**Optimization of Green Extraction and Purification of PHA Produced by Mixed Microbial Cultures from Sludge**

#### **Guilherme A. de Souza Reis 1,\*, Michiel H. A. Michels 1, Gabriela L. Fajardo 1,2, Ischa Lamot <sup>1</sup> and Jappe H. de Best <sup>1</sup>**


Received: 31 March 2020; Accepted: 17 April 2020; Published: 21 April 2020

**Abstract:** Sludge from municipal wastewater treatment systems can be used as a source of mixed microbial cultures for the production of polyhydroxyalkanoates (PHA). Stored intracellularly, the PHA is accumulated by some species of bacteria as energy stockpile and can be extracted from the cells by reflux extraction. Dimethyl carbonate was tested as a solvent for the PHA extraction at different extraction times and biomass to solvent ratios, and 1-butanol was tested for purifying the obtained PHA at different purification times and PHA to solvent ratios. Overall, only a very small difference was observed in the different extraction scenarios. An average extraction amount of 30.7 ± 1.6 g of PHA per 100 g of biomass was achieved. After purification with 1-butanol, a visual difference was observed in the PHA between the tested scenarios, although the actual purity of the resulting samples did not present a significant difference. The overall purity increased from 91.2 ± 0.1% to 98.0 ± 0.1%.

**Keywords:** polyhydroxyalkanoates; PHA; PHBV; mixed microbial culture; green extraction; dimethyl carbonate; purification; 1-butanol; wastewater valorization

#### **1. Introduction**

The constant search for environmentally-friendly alternatives for fossil-based materials has been backed up lately by an increase of research in the field of bioplastics such as polyhydroxyalkanoates (PHAs). PHAs are biodegradable polymers synthesized by a variety of bacteria in intracellular granules that serve as energy storage [1,2]. Industrial PHA production has been made feasible by using selected strains of pure microbial cultures to ferment refined substrates [3]. However, a much more sustainable, and perhaps cheaper, option can be found in the use of industrial residue streams as a source of bacterial feed [4–6]. VFAs are organic acids with an aliphatic chain of less than five carbons which can be present in or derived from a large variety of residue streams. VFAs have been shown to be an interesting and very feasible feedstock in the PHA production process by both pure and mixed microbial cultures [7].

Applying a mixed microbial culture (MMC) in the process could furthermore reduce the costs of PHA production, because sterilization of the substrate and reactors is not needed. It has been observed that activated sludge of municipal wastewater treatment plants can be used as a source of MMC with a good PHA-accumulating potential [8,9]. Certain fermentation strategies can be used to explore this accumulating potential and generate a PHA-rich biomass [10–12]. A very useful method is a dynamic fed-batch fermentation with alternating repeated periods of feast and famine, which can also be combined with an aerobic or anaerobic environment [13–15]. Through application of a pulsed VFA-feeding regime, it is possible to reduce the effects of too extreme pH variations caused by the addition of VFAs to the medium, as well as to favor the maintenance of the PHA-accumulating bacterial population over other non-accumulating species during the famine periods. This method also stimulates the PHA-accumulating bacteria to stockpile the biopolymer intracellularly throughout multiple feast and famine cycles, which highlights this feed-on-demand process amongst other approaches even on an industrial scale [8].

When studying the use of different feedstocks for PHA production, it is also important to understand the relationship between the feed composition and the monomeric proportions of the resulting polymer. When VFAs are the bacterial feed, acetic and propionic acid are the main precursors in a mechanism for the production of the monomers 3-hydroxybutyrate (3HB) and 3-hydroxyvalerate (3HV) in PHA [16–19]. Acetic acid can be converted into 3HB via acetyl-CoA, while acetic acid together with the odd numbered propionic acid are used to form 3HV via the conversion to acetyl-CoA and propionyl-CoA [17,20]. The combined production of 3HB and 3HV as monomers leads to the synthesis of copolymer poly(3-hydroxybutyrate-co-3-hydroxyvalerate) (PHBV) (Figure 1).

**Figure 1.** Representation of molecular structure of poly (3-hydroxybutyrate-co-3-hydroxyvalerate) (PHBV) (Drawn on ACD/ChemSketch Freeware, version 2019.2.1, Advanced Chemistry Development, Inc., Toronto, ON, Canada, www.acdlabs.com, 2020).

Extraction is the next step in the production process of PHA, which is done by solubilizing the intracellular PHA followed by separation of the extracted residual biomass and isolation of PHA from the solvent. Reflux and Soxhlet extractions, with and without biomass pretreatments, have been described in literature for many different solvents [21–24]. Non-halogenated solvents have been the focus of many researches for their reduced toxicity, although the chlorinated ones, such as chloroform and dichloromethane, are still considered reference solvents because of their high efficiency [22,25,26]. The non-halogenated solvent dimethyl carbonate performs much better than a range of solvents such as diethyl carbonate, propylene carbonate and ethyl acetate and it achieved satisfactory yields of PHA recovery when compared to dichloromethane [27,28]. Furthermore, dimethyl carbonate (DMC) is considered to be a green solvent for its low toxicity [29] when compared to chloroform and dichloromethane [30]. Therefore, DMC was chosen as the solvent for PHA extraction in the present work. 1-Butanol has also been shown to be efficient as a solvent for PHA extraction [9], with the advantage of leading to a simple separation process through gelation of the polymer when cooling down the mixture [31]. Due to this easiness in separating the solvent from the solid PHA, 1-butanol was chosen as a purification agent in this study.

Total PHA content in biomass and purity evaluation of the obtained polymer can both be done with various simple analytical techniques. Thermal gravimetric analysis (TGA) [32] can be applied for a quick overview of these parameters when focusing on the degradation temperature of the produced PHA. Gas chromatography combined with mass spectrometry (GC-MS) is a more accurate technique that allows the investigation of the monomer concentration, composition and purity of the product [33]. Some studies made a parallel in between these techniques, showing that, even though different PHA content values were obtained by each method, there is a direct correlation between them [34]. For the GC-MS analysis, a pre-treatment step has to be added for the PHA to be able to be analyzed.

In the present work, the aim was to optimize a green extraction and purification of PHA from a mixed microbial culture, using dimethyl carbonate and 1-butanol, respectively. The biomass to solvent ratio or PHA to solvent ratio and the extraction or purification time were the parameters to be optimized.

#### **2. Material and Methods**

#### *2.1. Fermentation Process for PHA Production*

The fermentation process was adapted from the work of Valentino (2015) and the patent of Werker (2013) [8,12]. The accumulation of PHA was done in two identical 2.5 L bioreactors (Infors™ MINI-2.5-BACT, Bottmingen, Switzerland) set at 25 ◦C and 200 rpm and aerated with 1.5 L min−<sup>1</sup> of air. Thickened secondary sludge from the wastewater treatment plant of water board Brabantse Delta in Bath, the Netherlands, was used as a source of PHA-accumulating bacteria. A synthetic feed was made with a solution of acetic and propionic acid in a molar proportion of 3:1, respectively, and a total chemical oxygen demand (COD) of 20 g L−1. This feed composition guarantees that the PHA produced is the co-polymer PHBV in an expected monomer distribution of 50% of each [17]. To start the process, 500 mL of sludge with volatile suspended solids (VSS) concentration of about8gL−<sup>1</sup> was mixed with 500 mL of tap water in the reactor. The mixture was then left for 2 h without any feed under the fermentation conditions in order to stabilize the biomass. A single feed pulse of 10 mL was then given, followed by the first starvation period of 1 hour. By the end of this period, the dissolved oxygen (DO) level was then considered to be the maximum achieved. The feed-on-demand was then automatically controlled by a DO value corresponding to 80% of the maximum as the set point as the condition for the next pulse feed. The feed pulses corresponded to a volume of 10 mL given over 1 min every time the mentioned condition was met. The whole process was set to a total of 22 h. The process was automatically stopped by dosing hydrochloric acid until a pH of at least 2 was reached in order to cease the bacterial metabolism. At that point, stirring and air inlet were shut down.

#### *2.2. Homogenization of Biomass*

The mixture inside the bioreactors was left for a few minutes to settle and the PHA-rich biomass was centrifuged (5810 series, Eppendorf™, Nijmegen, The Netherlands) and washed twice with tap water. The biomass was then freeze-dried (CHRIST Alpha 1-4 LD Plus) overnight. The dried biomass resulted from 20 runs (10 in each bioreactor) was mixed and made homogeneous using a mortar and pestle.

#### *2.3. Extraction Process*

The PHA was extracted from the biomass via reflux by adding 25 mL of dimethyl carbonate (DMC) as a solvent to different amounts of biomass in a round-bottom flask connected to a cooling column. Different biomass to solvent ratios, 0.01 g mL<sup>−</sup>1, 0.025 g mL−1, 0.05 g mL−1, and 0.1 g mL−1, were tested in duplicate by adjusting the amount of biomass. These biomass to solvent ratios will be further referred as 1%, 2.5%, 5%, and 10%, respectively. The round-bottom flask was immersed in a pan filled with glycerin, previously heated up to the boiling point of the solvent (90 ◦C for the DMC). Different times of extraction were tested (0.5 h, 1 h, 1.5 h, and 2 h). After the extraction, the pan was removed to allow the flask to naturally cool down to room temperature. A vacuum filtration with Whatman™ paper filters was then used to separate the biomass from the solution. A rotary evaporator (Hei-VAP Value, HeidolphTM, Schwabach, Germany) was used to recover the solvent and to separate the PHA in the form of a film attached to the wall of the flask. The obtained PHA was left to dry overnight and then weighed. The experimental data obtained under the different extraction conditions was compared using first an F-test for variances and then the adequate t-test for the p values. Chloroform and dichloromethane were used as reference extraction solvents [26–28], for which a biomass to solvent ratio of 1% and an extraction time of 1 h were used.

#### *2.4. Purification Process*

The extracted PHA was purified with 1-butanol (≥99%, Sigma-Aldrich™, Zwijndrecht, The Netherlands) via reflux. 20 mL of the solvent was added to different amounts of the polymer in a round-bottom flask which was connected to a cooling column. The different PHA to solvent ratios tested in duplicate were 0.01 g mL−1, 0.02 g mL−1, and 0.04 g mL−<sup>1</sup> (1, 2 and 4%). The flask was immersed in a pan with glycerin and heated up to the boiling point of 1-butanol (117.7 ◦C). The purification times tested were 0.5 h, 1 h, and 2 h. The round-bottom flask was then taken from the pan, closed with a cap and allowed to cool down at room temperature overnight. The solvent was then separated from the jelly-like polymer by physically pressing it out [31] with a piece of cloth. The PHA was allowed to dry at room temperature overnight. The 1-butanol was separated from the solubilized contaminants and recycled with a rotary evaporator (Hei-VAP Value, HeidolphTM, Schwabach, Germany). The experimental data obtained under the different purification conditions was compared using first an F-test for variances and then the adequate t-test for the p values.

#### *2.5. Analytical Methods*

#### 2.5.1. Thermal Gravimetric Analysis (TGA)

Samples of produced biomass were analyzed in duplicate for PHA content and every sample of extracted and purified PHA was analyzed for purity with TGA (TGA 500Q, TA instruments, Etten-Leur, The Netherlands). Because of the feed composition, all PHA produced was considered to be the co-polymer PHBV. A mass of around 10 mg was placed into platinum pans and a ramp mode of 5 ◦C per minute until 600 ◦C was set under nitrogen atmosphere, with flow rates of 40 mL min-1 on the balance, 60 mL min−<sup>1</sup> on the sample (adapted from Hahn and Chang, 1995) [34]. The mass of PHA in each sample was determined as the mass loss in the temperature range between 250 ◦C and 270 ◦C. A commercial sample of PHBV was used to establish this range.

#### 2.5.2. Gas Chromatography/Mass Spectrometry (GC-MS)

#### Sample Preparation

Following the procedure in Lo et al. (2009) [33], an acidic methanolysis reaction was used to hydrolyze the polymeric chain and to convert the monomers into their methylated ester form. A PHA sample of each experiment was weighed for a mass between 2 and 5 mg and brought into a 5 mL reaction vial. 1 mL of chloroform (99%, Sigma-Aldrich™, Zwijndrecht, The Netherlands), 0.95 mL of methanol (Technical grade, BOOM™, Meppel, The Netherlands) and 0.05 mL of sulfuric acid (95%–98%, Sigma-Aldrich™, Zwijndrecht, The Netherlands) were added into the reaction vial which was closed tightly and shaken. The reaction vial was placed into a heating block (Techne Dri-block™, Staffordshire, England, UK) at 100 ◦C for 6 h and shaken every 1 h. The vial was then left to cool down to room temperature. The solution was transferred from the vial to a centrifuge tube and 1 mL of 1 M NaCl (Extra pure, BOOM™, Meppel, The Netherlands) solution was added. The tube was shaken and centrifuged (5810 series, Eppendorf™, Nijmegen, The Netherlands) for 3 minutes at 4000 rpm. The aqueous layer was discarded and another 1 mL of the NaCl solution was added. The tube was shaken and centrifuged again. The organic phase was then taken from the tube with a needle attached to a 1 mL syringe. The solution volume was measured with the syringe graduation, filtered (0.45 μm × 13 mm PTFE filter) and transferred to a vial.

#### GC-MS Settings

A calibration curve was made using standards of methyl-(R)-3-hydroxybutyrate and methyl-(R)-3-hydroxyvalerate (>98%, Sigma-Aldrich™, Zwijndrecht, The Netherlands) using chloroform as a solvent to achieve different dilutions. The GC-MS used was a 7820A GC System/5977E MSD Agilent Technologies™ with a HP-5MS capillary GC column (30 m × 0.25 mm, 0.25 μm), flow of 2 mL min-1 helium, sample injection of 1 μL, temperature of 250 ◦C at the injector and detector with a heating rate of 10 ◦C min<sup>−</sup>1. The software NIST MS Search Program was used to identify the samples components through their mass spectrum.

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

#### *3.1. PHA Accumulation*

Although there were slight variations throughout the runs, the fermentation process was consistent, with an initial volatile suspended solids concentration of about 4 g L<sup>−</sup>1. A range of 23 to 25 pulse feeds were given during each of the 20 runs. Each fermentation run resulted in around 5.5 g of dry biomass. It was observed that the biomass easily settled to the bottom of the vessel.

The TGA analysis of the dry biomass revealed a content of around 40% of PHA in mass, represented by a degradation peak at the temperature range of 265 to 277 ◦C, which was proved to be the right degradation temperature by comparing the analysis with a commercial sample of PHBV. Hahn and Chang (1995) [34] discovered a correlation between the PHA content measured through TGA and the PHA content measure through GC analysis, where the result from GC analysis are considered to be more accurate. This correlation is expressed as a linear model:

$$y = 1.16x - 15.27\tag{1}$$

where *x* is the PHA content by TGA and *y* the real content. Using this correlation, the total PHA content is around 32%. This result is slightly lower, but still close, to the ones mentioned in the PHARIO report [9], for which the same MMC source was used and the PHA accumulation resulted in values around 39 g of PHA per 100 g of VSS. A difference, however, that might explain the higher production result in that work is the nitrogen and phosphorous supplement in the feed composition and the extra feeds during acclimation process of the biomass.

#### *3.2. Extraction*

The results of PHA extraction in all studied conditions were around 31 g of PHA per 100 g of biomass, with slightly higher extraction values at lower biomass to solvent ratios and longer extraction times (Table 1). This result indicates a very high polymer recovery, which contrasts with what is discussed in Samorì (2015) [27], where only about half of all the PHA inside the MMC biomass could be extracted with dimethyl carbonate without any cell pretreatment. It is important to mention, however, that the MMC used in that work for PHA accumulation has a different source and it was submitted to an extensive process of bacterial selection over time, which might affect the general composition of the biomass and, perhaps, the efficiency of the DMC as a solvent for PHA extraction.

**Table 1.** Extraction results of PHA with dimethyl carbonate. Freeze dried biomass after polyhydroxyalkanoates (PHA) accumulation was extracted in 25 mL dimethyl carbonate at its boiling point at different biomass to solvent ratios for different extraction times. The biomass to solvent ratio is expressed as a percentage of grams of biomass per 100 ml of solvent. Values represent averages ± standard deviation of duplicates in g of PHA per 100 g of biomass.


abcd Average values not sharing a common superscript were significantly different (p < 0.05).

The scenarios with a 10% biomass to solvent ratio presented some practical issues because of the relatively high amount of biomass that settled in the bottom part of the extraction flask in direct contact with the heating source and with low or no contact to the solvent.

The PHA extraction process with DMC was compared with chloroform [23] and dichloromethane [28] as the reference solvents (Table 2).

**Table 2.** Comparative extraction results for the different solvents. The reflux extraction of PHA with dimethyl carbonate at 1% biomass to solvent ratio for 1 h was compared with the reflux extraction with chloroform and with dichloromethane at the same solvent ratio and time. Values represent averages ± standard deviation of duplicates of g of PHA per 100 g of biomass.


Although the amount of extracted PHA seems higher with chloroform or dichloromethane compared with dimethyl carbonate as a solvent, when manually stretched, the polymer films produced with chloroform and dichloromethane were both very brittle and not much elastic. They would immediately break apart when pulled. The brittleness was caused by a higher percentage of impurities in the PHA extracted with chloroform and dichloromethane, as it is further discussed in the results obtained with TGA. The PHA plastic films produced in the process with dimethyl carbonate, on the other hand, had a much higher plastic deformation capability, similar to a common strong plastic bag. However, regardless of the solvent used, the resulting solid PHA had a green/brown color after the solvent recovery in all the produced samples (Figure 2).

**Figure 2.** A sample of PHA film produced directly after the solvent recovery. The dark green color was common to all of the samples produced.

A TGA of PHA samples obtained directly by extraction with dimethyl carbonate reveals a purity of 91.2 ± 0.1% versus a purity of 82.5 ± 3.3% for the extraction with chloroform and of 86.4% ± 3.7% with dichloromethane. These results could explain the higher yields obtained for the extraction process with chloroform and dichloromethane meaning that these reference solvents are solubilizing not only the PHA, but also higher amounts of other compounds present in the biomass, which results in lower overall purity in these samples.

#### *3.3. Purification*

Purification of the extracted PHA with 1-butanol revealed that the whitest product was obtained with a PHA to solvent ratio of 1% and after 0.5 h of purification time (Figure 3).

**Figure 3.** Comparison between the purified PHA at different experimental points.

It was expected that a lower PHA to solvent ratio led to PHA with less impurities. However, a longer purification time led to a darker-colored product, although not much difference was registered in the actual purity of the samples (Table 3).

**Table 3.** Purity results by TGA of the samples submitted to the purification process by 1-butanol. The PHA to solvent ratio is written as in grams of PHA per 100 ml of solvent Values represent averages ± standard deviation of triplicates of percentage of purity of purified PHA.


abcd Average values not sharing a common superscript were significantly different (p < 0.05).

The evaluation of PHA by TGA revealed an increase in purity from 91.2 ± 0.1% to 98.0 ± 0.1% after purification with 1-butanol (Figure 4). The peak degradation temperature of the PHA was identified to be 253.4 ± 7.3 ◦C which is comparable to literature about different monomer compositions of the PHBV copolymer [35–37].

**Figure 4.** Thermal gravimetric analysis (TGA) of extracted PHA before (**A**) and after (**B**) purification with 1-butanol.

#### *3.4. Analysis by Gas Chromatography-Mass Spectrometry (GC-MS)*

Samples of the extracted PHA before and after purification were analyzed with GC-MS for its monomeric composition and identification of impurities (Figure 5). Besides 3-hydroxybutyrate and 3-hydroxyvalerate, the monomer 3-hydroxy-2-methylvalerate was also present in minor quantities in the samples. This monomer has been reported already as a common component of polymers synthesized by enriched cultures of glycogen-accumulating organisms (GAO) [35,38,39].

The GC-MS analysis revealed a monomer composition of 35.6 ± 2.5% 3-hydroxybutyrate and 64.4 ± 2.5% 3-hydroxyvalerate. Given the feed composition, a monomer distribution of the produced PHA of 50% 3-hydroxybutyrate and 50% 3-hydroxyvalerate was expected [17]. However, less energy is needed to metabolize propionic acid than acetic acid [40], which explains the higher percentage of 3-hydroxyvalerate in the PHA.

The non-purified PHA (Figure 5A) contained a bigger variety of impurities than the purified PHA (Figure 5B). Not much can be said about the absolute concentration of impurities before and after the purification process, as no calibration curves were made for the non-PHA related compounds. However, a reduction of 71.4%, 71.6%, and 63.7% in the areas of the impurities III, IV and V, respectively, was calculated, indicating a significant reduction in the overall concentration of such impurities.

Although the quantity of impurities was reduced after the 1-butanol treatments, hexadecanoic acid, octadecanoic acid and dehydroabietic acid were still found in all purified samples. Hexadecanoic and octadecanoic acids have been reported as storage compounds produced by mixed bacterial cultures [41]. The source of the dehydroabietic acid is unknown.

**Figure 5.** Chromatograms of PHA before (**A**) and after (**B**) purification. The identified compounds are: I—3-hydroxybutyric acid; II—3-hydroxyvaleric acid; III—Hexadecanoic acid; IV—Octadecanoic acid; V—Dehydroabietic acid.

#### *3.5. Applicability*

In this work, a mixed microbial culture was used for a PHA accumulation procedure followed by an extraction and purification with different solvents for obtaining a high purity final product. However, the extraction and purification are a two-step process that can be very costly when it comes to an industrial setting. For certain PHA applications where high purity is not a major factor, a single extraction with DMC could be enough for the commercial feasibility of the process. For applications where high purity PHA is required, the purification step can be added, although higher costs should be expected.

#### **4. Conclusions**

The extraction of PHA from mixed microbial cultures can be successfully done with dimethyl carbonate via reflux extraction. Overall, a very small variance of PHA yield was observed for different extraction times or biomass to solvent ratios. A ratio of 0.05 g ml−<sup>1</sup> is considered to be ideal, as the use of higher amounts of biomass lead to practical difficulties. Although higher extraction yields can be obtained with chloroform or dichloromethane as solvents, that also leads to a decrease in the purity of PHA and a less sustainable extraction.

A purification of the extracted PHA with 1-butanol resulted in an increase in purity from 91.2 ± 0.1% to 98.0 ± 0.1%. Although the total purity is approximately the same for different purification times and PHA to solvent ratios, a 0.01 g mL−<sup>1</sup> ratio for 0.5 h of purification time led to a whiter PHA.

Dimethyl carbonate is a great alternative to conventional hazardous solvents in the extraction process of PHA and 1-butanol can be used to increase the purity of PHA if necessary, leading to a more commercially attractive product, although this could lead to higher production costs. In terms of a circular economy, the whole process opens new environmentally friendly possibilities for the bioplastic industry.

**Author Contributions:** Conceptualization, G.A.d.S.R., M.H.A.M., G.L.F. and I.L.; Data curation, G.A.d.S.R. and G.L.F.; Formal analysis, G.A.d.S.R., M.H.A.M. and G.L.F.; Funding acquisition, J.H.d.B.; Investigation, G.A.d.S.R. and G.L.F.; Methodology, G.A.d.S.R., M.H.A.M. and G.L.F.; Project administration, J.H.d.B.; Supervision, M.H.A.M. and I.L.; Writing—original draft, G.A.d.S.R. and G.L.F.; Writing—review and editing, G.A.d.S.R., M.H.A.M., I.L. and J.H.d.B. All authors have read and agreed to the published version of the manuscript.

**Funding:** This work was partially funded by Interreg North-West Europe (European Regional Development Fund) as a part of the WOW! Project (Wider business Opportunities for raw materials from Wastewater) [grant NWE619].

**Acknowledgments:** The authors would like to thank the technicians at Avans University of Applied Sciences for their technical assistance in this research and the Living Lab Biobased Brazil for connecting the authors. We are also grateful for the editors and the reviewers for their constructive comments which helped to improve the quality of this publication.

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

#### **Abbreviations**


#### VSS Volatile suspended solids

#### **References**


© 2020 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (http://creativecommons.org/licenses/by/4.0/).

#### *Article*

## **Quantitative PCR Detection of Enteric Viruses in Wastewater and Environmental Water Sources by the Lisbon Municipality: A Case Study**

#### **Pedro Teixeira 1,2,3,\*, Sílvia Costa 1, Bárbara Brown 1, Susana Silva 4, Raquel Rodrigues <sup>3</sup> and Elisabete Valério <sup>3</sup>**


Received: 10 January 2020; Accepted: 13 February 2020; Published: 15 February 2020

**Abstract:** Current regulations and legislation require critical revision to determine safety for alternative water sources and water reuse as part of the solution to global water crisis. In order to fulfill those demands, Lisbon municipality decided to start water reuse as part of a sustainable hydric resources management, and there was a need to confirm safety and safeguard for public health for its use in this context. For this purpose, a study was designed that included a total of 88 samples collected from drinking, superficial, underground water, and wastewater at three different treatment stages. Quantitative Polimerase Chain Reaction (PCR) detection (qPCR) of enteric viruses Norovirus (NoV) genogroups I (GI) and II (GII) and Hepatitis A (HepA) was performed, and also FIB (*E. coli*, enterococci and fecal coliforms) concentrations were assessed. HepA virus was only detected in one untreated influent sample, whereas NoV GI/ NoV GI were detected in untreated wastewater (100/100%), secondary treated effluent (47/73%), and tertiary treated effluent (33/20%). Our study proposes that NoV GI and GII should be further studied to provide the support that they may be suitable indicators for water quality monitoring targeting wastewater treatment efficiency, regardless of the level of treatment.

**Keywords:** norovirus; water reuse; water quality

#### **1. Introduction**

Monitoring every pathogenic microorganism potentially present in water, namely viruses, bacteria, protozoa, or fungi, is unrealistic given the number of resources necessary for that purpose. As an alternative, microbiological water quality assessment has been focused essentially on detecting fecal indicator bacteria (FIB), namely *Escherichia coli* (*E. coli*) and *Enterococcus* spp. [1–6]. FIB are used as surrogates for enteric pathogens in particular for monitoring fecal contamination in environmental waters, relying on the principle that FIB existence is concomitant with pathogen presence, as demonstrated in several studies [7,8]. However, there are also numerous studies showing that pathogens do not correlate significantly with FIB [9–18]. Since FIB, like *E. coli* and *Enterococcus* spp. are shed in most animal feces [4,19,20], the lack of suitability between FIB levels and human health

outcomes may be related to the FIB source. Amplifying the disconnection between FIB and pathogens is the ability for FIB not only to persist but to grow in environmental habitats like terrestrial soils, aquatic sediments, and aquatic vegetation [16,21–26]. Further studies are thus essential to assess the suitability of FIB as sole indicators not only in environmental waters but in an area that has been gaining more importance worldwide: treated wastewater use, as demand for water reuse, is increasing worldwide whether by necessity in developing countries or by environmental objectives in developed countries.

An efficient and sustainable hydric resources management allows non-potable uses for treated wastewater, such as irrigation, industrial processes, firefighting, recreational, or municipal services. Besides the existence of heavy metals, chemicals, hormones and endocrine disruptors in wastewater, it is still necessary to deal with the expected presence of resistant pathogenic microorganisms, many of which are not tested or included in current standards or legislation for water quality assessment. These microorganisms include viruses, bacteria, protozoa, and helminths, responsible for a significant number of potentially dangerous pathologies. There are some reports showing that wastewater treatment processes do not completely remove enteric viruses [27–29], even from effluents with adequate chlorine concentrations or UV treatment [30]. With an increase in reclaimed water use, the potential health impacts resultant from microbial contamination need to be further explored, as outbreaks of viral infectious diseases have been linked to insufficient treatment [31–33]. Several viruses, including Norovirus (NoV) and Hepatitis A (HepA), are listed in United States Environmental Protection Agency (USEPA's) drinking water Contaminant Candidate List (4-CCL 4), heightening waterborne viruses as a research priority [34]. According to a recent review by Teixeira et al. ([35], in press), numerous studies have focused on detecting enteric viruses in water samples, including treated and untreated wastewater, as well as environmental waters. Data on NoV and HepA viruses' concentrations, however, particularly in tertiary-treated wastewater, are scarce ([35], in press). Therefore, further studies are necessary to evaluate the adequate indicators for water reuse quality evaluation and support their application for regulatory purposes.

Globally, NoV is responsible for nearly 20% of all acute gastroenteritis cases [36], with 677 million cases per year and over 213,000 deaths [37]. In risk groups comprising children, elderly, or immunocompromised individuals, morbidity, and mortality rates of NoV infection are significant [38–42]. NoV exposure, and possible outbreaks, have been reported in schools, hospitals, cruise ships, nursing homes, swimming pools, and restaurants [43–47]. Transmission of the virus primarily occurs via fecal-oral contamination route, direct contact with an infected individual, and contaminated water or food consumption [47–51]. Contact with the virus may occur through drinking [52], recreational [44], or irrigation water [53], leading to waterborne outbreaks [54–56]. The vast contaminated aqueous sources linked to NoV indicate a ubiquitous distribution of the virus [57]. HepA virus transmission occurs mainly by the oral-fecal route (about 95%), and ingestion of contaminated water and food [58–62]. HepA virus has been detected in drinking water, surface water, groundwater, treated, and untreated wastewater [63–67]. With global distribution, HepA is a major causal agent of acute viral hepatitis, with approximately 1.4 million cases reported annually globally [60,68], and a high endemicity of hepatitis A in regions with low sanitation [60,69,70].

Taking benefit from an existing standardized method for detecting NoV and HepA viruses in food and bottled water using real-time quantitative PCR detection (RT-qPCR) (ISO/TS, 15216-1:2017), the aim of our study was to develop a method based on this standardized method and assess the suitability of these enteric viruses as water quality indicators for water use and reuse. Moreover, the method developed aimed to be suitable for several water matrices—drinking, underground, superficial, treated, and untreated wastewater, thus enabling its use in routine water quality testing for detecting and quantifying NoV GI/GII and HepA viruses, and allowing managing entities to evaluate more accurately potential public health risks.

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

#### *2.1. Sampling*

Fieldwork was carried out from February 2018 to December 2018, with a total of 88 samples collected. Sampling included: 1) underground water samples collected at two different sites in Lisbon (n = 19); 2) superficial water intended for drinking water production (n = 10); 3) drinking water from Lisbon´s supply system (n = 11); 4) wastewater collected at a Wastewater Treatment Plant in the Lisbon district, at three different stages – untreated influent (n = 15), effluent with secondary treatment (n = 15) and effluent with tertiary, sand filtration and UV treatment (n = 15). Additionally, blank assays were performed with distilled, sterilized water samples (n = 3). All samples were collected in sterile polyethylene containers and stored at 4 ◦C for less than 2 h until chemical and microbiological determinations were initiated. 0.5 mL of a 10% dechlorination agent Na2S2O3 was added to containers before sampling in order to neutralize possible existing residual chlorine from drinking water samples. Water samples for enteric viruses' determinations were collected in sterile glass containers. Enteric viruses were determined in different volumes according to the water source (25–5000 mL). Microbiological indicators were determined in 100 mL, and chemical indicators were determined according to standard procedures, from a total of 1000 mL [71]. The time lapse between sample collection and laboratory processing did not exceed 24 h.

#### *2.2. Detection and Quantification of Enteric Viruses*

The procedure established in this study was based on the international standard method for the determination of viruses in foods ISO/TS 15216-1:2017 [72]. Initially, 10 μL of Mengo virus strain vMC0 (ceeramTOOLS®, Biomérieux, France) were added to each sample, to be used as an internal process control to assure the RNA extraction efficiency. Samples with extraction efficiency, ≥1% were considered valid, as established in Norm ISO/TS 15216-1:2017 [72]. Viral particles were captured through filtration with a 0.45 μm pore size (47 mm diameter) positively charged nylon membrane (Amersham Hybond N+, GE Healthcare Life Sciences, UK), from 25 mL (untreated wastewater), 1000 mL (blank assay, secondary, tertiary treated wastewater and superficial water) and 5000 mL (drinking and underground water) samples. It should be noted that there are existing alternative protocols for virus collection [73], which stipulate significantly high volumes for filtration and virus collection in surface and groundwater—300 L and 1500 L, respectively. For practical reasons, namely to avoid filter clogging and the possible inexistence of such water volumes in some groundwater sources, and more importantly, aiming to establishing a reasonably achievable method for detecting and quantification enteric viruses in diverse water samples for laboratories that perform the routine water monitoring, in this study different volumes were tested and filtered according to the sample matrix. Filters were then transferred into a sterile tube, and 4 mL of tris/glycine/beef extract (TGBE) buffer were added and shaken at approximately 50 oscillations min−<sup>1</sup> for 20 <sup>±</sup> 5 min. An additional 4 mL of TGBE buffer was added, and the eluates pooled into a single clean tube. The pH was adjusted to 7.0 <sup>±</sup> 0.5 with HCl (≥5 mol/L). Subsequently, samples were concentrated using Amicon® Ultra-15 Centrifugal Filter Devices with a 100 kDa molecular weight cut-off (Merck Millipore, Darmstadt, Germany), through centrifugation 4000 g, at 4 ◦C, for 15 min. For RNA isolation, 500 μL TRIzol® reagent (Thermo Fisher Scientific, Waltham, MA, USA) was added to the samples and mixed for 5 min, 30 ◦C at 350 rpm in a thermomixer (Eppendorf, Germany). Afterward, 200 μL of chloroform were added and samples were mixed in a thermomixer (Eppendorf, Germany) for 3 min, 30 ◦C at 350 rpm. Centrifugation was then performed for 15 min, at 4 ◦C and 12,000 g. RNA was extracted from 140 μL final volume (approx.) of the concentrated sample (aqueous phase) using a QIAamp Viral RNA Mini kit (QIAGEN, Hilden, Germany), according to the manufacturer's instructions. Samples were stored at −80 ◦C until analysis. Quantitative Real-Time PCR (RT-qPCR) was performed (Applied Biosystems AB7500 qPCR) for the specific detection of NoV GI, NoV GII, and HepA viruses. An initial screening for the selected viruses was performed, in addition to confirmation for the absence of inhibitors with

10-fold dilutions. RT-qPCR was performed for NoV GI, NoV GII, HepA, and Mengo virus detection and quantification, with the use of commercial kits (ceeramTOOLS®, Biomérieux, France), according to the manufacturer´s specifications.

#### *2.3. Microbiological Analysis*

Detection of total coliforms, E. coli, enterococci, and fecal coliforms was performed through the use of Colilert and Enterolert with Quanti-Tray (IDDEX Laboratories, Westbrook, ME, USA). Samples were processed according to manufacturer´s instructions. For heterotrophic plate count at 22 ◦C and 37 ◦C, agar inclusion technique was performed using 1 mL aliquots of the water sample (after the necessary dilutions were performed) and adding 15 mL of Yeast Extract agar (VWR Chemicals, Radnor, PA, USA), and incubated at 36 ◦C for 44 h and 22 ◦C for 68 h. After the incubation period at each temperature, all colonies were quantified for each case.

#### *2.4. Physical and Chemical Assessment*

Temperature and pH were both determined (Thermo Scientific™ Orion™3-Star Benchtop pH Meter, Thermo Fisher Scientific, Waltham, MA, USA) according to standard methods, as well as free chlorine and total organic carbon (TOC) [71]. Conductivity was measured according to Standard Guideline NP EN 27888:1996 (MeterLab CDM 210) and Mohr´s Method was performed to assess chlorides.

#### *2.5. Statistical Analysis*

A descriptive analysis was made for numeric and categorical variables. For comparing different measures among the three wastewater treatment phases, a Kruskal–Wallis non-parametric or Fisher's Exact Test was used. When differences were found, multiple comparisons with Bonferroni correction was also performed. Correlation in each wastewater treatment phase was measured with the Pearson correlation coefficient. Statistical tests were performed bilaterally at a significance level of 5%. The statistical analysis of the data was performed using statistical software R, version 3.4.3. [74].

#### **3. Results**

#### *3.1. Environmental and Drinking Waters*

A statistical summary for virus, microbiological, physical and chemical results (median) is presented in Table 1, concerning blank assays (BA), superficial water intended for drinking water production (DW1), drinking water from Lisbon's public supply system (DW2) and underground water samples collected at two different sites in Lisbon (GW1 and GW2). None of the enteric viruses targeted in this study—NoV GI, NoV GII, and HepA—were detected in samples from superficial water intended for drinking water production (n = 10), drinking water from Lisbon's supply system (n = 11) and underground water samples (n = 19).

Microbiological, physical, and chemical results for drinking water samples (n = 11) were all in compliance with national legislation—Law Decree No. 306/2007, 27th August, and Law Decree No. 236/1998, 1st August, with no microbial contamination detected. Environmental - superficial and underground bodies of water - presented similar and reduced levels of fecal contamination, with no significant differences between them (Table 1). Apart from conductivity and TOC, which presented lower and higher values, respectively, in the superficial water samples. The physical and chemical results were similar between the superficial and underground water samples.

**Table 1.** Microbiological, physical, and chemical results (median) for blank assays (BA), superficial water intended for drinking water production (DW1), drinking water from Lisbon's supply system (DW2), and underground water samples collected at two different sites in Lisbon (GW1 and GW2). MNP—Most Probable Number, CFU—Colony Forming Units, GC—Genomic Copies. ND—Not detected. n = total analyzed samples. LD (limit of detection) = 1 MPN/100 mL.


#### *3.2. Wastewater*

Untreated wastewater influent (WW1), effluent with secondary treatment (WW2) and effluent with tertiary treatment (WW3) statistical results summary for virus, microbiological, physical and chemical results (median) is presented in Table 2. HepA virus was only detected in one untreated influent sample, with a 3.99 <sup>×</sup> 106 gc/100 mL concentration. NoV GI was detected in all wastewater treatment stages - untreated wastewater influent (n = 15), effluent with secondary treatment (n = 7) and effluent with tertiary treatment (n = 5). Similar results were obtained for NoV GII with the following positive samples: untreated wastewater influent (n = 15), effluent with secondary treatment (n = 11) and effluent with tertiary treatment (n = 3). It can be noted that NoV GI median concentration reductions are significant between stages; a 6.67 <sup>×</sup> 107 gc/100 mL initial concentration is reduced to 8.34 <sup>×</sup> 106 gc/100 mL (WW2) and to 8.81 <sup>×</sup> 105 gc/100 mL final concentration. NoV GII initial 4.86 <sup>×</sup> <sup>10</sup><sup>7</sup> gc/100 mL median concentration is also significantly reduced with secondary treatment to 2.29 <sup>×</sup> 10<sup>6</sup> gc/100 mL (WW2). With UV and sand filtration, the effluent's concentration is also diminished (9.69 <sup>×</sup> 105 gc/100 mL), although with no observed statistical significance.

Significant differences were found between all treatment wastewater phases for FIB and HPC at 37 ◦C and 22 ◦C concentrations. Notably, we can observe a significant FIB decay not only with secondary treatment but also with tertiary sand filtration and UV treatment—after which FIB have been eliminated upon tertiary treatment (results < limit of detection). According to current Portuguese national legislations—Law Decree No. 119/2019, 21st August—these effluents (with tertiary treatment) can be adequate for non-potable uses including irrigation, industrial uses, firefighting or street cleaning. Excluding chlorides, with no significant variations between wastewater treatments, the majority of the physical and chemical results, display a statistically significant reduction with secondary treatment activated sludge.

**Table 2.** Microbiological, physical, and chemical results (median) for untreated wastewater influent (WW1), effluent with secondary treatment (WW2) and effluent with tertiary treatment (WW3). \* Kruskall–Wallis non parametric test. \*\* Multiple comparisons tests with Bonferroni correction. a—Statistically significant difference between untreated influent and secondary treated effluent. b—Statistically significant difference between untreated influent and tertiary treated effluent. c—Statistically significant difference between secondary and tertiary treated effluent. MNP—Most Probable Number, CFU—Colony Forming Units, GC—Genomic Copies. ND—Not Detected. n = total analyzed samples. LD (limit of detection) = 1 MPN/100 mL.


Correlations between the microbiological parameters analyzed in each wastewater treatment phase (WW1, WW2 and WW3) were measured (Figure 1A–C). A positive correlation was observed between NoV GI and NoV GII in untreated influent (ρ = 0.54), and between NoV GI and fecal coliforms (ρ = 0.61). One of the highest correlations on this phase was found between Total Coliforms and Enterococcus (ρ = 0.68). Several positive correlations were detected for secondary treated effluent samples in microbiological parameters but none significant concerning NoV GI/GII. In tertiary treated effluents, only one moderate correlation was observed for HPC 22 ◦C/37 ◦C (ρ = 0.60).

**Figure 1.** *Cont.*

**Figure 1.** Graphical display of the correlation matrix of the microbiological parameters tested in each wastewater treatment phase (**A**) WW1, (**B**) WW2, and (**C**) WW3). The "?"—means "Not enough finite observations".

#### **4. Discussion and Conclusions**

In the case of environmental and drinking waters, enteric viruses NoV GI, NoV GII, and HepA virus were not detected in any of the samples in drinking, superficial, or underground water, suggesting an absence of fecal contamination from a human source in these types of water. However, the presence of total coliforms, *E. coli*, enterococci, and fecal coliforms were detected in samples from underground and superficial samples, and though in reduced concentrations, it suggests a natural and expected fecal contamination from animal origin. Studies showed that NoV, could be detected in particularly high concentrations in feces - up to 10<sup>11</sup> GC/g [75]. Consequently, the occurrence of NoV in surface waters can be linked to contaminated water sources (i.e., fecal contamination) and with environmental conditions determining the survival of the virus [14,76]. Since these viruses are highly infectious [49] and demonstrate a high resistance to environmental degradation in water [77–79], its presence in surface and groundwater intended for human consumption raises as a potential public health risk, by which frequent monitoring of enteric viruses such as NoV besides FIB, could be desirable and prudent.

On what concerns the treated wastewaters, our results demonstrate that the absence of FIB does not imply the absence of pathogenic microorganisms, namely NoV GI and GII in tertiary treated wastewater (obtained after sand filtration and UV). FIB monitoring and their correlations with enteric viruses have recently been questioned [17,80,81], hence the necessity for more adequate indicators in water quality monitoring, especially for wastewater use. Growing water resources management concerns arise, not only in areas affected by droughts but also importantly in urban areas, like the one targeted in this study—Lisbon, where water reuse represents one of the municipality's commitments as part of a sustainable hydric resources management, with self-evident safeguard for public health. The results here obtained for Nov GI and GII concentrations are in accordance with previous studies targeting untreated, secondary, and tertiary treated wastewater [82–85].

The results of our study support the hypothesis that NoV GI and GII might be suitable indicators for water quality monitoring regarding wastewater treatment efficiency, regardless of the level of the treatment—secondary or tertiary. A correlation between NoV (GI) and FIB (fecal coliforms) is observed in untreated effluent, but in subsequent treatment phases, there was no correlation between the targeted enteric viruses and FIB. While positive correlations observed between FIB and NoV in untreated wastewater could be expected and were indeed observed, the absence of correlations between FIB and secondary and tertiary treatment stages points out to a lack of a link between FIB and the targeted enteric viruses concentrations. Moreover, in several samples with an apparent total FIB elimination with UV and sand filtration treatment, NoV was still detectable. Our results point out to the need of further studies and expansion to different WWTPs, in diverse locations and including a higher number of samples, not only to further reinforce the results of this study and validate our hypothesis but also to expand detection methods to other viral pathogens and evaluate the best and most comprehensive indicators to use for water reuse situations. While cell culture methods have been the gold standard for the detection of infectious viruses, qPCR methods became essential for enteric virus detection in water samples due to shorter detection times, high sensitivity and specificity, and the ability to detect viruses that are not easily or at all culturable [86]. A significant limitation of the method applied in our study is the inability to differentiate infective viruses from non-infective viruses. Nonetheless, viral genomes are present in water samples, particularly in reclaimed wastewater. Considering an estimated 50% infectious dose (ID50) between 18 and 1,015 genomic equivalents [49], public health risks may arise even from non-potable uses like irrigation or street cleaning, as incorrect usage of treated wastewater has caused outbreaks of viral infectious diseases worldwide [31–33]. To this fact, it must enhance while qPCR delivers quantitative data with high accuracy; results obtained by this method should be interpreted prudently because of potential losses during the sample concentration/extraction/purification procedure, which may result in an underestimation in the detection/quantification process [86]. The whole process control used in our study – Mengo virus – allowed evaluating the efficiency for virus recovery in the different stages – from sample concentration and RNA extraction, to RT-qPCR. However, even within the established ISO values (recovery ≥1%), the obtained recovery rates for this control were usually low (<10%), which can underestimate NoV GI and GII concentrations, particularly in reclaimed wastewater. Importantly, this study shows that a procedure based on the international standard method for the determination of viruses in foods ISO/TS 15216-1 [72] could be applied for routine water quality monitoring of NoV and HepA, in different water matrices.

The importance of FIB is undoubtedly as they are responsible for a significant improvement in water quality assessment and safety management for many decades [1]. Nevertheless, current regulations and legislation require critical revision to determine safety in particular for water reuse, as we can observe samples determined compliant for bacterial indicators that are positive for NoV GI and GII presence. Representing a key to the global water crisis, water reuse should also be determined safe considering the different treatments applied, as enteric viruses' infections also represent a worldwide economic concern.

**Author Contributions:** Conceptualization, E.V. and P.T.; methodology, E.V. and P.T.; validation, E.V. and P.T; formal analysis, P.T., E.V. and S.S.; investigation, P.T., S.C., B.B. and E.V.; resources, P.T., R.R. and E.V.; data curation, P.T. and E.V.; writing—original draft preparation, P.T.; writing—review and editing, E.V., P.T., S.C., S.S., B.B and R.R.; visualization, P.T., S.S. and E.V.; supervision, E.V.; project administration, E.V. and P.T.; funding acquisition, P.T. and E.V. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded by the Lisbon Municipality (CML) in collaboration with Instituto Nacional de Saúde Doutor Ricardo Jorge (INSA).

**Acknowledgments:** The authors are grateful to Ângelo Mesquita and José Canêdo for their contributions to the implementation of this project. We would also like to thank Helena Rebelo and João Brandão for all the help provided. Any opinions, findings and conclusions expressed are those of the authors and do not necessarily reflect the views of the Lisbon Municipality.

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

#### **References**


© 2020 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (http://creativecommons.org/licenses/by/4.0/).

### *Article* **Wastewater Reclamation in Major Jordanian Industries: A Viable Component of a Circular Economy**

#### **Motasem N. Saidan 1,\*, Mohammad Al-Addous 2, Radwan A. Al-Weshah 3, Ibrahim Obada 4, Malek Alkasrawi <sup>5</sup> and Nesrine Barbana <sup>6</sup>**


Received: 13 March 2020; Accepted: 27 April 2020; Published: 30 April 2020

**Abstract:** Water scarcity remains the major looming challenge that is facing Jordan. Wastewater reclamation is considered as an alternative source of fresh water in semi-arid areas with water shortage or increased consumption. In the present study, the current status of wastewater reclamation and reuse in Jordan was analyzed considering 30 wastewater treatment plants (WWTPs). The assessment was based on the WWWTPs' treatment processes in Jordan, the flowrates scale, and the effluents' average total dissolved solid (TDS) contents. Accordingly, 60% of the WWTPs in Jordan used activated sludge as a treatment technology; 30 WWTPs were small scale (<1 <sup>×</sup> 104 m3/day); and a total of 17.932 million m3 treated wastewater had low TDS (<1000 ppm) that generally can be used in industries with relatively minimal cost of treatment. Moreover, the analysis classified the 26 million m3 groundwater abstraction by major industries in Jordanian governorates. The results showed that the reclaimed wastewater can fully offset the industrial demand of fresh water in Amman, Zarqa, and Aqaba governorates. Hence, the environmental assessment showed positive impacts of reclaimed wastewater reuse scenario in terms of water depletion (saving of 72.55 million m3 groundwater per year) and climate change (17.683 million kg CO2Eq reduction). The energy recovery assessment in the small- and medium-scale WWTPs (<10 <sup>×</sup> 104 m3/day) revealed that generation of electricity by anaerobic sludge digestion equates potentially to an offset of 0.11–0.53 kWh/m3. Finally, several barriers and prospects were put forth to help the stakeholders when considering entering into an agreement to supply and/or reuse reclaimed water.

**Keywords:** reclaimed water; circular economy; anaerobic digestion; biogas; reuse; water pricing; water depletion; industrial sector

#### **1. Introduction**

Water is becoming a limited resource in terms of quantity and quality due to the growing global economy and population, accelerating urbanization, and climate change effects [1–3]. Water reuse has been employed as an alternative water supply in arid and semi-arid regions [4,5]. In this context, wastewater and water reclamation plays a vital role in sustainable water resource management and mainly in various application such as agricultural irrigation, industrial processes, aquaculture, and for any non-human contact utilization, etc. [5–8]. Moreover, reclaimed wastewater is a resource that can be continuously produced unaffected by climatic conditions [9,10], especially in the Mediterranean region, one of the most vulnerable areas to climate change and with limited water resources [11–13].

However, the potential of recycling and reusing treated wastewater in a transition to a circular economy should be exploited thoroughly in arid and semi-arid areas, since it could synergize the wide adoption of water reuse as an alternate water supply [14–16].

#### *Wastewater Reclamation Overview in Jordan*

Jordan is classified as a semi-arid to arid country, with scarce water resources compared with other countries in the Middle East, and is ranked among the poorest countries in the world in terms of water availability [17–25]. Figure 1 shows the water resources in Jordan including locations of wastewater treatment plants (WWTPs). The Syria crisis is still adding strain on Jordan's economy and infrastructure and has put pressure on all sectors including water, municipal services, and electricity supply [26–34]. This problem is even more intense in areas with high population due to refugee influx that caused unsustainable over-exploitation of groundwater, and consequently led to increasing groundwater salinity and depleting resources (i.e., the water table was reduced by 5 m in areas like Dhuleil-Hallabat, area of the Amman-Zaraq basin, and tripled in salinity) [35].

Most of the published literature on water reuse in the Middle East focused on reclaimed wastewater uses in agricultural fields [36–40]. For instance, Hussain et al. (2019) reviewed 124 recent publications on the multiple aspects of safe use of treated wastewater for agriculture, landscape, and forestry and for non-conventional water resources management [40]. Moreover, it is also reported that approximately 20 million hectares of arable land worldwide is irrigated with wastewater [41].

Considering the water scarcity situation, Jordan has given top priority to the use of reclaimed wastewater in agriculture and industrial sectors [42–45], hence, the reuse of wastewater in agriculture has replaced freshwater resources, which were previously used for irrigation, allowing freshwater to be reallocated to the municipal sector where there is higher demand and quality water is needed for potable use. Despite of that, the agricultural sector accounts for 75% of all water consumption in Jordan and produces only 2% of the Gross Domestic Product (GDP) [46]. On the other hand, to reuse the reclaimed wastewater eco-efficiently in the industrial sector, most of the industrial facilities need to improve their wastewater management practices and upgrade their on-site treatment units to treat the wastewater before use [47].

The WWTPs play a vital role to decrease the environmental impacts of municipal and industrial discharges [48], while having an advanced (tertiary) treatment, the wastewater recycling and reuse can be promoted [49], as well as, enhancing the recovery of materials or energy [50]. Wastewater reclamation is one of the recommended solutions for the problem of water scarcity although the process may be complex, costly in terms of resources, and energy demanding depending on the quality of treated wastewater and the adopted technology for tertiary treatment [51]. However, shifting of the WWTP effluents from their application in agricultural irrigation to the industrial sector will require recognition of the fact that some agricultural activity would no longer have access to water for irrigation. Despite this, such a shift is recommended since most of the existing conventional treatment of WWTPs (mechanical chemical and biological treatment) does not eliminate emerging pollutants (i.e., pharmaceuticals and personal care products, hormones and steroids, persistent organic pollutants, etc.) from the wastewater, which can be induced into the food chain, subsequently causing adverse ecological and human health effects [52].

So far, wastewater reclamation and reuse in the context of water shortage in Jordan is not high, overall, whereas the potentiality of wastewater reuse is huge. The objectives of this paper are to compressively analyze the current status of wastewater reclamation and its reuse in major industries

in Jordan, and to summarize the opportunities and the challenges of expanding wastewater reuse, and then to put forth prospects for future wastewater reclamation and reuse in Jordan.

**Figure 1.** Water resources in Jordan.

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

#### *2.1. Wastewater Treatment Plants in Jordan*

Jordan has a fair operational capacity in wastewater treatment, although it is highly cost-intensive. The 34 central WWTPs are expected to treat 240 million m3 per year (MCM/year) by 2025 [18]. Increasing sanitation coverage is expensive, and the proposed shift in water sector expenditures from water supply to sanitation in 2011–2013 is a significant step toward increasing coverage. In 2013, collection costs amounted to JOD 47 million (1\$ is 0.71 Jordan Dinar (JOD)) and treatment costs to JOD 43.1 million [53]. Moreover, water and sanitation service costs are subsidized. Combined water and sewer bills amount

to less than 0.92% of the total household annual expenditures. With Jordan's population expected to almost double by 2050, water demand will exceed the available water resources by more than 26% [18].

Figure 2 shows the variety and distribution of 34 different processes in WWTPs in Jordan. The most widely used technologies are the activated sludge (AS) process with a share of 60%. Followed by the wastewater stabilization pond (WSP) process with a share of 19%. While the trickling filter (TF) and AS process, Membrane Bioreactor (MBR) and TF process, and oxidation sludge (OS) process were evenly having the same use share of 6%, respectively. The TF process was the least used technology with a share of 3%. Moreover, one of these WWTPs is of super-large scale (><sup>30</sup> <sup>×</sup> <sup>10</sup><sup>4</sup> <sup>m</sup>3/day), 4 WWTPs are of medium scale (1 <sup>×</sup> 104–10 <sup>×</sup> 104 m3/day), and 30 WWTPs are small scale (<1 <sup>×</sup> 104 m3/day), which are generally built in medium and small size cities and refugees camps.

**Figure 2.** The variety and distribution of different processes in wastewater treatment plants (WWTPs) in Jordan. AS stands for activation sludge; OS is oxidation sludge; TF is trickling filter; WSP is wastewater stabilization pond; MBR + TF is Membrane Bioreactor and TF process; and TS + AS is trickling filter and activation sludge process.

#### *2.2. Data Gathering and Analysis*

The analysis carried out in the present study is divided into four main steps as illustrated in Figure 3, which shows the methodological approach to addressing the specific objectives of this study.

A desk study was carried out for the available baseline documents (i.e., unpublished, monthly progress reports, internal memos, and minutes of meetings) and other references for collecting the technical data. The data and information used in the present study were gathered via semi-structured interviews with key stakeholders in the water (Ministry of Water and Irrigation, Ministry of Agriculture, Ministry of Environment, etc.) and industrial sectors (Ministry of Trade and Industry, Chambers of Industry, etc.), and with international funding agencies (i.e., USAID, GIZ, etc.) involved in the ongoing projects targeting integrated water resource management in Jordan. In addition, qualitative and quantitative data and information have been derived from unpublished government reports.

Moreover, before the interviews, a brief session was hosted to probe respondents for greater clarity in answers and consistency in relation to the objectives of the questions.

Information obtained through the interviews was crosschecked with the objective to reassess gaps and divergences of information.

**Figure 3.** Diagram of the framework.

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

#### *3.1. Wastewater Reclamation: Current Capacity and Potential Reuse*

#### 3.1.1. Reclaimed Wastewater Production: Overview and Potentials

Most of the WWTPs in Jordan provide secondary treatment with a variety of activated sludge processes followed by disinfection with chlorine. The exception is the Aqaba treatment facility, which provides tertiary filtration of the oxidized secondary effluent followed by ultraviolet disinfection and a chlorine residual. The total effluent of the wastewater flow from the WWTPs is around 166 million m3 based on data obtained from the Ministry of Water for the year 2018, as shown in Table 1.

The industrial sector mostly relies on fresh water, which could be used for domestic purposes. For instance, the industry uses 32.2 million m<sup>3</sup> groundwater, 4.8 million m<sup>3</sup> surface water, and 1.7 million m3 of treated wastewater [18]. Thus, this provides a great opportunity for groundwater-to-recycled water substitution.

In the present study, the WWTP effluents were classified according to their average total dissolved solids contents (TDSs) as follows: <1000 ppm; 1000 < TDS <1500; and >1500, based on wastewater analysis data (average data 2010–2016). Figure 4 shows the classification of WWTPs according to their effluents' TDS.

Table 1 shows the annual WWTP effluents' flow rate according to the TDS classifications. The first class (TDS < 1000 ppm), which relatively has the lowest TDS, can be reused several times in most industrial applications, especially in thermal units, cooling towers, etc. For instance, Aqaba recycled water, which has the lowest salinity among the WWTPs in Jordan (TDS = 587 ppm), is most readily usable in industrial applications. So potentially, this class represents 9 WWTPs distributed in different locations in Jordan, as shown in Table 1, and, in total, 17.932 million m<sup>3</sup> of treated wastewater of this class can be used directly with no or low cost of on-site treatment in the industrial sector depending on the fit-for-purpose water criteria.

However, the second class (1000 < TDS < 1500), which has medium TDS, has the highest annual effluent flow rate of 147.323 million m<sup>3</sup> in total out of 18 WWTPs distributed in widely different locations in Jordan, as shown in Table 1. The most effluent wastewater flowrate in this class is generated from Al Samra WWTP with 117.1 million m<sup>3</sup> per year by offering sanitation services to about two million in Amman and Zarqa governorates; the first and third most populated cities in Jordan, respectively [44,54]. With such large capacity and modern technology to ensure the highest purifications, Al Samra is considered as one of the largest plants in the region [40], which treats about 70.54% of total reclaimed wastewater in Jordan. This class represents 18 WWTPs distributed in different locations in Jordan, and, in total, 147.33 million m<sup>3</sup> of treated wastewater of this class can be used with medium cost of some necessary modification in the plant process in the industrial sector depending on the fit-for-purpose water criteria.

The third class has a TDS > 1500, the WWTP effluents in this class cannot be used without further intensive treatment such as: demineralization; blending with low-salinity water; and some change in the industrial process. This class represents three WWTPs with 0.788 million m3 of treated wastewater, as shown in Table 1. Therefore, due to the high capital cost of investment and relatively expensive operating cost, this class is excluded from the present study analysis.


**Table 1.** Annual WWTP effluent flow rate according to the total dissolved solids (TDSs) classifications.

Excluding food and pharmaceutical industries, the total groundwater abstraction for industrial purposes was approximately 26 million m<sup>3</sup> in Jordan in 2015 [55]. The major industries considered in the present study as the major groundwater abstracting industries are clarified in Table 2. Considering this, Figure 5 shows the total groundwater abstraction by major industries in Jordanian governorates, where the industries in Karak governorate were the most groundwater abstracting, with approximately 11.5 million m3 per year. Followed by the industries in Ma'an (4.38 million m3 per year). While the industries in Zarqa and Aqaba governorates were close to each other in terms of groundwater abstraction with 2.74 and 1.825 million m<sup>3</sup> per year, respectively. The industries in the Capital Amman were the least groundwater abstracting with 0.611 million m3 per year.

**Figure 4.** Classification of WWTPs according to their effluents' TDS.

**Figure 5.** Percentages of total groundwater abstraction by major industries in Jordanian governorates.



Hence, based on the data of first class and second class in Table 1, the potential reclaimed wastewater substitution in major industries in Jordanian governorates is shown in Figure 6. It is obvious that the reclaimed wastewater in Zarqa governorate can fully substitute the industrial demand of fresh water (Figure 6a) and the needs for irrigation of 3000 donums for 20–30 farmers adjacent to Al Samra WWTP as reported by Hussein (2018) [54] and Maldonado (2017) [44]. The full substitution of industrial demand is also noticed in both Amman and Aqaba governorates with 13.13- and 3.36-fold, respectively. However, the shortage of industrial demand substitution is significantly clear in both of Ma'an and Karak governorates with substitution amounts of 2.45 and 10.4 million m<sup>3</sup> per year, respectively, as clearly shown in Figure 6b. Therefore, for the WWTPs in the governorates with a substitution factor less than one (mainly Ma'an and Karak governorates) it is preferable to prioritize their effluents (reclaimed wastewater) for irrigation use where applicable.

**Figure 6.** Reclaimed wastewater substitution in major industries in Jordanian governorates: (**a**) substitution factors, and (**b**) substitution amounts (million m3 per year).

Figure 7 shows the responses of the interviewed industries (17 samples from those shown in Table 2). It is drastically indicated that low TDS (water salinity) is the major requirement that was requested by 35% of the responses. Interestingly, the sample responses showed willingness to accept to replace the groundwater with reclaimed wastewater.

**Figure 7.** Responses of reclaimed wastewater quality requirements by the major industries.

However, 6% of the responses requested advanced treatment to receive very low values of TDS, biological oxygen demand (BOD), and chemical oxygen demand (COD). Zero total suspended solids (TSS) was requested by 17% of the responses, and this was mainly required for the cooling of power generators.

#### 3.1.2. Environmental and Economic Benefits

Pintilie et al. (2016) studied the life cycle assessment (LCA) of substituting fresh water with treated wastewater obtained from tertiary treatment and concluded that it does not lead to a substantial

improvement of environmental impact for most of the indicators [48]. However, only water depletion (WD) and climate change (CC) were considered in the present study to compare the environmental impact between reclaimed wastewater reuse and no reuse scenarios. WD is recommended for water-stressed situations because a net saving of water from nature represents the most important effect of water reuse. The WD indicator values proposed by Pintilie et al. (2016) were considered in the present assessment as the following: 5.74 <sup>×</sup> 10−<sup>4</sup> m<sup>3</sup> per m3 entering the whole system for the no reuse scenario, and <sup>−</sup>4.39 <sup>×</sup> 10−<sup>1</sup> m<sup>3</sup> per m<sup>3</sup> entering the whole system for the reclaimed wastewater reuse scenario [48]. Negative values mean benefits to the environment, and positive values mean damages. Accordingly, using the data in Table 1, the annual wastewater effluent amounts (mainly the total flowrates of grouped WWTPs (million m3/year)) of both TDS less than 1000 ppm and 1000 < TDS < 1500 ppm were 17.93 and 147.33 million m<sup>3</sup> per year, respectively. The sum of them is 165.26 million m3 per year, and using the aforementioned WD indicators, the analysis revealed that 94,860 m<sup>3</sup> of fresh water are depleted for the scenario of no-reuse of reclaimed wastewater; however, 72.55 million m3 of water can be saved in reclaimed wastewater reuse in major industries in Jordan, as shown in Figure 8. Results of a similar tendency were founded in literature [48,56].

**Figure 8.** Water depletion (million m3 of water) and climate change (million kg CO2Eq) analysis for both reclaimed wastewater reuse and no reuse scenarios (negative values mean benefits to the environment, and positive values mean damages).

The CC indicator values proposed by Pintilie et al. (2016) were considered as stated above [48]. The CC indicators were with negative values (indicates benefits to the environment) according to Pintilie et al. (2016) are the following: <sup>−</sup>1.07 <sup>×</sup> 10−<sup>1</sup> and <sup>−</sup>3.20 <sup>×</sup> 10−<sup>2</sup> kg CO2Eq per m<sup>3</sup> reclaimed wastewater for both scenarios of reuse and no reuse, respectively [48]. Accordingly, using the data in Table 1, the annual wastewater effluent amounts (mainly the total flowrates of grouped WWTPs (million m3/year)) of both TDS less than 1000 ppm and 1000 < TDS < 1500 ppm were 17.93 and 147.33 million m<sup>3</sup> per year, respectively. The sum of them is 165.26 million m<sup>3</sup> per year, and using the aforementioned CC indicators, as shown in Figure 8, both scenarios showed beneficial impacts (negative values) to the environment in terms of climate change impacts. The no reuse scenario has relatively higher benefits with 17.683 million kg CO2Eq reduction compared to a 5.288 million kg CO2Eq reduction for the reuse scenario.

Normally, several factors influence the reclaimed wastewater provision and exploitation as a substitute [57,58]. According to the economic analysis of wastewater reclamation in Jordan, the difference between water price and reclaimed wastewater price plays a vital role in the willingness of the industries to accept the reclaimed wastewater as substitute. Therefore, for the low TDS (<1000) reclaimed wastewater (Table 1), the average cost of one m3 of reclaimed wastewater is estimated at

0.55 JOD (including the pipeline installation, pumping electricity, and operation and naintenance (O&M) costs), while the cost of fresh water is 1 JOD/m3. In this case, the reclaimed wastewater is competitive to some extent with regard to its price advantage. Moreover, based on experts' estimation, the environmental value of groundwater saved in the groundwater aquifer is 1.5 JOD/m3. Hence, the cost–benefit analysis of this case (water of TDS < 1000) is attractive for the consumer and the government.

While for reclaimed wastewater with TDS higher than 1000 ppm, a treatment is needed based on the application. Therefore, excluding the reuse of reclaimed in cement and concrete industries, the average cost of one m<sup>3</sup> of reclaimed wastewater is estimated at 2 JOD (including treatment, pipeline installation, pumping electricity, and O&M costs). It is worth mentioning that the long-distance pipelines from WWTPs to industrial zones and clusters, were the major cause for such costly per m3 water cost, especially in southern Jordan clusters. In order to overcome the hesitance of industries to reuse reclaimed wastewater when advanced treatment is required, subsidies by way of discounted cost of water should be provided in addition to fund allocation for capital cost coverage when on-site treatment is needed, as well as policy reforms to enhance the financial sustainability of the water sector.

#### *3.2. Energy Recovery from Wastewater Reclamation*

Wastewater treatment in WWTPs (mainly AS treatment process) requires around 0.38–2.74 kWh/m<sup>3</sup> in Jordan, as shown in Figure 9. Additionally, 0.95–1.25 kWh/m3 is needed for wastewater as reported in literature [59,60]. The difference in energy use needed for wastewater reclamation and supply can be reduced by recovering organic energy during the wastewater treatment process [59]. Currently, only in the Al Samra WWTP, biogas production from sludge treatment is undertaken in Jordan. As shown in Figure 10, the two types of thickened sludge are mixed in two covered tanks of 98 m3 volume before being pumped and introduced in seven anaerobic digesters of a capacity of 15,900 m3 each. In the digesters, the sludge is mixed thoroughly by Cannon®mixers (Trevose, PA USA) using the recycled compressed biogas. The sludge stays for three weeks at 35 ◦C in the digesters. Heating is done by hot water recovered from the cooling of the engines in a shell-and-tube heat exchanger. Through hydro energy and biogas production, the Al Samra WWTP has a potential energy recovery of 95% of its needs, only 5% is drawn from the national grid. Moreover, 300,000 tons of CO2 is saved per year through energy recovery and renewable energy utilization [61].

The introduction of anaerobic sludge digestion is generally expected to offset 25–50% of an aerobic wastewater treatment plant's energy needs [59,63,64], however, based on WWTP data gathered in Jordan, having anaerobic sludge digestion in the small- and medium-scale WWTPs (<<sup>10</sup> <sup>×</sup> 104 m3/day) can potentially produce electricity that would equate to an offset of 0.11–0.53 kWh/m3. Consequently, this may help in reducing the costs of reclaimed wastewater reuse with further treatment requirements mainly for reclaimed wastewater with TDS higher than 1000 ppm as stated before.

However, energy produced from anaerobic sludge digestion can be feasibly increased by co-digestion with kitchen or other organic wastes [65–69]. Currently, the co-digestion is only applied at a laboratory scale in Jordan. Al-Addous et. al. (2019) evaluated the potential biogas production from the co-digestion of municipal food waste and wastewater sludge at a refugee camp. Accordingly, a possible ratio to start with is 60–80% organic waste, which can produce 21–65 m<sup>3</sup> biogas ton−<sup>1</sup> of fresh matter [70].

Notwithstanding that co-digestion does not exist in Jordan yet, the anaerobic digestion systems tend to be well operated in Jordan (i.e., Al Samra WWTP). Hence, when co-digestion is utilized in Jordan, this will be a vital opportunity to make cost-effective use of existing facilities and improve sludge biogas potential [71–74].

**Figure 9.** WWTPs' electricity consumption data for the activated sludge (AS) treatment process.

**Figure 10.** Biogas production in anaerobic digesters at the Al Samra WWTP [62].

#### *3.3. Reclaimed Wastewater Reuse: Barriers and Prospects*

#### 3.3.1. Reclaimed Wastewater Quality and Industrial Needs

Industrial uses of reclaimed wastewater come in many different ways such as cooling-water, processing, and boiler feed water. Therefore, the process water requirements for water quality vary depending on the industry. Some of the concerns for industrial use of reclaimed wastewater are corrosion, scaling, and biological growth; however, these concerns are applicable to potable water as well. Most of cooling water is treated already to address these concerns. For instance, corrosion is a concern in cooling water no matter whether the facility uses potable water or reclaimed wastewater. Scaling from dissolved minerals such as calcium, magnesium, and phosphates can be controlled by monitoring and chemically treating the water to prevent scaling. Magnesium–phosphorus precipitation from sludge and the recovery of struvite after anaerobic sludge treatment process is conducted in order to prevent clogging in pumps and pipes in any further reuse applications [75–78]. Biological concerns can be addressed by adding chlorine to levels of 2.0 mg/L that will kill most microorganisms that causes corrosion or deposits in cooling systems [79].

To facilitate the use of recycled water in industrial applications, the information on the quality of the municipal recycled water should be provided and available to the industrial users. Moreover, opportunities to improve water quality for specific purposes, either by the supplier through additional treatment and/or source control, or the industrial user can improve treatment and control processes to levels specific to its process needs.

#### 3.3.2. Reclamation of Wastewater Technologies

Lyu et al. (2016) discussed all advances in technology by which wastewater may be treated to meet the most stringent quality requirements and be used for any purposes desired [49]. For instance, the technologies applied in wastewater reuse include: (1) oxidants for disinfection purposes using sodium hypochlorite; ultraviolet radiation [80]; and ozone for high bactericidal disinfection [81] and the removal of 90–99% for antibiotics and estrogens [82]; (2) biological treatments such as anaerobic, maturation ponds and constructed wetlands [83–85]; (3) physical separations such as membrane filtration for 81% removal of electroconductivity, 83% for Na+, and 80% for Cl<sup>−</sup> [86]; the removal of 95% of heavy metals [87]; the removal of >89% of pharmaceuticals [88]; (4) electrochemical treatments to completely remove *Escherichia coli* [89–91]; as well as, (5) solar photocatalysis with TiO2 for >90% removal of emerging pollutants (i.e., pharmaceuticals and personal care products) [92], and removals of 33% for Cd and 75% for Co [93].

#### 3.3.3. Reclaimed Wastewater Supply Continuity

The industry demands a constant non-interrupted flow of reclaimed wastewater throughout the day [94]. In Jordan, although the reclaimed wastewater supply volumes vary diurnally and seasonally, its continuity is not so critical since WWTP effluents have relatively uninterruptable higher flows than the demand flows needed by the nearby main industries. It is noteworthy that flow equalization and water conveying capacities should be investigated to match the supplies with the demands and vice versa [95].

#### 3.3.4. Willingness to Participate and Willingness to Pay

As deduced based on the results of these interviews, most of main industries considered in the present study expressed a positive stance toward reclaimed wastewater reuse, while they are willing to pay a significantly less amount of money than they already pay, for freshwater. Therefore, a comprehensive survey about the willingness of the industrial sector to switch to the use of reclaimed wastewater instead of groundwater is of high significance. Such surveys will help in providing more accurate data for the financial evaluation of the recycled water service and a basis for negotiation with the industries.

Factors that influence industrial user's 'willingness to pay' for reclaimed wastewater include: (1) price of alternative water sources (i.e., potable, surface water, and groundwater supplies); (2) perception of the scarcity of alternative sources; (3) capital and operating costs of switching to reclaimed wastewater supply; (4) reclaimed wastewater quality, quantity, and levels of service and reliability of supply.

#### 3.3.5. Pricing Systems

A range of pricing systems for reclaimed wastewater can be proposed in Jordan and assessed on a win–win situation. The pricing systems can be employed alone or in combination [96], which are, but not limited to: (1) A usage fee scheme in which the industries finance the infrastructure installation, and then the usage charge offsets the supply cost of the reclaimed wastewater. For instance, such type of pricing was adopted in 2003 by the Australian government under the national water reform process [97]. (2) A connection fee which is a once-off contribution toward the cost of infrastructure needed to deliver reclaimed wastewater to the industry's delivery point. This fee may be subject to negotiation between the supplier and the industries to agree on a financial arrangement where both parties may fully or partially cover the fee of the actual work to deliver the reclaimed water to the delivery point. (3) A flat fee regardless of use ("take or pay" arrangement). For instance, regardless of actual use, the industries are obliged to pay for 75–100% of the contracted recycled water volume, and for all water consumed by the industries above the contracted level. Although this pricing scheme provides the WWTPs with guaranteed income that sustains the financials of running the scheme, it may encourage overuse of reclaimed wastewater by the industry and improper discharges to the environment.

#### 3.3.6. Reclaimed Wastewater Agreements

Specific reclaimed wastewater guidelines are important in managing the supply and use of reclaimed wastewater particularly in relation to quantity and quality [98]. Through the agreement negotiations between the supplier of reclaimed wastewater and the customers (i.e., industries). Wherein, the parties agree to a set of obligations and responsibilities under which the reclaimed wastewater reuse scheme will operate [99]. Key issues that reclaimed wastewater agreements should cover include: (a) price, quantity, and quality of reclaimed wastewater; (b) security of the reclaimed wastewater supply; (c) measures to identify, allocate, and manage risks and ensure safe use of reclaimed wastewater; (d) liabilities and insurance for potential damages caused by supply and use; and (e) compliance with legislative and common law requirements.

#### **4. Conclusions**

The following findings can be concluded in the present study:


**Author Contributions:** Conceptualization, M.N.S. and I.O.; Methodology, M.N.S., I.O., and R.A.A.-W.; Formal analysis, M.N.S., M.A., M.A.-A., N.B., and I.O.; Investigation, M.N.S., R.A.A.-W., and I.O.; Resources, M.N.S. and I.O.; Writing—original draft preparation, M.N.S., M.A., M.A.-A., and I.O.; Writing—review and editing, M.N.S. and R.A.A.-W.; Project administration, M.N.S., N.B. and I.O. All authors have read and agreed to the published version of the manuscript.

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

**Acknowledgments:** The corresponding author does acknowledge with gratitude the University of Jordan for granting him sabbatical assistance to undertake this research study. The authors also thank the anonymous reviewers and the editor of the journal for their constructive feedback.

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

#### **References**


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