmosis into the treatment chain [66,70]. Verhuelsdonk, Glas [78] did an economic analysis on brewery wastewater reuse and reported that UASB wastewater could be treated to drinking water quality with a yield of 63% by using an MBR + UF + RO system.

A comparison study was done by Skoczko, Puzowski [59] to compare the modernized vs conventional treatment methods on a newly modernized wastewater treatment plant. On the basis of the conducted research, it was noted that the operation of the plant after modernization was more cost-intensive. There were additional electricity costs due to ensuring adequate pressure on the membranes. Nevertheless, the obtained results of the removal of contaminants showed BOD removal of over 99.0%, COD removal of 99.0%, TSS removal of 99.5%, and removal degree of biogenic compounds also increased and exceeded 98%. Although the membranes have been well researched and are still being improved, it continues to show high operational costs due to aeration and membrane fouling, which constitutes a major drawback.

#### **10. Membrane Preservation, Fouling and Cleaning Methods**

The accumulation of particulates such as fats, grease, protein, and organic matter can cause build-up on the membrane material resulting in membrane fouling and wetting which is a huge economic influence on the use of membranes as they account for 72% of the capital investment [10]. The types of foulants which may interfere with membrane performance include chemical foulants such as scaling, physical foulants such as deposition of particles, biological foulants such as microbes, and organic foulants which interact with the membrane material [79]. Membrane wetting is the process in which membrane materials lose their hydrophobicity and allow for liquids to penetrate the membrane pores resulting in a direct liquid flow from feed through the wetted pores, substantially deteriorating permeate quality [79]. The fouling and wetting of membrane materials impairs the membrane performance and shortens membrane lifetime, thereby reducing NH3 recovery from AWW.

To reduce fouling and wetting, membranes can be cleaned. Several chemical and physical cleaning methods were developed to remove membrane fouling. The membrane cleaning process is affected by different factors. The type and mode of cleaning for example, physical cleaning, doesn't really retrieve membrane permeability effectively as it only removes loose particles. Temperature is considered as another factor that may take effect

on the membrane cleaning strategy. Increasing temperature is substantial for cleaning the fouling membrane by increasing solubility due to reactivity of functional groups at high temperatures of the organic matters and increasing mass transfer dispersion with mechanical destabilization of biofilm layers on the membrane surface [58]. Increasing the pH also has a directly proportional relationship with membrane cleaning efficiency [80]. For instance, increasing pH from 4.9–11.0 will affect the cleaning percentage from 25%–44% and, at pH 11, it is very easy to break down the gel layer on the membrane surface when compared to the lower pH [58]. Table 4 below shows some of the membrane cleaning methods used to reduce fouling and improve membrane life in membrane technology.


**Table 4.** Membrane cleaning methods used in membrane technology.

Recommendations and future perspectives:


biological treatment, water from the mixed liquor is forced to transfer through the semipermeable membrane to the draw side under the osmotic pressure gradient. The pollutants, activated sludge, and solids are all rejected by the membrane. The superior performance of OMBR over conventional MBR has been demonstrated in previous research [58]. This OMBR can be integrated into the proposed system of this review instead of UF. This will reduce overall running costs incurred through high energy consumption, the cost for chemical cleaning, and membrane life which are limitations in pressure-driven membrane processes.

3. Several studies reported that chemical cleaning could achieve highly efficient membrane cleaning from organic foulants, which may have a strong interaction with the membrane surface [87–90]. Although chemical cleaning is a viable option, it does not provide the eco-friendliness and biological treatment options the world is moving towards and this might cause an environmental problem as the effluent stream may be discharged containing chemicals. Hence more physio-biological pretreatment options and parameter optimization can be a way to ensure limited fouling and maintain a minimal pollution footprint.

#### **11. Conclusions**

Whilst biological processes such as anaerobic and aerobic digestion provide the muchneeded benefits of being environmentally friendly and economical, they still fall short in nutrient removal, digesting FOGs, and removing suspended solids. The choice of reactor also affects the composition of effluent, the costs incurred during operation, and the space required. Anaerobic digestion is very sensitive, involving different bacterial groups (methanogenic, acetogenic, etc.), which all have different optimum conditions. These bacteria are inhibited by process parameters such as temperature, pH, VFA concentrations, etc. Therefore, it is paramount to maintain stable operating conditions in the digester. The DEGBR gives numerous advantages, such as ease of operation, and lowers energy requirements for pumping, as the water is aided by gravity and also provides turbulent mixing through the recycled stream. In contrast, the up-flow reactors such as the EGSB and the UASB experience poor reactor performance caused by a high up-flow velocity, biomass washout and higher energy requirements to oppose gravity and compensate for head losses to friction. The DEGBR has become more favorable for treating high strength wastewater. Adding a pre-treatment stage before anaerobic digestion, where enzymes are used to hydrolyze and break down FOGs increases biogas production, improves reactor performance, and results in ease of operation. Other post anaerobic digestion treatment stages such as nitrification, denitrification, membrane filtration, and ultraviolet radiation can be added to improve the removal efficiency of P, C, and N, as well as help meet the regulation standards.

**Author Contributions:** Writing—original draft preparation, L.G.; Supervision, M.B.; Funding acquisition, M.B.; Writing—review and editing, M.B. & L.G.; Reviewing—T.H., D.I., C.G., and M.N. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded by National Research Foundation, Thuthuka Funding, R017, the Cape Peninsula University of Technology University Research Fund (URF), and the Bioresource Engineering Research Group (CPUT, BioERG) subsidy cost centers RK45 and R971.

**Acknowledgments:** The authors wish to acknowledge the National Research Foundation Thuthuka Funding, R017, for their financial contribution to this work, the South African Breweries (SAB).

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

#### **Abbreviations**



#### **References**


### *Review* **Recent Advances of Nanoremediation Technologies for Soil and Groundwater Remediation: A Review**

**Motasem Y. D. Alazaiza 1,\*, Ahmed Albahnasawi 2, Gomaa A. M. Ali 3, Mohammed J. K. Bashir 4, Nadim K. Copty 5, Salem S. Abu Amr 6, Mohammed F. M. Abushammala <sup>7</sup> and Tahra Al Maskari <sup>1</sup>**


**Abstract:** Nanotechnology has been widely used in many fields including in soil and groundwater remediation. Nanoremediation has emerged as an effective, rapid, and efficient technology for soil and groundwater contaminated with petroleum pollutants and heavy metals. This review provides an overview of the application of nanomaterials for environmental cleanup, such as soil and groundwater remediation. Four types of nanomaterials, namely nanoscale zero-valent iron (nZVI), carbon nanotubes (CNTs), and metallic and magnetic nanoparticles (MNPs), are presented and discussed. In addition, the potential environmental risks of the nanomaterial application in soil remediation are highlighted. Moreover, this review provides insight into the combination of nanoremediation with other remediation technologies. The study demonstrates that nZVI had been widely studied for high-efficiency environmental remediation due to its high reactivity and excellent contaminant immobilization capability. CNTs have received more attention for remediation of organic and inorganic contaminants because of their unique adsorption characteristics. Environmental remediations using metal and MNPs are also favorable due to their facile magnetic separation and unique metal-ion adsorption. The modified nZVI showed less toxicity towards soil bacteria than bare nZVI; thus, modifying or coating nZVI could reduce its ecotoxicity. The combination of nanoremediation with other remediation technology is shown to be a valuable soil remediation technique as the synergetic effects may increase the sustainability of the applied process towards green technology for soil remediation.

**Keywords:** environmental ecotoxicity; nanoremediation; nZVI; CNTs; remediation process; soil remediation

#### **1. Introduction**

Contaminated soil and groundwater, especially in industrialized and urban areas, is a widespread problem that presents extreme risks to the environment and humans [1,2]. Numerous studies have focused on the remediation of soil, groundwater, wastewater, and landfill leachate polluted by various contaminants [3,4]. Soil and groundwater remediation can be broadly classified according to the place of remediation, which can be ex situ or in situ. For ex situ remediation, the polluted soil or groundwater is recovered from the subsurface and treated on the same site or transferred to another site for treatment [5]. In contrast,

**Citation:** Alazaiza, M.Y.D.;

Albahnasawi, A.; Ali, G.A.M.; Bashir, M.J.K.; Copty, N.K.; Amr, S.S.A.; Abushammala, M.F.M.; Al Maskari, T. Recent Advances of Nanoremediation Technologies for Soil and Groundwater Remediation: A Review. *Water* **2021**, *13*, 2186. https://doi.org/10.3390/ w13162186

Academic Editor: Sergi Garcia-Segura and Chin-Pao Huang

Received: 29 May 2021 Accepted: 5 August 2021 Published: 10 August 2021

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

**Copyright:** © 2021 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https:// creativecommons.org/licenses/by/ 4.0/).

in situ remediation is when the contaminated soil or groundwater is remediated directly in the subsurface. The in situ remediation process is often preferred because it is cheaper than the ex situ remediation process [2,6]. For example, according to Chany et al. [7], the remediation cost of removal and replacement of contaminated soil is very expensive (on the order of \$3 million/ha), which is considered a big challenge for developing countries in terms of environmental sustainability practice [7].

Reducing pollution to a desirable and safe level is the main target of soil and groundwater remediation processes. Physical, chemical, and biological technologies have been used to achieve this goal for soil and groundwater remediation. In general, several factors play a significant role in the selection of the optimal soil and groundwater remediation, including soil properties and contaminants and the nature of selected and designed remediation technology [8]. Conventional methods such as pump-and-treat involve pumping groundwater by wells and removal of contaminants from the extracted groundwater by ex situ methods such as carbon adsorption, air stripping, chemical oxidation/precipitation, or biological reactors. However, these methods are associated with high operating costs and contaminated waste production [4]. For groundwater and soil contaminated with organic contaminants in the form of dense nonaqueous phase liquids (DNAPLs), emergent remediation technology such as surfactant enhanced remediation (SER) has been shown to be effective. Nevertheless, these technologies are associated with risks; with the decrease in interfacial force of DNAPLs, uncontrolled vertical movement may occur [9].

In recent years, nanotechnology has been increasingly considered in a broad range of fields. Nanoparticles (NPs) have many essential and promising properties due to their enabled functions in many sectors [10–13]. NPs are produced by combining multidisciplinary fields such as molecular level manufacturing principles and engineering. Generally, nanotechnology is a technique that constructs particles in a size range of 1–100 nanometers, studies the physical phenomena related to those particles and applying these in many sectors [4]. Nanotechnology is being used in many sectors such as the chemical, electrical, biomedical, and biotechnology industries. While many industries produce and use various forms of nanomaterials, there are many attempts to use nanotechnology for environmental protection activities such as water and wastewater treatment, pollution control, and treatment/remediation of contaminated soil and groundwater [14].

Technologies that apply nanoremediation for contaminated sites have been used in recent years (2009 till now). Studies conducted to evaluate nanoremediation technologies are mostly bench-scale with few field-scale applications [2]. The main advantages of using nanoremediation for soil and groundwater remediation, especially for extensive site cleaning, are reduced cost and cleanup time, complete degradation of some contaminants without the need for the disposal of polluted soil and without the need to transfer the soil or pump groundwater [14,15].

Nanoremediation technologies involve the use of reactive NPs for conversion and detoxification of contaminants. The main mechanisms for remediation by NPs are catalysis and chemical reduction [14,16]. In addition, adsorption is another removal mechanism facilitated by the NPs since NPs have high surface-area-to-mass ratios and different distribution of active sites, increasing the adsorption ability [17]. Many engineering NPs have highly feasible characteristics for in situ remediation applications due to their innovative surface coating and minute size. In addition, NPs can diffuse and penetrate the tiny spaces in the subsurface and be suspended in groundwater for a long time; compared to microparticles, NPs can potentially travel long distances and achieve larger spatial distribution [14].

The physical movement of NPs and/or transport in groundwater is dominated by random motion or Brownian movement rather than the wall effect as a result of their nanoscale characteristics [18]. Thus, compared to microscale particles, which are strongly influenced by gravity sedimentation due to their density and large size, the movement of NPs is not controlled by gravity sedimentation, remaining suspended in groundwater during the remediation process. Thus, NPs afford a functional treatment approach allowing direct injection into the subsurface where pollutants are present [14].

Several studies have revealed the potential use of nanoremediation for soil and groundwater [19–22]. However, the environmental effects of those NPs are still unclear and need more investigation to understand the environmental fate and toxicity of NPs, as these issues are crucial for environmental protection practice.

The use of nanomaterials for soil and groundwater remediation has been widely tested at the laboratory level against a large number of contaminants, offering promising results [23,24]. However, nanomaterials may pose positive or negative impacts on living organisms, the environment, society, and the economy, which should be evaluated in a case-specific context. Appropriate documentation of nanoremediation risks, field-scale validation of remediation results, science–policy interface consultations, and suitable market development initiatives are ways to increase the popularity and acceptability of nanoremediation technologies [25]. Savolainen et al. [26] stated that the fundamental elements of risk assessment are likely to remain and will continue to include the elements designed for other chemicals and particles, notably (1) hazard identification, (2) hazard characterization, (3) exposure assessment, and (4) risk characterization, which are the four steps of the risk assessment process [26]. However, the environmental effects of those NPs are still unclear and need more investigation to understand the environmental fate and toxicity of NPs as these issues are crucial for environmental protection practice [15,27].

Various nanomaterials have been investigated for soil and groundwater remediation, such as metal oxides, nanoscale zeolites, enzymes, carbon nanotubes and fibers, titanium dioxide, and noble metals [14]. Generally, zero-valent iron (nZVI) is most widely used for soil and groundwater remediation as nZVI is considered a suitable electron donor and highly reactive [28]. The use of these different nanomaterials will be discussed in detail in this review.

The main objective of this review is to present the recent studies and development regarding the application of nanotechnology for the remediation of soil and groundwater that are contaminated by a wide range of compounds such as hydrocarbons and heavy metals. The focus is primarily on the developments of the last decade, which has witnessed a substantial increase relating in the number of studies examine nanotechnology for the remediation of soil and groundwater. The potential impact of NP use on the environment is also presented and discussed. Finally, the feasibility of combining nanoremediation with other remediation technologies is also discussed.

#### **2. Relationship between Soil and Groundwater: Contaminants and Remediation**

Soil and groundwater are susceptible to pollution by a wide array of pollutants such as petroleum hydrocarbon, chlorinated solvents, and heavy metals. [29]. Selecting a proper remediation technology for a contaminated environment usually depends on contaminant characteristics and contaminated site characteristics such as physical, chemical, and biological properties. All these factors should be considered during the remediation process, design, and implication. Moreover, the time/cost constraints, the regulatory requirements, and the remediation mechanisms should be considered in the selection process.

Nevertheless, adopting risk-based management approaches is increasingly a focus of environmental researchers due to the high demand for sustainable responses to environmental pollutions [14,30]. The polluted environments are usually surface water, sediments, soil, and groundwater, which are mainly contaminated with low and high molecular weight petroleum hydrocarbon compounds, semi-volatile organic compounds, volatile organic compounds, polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls, persistent organic pollutants, organochlorinated pesticides, NAPL, hydrophobic organic compounds, heavy metals, and xenobiotics [31–33]. These pollutants may migrate or spread far from the source and seriously affect flora, fauna, and the ecosystem [30]. Managing the polluted environments requires the selection of the proper remediation technology for the pollutants, destruction, and separation methods according to many ex situ and in situ remediation methods for surface water, sediments, soil, and groundwater comprising physicochemical, biological, chemical, thermal, electromagnetic, electric, and

ultrasonic remediation technologies [34–36]. Remediation in an aqueous environment includes remediation of groundwater and surface water polluted by contaminants, whereas soil remediation includes remediation of sediment subsoil and topsoil polluted by contaminates [37]. Soil and groundwater remediation could be conducted separately or together, depending on the concentration of contaminants and the extent of pollution. The efficiency of remediation technology depends on the design and implication based on the characteristics of polluted soil and the remediation technique. Combining one remediation technology with others sequentially or simultaneously may enhance the overall remediation process through combined or synergistic effects [38].

#### **3. Nanomaterials**

NPs are particles with a dimension between 1 and 100 nm, whereas nanomaterials are materials with a dimension of 100 nm or less in one dimension at least [39–46]. NPs have many reaction/adsorption sites on their surface due to the large proportion of atoms [10,47]. This unique property makes NPs highly reactive with surrounding contaminants cumbered to the composition materials in the macro scale [2]. Nanomaterials can be classified as natural or manufactured. Clay, iron oxide, and organic matter are examples of naturally occurring NPs in soil composition. Manufactured NPs are either developed or synthesized with a unique property to enhance their industrial or technological applications [32,48–50]. Generally, nanomaterials can be produced by two methods; the first method is from outside to inside or from top to bottom, whereby a significant part transfers into the minor parts. The second method is from bottom to top, whereby small parts are buildup into more extensive parts [51].

Several nanomaterials have been developed for contaminant remediation, such as nZVI, nanoscale zeolites, carbon nanotubes, metal oxides, bimetallic nanoparticles (BNPs), enzymes, and titanium dioxide (TiO2) [52–57]. Soil remediation using these three nanoremediation materials (nZVI, TiO2, and CNTs) can be found in [58].

The unique characteristics of the nanomaterials, such as high surface area, quantum size effect, ease of separation and recycling, etc., support their usage as adsorbents. For example, the ferromagnetism of iron-doped nanomaterials supports their recycling and reuse [11,59]. The potential for nanomaterials recyclability makes them economically attractive. The hydrophobicity of fullerene is the key factor for its adsorption properties and ease of recycling [60].

#### **4. Nanoremediation**

Many researchers have focused on the use and development of nanoremediation technologies for soil and groundwater remediation [4]. Nanoremediation is considered an eco-friendly technology. As a result, it is considered a feasible choice for conventional site remediation technology [4,17,61].

Nanoremediation may provide a cost-effective and faster solution for site remediation. Various NPs have been used for nanoremediation, such as metal oxides, nanoscale zeolites, nanometals, carbon nanotubes, and titanium dioxide. In this section, recently published studies relating to soil and groundwater remediation using four nanotechnologies are presented. The four technologies are nanoscale nZVI, carbon nanotube, metal nanoparticle, and magnetic nanoparticle.

#### *4.1. Soil Nanoremediation*

The first implementation of NPs on the field scale for soil and groundwater remediation was reported 20 years ago and revealed that NPs could remain active in injected soil for up to 56 days and move with groundwater for more than 20 m [62]. Zhang et al. [62] reported that more than 99% of trichloroethene (TCE) could be removed from contaminated sites within a few days.

#### 4.1.1. Nanoscale Zero-Valent Iron

Injection of nZVI is well-suited to soil remediation because of its limited disruption to the environment, fast kinetics, cost-effectiveness, and non-toxic nature [63]. According to Karn et al. [63], the first synthesis and use of nZVI were reported in the 1990s. Iron nanoscale was synthesized from Fe2+ and Fe3+ to produce particles ranging from 10 to 100 nm [14]. nZVI was used to remove many contaminants from water, mainly halogenated organic compounds that usually contaminate soil and groundwater. They reported for the first time the effectiveness of using nZVI for detoxification and transformation of many environmental contaminants such as chlorinated organic solvents, polychlorinated biphenyls, and organochlorine pesticides.

Moreover, the authors also showed that modifying nZVI may increase process speed and efficiency. In a recent study, Tian et al. [64] characterized and investigated the application of polyvinylpyrrolidone (PVP)-enhanced nZVI to remediate TCE-contaminated soil [64]. The results showed that the size of prepared PVP-nZVI was around 70 nm when the isoelectric point was around 8.5. In terms of TCE removal efficiency of the investigated system, the removal of TCE was around 84.73%. They concluded that the PVP-nZVI technology was promising to remediate TCE-contaminated soil. Subsequently, Reginatto et al. [65] investigated the performance of nZVI for the removal of hexavalent chromium (Cr(VI)) from clayey residual soil [65]. Five different nZVI materials to contaminant ratios were used, and three different nZVI injections pressure were applied. The result showed that the ratio between nZVI and Cr(VI) significantly affected removal efficiency. The removal efficiency at (1000/11) mg mg−<sup>1</sup> ratio was 98%, whereas at (1000/140) mg mg<sup>−</sup>1, it was 18%. As the pressure increased, the contaminant leaching increased; thus, the pressure of 30 kPa was more efficient.

In another study, Shubair et al. [66] investigated nitrate removal in porous media using nZVI and modified nZVI using Cu in upflow packed sand column containing a multilayer system [66]. The results revealed the optimal condition for high nitrate removal when 10 cm of nZVI/sand was used, where the nitrate removal efficiency was around 97%. On the other hand, for Cu-modified nZVI/sand, the best condition was noted when a double 5 cm layer was used, where complete nitrate removal was observed. The result suggests that using nZVI in a single layer or Cu-modified nZVI in a multilayer could achieve high nitrate removal. In a subsequent study, Xue et al. [67] investigated the performance of rhamnolipid modified nZVI (R-nZVI) to reduce lead (Pb) and cadmium (Cd) in river sediments by immobilization [67]. They demonstrated that after 42 days, R-nZVI transformed unstable Pb and Cd to stable fractions as the residual percentage of Pb and Cd increased to reach 43.10 and 56.40%, respectively. In a recent study, Blundell and Owens [19] investigated the performance of nZVI for 1,1,1-Trichloro-2,2-bis(4-chlorophenyl) ethane (DDT) removal from contaminated soil [19]. They compared the efficacy of nZVI to microscale zero-valent (μZVI). They found that samples treated with nZVI showed around 85% reduction in DDT concentration, whereas about 11% reduction in DDT was observed when μZVI was used. The result clearly shows the superiority of using nZVI over μZVI on DDT removal from contaminated soil. Table 1 summarizes the main results of the recent studies conducted for soil remediation by nanoremediation technologies. The mechanism of metal ions removal using nZVI which involve reduction, oxidation, adsorption and/or precipitation, as shown in Figure 1 [19].

*Water* **2021**, *13*, 



**Table1.**Recentstudiesthatemployednanoremediationmethodsforsoilremediation.

*Water* **2021**, *13*, 2186



**Figure 1.** Metal ions removal using nZVI (**A**) is the mechanism and (**B**) is the particles aggregations. Reprinted with permission from [19] (2021, Elsevier).

#### 4.1.2. Carbon Nanotubes

Since the beginning of their application in the water treatment industry, carbon nanotubes (CNTs) have received significant attention from many researchers due to their superior properties, especially their adsorption properties, since CNTs have a strong ability to be attached to the functional groups of pollutants [79]. CNTs can be classified as single-walled carbon tubes and multi-walled carbon nanotubes.

Many studies have been carried out to investigate the performance of CNTs in terms of soil and groundwater remediation. One such attempt has been conducted by Apul et al. [68], who evaluated the performance of microwave-assisted CNTs for removing total petroleum hydrocarbons (TPH) from the soil [68]. Results showed that after using microwave treatment for 60 s, an 82% removal efficiency of TPH was achieved. Zhang et al. [69] assessed the remediation of Cr(VI) contaminated soil using carboxylate or hydroxylated multi-walled carbon nanotubes (MWCNT-COOH or MWCNT-OH) [69]. In addition, the effect of their catalytic activity on the reduction of Cr(VI) by citric acid was evaluated. The results showed that at pH 5, Cr(VI) adsorption capacity was 8.09 and 7.85 mg g−<sup>1</sup> by MWCNTs-COOH and MWCNT-OH, respectively. In a subsequent study, Cheng et al. [70] studied the efficiency of modified carbon black NPs (MCBN) for petroleum biodegradation and heavy metal immobilization in contaminated soil remediated by plant– microbe combined remediation [70]. The result showed that 65% of petroleum degradation increased in petroleum-Ni co-contaminated soil, whereas in petroleum-Cd co-contaminated soil, the increase in petroleum degradation was 50%. Moreover, the result showed that heavy metals' availability could significantly decrease by using MNCB in Cd- and Nicontaminated soil, leading to enhancing the plant's growth.

In another study, Gong et al. [80] investigated the performance of single-walled carbon nanotubes (SWCNTs) and multi-walled carbon nanotubes (MWCNTs) for the reduction of dichlorobiphenyls- chloroethane (DDT) and hexachlorocyclohexane (HCH) [80]. The authors used different concentrations of SWCNTs and MWCNTs as well as different remediation times. The result showed that CNTs could effectively treat DDTs and HCHs. Optimum conditions for the SWCNTs were 0.29 wt% dosages for 4 months. In addition, the results suggest that the efficiency of CNTs remediation was highly dependent on dosage and sediment–sorbent contact time. Abbasian et al. [72] used MWCNTs to enhance the bioremediation of crude oil-contaminated soil [72]. They mixed different concentrations of

crude oil with MWCNTs for 30 days, and then the microbial diversity of these samples was identified by fluoxetine (FLX) pyrosequencing. The results revealed that using MWCNTs can enhance the degradation of hydrocarbons by increasing the total microbial population. MWCNTs were also used to examine the performance of carbon materials to remediate DDTs and HCHs from sediment. The results showed that sediment remediated with 2 wt% activated carbon (AC) and MWCNTs showed 93% and 59% decrease for DDTs, respectively, and 97% and 75% for HCHs in aqueous equilibrium [73]. The results suggest that the AC was more effective than MWCNTs due to its great specific surface area. These findings revealed the promising of using carbon materials as in situ soil remediation.

#### 4.1.3. Metal and Magnetic Nanoparticles

The use of metal NPs to remediate and immobilize the contaminant from soil and groundwater has attracted much attention recently [74]. Many recent studies examined the performance of using metal NPs for soil remediation. One such attempt has been presented by Peikam and Jalali [74], who studied the remediation of Zn, Ni, and cadmium (Cd) from two contaminated non-calcareous and calcareous soils by SiO2 NPs [74]. The result showed that the reduction of Cd was maximum with 3% SiO2 (56.1%) and 1% Al2O3 (38.3%) for the calcareous and non-calcareous soils, respectively. In terms of Zi, the highest reduction in calcareous and non-calcareous soil was 57.1% for 3% TiO2 and 28.8% for 3.0% Al2O3. In an earlier study, Qiao et al. [75] examined the performance of biocharsupported iron phosphate NPs stabilized by a sodium carboxymethyl cellulose composite for Cd remediation from contaminated soil [75]. A batch experiment with composite (soil-to-solution ratio 1 g: 10 mL) was used. The results indicated that after 25 days, 81.3% of Cd was reduced. The results suggest that the investigated composite could enhance the immobilization of Cd in soil by reducing bioaccessibilty and leachability. In a recent study, Baragano et al. [76] compared the performance of goethite nanospheres (nGoethite) and nZVI for contaminated soil remediation [76]. The result showed that for 2% nZVI dosage, the decrease was 89.5%. The soil phytotoxicity was reduced in general, and the soil parameters were not negatively affected by using nZVI to remediate the contaminated soil. On the other hand, the use of nGoethite showed an excellent result as a small dosage of nGoethite (0.2%) could decrease the As by 82.5%. However, at high dosage, the soil phytotoxicity increased as the electrical conductivity of the soil increased due to using high dosage. The results suggest that both nGoethite and nZVI are promising nanomaterials for As immobilization from contaminated soil.

Environmental remediation using magnetic NPs (MNPs) has received attention recently because of their facile separation using a magnet and special metal-ion adsorption. Several studies investigated the performance of MNPs for soil and groundwater remediation. Fan et al. (2016) examined new MNPs (core–shell Fe3O4@SiO2 NPs coated with iminodiacetic acid chelators) for contaminated soil remediation by non-magnetic heavy metals [77]. The mechanism of removal was chelation and separation by magnetic force. The results indicated that the removal rates of Cd and Pb were 84.9% and 72.2%, respectively. In addition, the results demonstrated that the organic content of the soil negatively affected the removal of the residual heavy metals, whereas the use of MNPs did not change the chemical composition of the soil.

#### *4.2. Groundwater Nanoremediation*

The use of NPs in water treatment started in the 1990s and is therefore considered one of the newer technologies. Gillham and Hannesin (1964) were the first researchers to use the idea of using NPs on decontamination of contaminated water. They used nZVI for remediation of the halogenated group [81]. Nevertheless, Wang and Zhang (1997) conducted the first study that used NPs to remediate organo-chlorines from contaminated groundwater. They observed complete and rapid removal of several aromatic chlorinated using bimetallic NPs [82].

#### 4.2.1. Nanoscale Zero-Valent Iron

Since 1997, various nanomaterials have been used for groundwater remediation. However, the application of nZVI for groundwater remediation has received more focus due to their low cost of production and low toxicity [2,63,83]. Figure 2 represents the remediation process by injection of nZVI for DNAPLs. It is reported that nZVI could be used for chlorinated organic compound remediation. Lin et al. (2018) studied the performance of polyethyleneimine (PEI)-coated nZVI (PEI-nZVI) to remediate three DNAPLs (perchloroethene (PCE), trichloroethylene, and 1,2-dichloroethene (1,2-DCE)) by direct injection in the field test [84]. The result showed that after one day of injection, a remarkable reduction in the DNAPLs was recorded. The result showed complete removal (>99%) of the three DNAPLs after one day from the (PEI-nZVI) injection [84]. In a recent study, Chen et al. (2020) investigated the performance of sulfide-modified nZVI (S-nZVI) supported on biochar (BC) for TCE removal from groundwater. In addition, the effect of many factors such as the mass ratio of S-nZVI to BC, pyrolysis temperature of biochar, and initial pH on the TCE removal were examined [20]. The results indicated that the mass ratio of S-nZVI to BC could satisfy the amount of degradation and adsorption of TCE. The pyrolysis temperatures could influence the TCE degradation and adsorption by changing the physicochemical properties of BC. The initial pH had no significant effect on the total TCE removal, whereas the degradation was enhanced at high pH. Moreover, the result showed that within 2 h reaction time, 100% of the TCE was removed at S-nZVI@BC500, where at S-nZVI@BC300 and S-nZVI@BC700, the removal efficiencies of TCE were 79.97% and 86.4%, respectively [20]. In a similar study, Chen et al. [20] studied the effects of Fe/S molar ratio, dissolved oxygen, initial pH, and particle aging on TCE remediations by S-nZVI [20]. The result indicated that Fe/S molar ratio and initial pH remarkably affected the TCE removal, where a higher TCE removal was obtained at Fe/S molar ratio of 60 at pH above 7. A slight decrease in TCE decolorization was observed when S-nZVI was aged up to 20 days, whereas a remarkable decrease was observed at an aging time of 30 days. Finally, dissolved oxygen had a small effect on TCE removal S-nZVI [85]. In another study, Zhu et al. [86] used green technology to synthesize nZVI/Cu from green tea for Cr(VI) contaminated groundwater remediation [86]. The result showed that the removal efficiency of Cr(VI) was enhanced by decreasing the initial Cr(VI) concentration and initial pH and increasing the temperature, while the presence of humic acids in groundwater decreased the activity of nZVI/Cu. In addition, the result indicated that at optimum conditions (pH = 5, temperature 303 K), the Cr(VI) removal efficiency was 94.7%. Finally, the results suggest that nZVI/Cu is a promising green technology for contaminated groundwater by Cr(VI) [86]. Díaz et al. [21] evaluated the performance of two dosages of commercial nZVI (1 and 5%) for Cu and/or Ni immobilization from water and acidic soil. The results showed that the presence of Cu affected the immobilization of Ni, whereases the presence of Ni did not affect the immobilization of Cu. The efficiency of nZVI was better in water than in soil. The use of 5% dosage completely removed Cu and Ni from water samples, where in soil samples, 5% dosage achieved 54% and 21% embolization for Ni and Cu, respectively [21].

#### 4.2.2. Carbon Nanotubes

In recent years, the use of CNTs for water and groundwater remediation has been increasingly attractive due to their high adsorption affinity. Many recent studies investigated the performance of CNTs for contaminated groundwater remediation. Mpouras et al. [22] investigated Cr(VI) removal from groundwater by MWCNTs. In addition, the effect of operating conditions such as MWCNTs and Cr(VI) concentration, pH, and contact time were examined [22]. The results showed that pH has a significant effect on the adsorption efficiency of MWCNTs; for pH higher than 7, the adsorption process remarkably increased. The adsorption process increased by increasing the MWCNTs concentration. At pH 8, the adsorption percentage increased from 85% to 100% as the concentration of MWCNTs increased from 10 to 50 g L−<sup>1</sup> [22]. In another study, Lico et al. [87] examined the performance of MWCNTS for the removal of unleaded gasoline from water [87]. They used a lab-scale

experiment by adding 20 mL of unleaded gasoline to 30 mL of water and adding different MWCNTs. The results indicated that a small amount of MWCNTs (0.7 g) could remove within 5 min a high percentage of unleaded gasoline. In another study, Liang et al. [88] investigated the efficiency of using alumina-decorated MWCNTs (Al2O3/MWCNTs) for simultaneous remediations of cadmium (Cd(II)) and TCE from groundwater [88]. They conducted a batch experiment for a wide range of conditions. The result showed that the maximum adsorption capacities achieved by Al2O3/MWCNTs were 19.84 mg g−<sup>1</sup> for Cd(II) and 27.21 mg g−<sup>1</sup> for TCE. The results suggest that Al2O3/MWCNTs could be a promising technology for Cd(II) and TCE-contaminated groundwater remediations [88]. Table 2 summarizes the recent works conducted in water and groundwater remediation by nanoremediation technologies.

**Figure 2.** Remediation of an aquifer contaminated by DNAPLs injecting nZVI suspensions directly at the sources of contamination. Reprinted with permission from [83] (2014, Elsevier).

#### 4.2.3. Metal and Magnetic Nanoparticle

The use of metal and magnetic NPs in water and groundwater remediation has received significant attention due to their unique properties. Ou et al. [90] studied the performance of iron-coated aluminum (Fe/Al) BNPs and aluminum-coated iron (Al/Fe) BNPs for the remediations of Cr(VI) from contaminated groundwater [90]. The results indicated that the Cr(VI) removal rate depended on reactive sites and the saturation concentration when (Fe/Al) was used. Moreover, the results showed that the investigated NPs could decrease Cr(VI) to Cr(III). The removal efficiency was 1.47 g g−<sup>1</sup> when (Fe/Al) BNPs were used and 0.07 g g−<sup>1</sup> when (Al/Fe) BNPs were used [90]. In a subsequent study, Wang et al. [91] examined the removal of Cr(VI) from contaminated groundwater using iron sulfide NPs (FeS NPs) [91]. The batch test results indicated that a high removal efficiency (1046.1 mg Cr(VI) per gram FeS NPs) was achieved when FeS NPS was used. This high removal efficiency could be attributed to three mechanisms: reduction, adsorption, and co-precipitation. In addition, they found that the pH significantly affected the Cr(VI) removal using FeS NPs. The results suggest that the synthesized Fe NPs could be a

promising remediation technology for in situ remediations of Cr(VI) contaminated soil and groundwater [91]. In another study, Xie et al. [92] investigated the immobilization of selenite (Se(IV)) in soil and groundwater using Fe-Me binary oxide NPs [92]. The results showed that high Se(IV) uptake was noticed at a pH range of 5–8, the typical groundwater range. According to Langmuir's maximum capacity, the adsorption capacity was 109 mg Se(IV) per g Fe-Me NPs [92]. In another study, Dong et al. [93] examined the effect of the aging time of Fe/Ni BNPs on particle activity [93]. Moreover, they investigated the reactivity of aged Fe/Ni BNPs by examining their performance in removing tetracycline (TC). The results showed that the aged time plays a significant role in TC removal. The removal efficiency of TC was in the range of 82.3–92.5%. As the aged time increased to 5–15 days, the removal efficiency of TC decreased by 20–50% to reach around 50%, due to oxidation and aggregation of the particles. Finally, the removal efficiency of TC by 90 days using aged Fe/Ni BNPs was around 30% [93].

Groundwater remediations using MNPs have received attention recently because of their facile separation using a magnet and unique metal-ion adsorption. Many studies recently investigated the performance of MNPs for groundwater restoration. Gong et al. (2017) investigated the performance of FeS-coated iron (Fe/FeS) magnetic NPs (MNPs) for the remediation of Cr(VI)-contaminated groundwater (Figure 3) [80]. The results showed that the molar ratio of S/F has a significant effect on the Fe/FeS MNPs. Increasing the S/Fe molar ratio to 0.138 decreased Cr(VI) removal by 42.8%. However, a further increase of 0.207 increased the removal efficiency by 63% within 72 h.

Moreover, the results indicated that the adsorption process of Cr(VI) by Fe/FeS at S/F molar ration of 0.207 fitted with a pseudo-second-order kinetic model and the sorption capacity was 69.7 mg g−1, which was simulated by the Langmuir isotherm model [80]. Huang and Keller [94] developed a regenerable magnetic ligand nanoparticle (Mag-ligand) to rapidly remove Cd and Pb from contaminated water [94]. The results showed high performance of mega-ligand as Cd and Pb were removed from contaminated water quickly, and Cd was removed in less than 2 h where Pb in less than 15 min. The performance of mega-legend in terms of Cd and Pb was not affected by pH (3–10). In addition, the whole regeneration process can be achieved by washed Mega-legend easily by 1% HCl. The results suggest that modified mega-legend is a feasible nanoparticle for efficient, rapid, and convenient removal of Cd and Pb from the contaminated aquatic system [94]. In a recent study, Alani et al. [96] successfully synthesized zero-valent Cu NPs and examined their performance for dye removal from water [96]. The results showed that the removal time was between 5 and 13 min and over 90% removal efficiency was achieved, indicating that the synthesized zero-valent Cu nanoparticle has a great catalytic ability [96]. In another study, Li et al. [95] examined the performance of magnetic mesoporous silica NPs (MMSNPs) for the remediation of uranium (U(VI)) from high and low pH [95]. The result showed that MMSNPs were efficient for U(VI) removal in the pH range of (3.5–9.6) for artificial groundwater. They found that MMSNPs adsorption capacity can reach 133 g U(VI) per g MMSNPs; these results indicate that MMSNPs are a promising solution for treating U(VI) contaminated groundwater at extreme pH [95]. In a recent study, Ari et al. [97] successfully synthesized α-Fe2O3 NPs via a biosynthesis method using leaf extracts of Azadirachta indica (neem) and a non-toxic precursor salt (FeCl3·6H2O). In addition, they investigated the potential of using α-Fe2O3 NPs as heterogenous catalyst for tetracycline degradation. The result showed that α-Fe2O3 NPs demonstrated excellent performance as a heterogenous catalyst for degradation of tetracycline aqueous solution by the synergistic effect of the UV/Fenton system [97].





**Figure 3.** Groundwater remediation using Fe/FeS nanoparticles. Reprinted with permission from [80] (2017, Elsevier).

#### **5. Environmental Risk and Ecotoxicology**

Although nanomaterials have been used effectively for soil and groundwater remediation, exposure to nanomaterials may have deleterious effects on humans and environments. Toxicological risk assessments need data on both uptake and exposure of nanomaterials and the immediate effects of NPs when they enter the human system. However, to form a conclusion and recommendations, there are limited data in this domain. The process and factors influencing ecotoxicity are complex. Thus, many factors may determine the impact of synthesized NPs on organisms, such as dissolution potential, particle surface properties, aggregation potential, exposure environment properties, and the physiological, biological, and organism behavior when exposed to NPs [14].

Many studies highlighted the impact of nanomaterials on both humans and environments. For example, iron oxide NPs has a mutagenic impact as it may damage organisms' ability to develop or reproduce [98]. Results indicated exposure to subinhibitory concentrations of amoxicillin-bound iron oxide NPs, in the presence of humic acid, and increased bacterial growth in pseudomonas aeruginosa and Staphylococcus aureus [99]. The joint effects of NPs and other contaminants on terrestrial plants are increasingly investigated but still limited. To provide a sound basis for risk assessment, more research should evaluate the joint effects under realistic conditions [100]. The size and shape of NPs ultimately determine the degree of toxicology. Therefore, not only is monitoring of NPs in soil–plant systems is not essential, but more information is needed on their size allocation and physical properties [101]. Most of the reviewed nano-risk assessment approaches are designed to serve as preliminary risk screening and/or research prioritization tools and are not intended to support regulatory decision making. Although the conventional risk assessment framework is a valuable approach, it may fail to adequately estimate the health and environmental risks from engineering nanomaterials in the near term due to overwhelming methodological limitations and epistemic uncertainties [102]. In this section, an overview of the recent studies about concerns related to the environmental risk of using nanomaterials for soil and groundwater remediation is presented.

Gómez-Sagasti et al. [103] conducted a 3 months experiment to investigate the influence of nZVI concentration (ranging from 1 to 20 mg L−1) on soil microbial properties in two types of soil: sandy-loam and clay-loam soils [103]. The results presented evidence that soil type may affect the degree of potential toxic effects on soil microbial communities by nZVI. The results showed that the accentuated inhibitory impact of nZVI on soil microorganisms in sandy-loam soil was more obvious than clay-loam soil. This can be attributed to the high organic content in clay-loam soil, which acts as a protective agent

when nZVI was added to the soil by rendering nZVI inactive, thus prohibiting interaction with soil microorganism cells. Bacterial biomass and arylsulphatase activity, diversity, and richness were negatively influenced by remediation of sandy-loam soil by nZVI. In terms of concentration, they found no obvious concentration–response effect on the soil by nZVI application. The study suggests that many investigations are required using a wide range of soil types and soil proprieties to have clear insight into soil properties' effect and type on the impact of nZVI on soil bacteria communities [103]. In another study, Dong et al. [104] investigated the effects of carboxymethyl cellulose (CMC) surface coating on the cytotoxicity and colloidal stability of nZVI towards Escherichia coli (*E. coli*) and studied the interrelation between cytotoxicity and particle stability [104]. In addition, they examined the influence of CMC ionic strength (Ca2+), concentration, and aging treatment on particle cytotoxicity. The results indicated that nZVI without coating harms *E. coli* and time- and concentration-dependent.

On the other hand, the cytotoxicity of nZVI decreased when the nZVI particles were coated with CMC. This can be attributed to the cell membrane that kept intact in CMCmodified nZVI, whereas cell membrane disruption could be observed when bare nZVI contact with *E. coli*. The aged nZVI and CMC-nZVI did not affect *E. coli* due to the Fe<sup>0</sup> transformation to less toxic iron oxide. The toxicity of nZVI and CMC-nZVI related to the existence of Ca2+ was concentration-dependent as it can either decrease or increase. The presence of Ca2+ could decrease the toxicity of nZVI by causing aggregation and settling of nZVI.

However, the presence of Ca2+ could also increase the toxicity of nZVI by facilitating the adhesion of NPs onto the bacteria surface, forming a more toxic effect [104]. In another study, Chaithawiwat et al. [105] studied the effect of nZVI on the bacterial growth phases on four bacterial strains [105]. The results showed that lag and stationary phases for all bacterial strains were resistant to nZVI, whereas the bacterial strains in exponential and decline phases showed less resistance than lag and stationary phrases. In addition, the results indicated that increasing the nZVI concentration increased bacterial inactivation. The results suggest that it is necessary to consider the bacterial growth phase and nZVI concentration when studying the influence of nZVI on the bacteria [105]. In a subsequent study, Cheng et al. [106] examine the toxicity of S-nZVI to *E. coli* in an aqueous solution [106]. The result indicated that the toxicity of nZVI could be reduced by sulfidation as S/nZVI showed less toxicity at a lower F/S molar ratio, coming out from the higher iron oxide and sulfate and lower Fe0 content. The results suggest that the typical groundwater contents (i.e., Ca2+, HCO3−, SO4 <sup>2</sup>−, and humic acid) could drop the toxicity of nZVI. In addition, in the presence of groundwater mix components, the S/nZVI toxicity was negligible. The results suggest that the implication of S/nZVI could present a low toxicity risk in the ecosystem [106]. In a recent study, Li et al. [107] conducted a long-term study to examine the effect of zeolite-supported nZVI (Z-nZVI) on farmland soils on bacterial communities during the remediation of metals (Cd, As, Pb) [107]. The result indicated that temporary shifts in pH-sensitive, iron resistance/sensitivity, metal resistance, and denitrifying bacteria after adding Z-nZVI were eliminated due to the soil characteristics that drove the reestablishment of the indigenous bacterial community Z-nZVI and restored the bacterial DNA replication and denitrification activity in the soil. The results suggest that Z-nZVI is a promising nanoremediation technology for long-term metal-contaminated soil remediation without ecotoxicity effects [107].

The toxicity of using CNTs in soil and groundwater remediation has been studied by many researchers [108]. However, there are insufficient data related to the effect of CNTs on both humans and the environment. Song et al. [109] studied the effects of MWCNTs different dosages (0.5, 1.0, 2.0, wt%) on bacterial communities, especially the metabolic function, in phenanthrene contaminated sediment [109]. The results indicated that the metabolic function of microbial communities could be significantly changed by the application of high dosage (0.5–2.0, wt%). This can be attributed to the utilization of carbon sources on Biolog ECO microplate. Remotion of phenanthrene-contaminated sediment with 0.5% MWCNTs presented the best microbial activity and Shannon–Wiener diversity index [109].

#### **6. Combined Nanoremediation with Other Remediation Technology**

The combination of nanoremediation technologies with other mitigation methods has attracted significant research in recent years. Synergetic studies can be characterized as combining multiple nanoremediation methods simultaneously or combined with soil flushing or with biotreatment. In this section, an overview of the recent work in this domain is presented.

Several studies combined many nanoremediation methods at the same time. Vilardi et al. [110] examined the combination of nZVI and CNTs for the remediation of Cr(VI), selenium (Se), and cobalt (Co) from aqueous solutions by conducted a batch experiment [110]. The result indicated that for Cr(VI), the main removal mechanism the reduction, whereas adsorption was the predominant mechanism for other metals. The results showed that the Cr(VI) removal efficiency was 100% when nZVI was used alone without pH change, whereas it decreased to around 90% when CNTs-nZVI nanocomposite was used. On the other hand, using CNTs-nZVI showed high removal efficiency for Se and Co at 90% and 80%, respectively. The results suggest that the CNTs–nZVI nanocomposite showed high adsorption efficiency for remediation of heavy metals-contaminated water [110]. In another study, Zhang et al. [111] studied the performance of carboxymethyl cellulose (CMC)-stabilized nZVI composited with BC (CMC-nZVI/BC) for remediation of Cr(VI)-contaminated soil [111]. The results indicated that, after 21 days, the immobilization efficiency of Cr(VI) was 19.7, 33.3, and 100% when the dosage of CMC-nZVI/BC was 11, 27.5, and 55 g Kg<sup>−</sup>1, respectively. The results suggest that the addition of BC to CMC-nZVI could decrease the Cr(VI) transformation slightly, as a small part of CMC-nZVI could be adsorbed to biochar. The Crtotal removal efficiency was high because the reduction reaction continued to remediation [69]. In a recent study, Qian et al. [112], for the first time, investigated the performance of biochar-nZVI for the remediation of chlorinated hydrocarbon in the field [112]. They used direct-push and water pressure-driven packer techniques. The field study results demonstrated a sharp reduction of chlorinated solvents in the 24 h after the first injection of nZVI, but within the next two weeks, a rebound of the concentrations in groundwater was observed. However, the implementation of biochar-nZVI highly improved the removal of the chlorinated solvent from groundwater for 42 days (Figure 4). The results suggest that biochar-nZVI is a promising combined technology for chlorinated solvent contaminated groundwater remediation [112].

Galdames et al. [29] developed a new approach combining nanoremediation with bioremediation for hydrocarbon and heavy metals remediation from contaminated soil [29]. Specifically, the method uses a combination of nZVI and compost from organic waste. The results indicated that the combination of nZVI and compost could decrease the aliphatic hydrocarbons concentration up to 60% even under uncontrolled conditions. In addition, they observed a remarkable decrease in ecotoxicity in the bio-pile of soil [29]. In another study, Alabresm et al. [113] studied the combination of PVP-coated magnetite NPs with oildegrading bacteria for crude oil remediation at the lab scale [113]. The result indicated that NPs alone removed around 70% of high oil concentration after 1 h. However, the removal efficiency did not increase due to the saturation of NPs. On the other hand, bioremediation by oil-degrading bacteria removed 90% of oil after 48 h. Finally, the combination of NPs and oil-degrading bacteria could completely remove the oil within 48 h. This was attributed to the sorption of oil components to NPs and following degradation by bacteria. Further investigation is needed to understand the oil removal mechanism when combining NPs with oil-degrading bacteria are used for oil remediation [113].

**Figure 4.** Step of injection procedure. Reprinted with permission from [112] (2020, Elsevier).

Recently, Czinnerova et al. [114] conducted a long-term field study that investigated the degradation of chlorinated ethenes (CEs) by using nZVI supported by electrokinetic (EK) treatment (nZVI-EK) [114]. EK may enhance the nZVI impact on soil bacteria and increased the migration and longevity of nZVI. The results indicated a rapid decrease in cis-1,2-dichloroethene (cDCE) at around 70%, followed by setting new geochemical conditions as a degradation product of CE (ethene, ethane, and methane) was observed. These new conditions enhanced the growth of soil and ground bacteria, such as organohalide-respiring bacteria. The results suggest that nZVI-EK remediation technology is a promising method for CE remediation from soil and groundwater and enhanced bacteria availability in soil and groundwater [114]. In another study, Sierra et al. [115] studied a combination of soil washing and nZVI for the removal and recovery of toxic elements (As, Cu, Hg, Pb, Sb) from polluted soil (Figure 5) [115]. The results showed that a high recovery yield was obtained for Pb, Cu, and Sb in the magnetically separated fraction, whereas Hg was concentrated in a non-magnetic fraction. Taking everything into account, the soil washing efficiency was enhanced by adding nZVI, especially for a larger fraction. The results suggest that the investigated methodology opens the door for NPs' use in soil-washing remediation [115].

**Figure 5.** Soil washing assisted nZVI nanoremediation. Reprinted with permission from [115] (2018, Elsevier).

Qu et al. [116] studied the implication of an activated carbon fiber (ACF)-supported nZVI (ACF-nZVI) composite for Cr(VI) remediation from groundwater [116]. In addition, they examined the effect of the operation condition such as nZVI amount on activated carbon fiber, initial Cr(VI) concentration, and pH value on the Cr(VI) removal by conducting a batch experiment. The results indicated that the aggregation of nZVI could be inhabited by ACF, which increases the nZVI reactivity and Cr(VI) removal efficiency. The removal efficiency of Cr(VI) decreased with increasing Cr(VI) initial concentration, whereas, in an acidic environment, complete removal (100%) of Cr(VI) was observed in 1 h reaction time. The proposed removal mechanism consisted of two steps: the first step was the physical adsorption of Cr(VI) on the ACF-nZVI surface area or inner layer, where the second step was a reduction of Cr(VI) to Cr(III) by nZVI [116]. In another study, Huang et al. [117] studied the activation of persulfate (PS) by using a zeolite-supported nZVI composites (PS-Z/nZVI) system and examined its efficiency for TCE degradation. The results indicated that Z/nZVI showed high ability towards PS activation (1.5 mM), and high removal efficiency (98.8%) of TCE was observed at pH 7 within 2 h. Moreover, the PS-Z/nZVI system showed high efficiency in terms of TCE for a wide range of pH (4–7) [117]. Table 3 summarizes the recent works conducted in soil and groundwater remediation by combining nanoremediation technologies with other remediation methods.

**Table 3.** Recent studies employed the combination of nanoremediation technologies with other remediation methods.



**Table 3.** *Cont.*

#### **7. Conclusions**

This review aims to present the latest advances in nanoremediation of contaminated soil and groundwater. The main advantages of using nanomaterials in soil remediation are reduction in cleanup time and overall costs, decreased pollutants to nearly zero in the site, and no need to dispose of polluted soil. The wide use of nZVI nanomaterials in environmental cleanup is due to their high reactivity and high ability to immobilize heavy metals such as Cd, Ni, and Pb. Modifying and/or coating nZVI may decrease the toxicity effects on soil microorganisms. The high adsorption capacity of CNTs is from the large surface area, which makes CNTs a great nanomaterial for organic and inorganic remediation. More studies are needed to investigate the effect of CNTs on the environment.

Soil and groundwater remediation using metal and MNPs is a promising technology due to the unique separation mechanism. Full-scale application of nanoremediation needs further evaluation, particularly in terms of efficiency and potential adverse environmental impacts. Combining nanoremediation with other remediation technology appears to be the future of soil remediation as the combination process increases the sustainable remediation practice towards green environmental protection practice.

#### **8. Recommendation and Future Prospective**

This review provides readers with a general overview of using nanoremediation for environmental cleanups, particularly soil and groundwater remediation. More work is needed to developing smarter nanomaterials for soil remediation. For instance, more advanced development could produce NPs with a high ability to work with several functions, such as interacting with hydrophilic and hydrophobic materials or catalyzing many pollutant reactions on the same particle. In addition, further research is needed to design and synthesize NPs that can remediate a wide range of contaminants; and enhance the injecting systems.

Most existing research on nanoremediation is confined to laboratory studies and modeling. Transferring these studies to in situ conditions is a challenge. Thus, more investigations are required in order to develop standard protocols and doses for the application of nanomaterials at the field level. Moreover, efforts should also focus on the application of nanoremediation in the field to understand nanoparticle's fate and transport behavior in soil, water, and sediments and how they affect the environmental variables.

Nanoremediation has been developed and evaluated over the last 20 years. There is, however, concern about its effects on both humans and the environment. With the rapid advancement of nanoremediation techniques, proper evaluation needs to be done to prevent or mitigate any potential environmental or ecological hazards.

In addition, the need for a more thorough understanding of the contaminants' removal processes and the nanomaterials behavior in nature has led to experimentation where no contaminant is present. Many researchers have examined the impacts of nanoremediation on the soil and groundwater bacteria, yet a clear insight into the interaction between nanoremediation materials such as nZVI and microbial activity is still unclear.

**Author Contributions:** Conceptualization and Funding, M.Y.D.A.; Writing—Original Draft Preparation, A.A. and M.Y.D.A.; Writing—Review and Editing, G.A.M.A.; N.K.C.; S.S.A.A.; M.F.M.A.; and T.A.M.; Validation, M.J.K.B. All authors have read and agreed to the published version of the manuscript.

**Funding:** The research leading to these results has received funding from Ministry of Higher Education, Research, and Innovation (MoHERI) of the Sultanate of Oman under the Block Funding Program, MoHERI Block Funding Agreement No. MoHERI/BFP/ASU/01/2020.

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

**Informed Consent Statement:** Not applicable.

**Data Availability Statement:** Data is contained within the article.

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

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