**Urban and Industrial Wastewater Disinfection and Decontamination by Advanced Oxidation Processes (AOPs): Current Issues and Future Trends**

Editors

**Stefanos Giannakis Ana Rita Lado Ribeiro Jorge Jesus Rodr´ıguez-Chueca**

MDPI • Basel • Beijing • Wuhan • Barcelona • Belgrade • Manchester • Tokyo • Cluj • Tianjin

*Editors* Stefanos Giannakis Universidad Politecnica de ´ Madrid (UPM) Spain

Ana Rita Lado Ribeiro University of Porto (FEUP) Portugal

Jorge Jesus Rodr´ıguez-Chueca Universidad Politecnica de ´ Madrid (UPM) Spain

*Editorial Office* MDPI St. Alban-Anlage 66 4052 Basel, Switzerland

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## **Contents**



A Pilot Study Combining Ultrafiltration with Ozonation for the Treatment of Secondary Urban Wastewater: Organic Micropollutants, Microbial Load and Biological Effects Reprinted from: *Water* **2020**, *12*, 3458, doi:10.3390/w12123458 ..................... **165**

## **Ra ´ul Acosta-Herazo, Briyith Ca ˜naveral-Vel´asquez, Katrin P´erez-Giraldo, Miguel A. Mueses, Mar´ıa H. Pinz ´on-C ´ardenas and Fiderman Machuca-Mart´ınez**


## **About the Editors**

**Stefanos Giannakis**, Doctor of Environmental Engineering from the Aristotle University of Thessaloniki, Greece, and the Polytechnic University of Catalonia, Spain, is currently a "Ramon´ y Cajal" Fellow at the Polytechnic University of Madrid, Spain. His main research lines include the development of novel, sustainable catalytic (waste)water treatment processes, unraveling the fundamentals governing microorganism disinfection, as well as elucidating the photochemistry and photobiology of natural waters. He has participated in 6 National and International projects on the chemical and microbiological quality of water and wastewater in Europe, Africa, and Latin America. Dr. Giannakis is currently the author of 75 peer-reviewed research articles, has contributed to 35 conference works and has an h-index of 25.

**Ana Rita Lado Ribeiro** holds a Researcher position at Laboratory of Separation and Reaction Engineering–Laboratory of Catalysis and Materials (LSRE–LCM) hosted by Faculty of Engineering of University of Porto and is Associate Editor of *Chemical Engineering Journal Advances*—Elsevier. Her main research interests include the analysis of contaminants of emerging concern in the environment; novel miniaturized sample preparation methodologies; water/wastewater treatment by advanced oxidation processes; and identification of reaction by-products. She is Principal Investigator of a funded project at LSRE-LCM (SAMPREP) and FEUP team leader of another one (ENANTIOTOX), having participated as team member in >15 research projects. She has co-authored 2 patent requests, 55 publications in JCR indexed journals and 8 chapters in books or in book series with peer review (h-index 25), presenting her work at highly ranked international conferences.

**Jorge Jesus Rodr´ıguez-Chueca**, Doctor of Chemical Engineering and Environmental Technologies from the University of Zaragoza, Spain, is currently Associate Professor I3 at the Polytechnic University of Madrid, Spain. Jorge has contributed 40 peer-reviewed research articles, 7 technical articles in non-indexed journals, 3 book chapters; in addition, he was the editor of the book of abstracts of IWA Young Water Professionals Conference 2019, with an h-index of 17. He has participated in 7 Regional, National and International projects on water treatment and management, being the Principal Investigator in three of them. His main research interests are the new technologies for environmental remediation, mainly focused on water and wastewater treatment (disinfection, removal of micropollutants, organic matter, etc.) especially through the use of Advanced Oxidation Processes.

## *Editorial* **Urban and Industrial Wastewater Disinfection and Decontamination by Advanced Oxidation Processes (AOPs): Current Issues and Future Trends**

**Ana Rita Lado Ribeiro 1, Jorge Jesus Rodríguez-Chueca <sup>2</sup> and Stefanos Giannakis 3,\***


**Citation:** Lado Ribeiro, A.R.; Rodríguez-Chueca, J.J.; Giannakis, S. Urban and Industrial Wastewater Disinfection and Decontamination by Advanced Oxidation Processes (AOPs): Current Issues and Future Trends. *Water* **2021**, *13*, 560. https://doi.org/10.3390/w13040560


Academic Editor: Anas Ghadouani

Received: 4 February 2021 Accepted: 19 February 2021 Published: 23 February 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/).

Without any doubt, the 21st century has kick-started a great evolution in all aspects of our everyday life. The tremendous advances in medicine, production of goods (such as chemicals, plastics), etc. have all contributed to an increase in the standards and the comforts of the modern society. However, combined with the growing demands in water due to the uneven population increase and product manufacturing needs, the rapid change has impacted the wastewater flows and composition of both urban and industrial origin. Environmental engineers have been facing a constantly changing landscape, where new problems seem to emerge in an alarming, "whack-a-mole"-like situation. From antibiotic resistance to microplastics, from the ever-growing list of contaminants of emerging concern to their transformation products, the need for proper, advanced wastewater treatment methods are dominating environmental engineering research to alleviate the shift of urgent needs of the water sector. Nonetheless, the multidisciplinary nature of this huge environmental problem spans a number of research topics, namely environmental, chemical and process engineering, water and sanitary engineering, environmental science, chemistry, biology and related disciplines, material science, toxicology, risk assessment, economy, social sciences, ecology and environmental policy, among others. Therefore, in order to develop efficient solution tools to tackle the urban and industrial wastewater challenges, it is of utmost importance to promote the knowledge exchange between all expertise fields involved, thus promoting a scientific, technological, and societal cooperation for a more sustainable environment and society.

It has also become clear over the past two decades that the perception of water as an infinite resource is distorted and a dangerous idea to hold on to. As a result, novel and more holistic approaches, such as effluents' reuse, the circular economy of water, the onehealth integration in wastewater treatment, have gradually replaced the archaic views on treatment and discharge of wastewater. The (well-deserved) price paid for adopting such viewpoints into policy and legislation has been partially transferred to the end of the line, namely wastewater treatment plants (WWTPs). Lower limits, recovery schemes, safer disposal and handling of wastes illustrate the reality of the treatment schemes nowadays.

In 2020, the Special issue "Urban and industrial wastewater disinfection and decontamination by Advanced Oxidation Processes (AOPs): current issues and future trends" was launched to gather the innovations of the research community on (waste)water treatment worldwide. Following the call, several articles on various aspects of (waste)water treatment have been published, both in form of original research papers and literature

reviews. Authors affiliated with institutions and companies from Colombia to Switzerland, from Portugal to India and from Spain to Iran, have contributed by disseminating their research results within this Special Issue; a brief summary of the published papers follows:

Guerra-Rodríguez et al. [1] overviewed the different strategies that can be adopted in the water sector in the context of the circular economy. Wastewater reuse is a mature and wellestablished strategy in many water-scarce regions of the world, but its reuse poses associated risks to ecosystems and humans due to the presence of pathogens, contaminants of emerging concern, and microbial resistance genes. However, there are other options for applying the circular economy in the water sector that are still largely unexploited, such as the recovery of materials from aqueous streams or energy recovery; Guerra-Rodríguez et al. reviewed all these options, highlighting the challenges and opportunities in their application.

Camargo-Perea et al. [2] reviewed extensively the data on the degradation of emerging pollutants by ultrasound AOP, with a special emphasis on pharmaceuticals, providing a deep understanding on the operation fundamentals and the parameters that most affect its efficiency, namely, the ultrasonic frequency, the electrical power, the pH and temperature of the solution, the nature of the target contaminant, the constituents of the water matrix, and the type and the geometry of the sonochemical reactor. In addition to the applications that were detailed, this review discussed the future perspectives and the cost implications of ultrasound AOP as a possible candidate to tackle water pollution by emerging pollutants.

Guateque-Londoño et al. [3] studied the removal of the antihypertensive losartan as a pharmaceutical model in simulated fresh urine through the application of ultrasound and UVC/H2O2. They reported higher selectivity of the sonochemical process traduced by a higher ratio between the degradation rate constants obtained in fresh urine and distilled water for the removal of losartan than that obtained in the UVC/H2O2 system. Although neither of the treatments reached the mineralization of the pollutant in distilled water, it was confirmed that the sonochemical system reduced the phytotoxicity of the sample, showing the potential of this technology on the regeneration of wastewater. Finally, the authors studied the generation of transformation products from the degradation of losartan, and the fitting of these results with theoretical methods to predict and rationalize the attack of degrading species. The computational analyses confirmed that the atoms on imidazole moiety of losartan were the most susceptible to transformation by the radical species, in agreement with the experimental results obtained.

The presence of antibiotics in wastewater effluents is a problem of emerging concern threating ecosystems and human health. Moles et al. [4] evaluated the behavior of five antibiotics (amoxicillin, enrofloxacin, sulfadiazine, trimethoprim, and azithromycin) in the influent and effluent of different WWTPs located in the north of Spain. Moles et al. determined that azithromycin was the antibiotic with the highest mass load followed by enrofloxacin, trimethoprim, sulfadiazine, and amoxicillin. Regarding the behavior of these substances in the WWTPs, they observed that biological treatments have a significant influence on removal, especially the use of trickling filters that showed the highest degree of removal. Besides, the authors studied the performance of a TiO2 photocatalytic treatment plant installed as tertiary treatment that minimized the Ti release into the environment and allowed catalyst reuse. This technology totally removed sulfadiazine, amoxicillin and azithromycin, while 80% of trimethoprim and enrofloxacin was removed.

Hospital effluents are the main emitters of pollutants of emerging concern such as pharmaceuticals and antibiotics. Rosero Parra et al. [5] described the treatment of hospital wastewater using two novel catalysts supported on graphene (SnO-GO and TiO2-GO). Both heterogeneous photocatalysts were demonstrated as alternatives for abatement of pollution in this kind of effluents, reducing chemical oxygen demand by 85%, the dissolved organic carbon by 94%, and 80% of phenols, allowing the discharge of the effluents in compliance with current legislation.

Karbasi et al. [6] studied a semiconductor photocatalytic material based on Bi and W oxide, as an alternative to the traditional TiO2 photocatalytic disinfection process. Owing to its large band gap and exceptional structure (micro/nanohierarchical, flower-like mor-

phology) it absorbs more photons and further into the visible range, while it reduces the kinetic limitations associated with semiconductors. After the complete characterization of the newly synthesized material, *E. coli* bacteria were used as a model microorganism and the disinfection capabilities of Bi2WO6 were found to match the benchmark TiO2 P25 under solar light but had superior bactericidal efficacy under visible light. This opens the possibilities of further studying this material and advance towards indoor applications, where sterility is a prerequisite and plain fluorescent lighting is provided.

The photocatalytic degradation of potassium hexacyanoferrate (III), as a model cyanocomplex of gold mining wastewaters, was studied by Arce-Sarria et al. [7] in a bench-scale compound parabolic collector (CPC) reactor assisted with a light-emitting diode (UV-A/LED) and a hydrothermally treated TiO2 P25, which showed a specific surface area 2.5-fold superior to the original catalyst, a slightly higher band gap energy, and a mesoporous structure. Although the TiO2 P25 performed better, the higher free cyanide release achieved with the modified material in this specific case may be beneficial for its reuse in the gold extraction process. This report opens new prospects for future research on the structural changes of the catalyst and their potentialities for photocatalytic applications.

Graça et al. [8] proposed a combined technology (ozonation followed by ultrafiltration) to simultaneously reduce organic contaminants and microbial load of secondary urban wastewater for its further reuse, complying with the legislation for water quality for irrigation. The estrogenic activity, cell viability and cellular metabolic activity were also monitored in order to assess the impact of the treatment on the biological effects. Although most organic microcontaminants were removed (except citalopram and isoproturon), the biological effects did not suggest the production of toxic by-products, and the reduction of the bacterial loads targeting the water quality standards for irrigation was effective, the authors highlighted that microbial regrowth was observed after storage, with the concomitant increase in the genes 16S rRNA and *intI*1.

Finally, Acosta-Herazo et al. [9] showed the main features of the PHOTOREAC tool developed, as well as the results in different simulations, thanks to the experience gathered by their research groups at Cartagena University (Cartagena de Indias, Colombia) and the Universidad del Valle (Cali, Colombia) during the last twenty years of research in heterogeneous solar photocatalysis, and on extensive literature research in photoreactor engineering. PHOTOREAC is an open-access application developed in the graphical user interface of MATLAB® that allows a user-friendly evaluation of the solar photoreactors operation. Furthermore, they showed the potential of the tool in several case studies such as the removal of dichloroacetic acid and methylene blue in different types of photoreactors, as well as the modeling of radiation in a Flat Plate Photoreactor.

Considering the above contributions and the current issues dominating literature, the recent challenges faced by WWTPs exceed their regular capacity, and the need for appropriate wastewater treatment calls for novel, sophisticated methods of decontamination prior to its discharge or reuse. Emerging threats such as antibiotic-resistant bacteria, antibioticresistance genes, and the contaminants of emerging concern (chemicals, microplastics) demand efficient, end-of-pipe solutions before their discharge into the environment or reclamation for reuse purposes. Under this scenario, AOPs have been procured as effective methods for wastewater disinfection and decontamination. Despite the huge amount of works developed in the last decade in this area, some research opportunities and future directions can be highlighted.

One important feature in any comparative analysis between different treatments or studies is the examination of the same endpoints for a safe effluent discharge or reuse, namely the abatement of chemical contaminants, mitigation of by-products, reduction of the toxicity of the effluent, bacterial inactivation, minimization of regrowth, control of antibiotic resistance determinants, and cost of treatment. This comparison is often difficult when different studies are focused in dissimilar endpoints, different matrices, and distinct operational conditions (typical at small-scale tests). It is important to highlight that largerscale studies than some published in this Special Issue and a further cost effectiveness

analysis of scaled-up processes are valuable for a comprehensive analysis of any given treatment option. In this regard, solar-driven AOPs are very promising when compared with other advanced technologies, being a research topic that has plenty room for further advances, for instance by developing and tailoring new photocatalysts aimed to be active under sunlight.

Regarding the endpoints, the development of tools to predict the fate and behavior of emerging contaminants during wastewater treatment would be a great asset. Given that the elimination of chemical and biological contaminants of emerging concern from wastewater by AOPs largely depends on the composition of the water matrix, studies conducted under environmental-like conditions are crucial to advance our knowledge on AOPs. In this sense, environmental relevant concentrations of mixtures of contaminants and realistic water matrices should be used instead of purely artificial conditions because the efficiency of each process results from a complex balance between inhibitory and promoting effects originated by the water components. The available data about such matrix effects on the elimination efficiency of contaminants in WWTPs is still limited and research towards the impact of the main water/wastewater constituents and the mechanisms governing it are required. These environmental-like conditions are also imperative for a better understanding of the competition effects between different contaminants; thus, the approach of studying spiked synthetic solutions spiked with one or few compounds at concentrations orders of magnitude above than those found in the environment tends to become obsolete, except for some specific applications such as performance studies of new catalysts, elucidation of degradation pathways, and characterization of chemical structures of transformation products.

As far as the fate of contaminants during treatment is concerned, the structural elucidation of the transformation products is pivotal for an accurate evaluation of the degradation pathways and for a deep understanding of their possible side-effects, like toxicity. The ecotoxicological studies provide a useful tool to clarify the harmful effects of both parent contaminants and their transformation products on the ecosystems and human health, as well as to understand the possible impact of the reagents and catalysts used. Combining other technologies can be an interesting option to deal with toxic transformation products, for instance by coupling an adsorptive post-treatment. On the other hand, AOPs are very promising as post-treatment to treat the resulting concentrate of membrane-driven processes.

Concluding, the water/wastewater treatment is a hot topic of research that has been challenged in the last decade by newly identified threats (e.g., antibiotic resistance, microplastics) that have provided new perspectives and recent opportunities for investigation. Besides the importance of bearing in mind the multidisciplinary of this topic and the need for knowledge exchange between different subjects in order to generate innovation and scientific advances, raising awareness within society and promoting individuals' involvement in the mitigation of wastewater problems is a prerequisite for a more proficient development and implementation of possible solutions.

**Author Contributions:** The authors made equal contributions to this editorial. All authors have read and agreed to the published version of the editorial.

**Funding:** Ana Rita Lado Ribeiro wouldlike to acknowledge the support of Base Funding—UIDP/50020/2020 of the Associate Laboratory LSRE-LCM—funded by national funds through FCT/MCTES (PIDDAC); and FCT funding under DL57/2016 Transitory Norm Programme. Jorge Rodríguez-Chueca acknowledges the Community of Madrid (Comunidad de Madrid) for funding the research project IN\_REUSE (APOYO-JOVENES-X5PKL6-88-KZ46KU) within the framework of the multi-year agreement with the Universidad Politécnica de Madrid. Stefanos Giannakis would like to acknowledge the Spanish Ministry of Science, Innovation and Universities (MICIU) for the Ramón y Cajal Fellowship (RYC2018-024033-I) and the "NAVIA" Project (PID2019-110441RB-C32).

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

**Informed Consent Statement:** Not applicable.

**Data Availability Statement:** Not applicable.

**Acknowledgments:** The editors would like to gratefully acknowledge the reviewers who kindly reviewed the submissions to this Special Issue, dedicating their time to provide precious scientific input towards the improvement of this collection of articles.

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

## **References**


## **Towards the Implementation of Circular Economy in the Wastewater Sector: Challenges and Opportunities**

## **Sonia Guerra-Rodríguez 1, Paula Oulego 2, Encarnación Rodríguez 1, Devendra Narain Singh <sup>3</sup> and Jorge Rodríguez-Chueca 1,\***


Received: 11 March 2020; Accepted: 13 May 2020; Published: 18 May 2020

**Abstract:** The advancement of science has facilitated increase in the human lifespan, reflected in economic and population growth, which unfortunately leads to increased exploitation of resources. This situation entails not only depletion of resources, but also increases environmental pollution, mainly due to atmospheric emissions, wastewater effluents, and solid wastes. In this scenario, it is compulsory to adopt a paradigm change, as far as the consumption of resources by the population is concerned, to achieve a circular economy. The recovery and reuse of resources are key points, leading to a decrease in the consumption of raw materials, waste reduction, and improvement of energy efficiency. This is the reason why the concept of the circular economy can be applied in any industrial activity, including the wastewater treatment sector. With this in view, this review manuscript focuses on demonstrating the challenges and opportunities in applying a circular economy in the water sector. For example, reclamation and reuse of wastewater to increase water resources, by paying particular attention to the risks for human health, recovery of nutrients, or highly added-value products (e.g., metals and biomolecules among others), valorisation of sewage sludge, and/or recovery of energy. Being aware of this situation, in the European, Union 18 out of 27 countries are already reusing reclaimed wastewater at some level. Moreover, many wastewater treatment plants have reached energy self-sufficiency, producing up to 150% of their energy requirements. Unfortunately, many of the opportunities presented in this work are far from becoming a reality. Still, the first step is always to become aware of the problem and work on optimizing the solution to make it possible.

**Keywords:** circular economy; wastewater reuse; zero waste; zero energy; substances recovery; sustainable development goals

## **1. Introduction**

It is evident that without water, there is no life. Water is not only essential for human survival, but also for all living organisms. On 28 July, 2010, water and sanitation were recognised as human rights by the United Nations General Assembly [1]. Moreover, water is essential for human activities in many sectors of the economy—and not only from a biological point of view. In the previous century, the increase of human activities in all economic sectors has favoured the development of society (e.g., increasing life expectancy). However, this was at the expense of the over-exploitation of planet resources, accompanied by the emission of greenhouse gases into the atmosphere, organic and biological pollutants into water bodies, and the generation of tonnes of solid waste [2]. In terms of

water, not only the quality of water bodies has worsened due to contamination from human activities, but there has also been increased withdrawal and consumption of water resources [3]. For these reasons, the number of regions in the world with hydric stress has grown considerably—even regions with a regular precipitation regime [4].

Under this paradigm, humanity must make decisions to increase the quality of water bodies and to increase sustainability in the management of the resources. The 2030 Agenda for Sustainable Development of the United Nations established a specific Sustainable Development Goal (SDG) to ensure availability and sustainable management of water and sanitation for all. This is the SDG 6, and it presents a specific target, 6.4 for instance, aimed to increase water-use efficiency in all of the sectors, ensuring sustainable withdrawal and supply of freshwater to address water scarcity. There are other SDGs not directly focused on water, but with specific targets involving it. This is the case of SDG 11 (to make cities and human settlements inclusive, safe, resilient, and sustainable) and 12 (to ensure sustainable consumption and production patterns). In December 2015, the European Commission adopted a Circular Economy Plan called "Closing the loop—a European Union (EU) action plan for the Circular Economy" [5]. It is a strategy to change the consumption model of European citizens towards a climate neutral, circular economy, which would minimise the impact on the environment. This plan is mainly focused on reducing, recycling, and recovering of waste, focusing on paper, ferrous metals, aluminium, glass, plastic, and wood. However, in context to the idea of "closing the loop", the circular economy can go beyond any sector, for instance, the wastewater treatment sector.

The circular economy can also be applied in the wastewater treatment sector. For example, thinking about the reclamation and reuse of wastewater seems to be an excellent option to increase water resources by reducing the environmental impacts [6]. The technology can treat wastewater to the point of being fit for use even by humans. As this idea is still unacceptable for most people, and is known as the "yuck factor" [7], the most common uses of reclaimed wastewater are agricultural, industrial, urban, environmental, or recreational purposes [6,8–10]. However, if we think not only about the value of water as a resource, but also about the substances contained in these effluents, we can observe a wide range of possibilities to apply the circular economy. For example, recovery of nutrients [11–14] and energy [15–18] from wastewater is well known, but there are more possibilities, such as the recovery of biomolecules [19–23], metals [24–27], organic and inorganic compounds [28,29], etc. On the other hand, there are other options to apply a circular economy in water, for example, the valorisation of sewage sludge. The traditional valorisation of sludge is through its reuse with agricultural purposes [30,31]. However, they can also be valorised for the preparation of carbonaceous materials with adsorptive [32–35] or catalytic abilities [36–39]. These adsorbents or catalyst can be applied at some stage of the same water or wastewater treatment, to remove organic pollutants by adsorption [40], or for the application of Advanced Oxidation Processes (AOPs) to remove organic or biological pollution [41].

The main goal of this review manuscript is to go deep in the state-of-the-art of the introduction of circular economy in the wastewater treatment sector, understood as the activities related to treatment, and valorisation and reuse of wastewater of any origin (municipal, industrial, or agricultural).

Throughout this manuscript, the degree of application of strategies to close the loop in the wastewater sector will be addressed. Actions with a high degree of maturation will be described, as well as the reclamation and reuse of wastewater, or the nutrients and energy recovery and others in a nascent level of development, in the form of recovery of biomolecules or valorisation of sewage sludge.

#### **2. Wastewater Reclamation and Reuse**

#### *2.1. Definition and Overview of Reclamation around the World*

The reclaimed wastewater is treated when the residual effluents are subjected to an additional or complementary treatment to increase or make adequate its quality to the final reuse purposes. Additionally, it is considered as a non-conventional water resource.

The reuse of these reclaimed effluents lies in their application in different purposes, instead of discharging them to water bodies after conventional treatment. According to current national legislations around the world, the reuse purposes are varied, such as agricultural, industrial, urban, recreational, environmental, or even human consumption [6]. The reclamation wastewater treatment plant (RWWTP) is normally a wastewater treatment plant equipped with a tertiary treatment composed of different technologies, depending on the quality requirements of the final effluent regarding the final use.

Besides, there must be a distinction between the indirect and direct reuse of treated wastewater, also called as unplanned or planned reuse respectively [42]. The indirect reuse is defined as the discharge of treated effluents to the watercourses, and after passing through the public hydraulic domain, is subject to its subsequent use. The direct reuse of treated wastewater is that in which the second use occurs after the first, without water being incorporated into the public hydraulic domain between them [43]. This review mainly describes the experiences in direct reuse of the treated effluents.

It is difficult to talk about reclamation and reuse of wastewater when approximately 80% of wastewater is discharged into the world's waterways without any kind of treatment [44]. The construction of RWWTPs would solve the problems associated to the uncontrolled discharge of wastewaters, as well it would help closing the loop reducing the withdrawal of water resources (e.g., with agricultural purposes), thereby, doing more sustainable management of this resource.

Until now, the reclamation and reuse of wastewater has been a priority exclusively for countries with significant problems of hydric stress or water scarcity. Figure 1 shows the top countries reusing wastewater, dividing the figures regarding the total use (m3/d; Figure 1A), total reuse per capita (m3/d; Figure 1B), and the ratio reuse:water extraction (Figure 1C) in 2008. Despite the outdated data shown in Figure 1, due to the absence of a current ranking, and the difficulty of individual search for accurate data from each country, these can be used as a basis for understanding which countries have a need to reuse wastewater. In this paradigm, during 2018, China and Mexico were at the top of the list of the countries reusing wastewater, with almost 15 hm3/d; however, in both cases, this regeneration is happening without additional treatment. If this point is considered, the top country to reuse the reclaimed wastewater was the United States of America, with almost 8 hm3/d. However, this ranking changes if the volume per capita is considered, or the percentage of reuse (considering the extraction of water resources). Under this classification, Qatar, Kuwait, and Israel would be the top countries, Kuwait being the country with the largest ratio of reuse regarding the withdrawal of water resources (35.2%).

**Figure 1.** Global situation of wastewater reuse around the world. Adaptation from [45]. (**A**) Total reuse of wastewater (hm3/d); (**B**) reuse (hm3/d per million capita); (**C**) percentage of reuse/extraction.

### *2.2. Legislation and Guidelines Around the World*

It is quite difficult to compile all the legislation and guidelines around the world about water reclamation and reuse, for this reason, this section tries to summarize the most important guidelines, as well as the legislation of some leading and emerging countries in this field.

In 2004, the Environmental Protection Agency of the United States of America (US EPA) published "*Guidelines for water reuse*". This guideline was revised later, in 2012. In this document, different physico-chemical and biological parameters are established regarding the final use of the reclaimed water. In this way, for instance, reuse of reclaimed water is recommended for agricultural purposes only when the concentration of *Escherichia coli* is below 1 CFU/100 mL. This recommendation contrasts with the ones given by the World Health Organization (WHO) in 2006. Thus, the WHO published the "*Guidelines for the safe use of wastewater, excreta and greywater*". This document tried to establish the maximum admissible values for different physico-chemical and biological parameters posing a threat to human health, considering the reuse of reclaimed wastewater in agriculture and aquaculture. In this case, the WHO considered that reclaimed water can be reused as long as the concentration of *E. coli* is below 1000 CFU/100 mL. This difference between both guidelines is evidently the reason why WHO guidelines are applied in developing countries. Usually, the populations in these countries have limitations (economical, lack of government support, lack of qualified personnel, etc.) in terms of access to technologies for water treatment.

In the European Union, the reclamation and reuse of water encounter numerous barriers in most of the member countries, for example, a supportive and coherent framework in terms of legislation. This is the reason why the Commission is working on common legislation for all of the members in the frame of the Circular Economy Plan. Nowadays, some European countries have the legislation to regenerate and reuse wastewater (e.g., Spain, Portugal, France, etc.); others are preparing legislation, while others have none at all. Table 1 summarizes the sectors in which reclaimed water is currently

applied on by country, and the regulations or guidelines of each country (European Union countries and outside EU countries).

Spain was ranked as the country in Europe with the highest rate of wastewater reuse, and, as can be observed in Figure 1, is in the top ten worldwide [46]. The Royal Decree 1620/2007 (RD 1620/2007) establishes the legal regime for the reuse of reclaimed water in Spain, thereby setting up different maximum admissible values for different physico-chemical and microbiological parameters, depending on the final use for the reclaimed water. In relation to the microbiological parameters, *Escherichia coli* and intestinal Nematodes (*Ancylostoma*, *Trichuris*, and *Ascaris* genera) must be controlled for all established uses. *Legionella* analysis is also mandatory for all of the established water uses, except for environmental use, which does not contemplate its control. *Salmonella* spp. should be controlled when the uses are agricultural and industrial. Finally, the control of the parasitic plathyhelminthes *Taenia saginata* and *Taenia solium* is mandatory if the water is used to irrigate pastures for consumption by meat-producing animals, since they are intermediate hosts of these parasites.

In addition to the existing quality criteria in RD 1620/2007, depending on the use to which the reclaimed water is going to be destined, others included in specific matters should be considered. For example, it needs to be in commensuration with the parameters established in Directive 2006/118/CEE, regarding the protection of groundwater against pollution and deterioration, when the reclaimed water is used for environmental purposes. This standard includes Environmental Quality Standards (NCAs) related to nitrates, salinity, pesticides, metals or metalloids, trichloroethylene, tetrachloroethylene, and other hazardous substances. Likewise, Directive 80/68/CEE establishes that, to guarantee effective protection of groundwater, it is necessary to prevent the discharge of List I Hazardous substances and limit the discharge of List II Hazardous substances. Along the same lines, RD 60/2011 on environmental quality standards in the field of water policy establishes NCAs for the so-called preferred, priority, and other pollutants [47].

Regarding outside European Union countries, Table 1 summarizes some information related to the reuse of treated wastewater in India, Mexico, Australia, Jordan, Singapore, South Africa, China, and Namibia. Some of them allow the reuse of treated wastewater as drinking water, as is the case of Singapore or Namibia. However, there are important differences between both countries in terms of legislation and technological infrastructure. In both cases, water is a national priority and both governments have solid strategies to promote water reuse [48,49]. Other countries are promoting water reuse in many purposes except for drinking water; this is the case of Mexico, China, Jordan, Australia, India, and South Africa. Almost all of them have solid legislation or guidelines with limits or restrictions in the use, according to quality water parameters.


#### *Water* **2020**, *12*, 1431



#### *2.3. Risks of Reclaimed Wastewater*

Wastewater discharges, even after treatment, contains alarge number of pollutants potentially dangerous to human health and ecosystems. Some of these pollutants, which are being increasingly analysed and found in reclaimed wastewater [51,52], are pathogens [53], micropollutants [54], antibiotic resistant genes [55], nanomaterials, by-products of disinfection, personal care products, and pharmaceuticals.

In terms of reuse of water, the presence of pathogen germs, such as bacteria, virus, or protozoa, represents the most severe threat for human health [56], as they are responsible for hydric waterborne diseases. However, chemical pollutants, such as medicines, heavy metals, hydrocarbons, etc. could also pose a risk to human health. The population could be exposed to these pollutants directly by consumption of polluted water, inhalation, or by skin contact, or indirectly by the consumption of food dropped or produced with polluted water effluents.

Studies are reporting the emission of different pathogens after the discharge of treated effluents. López et al. [57] and Mosteo et al. [58] reported the discharge of different pathogen bacteria, such as *E. coli*, *Enterococcus sp.*, *Pseudomonas sp.*, *Clostridium perfringens*, or *Staphylococcus aureus* in different wastewater treatment plants (WWTPs) located in Spain, with different secondary treatments, and even with, for some of them, tertiary treatment based on ponds. In the research by Mosteo et al. [58] pathogen parasites, such as *Giardia* and *Cryptosporidium* were not detected, but free-living amoebae, which are potentially pathogenic, were observed. However, other authors, such as Ajonina et al. [59], have reported the presence of *Cryptosporidium parvum* in the effluent of WWTP. As the presence of any of these pathogens would be responsible for possible waterborne diseases, the total or partial disinfection of wastewater becomes mandatory in order to be reused. Chlorination is the most common disinfectant agent because it satisfies most requirements as an ideal disinfectant. However, its use as a disinfectant to treat wastewater has the main disadvantage of generating disinfection by-products, such as trihalomethanes (THMs), among others, a consequence of the reaction between organic matter and chlorine [60]. THMs are carcinogenic, and their presence in water bodies represent a new threat for humans and ecosystems. For this reason, in the previous decade, new disinfection treatments have been developed in order to prevent the generation of these disinfection by-products, as well as to increase disinfection efficiency compared with chlorination or traditional technologies. Table 2 shows a summary of the recent studies related to the removal of pathogens in water reclamation. As can be observed in Table 2, the application of AOPs based on the generation of free radicals with a high oxidation potential has increased the efficiency of disinfection of wastewater and, hence, it has become a credible alternative to chlorination. Technologies, such as heterogeneous photocatalysis, Fenton, and photo-Fenton reaction, oxidation by persulfate salts, combination of ozone with UV-C radiation, or H2O2, etc., have been widely studied and reported as efficient in the removal of different microorganisms. Besides AOPs, there are other efficient technologies to treat wastewater, such as nanofiltration, reverse osmosis, electrochemical technologies, and coupling between different systems.


**Table 2.** Recent studies on the detection and/or removal of pathogens in treated effluents.

PS: persulfate; PMS: peroxymonosulfate; UV: Ultraviolet radiation; EDDHA: ethylenediamine-N,N- bis(2-hydroxyphenylacetic acid).

The increase of human activities poses a growth in the synthesis and consumption of new chemical substances [76]. These substances are released into the municipal sewer system via human excretion and direct/indirect disposal of medicine and chemical, and without the appropriate treatment, they reach the water bodies at trace level (ppb or ppt). In most cases, the lack of specific regulations and methodologies to detect them at trace level, as well as the largely unknown long-term effects on aquatic ecosystems and human health, making them the contaminants of emerging concern [77]. As previously mentioned, though there are various sources of contaminants of emerging concern, most of them belong to hospitals and veterinary, industrial, and domestic activities. They can be classified in different groups [78]: brominated flame retardants, chlorinated paraffins, pesticides, organofluorine compounds, personal hygiene products, pharmaceuticals, antibiotics, and drugs. These substances are characterized by high chemical stability and a low biodegradability. For this reason, conventional WWTP are not designed to remove them, the occurrence of these substances in the treated effluents being an usual affair [79]. Gros et al. [80] and Ibáñez et al. [81] reported the occurrence of medicines/antibiotics and other micropollutants in wastewater effluents in Greece and Sweden, respectively. However, further research is being done to know if the presence of these substances pose a risk if the reclaimed water is used. Some authors have reported the effects of micropollutants on aquatic ecosystems [82,83], and the World Health Organization (WHO) reported the possible impacts

from the consumption of low concentrations of pharmaceuticals present in drinking-water [84,85]. Some authors have verified the WHO conclusions with further research [86,87]. Table 3 shows recent literature references about the occurrence and reduction of micropollutants in wastewater. Between the degradation strategies, AOPs are feasible treatments, because of their capacity to degrade any kind of chemical, without forgetting their combination with other remediation treatments. Table 3 also summarizes other techniques, such as adsorption, biological treatments, filtration technologies, etc. Their inclusion, as a tertiary treatment in WWTP, will solve the threat.


**Table 3.** Recent studies on the detection and/or removal of organic micropollutants in treated effluents.

MPs: micropollutants; DCF: Diclofenac; SMX: sulfamethoxazole; CBZ: carbamazepine; ATN: atenolol; TCS: Triclosan; SCL: sucralose; BPA: Bisphenol A; IBP: Ibuprofen; ACMP: acetamiprid; CFN: caffeine; DDVP: dichlorvos; TCEP: tris (2-carboxylethyl) phosphine; ATZ: atrazine; DEET:N,N-diethyl- m-toluamide.

## *2.4. Tertiary Treatment in WWTP: Technologies for Wastewater Reclamation*

As reported in Section 2.3, the removal of the main risk associated with the reuse of reclaimed wastewater is well known, it is important to understand that there is not just a single technology to regenerate wastewater, and depending on the final quality required for the reuse, different possible configurations of treatment can be adopted. If the reclamation is considered as a tertiary stage in a WWTP, omitting the explanation of primary and secondary treatment, the regeneration line can be divided into four kinds of treatments with different efficiency in terms of microbiological pollution removal. Figure 2 shows the different configuration of tertiary treatments [107]. The most demanding treatment (treatment 1), can be composed of a coagulation–flocculation treatment followed by filtration or decantation; the whole stage has the aim of removing colloidal material and, for that, reducing organic matter concentration and turbidity. Then, it can be incorporated—a separation process based on the use of membranes (i.e., ultrafiltration (UF) or reverse osmosis (RO)). It is important to note that reverse osmosis can separate dissolved solids/salts. In this way, this treatment allows obtaining almost ultrapure water free of organic/inorganic and biological substances. Finally, although reverse osmosis can remove biological pollution, a disinfection step is required to introduce a residual amount of disinfectant to guarantee the quality of water until the point of use. Logically, the application of this treatment will have a higher investment, maintenance, and operating cost. When water with ultrapure quality is not required, the configuration of the line can be slightly modified, substituting, in this case, the UF/RO stage for a simple pre-disinfection based on UV-C radiation (treatment 2). Treatment 2 allows to obtain a high quality effluent, with very low content of organic matter, suspended solids, and turbidity thanks to the coagulation–flocculation stage, and with very high quality in terms of biological pollution, but without guaranteeing the absence of microbiological pollution. Finally, when the required quality for reuse is not high, the treatment can be simplified, up to a point, to a simple filtration and disinfection (treatment 3), or even only a filtration (treatment 4). As mentioned earlier, the reuse could be directed from the secondary effluent of the WWTP in the case where the quality of this effluent is within the permissible limits of the physico-chemical and biological parameters established in the legislation.

**Figure 2.** Proposal of wastewater reclamation line and biological risk associated with reclaimed water. Adaptation from [107].

#### **3. Resources Recovery**

#### *3.1. Nutrients*

Nutrient recovery can be seen as a loop or cycle comprising the following stages: (i) agricultural production; (ii) processing; (iii) consumption; (iv) collection; and (v) treatment of waste and wastewater. Thus, the return of the nutrients reclaimed from wastewater as mineral and organic substances to

agriculture closes the cycle. It is for this reason that municipal wastewater is an essential alternative source of nutrients [108].

The development of recovery processes for nutrient recycling is gaining increasing attention due to both economic and environmental reasons, such as improved cost-effectiveness of water treatment plants and reduced drinking water consumption [109].

The modern sanitation approach is focused on the recovery of valuable resources from wastewater by implementing energy recovery processes, maximizing water reuse, and prioritizing the recycling of nitrogen, phosphorus, and organic matter present in sewage [110]. By doing so, wastewater treatment facilities will be converted into resource recovery centres.

Different techniques can be employed for nutrient recovery, including chemical, physical, and biological processes. However, the most widely used method for the simultaneous recovery of nitrogen and phosphorus is the formation of struvite (magnesium ammonium phosphate), through crystallization/precipitation. Struvite is an excellent fertilizer due to its low solubility in water and slow rate of release. Besides, it can be applied directly to soil [111].

Struvite formation requires the presence of ammoniacal nitrogen (NH4 <sup>+</sup>) and phosphate (PO4 <sup>3</sup>−), which reacts according to reaction 1 [112]:

$$\mathrm{NH\_4^+} + \mathrm{PO\_4^{3-}} + \mathrm{Mg^{2+}} + 6\mathrm{H\_2O} \to \mathrm{MgNH\_4PO\_4} \cdot 6\mathrm{H\_2O} \tag{1}$$

It should be noted that its formation could be accelerated by the addition of magnesium (Mg2<sup>+</sup>) in excess [113].

Besides struvite, other products that can be recovered from secondary streams (generated in primary sludge thickening or after dewatering digested sludge) are ammonium sulphate or nitrate and calcium phosphate (see Table 4). To that extent, different technologies, such as Pearl®, PhosNix®, AirPrex®, PHOSPAQ® and Crystalactor® can be used [113,114]. Slow crystallization, which can be found in Ostara® DHV, Crystalactor®, and P-RoC® technologies, and instant precipitation (PRISA) are always required in liquid effluents. This allows recovery efficiencies around 85%–90% of the phosphorus contained in the supernatant [115]. Considering that heavy metals are present in low concentrations in the supernatant of the digester, it is unlikely that they produce soil contamination if the product is applied in agriculture.

The main challenge related to struvite precipitation is the recovery of phosphorus from wastewaters with low content of this element (<50 mg/L) and a high concentration of suspended solids (<2000 mg/L) [116]. For this reason, chemical precipitation of phosphorus with calcium (Ca2<sup>+</sup>) is also very common. Other reagents, such as aluminium (Al3<sup>+</sup>) and iron (Fe3<sup>+</sup>) can also be employed, although its use is not generalised because Al is toxic to most plants and Fe is strongly bound to phosphorus, making it less available [112].

There are also physical processes that can be used for nutrient recovery, such as ion exchange, adsorption, or membrane processes. Thus, Johir et al. [117] coupled a membrane bioreactor (MBR) to an ion-exchange unit, recovering the nutrients contained in the effluent of the MBR effectively. However, the operating cost of this technology is high because chemicals are needed for the regeneration of the spent material.

Membrane processes are adequate to produce a concentrate (including phosphorus, nitrogen, and potassium) that can be used in irrigation or as a raw aqueous stream for nutrient recovery by crystallization or precipitation [112]. The disadvantages associated with these processes are high-energy consumption and membrane fouling due to the accumulation of pollutants or salts, thus, being necessary to be cleaned regularly [116].

Zeolites can also be used as an adsorbent for the treatment of digestate due to their high cation-exchange capability, mainly for potassium and ammonium [118]. Natural zeolites, such as clinoptilolite, can also be employed. In this case, ammonium is adsorbed onto the mineral and when the material is spent, it can be used as a slow-release fertiliser or soil conditioner.


**Table4.**Processemployedfornutrientrecoveryfromdifferentwastewaterstreams.



#### *3.2. High Added-Value Products*

The food-manufacturing sector produces significant amounts of nutrient-rich effluents, which can be used as a resource to recover high added-value products, thus, developing more environmentally friendly processes. In this sense, during the extraction of olive oil, polyphenols are partitioned between water and oil phases, the main fraction being present in wastewater due to their high polarity and water solubility. The typical concentrations of polyphenols in olive oil wastewaters varied from 5 g/L to 25 g/L [122]. Thus, their recovery is promising since they are natural antioxidants of great interest, for food, pharmaceutical, and cosmetic industries [123]. Different technologies have been proposed for the recovery of olive polyphenols. Yangui and Abderraba [124] evaluated the use of activated carbon coated with a natural agent (milk proteins) for their recovery. An extraction yield of 80% and overall efficiencies for total phenols of 75.4% were obtained. Besides, the recovery of the most valuable compound, hydroxytyrosol, achieved a value of 90.6%. Kiai et al. [125] studied the Cloud Point Extraction (CPE) methodology to pre-concentrate phenolic compounds from olive processing wastewaters. The CPE process is a separation technique with low energy requirement, which is based on the ability of non-ionic surfactants to form micelles in aqueous solutions. Three non-ionic surfactants were analysed: (i) Triton X-100, (ii) Tween 80, and (iii) Genapol X-080. The optimum conditions were found to be a temperature of 70 ◦C, acidic pH (2.0), 30 min of equilibrium time and 10% (*w*/*v*) of surfactant. The phenolic compound recoveries achieved under these conditions were 65%, 62%, and 68% for Triton X-100, Tween 80, and Genapol X-080, respectively. Kalogerakis et al. [126] assessed the recovery of total phenols, hydroxytyrosol, and tyrosol by means of liquid–liquid solvent extraction. The effect of different organic solvents, such as diethyl ether, ethyl acetate, and a mixture of isopropyl alcohol and chloroform on efficiency recovery, was studied. The performance of the extraction agents followed this order: ethyl acetate >isopropanol/chloroform/> diethyl ether. It was calculated that around 0.25 kg of hydroxytyrosol, 0.062 kg of tyrosol, and 3.44 kg of total phenols could be attained after treating 1 m3 of olive wastewater with ethylacetate.

Wastewaters from dairy industries are also characterised by a high content in macronutrients with nutritional value, such as proteins and lipids, which can be employed as animal feedstock. For example, the Gopinatha et al. [127] study focused on the use of a low-cost polymer (lignosulphonate) to recover proteins and lipids from dairy effluent. Under optimum conditions (pH = 3.5, concentration of lignosulphonate = 0.016% (*w*/*v*), and temperature = 22 ◦C, a lipid recovery of 96% and a protein recovery (mostly caseins) of 46% was obtained.

Soybean wastewaters contain proteins and anti-nutritional factors (ANF), including the Bowman–Birk protease inhibitor, which showed anti-cancer or cancer-preventive activity [128], Kunitz trypsin inhibitor, and soybean agglutinin. Therefore, the recovery of both proteins and ANF can become an interesting alternative with potential therapeutic value. Li et al. [129] analysed the use of a quick-shearing system to obtain homogeneous electrostatic complexes between soybean whey protein and a high concentration of polysaccharide (4%, *w*/*v*). Two types of complexing agents, sulphated polysaccharides and carboxylated, were evaluated for protein recovery and trypsin inhibitor activity, paying special attention to the effect of pH. Sulphated polysaccharides were found to be the most suitable complexing agents, achieving a recovery of proteins of 90% and removal of ANF of 80% at pH 2.5.

Coal gasification wastewater is polluted with phenols, including monohydric phenols and dihydric ones, their concentrations ranging from 2900 to 3900 mg/L and 1600 to 3600 mg/L, respectively [130]. Different solvents (diisopropyl ether (DIPE) and methyl isobutyl ketone (MIBK)) can be used for phenol removal, depending on the type of phenols contained in the coal wastewater. Guo et al. [131] assessed the removal of total phenols and hydroquinone (dihydric phenol) using three different solvents: (i) methyl propyl ketone (MPK), (ii) DIPE, and (iii) MIBK. Besides, one-stage, two-stages, and three-stages of countercurrent extraction were also performed. The efficiency of removal using MPK was greater than those achieved using DIPE or MIBK. Besides, the use of three-stage extraction with MPK allows a total phenol recovery of 99.6%, thus, obtaining a final concentration of phenols

lower than 100 mg/L. A comparison between MIBK and MPK showed that the decrease in phenol emissions was as high as 52.5% by using MPK. Gui et al. [132] have developed a novel method for phenol and ammonia recovery based on an extraction process using a self-developed reagent, named as IPE. This methodology implies the following stages: (i) pre-extraction with IPE, (ii) stripping out of sour gas and ammonia, and (iii) deep extraction with IPE. Besides, the use of MIBK as an extracting agent was also evaluated for comparison purpose. The removal of total phenols and ammonia was similar with both extraction agents, the values being 90% and 97%, respectively, for IPE and 90.6% and 98%, respectively, for MIBK.

Wastewaters produced from pharmaceutical industries present a variable composition depending on the raw materials, the processes employed in the manufacturing, the season, and the location of the facility, which mainly affects the water quality. Therefore, it is complicated to define a treatment system for such a diverse industry [133]. However, pre-treatment and recovery of various valuable by-products, such as acids, heavy metals, solvents, and active pharmaceutical ingredients (APIs), which are contained in waste streams, is a control strategy of the waste, thus, making the pharmaceutical industry more sustainable. It is worth noting that the recovered waste stream can be used elsewhere in the process and the water can be employed in cooling towers or boiler feed, reducing the consumption of freshwater and, consequently, the operating costs. Taking into account that pharmaceutical wastewaters are composed of substances with molecular weights higher than 250 Da, membrane technologies can be effectively used to recover the product, provided that these are the only substances present in the effluent. In this sense, the application of nanofiltration has rapidly increased in the past decade. It has been commonly employed in aqueous solution with antibiotics [134]. For example, amoxicillin (MW = 365.4 Da), a widely used antibiotic in human and veterinary medicine to prevent respiratory, gastrointestinal, and urinary bacterial infections, can be separated and recovered by nanofiltration. This will mitigate the harm of amoxicillin to the environment as well as enhance the economy of the process.

Shahtalebi et al. [135] observed recoveries of amoxicillin as high as 97% and permeate flux of 1.5 L/min·m2, using a spiral nanofiltration membrane composed of polypiperazine amide.

Heavy metals, such as Cd and Ni, were also found to be present in pharmaceutical wastewaters, the concentration being around 35 mg/L [136]. Studies at the University of Alicante developed an electrochemical treatment for recycling and recovery metals (Ni, Zn, Pb, among others). This technique allows obtaining metals with high purity, while being less polluting than the classic pyrometallurgy, because it avoids gas emission, sulphur, and metal particles.

Table 5 summarises the high added-value product recovered, the technology employed, and the main results achieved.


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#### **4. Sewage Sludge Valorisation**

The treatment and disposal of sewage sludge is an environmental subject of prime concern due to its high organic load and low dewaterability, which implies great operational costs. Besides, it is a growing problem worldwide since the number of wastewater treatment plants (WWTPs) is rising all over the world, thus, causing a rise in sludge production. For instance, in the EU, more than 10 million tons of dry solids of sewage sludge were produced in 2015 [137]. This fact, together with the tightening of environmental quality standards, makes correct management of this waste essential. Therefore, its comprehensive management is a key point to reduce the operating costs and make wastewater treatment a more environmentally friendly process. Most countries are focused on the recycling of the sludge, for example, using it to produce energy by incineration or anaerobic digestion [138].

The traditional uses of sewage sludge include land application (with or without previous composting), landfilling or energy recovery [31]. Nevertheless, it should be noted that: (i) land application has to compete with other waste streams, the demand is variable and the legislation about this topic is becoming increasingly strict, (ii) landfilling is an unsustainable alternative due to concerns over pollution, loss of recyclable materials and loss of void for waste that cannot be recycled, and (iii) incineration is a high cost/technological option, and is currently only likely to be cost-effective for large cities. Besides, it lacks a high level of public acceptability due to concerns over gas emissions.

In light of the above, the need to develop new sludge management strategies becomes clear [139]. Nowadays, new perspectives in this topic are being opened, especially those focused on recovering valuable compounds, such as nutrients (mainly phosphorus and nitrogen), heavy metals, bioplastics, materials (adsorbents), proteins, and enzymes. In fact, sludge is now recognized as one of the future key raw materials, as stated in the Spanish Bioeconomy Strategy and in the Bio-Based Industries (BBI) European Strategy. Obviously, these new sewage sludge strategies should be designed coherently with the waste hierarchy introduced by the Waste Framework Directive (EU Directive, 2008), which proposes an order of preference for action to reduce and manage waste, i.e., prevention, minimisation, reuse, recycling, energy recovery and, as a final and undesired option, landfilling.

#### *4.1. Nutrients*

Sewage sludge is composed of significant amounts of nutrients, such as phosphorus (0.5%–0.7% of total solids) and nitrogen (2.4%–5% of total solids), in the form of protein compounds, which can be employed to produce fertilisers [140–142]. Taking into account that phosphorus is a finite resource extracted mainly from rocks located in a few regions of the world, and the high cost of commercial fertilizers, the recovery of nutrients from sewage systems is gaining attention to improve sustainability in agricultural production [112].

Chemical, physical, and biological processes can also be used to recover nutrients from sewage sludge in the same way as with aqueous effluents (see Table 6). Again, the simultaneous recovery of nitrogen and phosphorus as struvite by means of crystallization/precipitation is the main alternative [111]. Thus, it can be precipitated in the sewage sludge through acidic leaching followed by the chemical precipitation of phosphorus [143]. However, heavy metals can be present in the sludge, thus, requiring a pre-treatment to avoid its presence in the fertilizer. Chemical extraction processes, such as Gifhorn and Stuttgart, in combination with sulphidic precipitation or complexation of interfering ions and precipitation, can also be used to recover phosphorus as struvite and other phosphorus-based compounds (calcium phosphate and iron phosphide) from thickened, dewatered, or digested sludge. A high percentage of recovery from 35% to 60% can be achieved with these processes [11,115,119]. Another chemical process, named ExtraPhos, from the company Budenheim, allows the recovery of phosphorus as dicalcium phosphate, the yielding ranging from 40% to 60%. This is a four-stage technology, which implies: (i) wet chemical leaching, (ii) extraction of CO2, (iii) solid/liquid separation, and (iv) precipitation of phosphate [115,119,144].


**6.**Processemployedfornutrientrecoveryfromsewage

Oxidative processes, including Aqua Reci and LOPROX/PHOXNAM, can also be employed to recover phosphorus. In the first one, supercritical water oxidation followed by acid/alkaline leaching and precipitation were used, thus, recovering phosphorus as calcium phosphate, iron phosphide, and aluminium phosphide from thickened, dewatered, or digested sludge. In the second technology, a combination of low-pressure wet oxidation, nanofiltration, acid, and hydrothermal digestion for heavy metal separation was employed to obtain struvite and phosphoric acid from thickened sludge [11,115,121]. The percentage of recovery of these techniques varied from 40% to 60% for Aqua Reci and between 40% and 50% for LOPROX/PHOXNAM.

Another technology used for the recovery of phosphorus as P-rich slag is MEPHREC. This technique is based on a metallurgic melt-gassing, and utilizes dewatered sludge briquettes, the yield being around 65%–70% [11,115,144].

Sewage sludge after incineration (sewage sludge ash) can also be employed as a raw material for phosphorus recovery. Two methodologies are mainly employed: (i) wet chemical, and (ii) dry thermal. In the first one, an organic acid or solvent is added in order to extract phosphorus, thus, being recovered from this organic solution. In the second one, phosphorus is recovered by dissolving the ashes [120,145]. The recovery percentages are higher (from 65% to 99%) than those obtained with sewage sludge [11,115].

#### *4.2. Heavy Metals*

Various heavy metals, including Cu, Ni, Zn, Cd, Pb, Cr, and Hg are contained in activated sludge, thus, limiting its use for land application due to possible soil and groundwater contamination, which can also have a detrimental effect on human and animal health [140]. It should be noted that the concentration of these elements varies greatly depending on the origin of sewage sludge, the highest values being found for Cu (468 mg/kg) and Zn (803 mg/kg) in biosolids from WWTP of USA [146]. Thus, sludge must be treated for recovering heavy metals prior to its valorisation.

The extraction of heavy metals was analysed by different authors using various extracting agents, such as chelants: EDDS (Ethylenediamine-N,N'-disuccinic acid; biodegradable) and EDTA (Ethylenediaminetetraacetic acid; non-biodegradable), and organic acid (citric acid). It was found that Cu recoveries were approximately 70% at pH values higher than 4.5 with a molar ratio of EDDS to the total heavy metals in sewage sludge of 10:1. Similar Cu recoveries (72%) were obtained with EDTA when the same conditions were used [147]. When citric acid (0.1 M) was used at pH 3–4, Cu and Zn recoveries of 60%–70% and 90%–100% were achieved [148].

The removal of heavy metals from leaching effluents of sewage sludge using supported liquid membranes was studied by Yesil and Tugtas [149]. The highest removal efficiencies were obtained through 20% Aliquat 336-filled PVDF membrane at 35 ◦C when 1.0 M HNO3 was used as the permeate solution, the values being 27%, 22%, 30%, and 32% for Cr, Cu, Ni and Zn, respectively.

Medium volatile heavy metals, i.e., Cu and Zn, were removed from sewage sludge by means of calcination in an inert (N2) or oxidising atmosphere (air) in presence of Cl-donor (MgCl2) as an additive [150]. Thus, a Cl/sewage sludge ratio of 5% was enough to obtain a Cu removal of 80% in air and 88% in N2. For significant Zn removal (90% in both atmospheres), a higher Cl/sewage sludge ratio (15%) was needed.

Electrokinetic treatment was also studied as a potential technique for heavy metal removal [151]. It was found that the combination of this treatment with the addition of a chelating agent (tetrasodium of N, N-bis (carboxymethyl) glutamic acid), and a biodegradable biosurfactant (rhamnolipid) as electrolyte, results in a high removal efficiencies for Cu, Zn, Cr, Pb, Ni, and Mn, the values being 70.6 ± 3.41%, 82.2 ± 5.21%, 89.0 ± 3.34%, 60.0 ± 4.67%, 88.4 ± 4.43% and 70.0 ± 3.51%, respectively.

Table 7 summarises the recovered heavy metals, the technology employed, and the main results achieved.


#### *4.3. Adsorbents*

Adsorption processes are extensively studied due to its flexibility and ease of operation. Moreover, increasing attention is being given to the use of adsorbents from waste in order to develop low-cost strategies for water treatment [152,153]. In this sense, sewage sludge can be an interesting precursor for the synthesis of adsorbents due to its high content in carbonaceous matter. It should be noted that anaerobically digested sewage sludge, dewatered sewage sludge, and that from municipal/urban or WWTPs, can be used as precursors (Smith et al., 2009). Different methodologies have been proposed for adsorbent preparation, the most widely employed being: (i) carbonisation, (ii) physical activation, (iii) chemical activation, and (iv) combination of physical and chemical activation [154] (see Table 8).

Regarding carbonisation, various conditions were employed (temperatures varying from 650 ◦C to 1000 ◦C, heating rates between 5 ◦C/min and 20 ◦C/min, and time ranging from 0.5 h to 2 h), with the aim of preparing adsorbents with high BET (Brunauer–Emmett–Teller) surface areas, since this parameter significantly influences the adsorption process. Several authors reported that the increase in the aromatization, which occurred at high temperatures, is responsible for the increase in the BET area [155,156]. Nevertheless, the optimum conditions differed among the authors. In this sense, Zhai et al. [155] reported 850 ◦C as the most suitable temperature, whereas Bagreev et al. [156] and Inguanzo et al. [157] reported 950 ◦C and 650 ◦C, respectively, as the optimum temperatures. This is due to the great variability in the composition of sewage sludge. Another parameter of interest is macroporosity and mesoporosity, which can also be enhanced by raising the carbonisation temperature [158]. Moreover, the use of citric acid for sludge demineralisation also improved the BET area due to the removal of the inorganic content. In this sense, Kong et al. [159] increased the BET area from 6.8 m2/g to 385.8 m2/g when citric acid washing was applied.

Regarding physical activation, it implies the gradual burn-off of the carbonaceous matter, thus, leading to a significant increase of the inorganic content, which is mainly non-porous. Therefore, the adsorbents obtained by physical activation generally exhibit low BET surface areas. In order to improve their properties, different activation conditions were used. Various authors reported the use of steam, air, CO2, O2/N2 and air/N2 as gasifying agents, employing temperatures between 275 ◦C and 800 ◦C and times between 0.5 h and4h[154]. Among them, the best results in terms of BET area (226 m2/g) were obtained by Rio et al. [160], who employed a two-stage process, based on a carbonisation at 600 ◦C for 1 h, followed by activation with steam at 760 ◦C for 0.5 h.

Concerning chemical activation, several dehydrating compounds can be used, such as KOH, NaOH, ZnCl2, H2SO4, K2S, and H3PO4, alone, or in a combination (of two of them) in order to obtain high values of BET surfaces areas. On some occasions, carbonation, followed by chemical activation with KOH or ZnCl2, was employed [154]. Several studies reported the great effectivity of KOH as dehydrating agents in combination with carbonisation. In this sense, BET areas from 1058 to 1882 m2/g were obtained when different types of sewage sludge were carbonised at 700 ◦C for 0.5 h and, subsequently, treated with KOH in a ratio of 1:1 at 700 ◦C for 1 h [161–164].


*Water* **2020**, *12*, 1431

#### *4.4. Bioplastics*

Polyhydroxyalkanoates (PHAs) are biodegradable polymers with similar characteristics to those of petroleum-based polyolefins and can be considered a sustainable alternative to petroleum plastics. These biopolymers are naturally synthesised by bacterial fermentation of sugar or lipids. Besides, several microorganisms can accumulate PHA; however, its application is limited due to high production costs. In that sense, around 50% of the cost of microbial PHA is related to the cost of substrate or carbon source. This is why the growth of the PHA industry is low [165]. Therefore, it is necessary to find sustainable feedstock. Hence, microorganisms in activated sludge are able to accumulate PHA in the range of 0.3 to 22.7 mg biopolymer/g activated sludge [140]. It should be noted that PHA recovery using sewage sludge depends on various operating parameters. According to Takabatake et al. [166], the capability of sewage sludge to accumulate PHA was higher in the conventional process than in the aerobic–anaerobic ones, achieving an average content of PHA around 19% after the aerobic process. Thus, the addition of small amounts of oxygen in the anaerobic zone can improve the accumulation of PHA. Temperature is another factor that significantly affects the PHA production, which is favoured at low temperatures [167]. Besides, the type of reactor used for its production is also essential. It has been reported that the sequencing batch reactor (SBR) is the most adequate to obtain high production thanks to its flexible operation and simple control [168].

PHA production can be integrated during the treatment of urban water and sludge management in WWTPs [169]. This implies a process with four stages: (i) removal of readily biodegradable carbon from wastewater coupled to the selective biomass growth with capacity for PHA storage, (ii) acidogenic fermentation of sludge in order to produce a liquid stream with high content of volatile fatty acids (VFA), (iii) accumulation of PHA from the liquid stream with high VFA concentration using enriched biomass from stage 1 and, (v) recovery and characterization of PHA (Figure 3).

**Figure 3.** Schematic process flow diagram of municipal wastewater and sludge treatment together with the production of polyhydroxyalkanoates (PHA) (adapted from [168]).

Since PHAs are biodegradable compounds, it is crucial to generalize their use as packaging films or disposable products, such as bottles, cups, cutlery, and diapers, among others, in order to reduce the detrimental effect of petroleum-based plastics in the aquatic environment [170].

#### *4.5. Construction Materials*

Taking into account that the content of oxides of sewage sludge (Al2O3, CaO, SiO2, and Fe2O3) is similar to that of Portland cement or clay, it has been proposed that it can be used for the production of building or construction materials, including eco-cement, bricks, ceramic materials, supplementary cementitious materials (SCMs), or lightweight aggregates (LWAs) [171]. Nevertheless, a high amount of organic matter is also present in sewage sludge, which can affect the cementitious properties, causing low bonding strength. Therefore, pre-treatment of sewage sludge before being employed in the preparation of cement or concrete is a must.

Considering the studies of various authors [172], the valorisation of sewage sludge as raw material for the preparation of cement-like materials is viable. However, depending on the characteristics and origin of the sludge, different conditioning steps may be needed. In that sense, Xu et al. [172] and Lin et al. [173] reported the use of lime-dried sludge as a substitute of limestone in the production of cement at 1400 ◦C, the optimum amount of limestone being below 18 wt% in order to facilitate the formation of crystalline phase in the cement clinker. Tay and Show [174] found that a cement-like material, which satisfies the American Society for Testing and Materials (ASTM) requirements for mortars, could be synthesised by incineration at 1000 ◦C for 4 h using dewatered sewage sludge mixed with lime at a ratio of 1:1. Rezaee et al. [175] reported that dry sewage sludge could be used to partially replace (from 5% to 15%) the traditional raw materials in order to obtain an eco-cement with properties similar to that of Portland cement.

Besides, sewage sludge can also be used to partially replace clinker in cement due to its properties as a self-cementing agent. Thus, Valls and Vazquez [176] partially substitute Portland cement with sewage sludge in a range of 25% to 50%, obtaining different pastes with hydration products, such as those of Portland cement. However, the organic matter contained in the sludge is degraded due to the alkalinity of the cement. Thus, in order to accelerate this process, CaCl2 and Ca(OH)2 can be added as additives. According to Hamood et al. [177], sewage sludge with high moisture content can provide the required water for the preparation of mortars through its conditioning with unprocessed fly ash, thus, avoiding both dewatering and drying processes.

The partial replacement of Portland cement by raw sewage sludge in cement is not a feasible option due to the low pozzolanic activity of the sludge, which needs pre-treatment to improve its activity. Thermal treatments (incineration) have been widely used since the organic matter is decomposed, and high temperature favours the activation of inert minerals, such as kaolin, which is transformed into metakaolin [171].

For the preparation of LWAs, it is necessary to pelletize and thermally treat the raw sewage sludge. Depending on the proportion of sewage sludge, different materials can be obtained. Thus, a mixture of 10% sewage sludge and 90% clay after sintering at 1150 ◦C for 0.5 h, resulted in an LWA with lower density, higher porosity and reduced compressive strength, in comparison to the traditional one [178]. Blends of waste glass and sewage sludge can also be employed to produce LWA, obtaining higher compressive strength in LWA as the amount of glass increased [179].

#### *4.6. Proteins*

Proteins can be considered high-added-value products due to their use as liquid fertilizers, protein foaming agents, or adhesives or animal feed [180]. Based on the high proportion of proteins (up to 61%) in the activated sludge and considering that around 50% of the dry weight of bacteria cells are due to protein content, their recovery is an issue of great interest [181,182]. In this sense, different methodologies have been proposed for protein recovery, the first stage always being the solubilisation of the intracellular material in the sludge [179]. Various treatments, such as physical [183], chemical [184,185], physico–chemical [180,186,187], biochemical [184], and hydrothermal ones [188] have been reported to effectively break the floc structure of the sludge; thus, releasing the intracellular content into the aqueous phase. The physical methods rely on centrifugal, ultrasonic, and membrane processes. The chemical or physico–chemical ones include acid or alkaline treatment alone, or in combination with hydrolysis [184,189]. The biochemical methods consist of enzymatic or enzymatic assisted treatment [184,190]. The hydrothermal methods are based on solubilization using oxidizing (air or oxygen) or inert (nitrogen) atmospheres. It should be noted that after solubilisation, it is necessary to precipitate proteins from the aqueous solution, dried and, on some occasions, purify them, before obtaining the protein product. Table 9 summarises the technology employed, main results achieved, and advantages and disadvantages inherent to the methodology used. It is common that the recovery process causes the solubilisation of heavy metals contained in the sludge, together with the release of intracellular compounds (mainly proteins, carbohydrates, and humic acids). Therefore, it is needed to detoxify the sludge by means of sterilization and remove heavy metals or toxins before its valorisation to obtain proteinaceous products for animal feed to avoid compromise of animal health [191].


Processemployedforproteinfrom


**Table9.***Cont.*

#### *4.7. Hydrolytic Enzymes*

Hydrolytic enzymes, such as amylase, phosphatase, lipase, protease glucosidase, and aminopeptidase, among others, have been found to be present in activated sludge and are responsible for the biodegradation of organic matter during the aerobic digestion [192]. Besides, enzymes are widely used in different fields, including agriculture, detergents, pulp and paper, cosmetics, dairy, etc. For example, lipases hydrolyse lipids in order to obtain glycerol and fatty acids, which can be used in the production of soap [193]. Protease can be employed as a target for therapeutic agents against important diseases, such as malaria or cancer, and amylase can be employed in starch saccharification, because it can hydrolyse starch into glucose [194]. Therefore, the valorisation of sludge to produce enzymes is a promising strategy to alleviate the environmental impact of its generation. There is no standardised methodology to extract valuable enzymes from activated sludge. In that sense, different chemical and physical methods have been proposed, depending on their nature (see Table 10).

Regarding lipase, Nabarlatz et al. [195] carried out its extraction from activated sludge using stirring or ultrasonication in combination with additives: non-ionic detergent (Triton X-100), cation exchange resins (CER), or buffer. It was found that the use of additives improved enzyme recovery when stirring was used. When ultrasonication was employed, extraction time and operating conditions were key factors, whereas the addition of a detergent was less influential.

Considering protease, Frolund et al. [196] were able to extract exopeptidase using CER and Triton X-100. CER was highly effective in the extraction of the enzyme in the floc matrix. Jung et al. [197] studied the use of a disrupting chamber (Dyno mill) for the extraction of two differently activated sludge: (i) cultivated in laboratory, and (ii) from WWTP, giving special attention to the effect of temperature on the extraction. The enzyme was recovered by means of precipitation with ammonium sulphate. Proteases activity varied from 75.8 ± 10.2 U/mg to 116.1 ± 4.2 U/mg at 75 ◦C as a function of the origin of the sludge. Gessesse et al. [198] analysed different methodologies: (i) Triton X-100, (ii) CER, (iii) ultrasonication in buffer, and (iv) ultrasonication with Triton X-100, obtaining the highest activity in this case. Therefore, it seemed that the combination of ultrasounds with additives is the ideal methodology for protease extraction.

In the case of amylases, it was recovered by ultrasonic CER and an additive. If CER is employed, around 2000 U/g VSS (volatile suspended solids) can be recovered, whereas, with Triton X-100, its concentration is a crucial factor, since 1600 U/g VSS were obtained with 0.1% Triton X-100 and 1300 with 2%Triton X-100 [199].



#### **5. Towards Energy Self-Su**ffi**ciency**

A wastewater treatment plant (WWTP) consumes from 0.45 to 1.25 kWh/m3, depending on its size and the treatments taking place in it. This energy is usually supplied to the plant externally, from conventional energy supply stations [200,201]. However, according to Gude [202], municipal wastewater (WW) has a total amount of energy of up to 9.7 kWh/m3. Assuming that the average energy consumption in WWTPs is 0.85 kWh/m3, the water can contain up to 12 times more energy than what is needed for its treatment [202]. This energy contained in wastewater can be divided into chemical (from the organic load present in the water), thermal, and potential energy, and it can be recovered from the water through different technologies (Table 11).

In addition, energy can be recovered from the sewage sludge. The sludge generated during the treatment processes contains about 60% of the energy initially present in WW, so it is frequently used for energy recovery [203].

Currently, energy recovery technologies are used, primarily, in large-scale treatment plants, biogas recovery being the most widely used one [204].



#### *5.1. Biogas Recovery*

Anaerobic digestion is one of the most widespread sludge stabilization processes in the current market. This process is capable of transforming the organic load present in the sewage sludge into biogas, which is a mixture of methane (50%–70%), carbon dioxide (30%–50%), and traces of other gases, such as nitrogen or hydrogen [210]. This process is usually carried out in an anaerobic environment, although aerobic digestion is also possible.

Methane, which is the significant component of biogas, is considered a greenhouse gas. Sending the sludge produced in a WWTP directly to a landfill would result in the release of this gas into the atmosphere by natural routes [140]. Moreover, thanks to its calorific value, biogas can be used for electricity generation, heat production, and as a fuel for vehicles among other uses. Due to that, it seems logical to implement anaerobic digesters that facilitate its collection which is highly beneficial [211].

The cost of electricity in a treatment plant is about 80% of the total operating cost. Even if the biogas can be sold, it is interesting to use the methane produced in the digesters in the treatment plant itself, to feed the gas engines and produce both electrical and thermal energy. In some cases, it can become one of the primary sources of energy in the WWTP [140].

The conversion of organic matter into biogas is a complex process that involves four different phases of biochemical reactions: hydrolysis, acidogenesis, acetogenesis, and methanogenesis. In case of the latter, the products of all the previous phases are transformed into the final products: methane and carbon dioxide [212].

Currently, anaerobic digestion is combined with other techniques to promote the generation of biogas and/or increasing the percentage of methane in it [203]. Within this framework, the co-digestion of food waste and wastewater sludge has been proposed. Some studies on this technique carried out through the concept of circular energy indicate that this technique is a right solution, generating significant energy savings in the overall operating cost of the plant [213].

However, the use of pre-treatments that increase methane production is most common. In fact, the objective of such pre-treatments could be described as: (1) to increase the digestibility of organic matter; (2) increase the rate of hydrolysis; (3) increase the percentage of methane in the mixture; (4) reduce sludge viscosity and reduce pumping costs [212].

A large variety of pre-treatments that can be applied, but all of them can be encompassed in one of the following categories:


Among these, the most commonly applied are physical and thermal pre-treatments. [219] found out that methane production was increased by 21%–31% when using a heat pre-treatment. Moreover, by means of thermal pre-treatment, Liu et al. [220] achieved an increase in methane production of 51%. Moreover, by combining the use of microwaves (600 W) and hydrogen peroxide (0.2 g/g total solids), methane production increased by 20%, while reducing fluid viscosity [220].

#### *5.2. Biodiesel Production*

The growing scarcity of fossil fuel reserves, together with the environmental benefits of biodiesel (such as the decrease in SOx and CO2 emissions), is the reason why the interest in this biofuel has grown significantly in recent years [140].

Commonly, vegetable oils are used for biodiesel (BD) production. However, the cultivation of plant species with the only purpose of producing BD has raised awareness about the competition of these crops with the food industry [205]. In addition, the cost of BD production is high (mainly due to the cost of raw materials), which does not make it competitive with petroleum-derived diesel.

Therefore, finding a non-edible and low-cost raw material is necessary to increase biodiesel competitiveness. Among the available options, sewage sludge, which might seem as industrial waste, increasingly receives more attention. Its high lipid content, the large quantities generated in water treatment plants, and its low cost, make it an ideal raw material for this purpose [221].

This biofuel is produced by transesterification of refined triglycerides and methanol (MeOH). This reaction usually takes place in the presence of a homogeneous catalyst, either acidic or basic (e.g., H2SO4, KOH or NaOH) [222]. The higher the oil content of the sewage sludge, the more BD it can produce; hence, it is recommended to select microorganisms for the treatments in the WWTP based on their ability to produce oil [211].

Depending on the origin of the wastewater, the sludge will have a different lipidic composition. Therefore, to know the specific treatments that will work best in each case, it is necessary to analyse the lipid content every time [223]. Finally, a high proportion of free fatty acids (FFAs) (which can account for up to 70% of the lipid content) causes problems during transesterification due to soap formation when conventional basic catalysts, such as NaOH, are used. Thus, acid catalysts are normally preferred in these cases [224]. However, by using these catalysts, the reactions slow down considerably. Although some studies that focus on the synthesis of biodiesel through non-catalytic transesterification [222], the vast majority of researchers strive to find new catalysts that allow for high efficiency in the production of biodiesel in a short time [225].

#### *5.3. Hydrogen and Syngas Production*

Along with biogas, biohydrogen is one of the biofuels that can be recovered from wastewater, being a great alternative to conventional fuels, as its energy content is 2.75 times higher than that of hydrocarbon fuel [140].

Hydrogen can be considered the most environmentally friendly fuel since its combustion only generates water. Although it is the most abundant element in nature, it cannot be found in molecular form (H2) due to its high reactivity, so it must be generated industrially [226].

Despite its great properties, very few processes can produce only the hydrogen. On the contrary, most recovery techniques produce syngas, which is a gaseous blend in which H2 and CO are the major constituents [206]. There are currently various techniques for obtaining syngas, gasification and pyrolysis being the most widely used ones [227]. In general, due to the high water content in the sludge (80%), drying pre-treatments are required before these treatments, which considerably increases the cost of this process and involves an energy input that is counterproductive, given the objective of energy recovery [228]. To avoid this previous step, many investigations are currently focused on using the supercritical water gasification (SCWG) process. Water, when in extreme conditions beyond the critical point (Tc ≥ 374 ◦C, Pc ≥ 22.1 MPa), has unique properties, which makes it an excellent non-polar solvent that can dissolve various organic substances present in the sludge, thus, increasing the effectiveness of hydrogen production. Achieving optimal operating conditions may be quite costly, so the use of catalysts is a key factor in reducing the activation energy of the reactions. The catalysts could also increase the effectiveness of gasification by being H2 selective [226].

Other processes that can be used to generate H2 from wastewater, although not so widespread, are electrohydrolysis [229] and biological treatments, such as microbial fermentation under dark or photo fermentation [230].

As previously discussed, the most common way to obtain biohydrogen is through the production of syngas, which implies that, nowadays, most of the efforts in this field are concentrated in obtaining a gaseous mixture with the highest proportion of hydrogen possible. Sludge pre-treatment has been one of the most commonly used methods for this purpose. Wang et al. [231] employed ultraviolet radiation as pre-treatment, and an 80.6% increase in the amount of hydrogen as compared to the untreated sludge has been observed. Ultrasound and heat have also been tested for this purpose, obtaining excellent results. Elbeshbishy et al. [232] applied an ultrasonic pre-treatment, achieving a 120% increase in hydrogen production.

To enhance the efficiency of gasification, it has been proven that including an oxidizing agent, such as a vapour stream, increases hydrogen production [233]. In addition, using catalysts has also shown high efficiency in syngas production. Zhang et al. [234] tested the impacts of Fe/Ca-based conditioners on syngas generation and have found that the final H2 yield was improved by 51.2% when iron was used, and 76.5% for CaO, catalysed process at 1273 K, compared to catalyst-free treatment.

Catalysts can also be used in combination with the SCWG process. Hantoko et al. [235] showed that activated carbon significantly improves syngas production. Additionally, Yan et al. (2019) studied the effect of the addition of Ni/Al2O3 catalysts. In this case, both the overall efficiency of the reactions and the proportion of H2 in the final mixture improved substantially, as the later increased by 146% compared to the experiment without catalyst [236].

Other processes, such as the use of adsorbents to retain CO2 and promote hydrogen generation, are also being investigated with considerable success. [233]. Although appreciable advances have been noticed in the recent years, it is still necessary to work hard to obtain a high-quality biohydrogen that can be used as fuel, which indeed would be a great alternative to currently used fossil fuels.

#### *5.4. Microbial Fuel Cell*

In addition to the production of biofuels and biogas, it is possible to recover wastewater energy from organic matter directly as electricity thanks to bioelectrochemical systems [202,237]. These processes, directly applicable both on water and sewage sludge, lead not only to electricity generation, but also to the removal of contaminants [238–242]. Among these processes, several authors have highlighted Microbial fuel Cell (MCF) as a potential solution for problems related to water and energy [243].

MFCs are composed of an anaerobic chamber (known as anode) and an aerobic chamber or cathode. Both chambers have an electrode and are separated by a membrane known as the cation exchange membrane (CEM). CEM has the function of regulating the passage of protons from one chamber to the other. The substrate, rich in organic components, is introduced into the anode and oxidized by microorganisms, leading to the generation of electrons and protons. The electrons generated in the process are transferred to the anode and, subsequently, travel to the cathode through an external circuit generating bioelectricity. Simultaneously the protons move from one chamber to the other through a polymer electrolyte membrane. Finally, electrons and protons combine with oxygen in the cationic chamber to produce a water molecule [244]. The voltage generated in an MFC is of the order of several hundred of millivolts. When the fraction of electron charge that contributes to electricity generation is 40% and the hydraulic retention times is 20 h, the potential of energy that can be recovered from wastewater by MFC can reach 0.65 kWh/m3 [211].

The efficiency of this process depends on many different parameters, such as pH, temperature, the type of substrate that is degraded, and its quantity, salinity, type of microorganisms, electrode material, etc. [140]. Among these factors, the selection of microorganisms is a fundamental design factor in CFM. More than 35 different species have been reported as exoelectrogen in MFCs, including *Escherichia*, *Pseudomonas*, *Geobacter*, *Clostridum*, *Shewanela*, and *Lactococcus lactis* [245]. Nowadays, many efforts are focused on obtaining a biofilm with the optimal conditions of conductivity and penetrability through the genetic alteration of microorganisms.

Despite having many strengths, this technology has some important disadvantages that must be addressed to generalize its adoption in the real life. The main drawback of the MFCs is their investment cost, about 30 times higher than that of other sludge treatments, as well as the internal resistance that restricts power generation [207]. This high cost is mainly due to the high price of the materials used to manufacture the electrodes and membranes. Another drawback is that, although it can operate in a wide range of temperatures, very low temperatures imply a significant decrease in its efficiency since the metabolic rate of the microorganisms decreases [213,246]. Finally, the electricity generated by this system is not capable of supplying enough energy to guarantee the continuous operation of

an electrical instrument. This issue, however, can be solved by linking MFCs to increase voltage or current generation [247].

To overcome the disadvantages, research in this field is focused on genetic modifications of microorganisms and development of electrodes and membranes, by aiming at better and cheaper materials. As an example, Zhang et al. [18] have demonstrated the efficiency of the cylindrical cell, instead of cubic ones, with a larger cathode area, increasing power density by 52% while reducing internal resistance by 60.9%.

#### *5.5. Heat Pumps: Thermal Energy Recovery*

The energy extracted by a heat pump can come from different sources, such as air, water, or earth. In recent decades, wastewater has been positioned as a strong candidate for the source of thermal energy in heat pumps, although the available heat is described as low-quality heat. In this case, the system is called wastewater source heat pump (WWSHP) [248]. Currently, more than 500 WWSHP are used around the world. The amount of energy that can be obtained in this way is much greater than that obtained from chemical energy [249].

A WWSHP is composed of a heat/cold distribution system, the main heat pump cycle and a heat exchanger. A basic scheme of the heat pump cycle, which consists of a compressor, a condenser, an expansion device, an evaporator, and auxiliary equipment. The heat is extracted from the wastewater through the evaporator and transferred to the medium, which is heated through the condenser [249].

The place where the recovered heat is used must not be more than 3–5 km away [248]. Thus, the best option is onsite consumption of the recovered energy. It can be used to heat digesters or even for sludge drying. Moreover, it can be used outside to heat nearby buildings or greenhouses [250]. Another factor that significantly affects the efficiency of the process is the variation in the flow and temperature of the effluent since, currently, the most widespread heat pumps have a single operating speed. To overcome this problem, Chae and Ren have created a fixed inverter hybrid heat pump that has managed to increase energy efficiency by more than 15% by adjusting the operating speed to match the flow rate [251].

Finally, the most important problem while applying this technology is the fouling of the surface of the exchanger due to the composition of the wastewater, which is primarily responsible for lowering the efficiency of the equipment considerably. Therefore, the outlet water is the most viable option for energy recovery, since it is relatively clean. Despite working only with treated water, fouling is still significant, so most of the efforts are focused on developing materials and configurations that prevent or delay the fouling of heat exchangers [252,253].

#### *5.6. Hydropower*

The kinetic energy contained in the wastewater can also be recovered in its path along with a WWTP. This is possible thanks to the installation of hydropower stations in the system. Their main elements are a turbine placed perpendicular to the water flow and an associated energy generator, to transform mechanical energy into electrical power. Due to the lower water flow available in a WWTP compared to rivers or waterfalls, the most appropriate systems would be mini and micro-hydropower (MHP) [254].

The generation of hydraulic energy has several advantages that must be considered. First, it is an environmentally friendly option, since it is a renewable energy source that has no associated greenhouse effect emissions of any kind. Additionally, compared to other renewable energies, such as solar or wind, it stands out because it can generate non-stop energy throughout the year, regardless of weather conditions. On the other hand, it is considered a cost-effective option since it allows to take advantage of the existing infrastructure for its installation, which reduces the dimensions of the necessary civil works [255]. Finally, the generation of hydraulic energy can be adjusted to the WWTP demand curve since consumption peaks correspond to the times when the highest water flow is collected and vice versa [209].

Despite the good prospects, the number of WWTP plants that have hydraulic power generators is quite small, due to the number of suspended solids contained in WW, which can damage the mechanism [209]. This problem can be minimized by locating the power generation system at the exit of the plant, where the water is cleaner [254].

However, the main issue faced by this technology is that most of the already installed MHP have very little flexibility to face significant flow drops, thus, hampering operation under the severe diurnal and seasonal flow variations typical of many WWTPs. To deal with this problem, new turbine models have been developed, as reported by [256], who have studied the behaviour of a variable flow turbine throughout a year. This system is capable of working in a flow range that goes from 53% to 123% of the design one, which allowed to take advantage of almost 96% of the effluent generated in 1 year, which yielded more than 68.1 MWh.

A study conducted on 100 plants in Ireland and the United Kingdom concluded that only larger and high-flow plants are viable for the installation of a hydraulic energy recovery system [257].

#### *5.7. Real Examples of Self-Su*ffi*cient WWTP*

All of the processes described so far can help to achieve energy self-sufficiency in a water treatment plant. Although most plants that apply these technologies do so only to provide a percentage of the total energy needed to carry out the treatments, it is possible to achieve energy self-sufficiency and even positive energy balance (especially in medium/large-sized plants). Table 12 shows some WWTP examples with these characteristics.

Nowadays, plants that have reached (or are close to achieving) energy self-sufficiency mainly rely on the production of biogas from anaerobic digestion [258].

Shen et al. [210], have performed an analysis of the current landscape by detecting 12 treatment plants in Europe and the United States that have achieved energy self-sufficiency. Among those, two plants located in Austria stand out as they exhibit a positive energy balance, which allows them to sell part of the energy produced. The WWTPs of Strass TP and Wolfgangsee-Ischl, respectively, generate 6.3% and 7% more energy than they use [259].

East Bay Municipal Utility District (EBMUD) wastewater facility (located in Oakland, CA) became the first energy-neutral WWTP in North America in 2012. Nowadays, EBMUD is producing electricity to meet 126% of the WWTP's electric power demand and the surplus electricity is supplied to the grid. Moreover, in the USA, the Sheboygan Wastewater Treatment Plant is nearly 100% energy self-sufficient thanks to its co-digestion program and cogeneration capacity [258].


**Table 12.** Energy self-sufficient WWTPs. Adapted from (Gu et al., 2017).

More surprising is the case of the Marselisborg WWTP in Denmark, since it generates more than 150% of its energy requirements, so it can also be considered as a power plant. This plant, with a treatment capacity of 12 million cubic meters per year, invested € 3M to optimize its treatment process and improve biogas production [204].

#### **6. Outlook and Concluding Remarks**

The main purpose of this review is to discuss the different opportunities and challenges that are prevalent to apply circular economy in the water sector. The consumption model of modern society is the first step towards avoiding the exertion of pressure over natural resources, and this change can be applied in all of the industrial activities. A circular economy has been applied to water, for some decades, when wastewater has been reclaimed and reused. However, this is always done by countries with hydric stress in order to increase the water resources. In this regard, the application of reclamation wastewater facilities, even in regions or countries without water scarcity, is recommended. This is mainly due to the positive environmental impacts of reclamation, which also minimizes the capture of water from natural sources. Despite the benefits of reclamation, possible associated risks might appear as a consequence of the presence of pathogen germs, emerging micropollutants, and antibiotic resistance genes. The control of these risks will ensure the increased use of reclaimed water. There are some other consolidated applications of a circular economy in the water sector, such as nutrients and energy recovery. Both strategies are widely implemented in new generation treatment plants, although research is mainly focused on the complete recovery of nutrients and the accomplishment of energy self-sufficiency. Despite the fact that the above-mentioned aspects are not only opportunities, but also realities in the water sector, future wastewater treatment plants should be designed as biorefineries, which do not only reclaim water as the product, but also facilitate valorisation and recovery of high-added value substances. Hence, society must change its view on wastewater, from being a waste effluent to a "stream rich of different valuable substances" with a tremendous economic impact and value. Moreover, sewage sludge is also potentially recoverable, by the extraction/recovery of substances or the generation of new products, such as adsorbents, bioplastics, or construction materials. In all of the cases, the recovery of substances or valorisation of the effluents help to decrease the pressure over natural resources. Nevertheless, efforts should be made to evolve technologies that would be able to bring down the high costs linked with these processes. However, a change of mindset, together with appropriate policies, should help achieve a fully circular economy in the water sector.

Finally, to conclude, Table 13 sums up the current situation of the wastewater sector regarding the implementation of a circular economy. With the help of an analysis of strengths, weaknesses, opportunities and threats (SWOT analysis), some challenges and opportunities have been identified. The strengths of each available technology, which have been discussed in the review, consist mainly in the possibility of obtaining benefits from something that has always been considered a waste, while decreasing pressure over natural resources. Even so, some weaknesses must be addressed, such as the considerable initial financial investment and operational costs, or the lack of adaptability of some treatments to different working conditions. However, the constant and quick development of new technologies driven by the global trend to promote greener processes is expected to help to overcome these deficiencies. Finally, it must not be forgotten that, for a circular economy to become a reality, a change of mindset is needed, together with the commitment of governments and private companies. **Table 13.** SWOT analysis of the wastewater sector regarding the implementation of circular economy.


**Author Contributions:** S.G.-R.: Writing—review & editing, Conceptualization. P.O.: Writing—review & editing, Conceptualization, Validation. E.R.: Conceptualization, Validation and Supervision. D.N.S.: Review & editing, Conceptualization. J.R.-C.: Writing—review & editing, Supervision, Project administration, Conceptualization, Formal analysis. All authors have read and agreed to the published version of the manuscript.

**Funding:** Universidad Politécnica de Madrid through the seed research project with Asian Institutions (ASIA-18-Z77QCN-2-CWL7CN).

**Acknowledgments:** Authors acknowledge the financial support from the Universidad Politécnica de Madrid through the seed research project with Asian Institutions (ASIA-18-Z77QCN-2-CWL7CN). Jorge Rodríguez-Chueca acknowledges Universidad Politécnica de Madrid through the research project APOYO-JOVENES-X5PKL6-88-KZ46KU. Likewise, S. Guerra-Rodríguez also acknowledges UPM for support through the pre-doctoral contract of the "Programa Propio".

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

#### **References**


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

#### *Review*

## **Use of Ultrasound as an Advanced Oxidation Process for the Degradation of Emerging Pollutants in Water**

## **Ana L. Camargo-Perea 1, Ainhoa Rubio-Clemente 2,3,\* and Gustavo A. Peñuela <sup>2</sup>**


Received: 4 February 2020; Accepted: 11 March 2020; Published: 9 April 2020

**Abstract:** Emerging pollutants are compounds of increased environmental importance and, as such there is interest among researchers in the evaluation of their presence, continuity and elimination in different environmental matrices. The present work reviews the available scientific data on the degradation of emerging pollutants, mainly pharmaceuticals, through ultrasound, as an advanced oxidation process (AOP). This study analyzes the influence of several parameters, such as the nature of the pollutant, the ultrasonic frequency, the electrical power, the pH, the constituents of the matrix and the temperature of the solution on the efficiency of this AOP through researches previously reported in the literature. Additionally, it informs on the application of the referred process alone and/or in combination with other AOPs focusing on the treatment of domestic and industrial wastewaters containing emerging pollutants, mainly pharmaceuticals, as well as on the economic costs associated with and the future perspectives that make ultrasound a possible candidate to solve the problem of water pollution by these emerging pollutants.

**Keywords:** emerging pollutants; advanced oxidation process; water pollution; ultrasound

## **1. Introduction**

Emerging contaminants (ECs) are chemical products, both natural and synthetic ones, that comprise a wide range of chemical compounds, including medical and recreational drugs, personal care products, steroids, hormones, surfactants, perfluorinated compounds, flame retardants, dyes, plasticizers and industrial additives [1–3]. The presence of ECs in the environment was not measured or controlled in the past because they did not cause concern and, in general terms, there were no studies demonstrating a health risk to humankind and living beings. Additionally, the use of ECs was not as high as it is currently; and they were not detected in water, since advances in instrumental analytical chemistry have only recently permitted their quantification at ultra-trace and trace concentrations [4,5], i.e., at concentrations from ng L−<sup>1</sup> to μg L−<sup>1</sup> [3,6]. Indeed, in the last years, ECs have been identified and quantified in effluents from wastewater treatment plant effluents, surface water, groundwater and even drinking water [3,5,7,8].

It is important to note that ECs can have harmful effects both on the environment where they are located and on human health. Nowadays, the toxicity ascribed to the presence of these pollutants on the environment has not been fully evaluated [8]; nevertheless, more and more eco-toxicological studies are being conducted [9]. In fact, the presence of ECs has been reported to represent a serious risk to both the environment and human health due to direct and/or indirect exposure [3,10], since they can negatively influence algae, invertebrates and fish, as well as ecosystem dynamics and community structure [11,12]. It has been found that ECs can act as endocrine disruptors and alter the reproduction cycles, water transport and osmoregulation processes of biota [13,14]. Other emerging pollutants have antimicrobial activity, leading to bacteria resistance to commonly used antibiotics [5] and, subsequently, resulting in worldwide spread of diseases. Additionally, ECs can be bioaccumulated [8], changing cellular reactions in vital organs, such as liver, kidney and gills [15]. Other studies have reported gene expression changes in organisms exposed to ECs [16].

It has been proven that some ECs are persistent pollutants that are hardly degraded by conventional processes [8,17,18]. For this reason, the implementation of new technologies to guarantee their removal is proposed [3,7,17,19].

Advanced Oxidation Processes (AOPs) have been evaluated as an option for the degradation of a variety of organic pollutants in waters [1,20]. These processes are characterized by a wide number of radical reactions, most of which involve chemical agents along with a source of ultraviolet (UV) radiation [21]. These radicals attack a large number of recalcitrant organic compounds such as ECs and, since they are not very selective, they become an excellent precursor to the conversion of a wide range of pollutants.

Several works have been carried out assisted by AOPs in order to evaluate their efficiency in degrading CEs. AOPs consist of the formation of the free hydroxyl radicals (HO•), which are capable of oxidizing toxic and/or recalcitrant organic compounds into more biodegradable and less dangerous products, such as oxidized species and short chain hydrocarbons of low molecular weight like formaldehyde and aliphatic acids [22], among other innocuous products; thus, they provide an improvement to the treatability of AOP effluents [17]. In fact, photocatalytic degradation has been conducted in the presence of UV radiation and photosensitizers including TiO2, H2O2 and persulfate, among other chemical agents, obtaining very positive results [23–25]. Likewise, photo-Fenton and ozonation at basic pH have been proven to be highly efficient in the degradation of this type of pollutants [26,27]. These advanced systems, therefore, offer a solution to the problem of EC environmental accumulation and resistance to biological degradation, in contrast to other processes, such as conventional physical or chemical processes [17,25].

Nevertheless, it should be noted that, among the different AOPs used in the treatment of ECs present in water, the use of ultrasound (US) has been reported to be a highly efficient process, not only in the removal of this kind of contaminants, but also in their degradation [28,29] and the conversion of other recalcitrant pollutants [26] and microbial load [30] in water. Likewise, the use of US, as an advanced oxidation process, is environmentally "clean" since it does not require the addition of chemicals to the aqueous medium in order to achieve its EC degradation target, and does not generate waste [31] like Fenton and photo-Fenton. Consequently, the use of US waves is an alternative option for the conversion of recalcitrant ECs.

Considering the above, this article reviews and discusses the contributions of researches on the degradation of ECs, especially pharmaceuticals, due to their potential risks to human and other living beings, in aqueous media through US, as an advanced oxidation technology, considering the presence or absence of catalysts or dissolved gases, among other parameters, influencing the efficiency of the aforementioned process. Additionally, the application of this process is described focusing on domestic and industrial wastewater containing ECs, as well as the economic cost estimation associated with the future perspectives related to its implementation alone or in combination with other AOPs.

#### **2. Ultrasound Process**

The US process has been reported as a very efficient AOP for the degradation of ECs present in water [32–36]. Additionally, it can overcome the limitations ascribed to the use of other AOPs commonly used for water treatment. It is noteworthy to mention that, by using the US process, mass transfer within the reaction medium is improved, as well as the EC degradation reaction rates. Additionally, the consumption of chemicals, such as oxidizing and catalyzing agents, is reduced and no sludge is generated [22,37].

As part of this review, the state-of-the-art of the implementation of US, as an advanced oxidation process, is analyzed based on several works reported in the literature. First of all, the fundamentals of the process are described to continue with the factors mainly influencing the efficiency of the process. Afterwards, a number of examples are provided in order to have a general idea of the versatility of the advanced oxidation technology alone and/or in combination with other AOPs to efficiently degrade persistent compounds such as ECs. Finally, the capital and operation and maintenance costs are mentioned, and the future perspectives related to the application of the process are highlighted.

#### *2.1. Operation Fundamentals*

Aqueous medium sonolysis involves the production of waves through sound at a specific frequency, with compression and expansion cycles, leading to the formation of cavitation bubbles. These bubbles grow by the diffusion of vapor or gas from the liquid medium, reaching an unstable size that provokes their violent implosion, which in turn generates very high temperatures and pressures, approximately 4200 degrees K and 975 bar, producing the so-called "hot spots" that allow the decomposition of the water molecule to generate HO• [9], which is capable of oxidizing recalcitrant pollutants such as ECs with its high oxidation potential (2.8 V) [38], leading to the degradation of the toxic compounds and producing innocuous products, such as H2O, carbon dioxide (CO2) and inorganic ions.

Equations (1)–(4) show the decomposition of water and other molecules commonly dissolved in water by sonochemical waves [9,39], being the HO•, as well as the hydroperoxyl radicals (HO2•), the main species that oxidizes the organic compounds present in the aqueous medium.

$$\mathrm{H}\_{2}\mathrm{O} \xrightarrow{\mathrm{(l)}\mathrm{)}} \mathrm{H}^{\bullet} + \mathrm{HO}^{\bullet} \tag{1}$$

$$\text{Co}\_2 \xrightarrow{()()} 2\text{O}^\bullet \tag{2}$$

$$\text{N}\_2 \xrightarrow{()()} 2\text{N}^\* \tag{3}$$

$$\rm{^{1}H^{\bullet}} + \rm{O\_{2}} \xrightarrow{\rm{\text{()}\backslash}} \rm{HO\_{2}^{\bullet}} \tag{4}$$

The cavitation bubbles are produced in two ways, symmetrically and asymmetrically. The difference between these is the support provided by a rigid surface (for instance, the surface of the reactor) for the bubbles to be formed. This difference has a direct influence on the way in which the bubbles implode, and thus on the release of pressure and temperature into the medium, resulting in the rupture of the water molecule and the formation of HO• [9]. The symmetrical bubbles release energy in all directions around their surface, while the asymmetrical ones generate an eruption of the liquid, mainly on the parts of the bubbles that are far away from the surfaces, forming long-range "micro-jets" that go to the solid surfaces [28].

There are three reaction zones in the solution during the ultrasonic treatment process: (a) inside the cavitation bubble, (b) the bubble/water interface and (c) within the bulk solution [28,32,40]. In each of these zones, different reactions occur that favor the decomposition of pollutants. Hydrophobic, non-polar and/or volatile compounds react inside the cavitation bubbles and at the bubble/water interface, while hydrophilic and/or non-volatile pollutants react within the bulk solution [28,41–43].

Inside the cavitation bubbles, the reaction of the pollutant can occur in two ways: pyrolysis of the highly volatile compounds, or chemical reaction with the free HO• formed. At the bubble/water interface, the reaction occurs by pyrolysis and, fundamentally, by a reaction with the HO• that are formed from implosion and tend to diffuse throughout the solution medium, reacting with the compounds that are present at the interface. Within the solution, decomposition occurs only by reaction with HO•, which are released into the aqueous medium through implosion of the cavitation bubbles [9].

*Water* **2020**, *12*, 1068

When free radicals reach the aqueous solution, they can recombine, as expressed in Equations (5)–(7), or react with hydroxyl ions (HO−) (Equation (8)), resulting in a decrease of the system oxidation potential.

$$\rm HO\_2^\bullet + HO\_2^\bullet \to H\_2O\_2 + O\_2 + O\_2 \text{(a}^1\text{\AA g}\text{)}\quad k = 8.3 \times 10^5 \tag{5}$$

$$\rm HO^{\bullet} + HO^{\bullet} \rightarrow \rm H\_2O + 1/2 \left(O\_2 + O\_2 \text{(a}^1 \text{ g)}\right) \quad k = 5.5 \times 10^9 \tag{6}$$

$$\rm HO^{\bullet} + \rm HO\_{2}^{\bullet} \rightarrow \rm H\_{2}O + O\_{2} + O\_{2} \text{(a}^{1}\text{ }\Delta\text{g)} \quad k = 7.1 \times 10^{9} \tag{7}$$

$$\rm HO\_2^\bullet + HO^- \rightarrow O\_2^\bullet + H\_2O \quad k = 10^{10} \tag{8}$$

However, from Equation (8), superoxide radicals (O2•−) are formed, as well as from the decomposition of HO2•, as described by Equation (9), which also contribute to the degradation of emerging organic compounds, although in a smaller proportion than by HO• [38]. Additionally, in acidic medium, O2•<sup>−</sup> can react with protons (H<sup>+</sup>) to form HO2• (Equation (10)). Both of the free radicals can recombine, as represented in Equation (11), resulting in the production of HO2 −, which in turn can be involved in HO• quenching (Equation (12)).

$$\rm{HO}\_2^\bullet \to \rm{H}^+ + \rm{O}\_2^\bullet \quad k = 7.5 \times 10^6 \tag{9}$$

$$\text{H}^+ + \text{O}\_2^\bullet \rightarrow \text{HO}\_2^\bullet \quad \text{k} = 5.1 \times 10^{10} \tag{10}$$

$$\rm{HO}\_2^\bullet + \rm{O}\_2^\bullet \rightarrow \rm{HO}\_2^- + \rm{O}\_2 \quad k = 9.7 \times 10^7 \tag{11}$$

$$\rm HO^{\bullet} + \rm HO\_{2}^{-} \rightarrow \rm HO\_{2}^{\bullet} + \rm HO^{-} \quad k = 7.5 \times 10^{9} \tag{12}$$

Hydrogen peroxide (H2O2) can also be formed in the US process, as described in Equation (5). In spite of the fact that H2O2 can scavenge HO• or be decomposed (Equations (13)–(15), respectively), it can be involved in the oxidation of ECs, as well as on the production of a higher amount of HO•, when US process is combined with UV radiation.

$$\text{HO}^{\bullet} + \text{H}\_2\text{O}\_2 \rightarrow \text{H}\_2\text{O} + \text{HO}\_2^{\bullet} \quad \text{k} = 3 \times 10^7 \tag{13}$$

$$\rm H\_2O\_2 \rightarrow HO\_2^- + H^+ \quad k = 2 \times 10^{-2} \tag{14}$$

$$\rm HO\_2^- + H^+ \to H\_2O\_2 \quad k = 10^{10} \tag{15}$$

The reaction rate constants for the reactions expressed in Equations (5)–(15) were taken from Pavlovna et al. [44], demonstrating that, in general terms and according to the values of the reaction rate constants, the free radicals are easily formed through the US waves. As mentioned previously, these free radicals can react with the target pollutant; however, they can also recombine or be quenched by other compounds found in water such as the natural constituents of the matrix, making the reaction of the hydrophilic compounds within the solution less efficient and slower [45]. In this regard, in order to avoid side reactions of the US oxidation system, the optimization of the operating parameters or factors influencing the most the oxidation potential of the system must be conducted. This would subsequently allow the reduction of the economic costs associated with the studied advanced oxidation process for a more efficient degradation of the ECs of interest.

#### *2.2. E*ffi*ciency of the Ultrasound Process*

The US process must consider the control and variation of the different operating parameters, including the ultrasonic frequency, the electrical power and the pH and temperature of the solution [46,47], in order to be optimized with the subsequent reduction in the costs associated with the process performance. The nature of the contaminant of interest and the constituents of the water matrix must also be considered during the US-assisted AOP optimization procedure since they are involved in the efficiency of the process. In addition to these factors, the type and the geometry of the sonochemical reactor must be considered.

#### 2.2.1. Reaction Zones—The Nature of the Emerging Pollutant

In the ultrasonic radiation process, as indicated above, three reaction zones are recognized for the degradation of compounds: the cavitation bubble, the bubble–water interface and the bulk solution [28,43,48]. The process by which degradation occurs differs from zone to zone. Hydrophilic substances are located within the solution, non-volatile hydrophobic compounds are mainly housed in the bubble–water interface, and volatile substances are commonly located within the cavitation bubble [22].

Inside the cavitation bubble, the degradation reaction of the contaminant occurs by pyrolysis; on the other hand, in the bubble–water interface, the main reaction mechanism is by the attack of free radicals, such as HO•, which are immediately formed by the implosion of the cavitation bubbles; finally, in the bulk solution the reaction occurs directly with the free radicals that reach this zone [28].

According to different investigations, in the US process, the degradation of volatile compounds occur in two zones: in the bubble–liquid interface, through the reaction with the HO• released from the implosion, and/or inside the bubble, directly by pyrolysis [22,41]. The rates of destruction of volatile contaminants depend on the physical and chemical conditions within the bubble, specifically the hydrophobic and volatile nature of these compounds [31,41]. On the other hand, it has been shown that the reaction rate constant of US degradation of volatile compounds decreases with increasing initial concentration, indicating that the relationship between the concentration of a volatile compound in the cavitation bubble and its concentration in the solution will influence the rate of ultrasonic reaction, considering that the collapse temperature depends on the specific heat ratio of the gas mixture [41,48–51].

Hydrophobic compounds such as carbamazepine (CBZ), which has a Henry's constant of approximately 1.08 <sup>×</sup> <sup>10</sup>−<sup>10</sup> atm. m<sup>3</sup> mol−<sup>1</sup> and a moderate solubility in water [9], can be mainly housed in the bubble–water interface, but it is also found within the solution, allowing the protagonist of its degradation to be the HO•, which are immediately formed from the implosion both of the cavitation bubbles and the bubbles that travel within the solution [9].

To evaluate the zone and the way in which a compound is degraded, Nie et al. [28] have implemented the so-called "scavengers" of the HO•. In an experiment where the US process was used to degrade the pharmaceutical diclofenac (DCF), isopropyl alcohol and terephthalic acid were used to inhibit the reaction of the target compound with HO•, functioning as quenchers. The acid was considered to react with free radicals in the bulk solution, while the alcohol reacted both at the bubble–water interface and in the bulk solution. In this regard, the authors verified that when only the acid was added, the degradation of the compound was inhibited. However, when the alcohol was used exclusively as an inhibitor, degradation of the target EC was considerably reduced. It was, therefore, concluded that oxidation of DCF occurred mainly by HO• in the supercritical interface, especially when water was saturated with air and oxygen (O2). Nonetheless, under argon (Ar)- and nitrogen (N2)- saturated conditions, DCF degradation occurred within the cavitation bubbles and/or the bulk solution.

In a study carried out by Kidak and Dogan [52], where the degradation of alachlor through the US process was evaluated, it was concluded that due to the physical properties of the compound, such as the water solubility limit (140 mg L−<sup>1</sup> at 20 ◦C), vapor pressure (negligible), Henry's constant (3.2 <sup>×</sup> <sup>10</sup>−<sup>8</sup> to 1.2 <sup>×</sup> <sup>10</sup>−<sup>10</sup> atm-m3 mol<sup>−</sup>1), octanol–water partition coefficient (Log Kow <sup>=</sup> 2.63–3.53) and its positive ionization, the compound was housed in the bubble–liquid interface, indicating that the degradation was due to the HO• recently formed from the implosion of the cavitation bubbles. The degradation obtained of the target compound was near 100% with a frequency of 575 kHz and an electrical power of 90 W.

Adityosulindro et al. [53] evaluated the degradation of ibuprofen (IBU) in order to ascertain the reaction zone in which the degradation of IBU was established, and whether it was due exclusively to HO•. For this purpose, they tested the sequestration of these radicals through two compounds, n-butanol, which is a short chain alcohol with partial solubility in water that is expected to react with the radicals housed in the bubble–liquid interface; and acetic acid, which should react with the free radicals in the bulk solution due to it is a completely miscible compound. The results obtained indicated that, indisputably, IBU reacted with the HO• recently formed during the implosion of the cavitation bubbles, which means that it is a compound housed in the interfacial zone [53]. The same conclusion was reached by Méndez-Arriaga et al. [42], who attributed the degradation of IBU to the HO• recently produced, since IBU is considered to be housed at the bubble–water interface due to its Henry's constant (1.5 <sup>×</sup> 10−<sup>7</sup> atm m3 mol−1), low solubility in water (21 mg L−1) and octanol–water partition coefficient (3.9).

In the case of acetaminophen (ACP), a polar compound with high solubility (12.5 mg mL−1), Villaroel et al. [54] reported that this contaminant was degraded in a greater proportion within the bulk, estimating that its behavior would be that of a hydrophilic substrate. Nonetheless, in this investigation, it was concluded that ACP can be housed both in the bulk solution and in the bubble–water interface, attributing its degradation to the HO• formed during the implosion of the cavitation bubbles. Based on the aforementioned authors' estimations, the hydrophilic or hydrophobic behavior of the target compound was more related to the initial pH value of the solution at which the study was carried out.

#### 2.2.2. Ultrasonic Frequency

The frequency with which ultrasonic waves are produced can range from 20 to 10,000 kHz, and the US process is divided into three regions: low, high and very high frequency [22]. In Table 1, the frequency ranges used in the ultrasonic oxidation process are listed.


**Table 1.** Frequency ranges used in the ultrasonic process. Taken from [22].

Ultrasonic frequency is a fundamental parameter in the performance of US process, since the size and duration of the cavitation bubble, the violence of the implosion and, therefore, the production of HO• depend considerably on it [9,55].

The number of cavitation bubbles and bubble collapses increases with rising frequency. However, it is important to note that the bubbles generated at high frequencies are small, and release less energy than low frequency bubbles generated by a single pulse [6,56,57]. In addition, the escape of more HO• is inferred, before recombining, when faster collapses occur [9,58]. In this sense, the optimal frequency is determined by the integral efficiency of the energy discharge, which depends on the quantity, size and lifetime of the bubbles. It is noteworthy to mention that the optimal frequency varies according to the different compound to be treated [52,59].

Rao et al. [9] chose two frequency values (200 and 400 kHz) to determine the optimal one for the degradation of CBZ. The first of these values was more effective for the degradation of the target compound. This result was ascribed to the differences in calorimetric powers obtained for both frequencies under the same electrical power (100 W), resulting in a higher calorimetric power for the 200 kHz frequency. This can be attributed to what was previously explained, i.e., each EC requires an optimal frequency at which its degradation will be favored, which depends on its physicochemical properties. This optimal frequency will also be influenced by the geometry of the reactor since, as mentioned above, it will depend on the formation of symmetrical or asymmetrical cavitation bubbles.

On the other hand, in the research carried out by Güyer and Ince [23], different levels of ultrasonic frequency were evaluated in the US process of the DCF. The results obtained allowed the conclusion that the maximal rates of DCF degradation were reached at a frequency of 861 kHz and the minimal ones at 1145 kHz (carrying out tests with values of 577, 861 and 1145 kHz). The improvement between the 577 and 861 kHz was due to the fact that the latter reduced the size of the bubbles, leading to a greater number of bubbles and active oscillations, which contributed to the generation of HO• improvement. However, the highest frequency evaluated this efficiency was reduced due to the fact that the "optimal" frequency related to the reactor configuration was surpassed [53,56].

#### 2.2.3. Electrical Power

The electrical power supplied to the ultrasonic transducer is a critical parameter that can largely determine the performance of the US process [9].

For Jiang et al. [41], the increase in ultrasonic power in the degradation of volatile compounds such as chlorobenzene, 1, 4-dichlorobenzene and 1-chloronaphthalene caused an increase in the cavitation energy, decreasing the cavitation limit and increasing the amount of bubbles produced. This resulted in a rise in the rate of degradation of this type of compounds, considering that the bubbles formed had enough energy to pyrolyze the tested pollutants. This is justified by the fact that volatile compounds are pyrolyzed within the cavitation bubbles, so the more bubbles formed, the more spaces for these compounds to react.

In a study carried out by Tran et al. [18], sonochemical efficiency was evaluated by means of calorimetric tests to determine the optimal power and to propose an experimental design in order to degrade the drug CBZ. It was determined that powers between 20 and 40 W favored sonochemical efficiencies, unlike what happened with powers of 10 W. This finding was attributed to the fact that as the power increased, so did the ultrasonic energy of the reactor, which caused the pulsation and collapse of the bubbles to be generated at a faster rate, resulting in a greater number of cavitation bubbles.

It is important to note that the effect of ultrasonic power and oxidizing species can be influenced by bubble dynamics [18]. The results reported by Gogate et al. [60] indicated that the size, number, lifetime and pressure of the bubbles were a complex function of the power dissipation rate. The research conducted by these authors explains the results obtained by Tran et al. [18], since they found that by increasing the power, the number of cavitation bubbles rose and, consequently, the production of HO• increased. In this way, the degradation of the target compound, CBZ in this case, was directly increased. Similar results were observed in the work carried out by Madhavan et al. [61] for DCF, who studied the degradation of this compound under a frequency of 213 kHz, a temperature of 25 ◦C, a variation of power density between 16–55 mW mL−<sup>1</sup> and a concentration of the pharmaceutical compound of interest of 0.07 mM. The same conclusion was also reached by Rao et al. [9], who studied the degradation of CBZ at pH 6, a frequency of 200 kHz and a power variation between 20 and 100 W, obtaining a higher degradation of CBZ at 100 W.

In the work carried out by Naddeo et al. [6], the degradation of DCF during the US process was evaluated. It was determined that, by increasing the power density from 100 to 400 W L−1, the concentration of the contaminant decreased, making US the most efficient process. This result supports the theory developed in the work described above, i.e., the greater the potency, the greater the degradation percentage of the compound under study.

On the other hand, the combination of different levels of the parameters that influence the degradation of a compound in the US process must be considered. In this context, when the response surface methodology was used to determine the optimal operating levels of CBZ ultrasonic treatment, [18] it was observed that the treatment time had a more statistically significant impact on the efficiency of CBZ removal in comparison with the electrical power, as efficient degradation of the contaminant at lesser powers (10–40 W) required more treatment time. This fact is ultimately reflected in the use of electricity and, therefore, in higher operating costs associated with the application of the oxidation process.

Meanwhile, Kidak and Dogan [52] stated that increasing electrical power also increased the number of the bubbles formed, and that better results were expected in the degradation of the pollutants of interest. This assertion was supported by the results obtained in the experimentation with alachlor, where near 100% degradation was obtained through the US process (initial concentration of alachlor 100 μg L<sup>−</sup>1, frequency of 575 kHz and powers of 45, 60 and 90 W). In addition, they observed an increase in the reaction rate constants as the ultrasonic power increased.

However, in the work carried out by Ince [62], it was evidenced that the degradation of paracetamol (PCT), also known as ACP, DCF and IBU was reduced when operating with a high frequency (861 kHz). The author attributed that fact to the formation of clouds of bubbles when exceeding the threshold power (optimal), which increased the sound waves and, as a result, decreased the cavitation activity. The same author pointed out that below the power threshold, when the power was increased, the efficiency of the process rose.

Adityosulindro et al. [48] evaluated the degradation of IBU by the US process and the influence of the power density in the conversion of the target pharmaceutical. It was determined that increasing the power in a range between 25−100 W L<sup>−</sup>1, over 180 min of treatment, contributed to a greater formation of HO•. However, the authors stated that above a critical or optimal power density value, a cloud of bubbles would be formed, dispersing the formation of sound waves, which would in turn decrease the efficiency of the process [53].

#### 2.2.4. Solution pH

The pH of the solution is a fundamental parameter in oxidation-reduction reactions. In the US process, the pH indicates the hydrophobic or hydrophilic nature of the target compound behavior, depending on whether the structure in which the pollutant is found is ionic or molecular. This property will allow the position to be determined in which the contaminant is housed in the US process, i.e., in the bulk solution (hydrophilic, non-volatile compounds), in the bubble–water interface (semi-volatile hydrophobic compounds), or within the cavitation bubble (hydrophobic, volatile compounds) [9]. This position, in turn, will determine whether the degradation pathway of the contaminant is by pyrolysis or by reaction with the HO• formed by implosion of the cavitation bubbles.

In the research carried out by Tran et al. [18], CBZ was degraded through the US process, considering the effects of the electric power, initial drug concentration, treatment time and pH of the solution (7–10). In this work, no significant influence on the part of pH was evidenced through an experimental factorial design, as a response surface methodology. This can be attributed to the fact that no tests were performed with acid pH values, which influence the structural form in which the compound is found in the aqueous medium and, therefore, the reaction zone in which it is found. Specifically, the pH values will favor or disfavor the hydrophobicity of the compound, with more hydrophobic compounds that are closer to the bubble–water interface reacting with the HO• that have just been formed from the implosion of the cavitation bubbles, whereas those compounds further away from the bubble–liquid (hydrophilic) interface possibly reacting with the HO• that reach the solution.

On the other hand, in the work carried out by Rao et al. [9], the influence of pH on the degradation of CBZ was evaluated, using levels between 2.0 and 11.0. The results showed that, at pH values between 4.5 and 11.0, the degradation remained constant and decreased in equal proportion, but with pH values close to 2.0 there was a small decrease in the degradation efficiency. This was ascribed to the fact that CBZ reacts at the bubble–water interface whenever hydrophobicity is favored—a result that was achieved with pH values between 4.5 and 11.0—whereas with pH values close to 2.0 the ionic structure of the compound, and thus its hydrophilicity, was favored. The compounds that can lodge very close to the cavitation bubbles can react with a greater amount of HO• than those ones that are in the bulk solution, which must wait for these oxidizing agents to reach them, being able to react with another compound along the way, such as the natural constituents of the aqueous matrix tested.

Meanwhile, Huang et al. [46] evaluated the degradation of DCF through US process in the presence of Zn0, performing an analysis of the influence of pH on this type of process. It was found that at pH higher than 2 the degradation of DCF was very small, while at equal or lower values, the degradation of the tested compound reached percentages higher than 80%. The authors attributed this behavior to the fact that the pKa of DCF is 4.15, considering that aqueous media with a pH lower than this value will manage to maintain the molecular structure of this compound, and concluded that this form favored the absorption reaction of DCF by Zn0.

In the degradation of IBU through the US process, the influence of pH was evaluated, experimenting with values higher and lower than the pKa of the compound (4.9). It was found that, at lower values (2.6 and 4.3), the compound remained unprotoned and its degradation slightly increased, while the opposite occurred with an alkaline pH value (8.0), where IBU degradation was affected. However, the authors argued that under its ionic form, IBU should accumulate less at the bubble–water interface, which is where the HO• attack mainly occurs [53].

Al-Hamadani et al. [31] evaluated the degradation of sulfamethoxazole (SFX) and IBU under three pH conditions: acid (3.5), below the pKa values of the target compounds; basic (7), above the pKa values; alkaline (9.5), well above these values. The results showed degradations near 100% of the compounds in 1 h of treatment for a pH below pKa, while degradation was significantly affected above these values. This is attributed to the molecular form of the compounds, i.e., when the pH of the solution was below pKa, the hydrophobicity of the drugs and, therefore, their position in the bubble–water interface is improved, favoring a rapid reaction with the HO• recently formed during the implosion of the cavitation bubbles.

#### 2.2.5. Constituents of the Water Matrix

Various investigations related to the degradation of ECs in water through AOPs have been carried out in aqueous matrices with different constituents. On one hand, some researches have been developed with synthetic waters which, in general, involve the use of distilled water doped with the chemical components offering the specific characteristics with which the researcher wishes to work. On the other hand, there are works operating with real wastewater or in which the efficiency of the process for natural surface and drinking water is evaluated.

The research carried out by Tran et al. [27] identified the levels of the operating parameters at which IBU could be degraded by 65% through a sono-electrolytic process under controlled conditions in synthetic water, using a statistical optimization procedure. These same conditions were evaluated with sewage from a municipal treatment plant, with organic and inorganic compounds, as well as microbial load, which was doped with a specific concentration of IBU. The result obtained was a greater degradation of the compound of interest (90%) than that statistically estimated with synthetic water. This result was ascribed to the apparent presence of the chloride ion (Cl−), which favored electrolysis, and might also favor the formation of hypochlorous acid (HClO), which can improve IBU oxidation. This demonstrates the importance of studying the organic and inorganic content of the water to be treated, as this may favor or limit the degradation of the target compounds.

It has been reported that Cl− have different effects on the elimination of ECs present in water treated by means of AOPs [63,64]. Rao et al. [9] evaluated the degradation of CBZ (0.025 mM) using 200 kHz 100 W US. These authors investigated the presence of different inorganic anions to determine their influence on the process. The anions evaluated were Cl−, SO4 <sup>2</sup><sup>−</sup> and NO3 −, and it was found that Cl− slightly restricted the degradation of the investigated drug, while the others did not have a significant impact on the degradation of the compound of interest. This slight inhibition in CBZ degradation due to the presence of Cl<sup>−</sup> can be attributed to the reaction of this ion with the HO• dispersed in the solution, resulting in the formation of ClOH•−.

In the work reported by Adityosulindro et al. [53] on the Fenton, US oxidation system and US-Fenton process, the efficiency of the degradation of IBU in distilled water and in wastewater from a municipal treatment plant was compared. The results showed a negligible difference between the degradation capabilities of all the evaluated processes in both distilled water and wastewater. In this context, the authors stated that the organic and inorganic content of the sewage effluent did not compete with IBU for HO• and that the latter was capable of reacting first with the oxidizing agent. It is important to highlight that the experimentation was carried out at acid pH, which could favor the location of IBU in the interface zone, making it more competitive when reacting with the HO• formed from the implosion of the cavitation bubbles.

Rao et al. [9] compared the degradation of CBZ contained in synthetic water with that of an effluent from a municipal wastewater treatment plant, evaluating the efficiency of two processes: ultrasonic irradiation alone, and in combination with photolysis using UV radiation emitting at a wavelength of 254 nm. The results showed that, for the US process, the constituents of the real wastewater matrix had no influence on the degradation of CBZ when compared with the results for distilled water. On the other hand, in the combined process, the wastewater matrix increased the efficiency of the studied drug degradation. This can be attributed to the photolysis of certain compounds contained in the wastewater that provide the oxidizing agent and favor the degradation of CBZ. As a matter of fact, the referred authors gave the example of nitrate ions (NO3 −).

In the research conducted by Villaroel et al. [54], the influence of ionic constituents of water on the degradation of ACP (82.69 μmol L−<sup>1</sup> and 1.65 μmol L<sup>−</sup>1), at a power of 60 W and ultrasonic frequency of 600 kHz, was evaluated. The results obtained in distilled water and in synthetic water containing calcium ions (Ca2<sup>+</sup>), magnesium ions (Mg2<sup>+</sup>), sulphates ions (SO4 <sup>2</sup>−), bicarbonates ions (HCO3 −), Cl−, potassium ions (K+) and fluorides ions (F−) were compared. The results indicated that, for the lowest concentration of ACP, a more pronounced acceleration of degradation was observed when this occurred in water with similar ion content than in distilled water. The authors attributed this to the high content of HCO3 −, which was likely to be the protagonist in the formation of the carbonate radical (HCO3•) when reacting with HO• radicals, being HCO3•, a contributor to the degradation of the target EC.

With regard to the use of dissolved gases and their influence on the degradation of organic ECs, in the work conducted by Nie et al. [28], whose objective was to degrade DCF through the US process, it was observed that under saturated air, O2 and Ar, a complete mineralization of nitrogen and a partial mineralization of carbon was achieved. When oxygenation was added to the reaction solution, HO2• was formed. Although these radicals do not have an oxidation potential as high as HO•, as mentioned previously, HO2• can contribute to the degradation of the compounds of interest [6].

On the other hand, it must be highlighted that when chlorine atoms are part of the target EC structure, they are transformed to Cl− through the reaction of the pollutant of interest with HO• or by pyrolysis in the US process [41,49]. Therefore, the release of Cl<sup>−</sup> occurs during the sonochemical degradation of chlorinated compounds, which was attributed to the rapid excision of the carbon–chlorine bonds by high temperature combustion occurring within the cavitation bubbles or at the bubble–liquid interface. Cl−, as indicated above, can reduce the oxidation potential of the process.

Under this scenario, studies aiming at examining the efficiency of the US process are required to be conducted by using real matrices due to the natural constituents of the water matrices can positively or negatively influence the degradation percentages and reaction rates of the ECs of interest.

#### 2.2.6. Temperature of the Solution

According to some authors, temperature variation in the US process directly influences cavitation intensity due to the changes in the physicochemical properties of the compound and the type of cavities formed, which can affect the kinetic velocity constant of the degradation reaction [65].

Al-Hamadani et al. [31] indicated that certain parameters were affected by increasing the temperature in the US process. First, it was found that cavitation energy decreased, as well as the threshold limit of the energy required to produce cavitation. In addition, it was found that the amount of dissolved gas was reduced, leading to the transfer of organic molecules from the bulk solution to the bubble–water interfacial region. Finally, the vapor pressure increased, causing the cavitation bubbles to contain more water vapor. Furthermore, the aforementioned authors, who evaluated the degradation of SFX and IBU through US, evidenced the temperature influence on the oxidation process. Temperatures between 15 and 55 ◦C were tested and it was concluded that, when this parameter was increased, the degradation of the studied compounds rose, as a rise in the temperature of the bulk caused the cavitation threshold to lower, which contributed to the formation of a greater number of cavitation bubbles and, therefore, to a greater amount of HO•. However, these authors pointed out that other works have shown an adverse effect of temperature on the degradation of the contaminant. These findings can be attributed to the fact that the surface tension and viscosity of the solution increase, generating cavitation bubbles with less intensity due to a rise in the vapor pressure of the liquid.

#### *2.3. Application of Ultrasound Process to Water Treatment*

Table 2 compiles several relevant research works related to the treatment of ECs through ultrasound as an AOP alone or in combination with other physical-chemical and advanced oxidation technologies.


**Table 2.** Summary of works related to the removal of emerging pollutants through the ultrasound process and its combination with other physical-chemical and advanced oxidation processes.


**Table 2.** *Cont.*

**Table 2.** *Cont.*



**Table 2.** *Cont.*


**Table 2.** *Cont.*

Due to the demonstrated efficiency ascribed to the use of US-assisted AOPs in the degradation of ECs in water, it has been widely applied for tackling the problem of water pollution with these pollutants of growing concern [32,36,64]. As stated previously, it is highlighted that the water matrix is a topic of utmost importance when it comes to the evaluation of the pollutant removal capability through AOPs. In fact, in the literature, different works have been reported based on the elimination through US waves of various ECs commonly present in water matrices of different nature, from drinking water effluents to natural surface water, with domestic and industrial wastewaters being highly studied [40,69,70] due to the vast variety of compounds that can be found in these kinds of aqueous matrices.

For instance, Cetinkaya et al. [69] investigated the decolorization of textile waters using the sono-Fenton process, obtaining better results at pH 3, achieving 96% of color removal. The influence of ferrous ions (Fe2+) concentration was analyzed, testing its variation between 0.05 g L−<sup>1</sup> and 0.2 g L<sup>−</sup>1. A color removal of 90% and 99% was observed with the lowest and the highest Fe2<sup>+</sup> concentration, respectively. These results indicated that the sono-Fenton process required small amounts of Fe2<sup>+</sup> to achieve high removals of the dyes. Additionally, H2O2 consumption was reduced by about 30% with the sono-Fenton process compared to the classic Fenton process. Furthermore, authors optimized operating parameters involved in the investigated AOP, achieving the highest removal of color at a frequency of 35 kHz, pH 3, 0.05 g L−<sup>1</sup> of Fe2+, 1.65 g L−<sup>1</sup> of H2O2 and a treatment time of 60 min.

The removal of tetracycline (TC) has also been evaluated by Nasseri et al. [40] in a wastewater effluent by applying the US process. Some of the natural characteristics of the studied wastewater were: pH 7.9, chemical organic demand (COD) of 25 mg L−1, HCO3 <sup>−</sup> content of 164 mg L−1, Cl<sup>−</sup> of 92 mg L<sup>−</sup>1, NO3 <sup>−</sup> of 24 mg L−<sup>1</sup> and Na<sup>+</sup> of 50 mg L<sup>−</sup>1. A lower removal rate of TC, but in the same order of magnitude, in wastewater (1.25 <sup>×</sup> 10−<sup>2</sup> min−1) compared to that one obtained in ultrapure water (1.75 <sup>×</sup> <sup>10</sup>−<sup>2</sup> min<sup>−</sup>1) was observed. These results may be ascribed to the negative influence of the water constituents, as explained previously; in this case, due to the high levels of organic matter, in terms of COD, which can prevent the formation of OH• and, subsequently, reduce the rate of TC degradation.

In turn, Serna-Galvis et al. [71] experimented with wastewater from El Salitre Treatment Plant, located in Bogotá (Colombia), with the objective of applying the sono-photo-Fenton/Oxalic Acid AOP for the removal of the following pharmaceuticals: DCF, CBZ, venlafaxine, ciprofloxacin, norfloxacin, valsartan, losartan, irbesartan, SFX, clarithromycin, azithromycin, erythromycin, metronidazole, trimethoprimine and clinimetropimine, as well as cocaine and its main metabolite benzoylecgonine. The operating conditions were: 300 mL of working volume, 88 W L−<sup>1</sup> of power density, 375 kHz of frequency, 20 ◦C of temperature, a UVA lamp of 4 W, a Fe2<sup>+</sup> content of 5 mg L−<sup>1</sup> and an oxalic acid concentration of 2 mg L<sup>−</sup>1. It was observed that the application of the sonochemical process alone led

to the release of contaminants from suspended solids. The addition of Fe2<sup>+</sup>, UVA light and oxalic acid to the US process significantly increased the elimination of the studied ECs in the effluent, thanks to the production of additional HO• through reactions between iron and the sonogenerated H2O2. It is important to note that the presence of oxalic acid makes iron more available for the formation of additional free radicals within the solution, causing the improvement of EC degradation.

With the aim of comparing the findings of degradation reported by US in wastewater, the work conducted by Vilardi et al. [70], where the efficiency of conventional and heterogeneous Fenton for the degradation of contaminants present in the wastewater of a tannery in terms of COD, total phenolic compounds (TP) and Cr(VI), is presented. The authors carried out the experimentation at large laboratory scale using a reactor with a volume of 7.4 L. It was concluded that the heterogeneous Fenton process was significantly more efficient with respect to the conventional one for the elimination of COD and TP, once the optimal values of the operating parameters were found. The percentages of COD and TP removal for the heterogeneous Fenton were 75.5 ± 2.1% and 85.1 ± 0.7%, respectively. Likewise, it was observed that a smaller amount of iron sludge was produced due to the heterogeneous Fenton process (17.5%) compared to that one achieved through the conventional Fenton process (21.6%), which is a key aspect for the feasible implementation of the process at industrial scale.

Although the heterogeneous Fenton process implemented above was demonstrated to produce relatively low amounts of sludge, a more environmentally safe process must be required to overcome the pollution of aqueous resources with recalcitrant contaminants. In this regard, the use of US as an AOP alone or in combination with other advanced oxidation technologies seem to be an attractive treatment option.

#### **3. Future Perspectives**

Although the application of US alone as an advanced oxidation technology to overcome the critical situation ascribed to ECs in aqueous environments has been demonstrated to be efficient, the coupling of US with other AOPs could improve the mineralization of emerging organic compounds [72,73] within a further reduced time of treatment. For this reason, the use of US hybrid techniques has been recently studied to improve EC mineralization results [6,23,45,73]. A clear example of this is the combination of sonolysis with the Fenton process. This combination, which is so-called sono-Fenton, could stimulate a faster conversion and/or mineralization of ECs. This is achieved through: firstly, higher generation of HO• [74]; secondly, an improved mixture and contact between HO• and the pollutants of interest [72,73], and thirdly, improved generation of Fe2<sup>+</sup> [75].

Different strategies in addition to the combination of the Fenton process with sonolysis have been tested in the last years. An example of this is the work developed by Tran et al. [20], where the electro-oxidation (EO) process was combined with US. This combination was based on the fact that, initially, the formation of HO• is achieved on the wall of an electrode made up of a non-active material through the EO process, and the chemical exchange of these HO• with contaminants could then be improved due to the formation of the US waves and cavitation bubbles resulting from the US process. In this study, a higher kinetic velocity constant and a greater efficiency in the removal of IBU was obtained with the combined process of EO/US in comparison with the results obtained in each process independently. As a result, 90% of the IBU contained in samples of municipal sewage was removed using optimal parameter levels, such as the treatment time, the current intensity and the US power, which were determined through the response surface methodology. The beneficial results of the exposure of electrochemical cells to the effects of US power are related to the improved mass transport, increased current efficiencies, and continuous electrode surface activation [20,76,77]. These effects can be attributed to the rapid generation and collapse of the micro-bubbles within the electrolyte medium or near the electrode surface [20,77,78].

In turn, Ince's 2018 study [62] evaluated the degradation of toxic ECs through US in combination with other AOPs. In this study PCT, DCF and IBU were analyzed, finding that the degradation of the selected ECs was more efficient at high frequencies and acid pH. Degradation was further improved with the presence of solid catalysts, which provided surfaces that enhanced the formation of cavitation bubbles and, therefore, the performance of the oxidation processes. In the referred research, the use of iron nano- and micro-particles resulted in a higher rate of DCF elimination by using nano-particles [62], which was attributed to the synergy of US with these particles through the enrichment of massive surfaces with excessive sorption sites and cavitation nuclei. In addition, reactions at the bubble–liquid interface were intensified by the distortion of asymmetric shapes, the degree of which increases as particle size decreases [62]. On the other hand, the coupling of an ozonation system with UV radiation and sonication, with the optional addition of FeSO4, completely degraded DCF [62]. Finally, this work compared the efficiency of the following AOPs: US, O3/US, UV/US and O3/US/UV. High removals of the drugs of interest were found in all the tested processes, reaching about 100% elimination accompanied by a mineralization between 40 and 60% of all the ECs with the combination of US, O3 and UV radiation.

In the work developed by Rao et al. [9], sonolytic and photolytic AOPs were combined for the degradation of CBZ. The result was a significant improvement in the drug degradation compared to the results obtained when the processes were individually implemented. The reason for this fact was related to the formation of H2O2 resulting from the recombination of HO• from sonolysis. This oxidizing agent can be photolized by UV light and more HO• can be produced, which are the main contributors to CBZ degradation.

CBZ removal was also studied by Mohapatra et al. [79], through the US process, Fenton and ferro-sonication (a combination of FeSO4 with the US process). It was found that the most efficient AOP was the Fenton process, with elimination percentages between 84–100%; this was followed by ferro-sonication, with values between 62–93%, while sonolysis only achieved CBZ elimination percentages between 22%–51%. The authors concluded that the higher the radiation intensity (5.8, 12.4 and 16 W cm<sup>−</sup>2), the greater the elimination of the target drug. Moreover, according to their research, the resulting ranges of efficiency between one process and another were because FeSO4 contributed to the formation of a greater amount of HO•.

Although Fenton process has been proven to be an efficient technology for the degradation of some ECs [70,79], residual sludge is produced, especially when the homogeneous Fenton process is applied [70]. In this regard, further studies are needed to give an alternative use to such as sludge, contributing to the so-called principles of the circular economy. In this regard, Vilardi et al. [80] treated a tannery wastewater with mixed-iron coated olive stone bio-sorbent particles in combination with H2O2. They found a COD removal efficiency of 58.4% and a TP removal of 59.2%, at H2O2/COD (w/w) equal to 0.875. The coated olive stones were regenerated with sodium hydroxide (NaOH) and oxalic acid (C2H2O4) solutions after five cycles in order to enable their reuse.

In addition, considerate the circular economy principles, economic costs analysis must be carried out in order to discern whether an AOP tested at laboratory or pilot plant can be scale up for industrial application in real water effluents.

#### **4. Cost Consideration**

As reviewed, the efficiency of ultrasound has been demonstrated to degrade any kind of recalcitrant pollutants. However, there are limitations related to the economic costs associated with the use of this advance oxidation technology for the treatment of water containing toxic pollutants [39]. One such limitation is the cost, which can be divided into two groups: the capital or inversion costs, which consists of those costs associated with the manufacture of the sonochemical reactors and can be amortized over a span of years at a considered amortization rate [39], and the operation and maintenance costs. The economic cost estimation linked to the operation and maintenance labor include the part replacements, which mainly consists of the transducer element replacement and the tip or electronic circuit replacements. In fact, according to Mahamuni and Adewuyi [39], the part replacement costs are assumed to be 0.5% of the capital costs. Labor and analytical costs must also be considered when operation and maintenance costs are estimated. Labor costs include inspection, repair and replacement based on hours of service life of control panels, leakages and pressure gauge, among others. In turn, analytical costs consist of the costs related to the analysis of samples and, subsequently, the costs associated with the reactants and chemicals used for the sample analysis. Additionally, electrical costs, which can be based on the power consumption of the referred AOP, are of utmost importance since they are usually very high, especially in those countries where the cost of each kWatt is high. Hence, the use of renewable resources for generating electrical energy is an attractive option that is emerging for the advanced oxidation system to be implemented. As a matter of fact, Rubio-Clemente et al. [81] assessed the efficiency of the UV/H2O2 system powered by a photovoltaic (PV) system in a photochemical reactor at laboratory scale. According to the results reported by the authors, similar efficiency was observed between the oxidation system powered with energy from the electrical grid and that one generated using the PV cells implemented.

Another alternative for reducing the economic costs related to the use of ultrasound for treating polluted water is utilizing hybrid oxidation techniques by combining US with other AOP, including the use of oxidizing or catalyzing agents, such as ozone (O3), H2O2, iron, titanium dioxide (TiO2), wolfram trioxide (WO3), zinc oxide (ZnO), etc., and electrochemistry to name just a few. In this regard, Expósito et al. [25] evaluated the efficiencies of mineralization in terms of total organic carbon (TOC) and CBZ removal by using the US/UV/H2O2/Fe oxidation process at laboratory scale in a thin film UV reactor coupled to a 24 kHz 200 W direct immersion horn-type sonicator, obtaining efficiencies around 90%, which are higher than the efficiencies reached by the processes alone. In fact, a synergistic effect higher than 55% was found between the US process and UV irradiation.

However, although application of US hybrid techniques in some occasions can be more attractive for water treatment, Mahamuni and Adewuyi [39] reported that the costs associated with these treatment techniques are one to two orders of magnitude higher than when US is implemented alone. This can be ascribed to the costs linked to the additional chemicals used for the hybrid process to occur, i.e., the use of oxidizing agents such as O3 and H2O2, or the catalyzing agents as iron salts, TiO2, ZnO or WO3, among others, as well as the adjustment of the pH of the solution if needed. Moreover, when US is used along with UV radiation, the costs associated with the replacements of the lamps and the electrical consumption of the lamps must be considered, as well as those ones related to the O3 generator repair when O3 is combined with US.

With this in mind, it can be concluded that the cost estimation studies based on pilot plants would be of high importance for to discern both the capital and the operation and maintenance costs related to the implementation of the US process. Furthermore, although high efficiencies can be obtained in a short period of time by using hybrid techniques with US, the economic costs associated with it are higher; therefore, further studies are needed to discern if the combination of US with another AOP is worth to be implemented under any circumstances. On the other hand, the type of pollutant plays a crucial role on the cost estimation procedures, since treating water containing hydrophobic pollutants has lower costs ascribed in comparison with those ones for treating compounds of hydrophilic nature [39].

#### **5. Conclusions**

After a critical review of the results found in the literature concerning the US process for the elimination of ECs, it is important to highlight the following conclusions:


**Author Contributions:** Conceptualization, A.L.C.-P., A.R.-C., G.A.P.; investigation, A.L.C.-P., A.R.-C., G.A.P.; writing—original draft preparation, A.L.C.-P.; writing—review and editing, A.R.-C., G.A.P.; supervision, A.R.-C., G.A.P. All authors have read and agreed to the published version of the manuscript.

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

**Acknowledgments:** To the Universidad de Antioquia for its commitment to education in the country and its stimulus "Student Instructor" that contributed to the development of this article. To the Research Group "Diagnóstico y Control de la Contaminación-GDCON" for its support in the development of this review.

**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* **Degradation of Losartan in Fresh Urine by Sonochemical and Photochemical Advanced Oxidation Processes**

## **John F. Guateque-Londoño 1,2, Efraím A. Serna-Galvis 1, Yenny Ávila-Torres 3,\* and Ricardo A. Torres-Palma 1,\***


Received: 30 October 2020; Accepted: 29 November 2020; Published: 3 December 2020

**Abstract:** In this work, the degradation of the pharmaceutical losartan, in simulated fresh urine (which was considered because urine is the main excretion route for this compound) by sonochemistry and UVC/H2O2 individually, was studied. Initially, special attention was paid to the degrading action of the processes. Then, theoretical analyses on Fukui function indices, to determine electron-rich regions on the pharmaceutical susceptible to attacks by the hydroxyl radical, were performed. Afterward, the ability of the processes to mineralize losartan and remove the phyto-toxicity was tested. It was found that in the sonochemical treatment, hydroxyl radicals played the main degrading role. In turn, in UVC/H2O2, both the light and hydroxyl radical eliminated the target contaminant. The sonochemical system showed the lowest interference for the elimination of losartan in the fresh urine. It was established that atoms in the imidazole of the contaminant were the moieties most prone to primary transformations by radicals. This was coincident with the initial degradation products coming from the processes action. Although both processes exhibited low mineralizing ability toward losartan, the sonochemical treatment converted losartan into nonphytotoxic products. This research presents relevant results on the elimination of a representative pharmaceutical in fresh urine by two advanced oxidation processes.

**Keywords:** advanced oxidation process; elimination routes; fresh urine; pharmaceutical degradation; processes selectivity; theoretical analysis

## **1. Introduction**

Losartan was the first commercialized angiotensin II antagonist pharmaceutical. This is an antihypertensive consumed widely around the world [1]. Urine is the main route of excretion of losartan from the human body, ≈35% of the oral dose is expelled without alterations [2], reaching the wastewater systems. In fact, losartan has been determined in ranges of 0.0197–2.76 μg L−<sup>1</sup> in wastewater treatment plants influent (WWTP) [3,4]. This indicates that losartan is not effectively removed by the conventional systems in WWTP.

In the aquatic environment, losartan can promote noxious effects on organisms, and it can be transformed into more toxic and persistent substances [5–7]. The recalcitrance to conventional treatment systems, negative environmental impact, and high excretion of losartan in urine lead to consider alternative options to eliminate this pharmaceutical from aqueous media. Particularly, the application of degradation processes should be focused on primary contamination sources, such as human fresh urine.

Advanced oxidation processes (AOPs, which are based on the production and utilization of radical species to attack pollutants) are interesting options for losartan elimination from urine, to avoid entering into the wastewater systems. Indeed, AOPs such as UVC/H2O2 and sonochemistry have been successfully applied for the elimination of different pharmaceuticals in diverse aqueous matrices [8].

In the UVC/H2O2 process, UVC light (e.g., photons of 254 nm) promotes the homolysis of hydrogen peroxide, generating hydroxyl radicals (Equation (1)) available to degrade organic contaminants (Equation (2)) [9].

$$\text{H}\_2\text{O}\_2 + \text{hv}\_{(\text{<290 nm})} \rightarrow 2\text{HO}^\bullet \tag{1}$$

$$\text{HO}^{\bullet} + \text{Pollutant} \rightarrow \text{degradation products} \tag{2}$$

Meanwhile, the sonochemical process, which uses high-frequency ultrasound waves ")))", produces hydroxyl radicals from the breaking of water molecules and dissolved oxygen (Equations (3)–(5)) [10].

$$(\text{H}\_2\text{O} + \text{})\text{)} \to \text{HO}^\bullet \, + \,\text{H}^\bullet \,\tag{3}$$

$$(\text{O2} + \text{()})) \rightarrow 2\text{O}^\bullet \tag{4}$$

$$\left(\mathrm{H}\_{2}\mathrm{O} + \mathrm{O}^{\bullet}\ \mathrm{I}\right)\mathrm{O} \to 2\mathrm{HO}^{\bullet}\tag{5}$$

It should be mentioned that some previous works have evidenced the high potentiality of AOPs to eliminate pollutants in urine [11–20]. However, until now, the treatment of losartan in fresh urine, considering the intrinsic degradation abilities of UVC/H2O2 and sonochemistry has not been reported. Moreover, computational analyses about the reactivity of this pharmaceutical toward hydroxyl radical species or phytotoxicity tests of the treated water have not been considered. Thereby, the present research was focused on the losartan treatment in fresh urine by UVC/H2O2 and ultrasound individually. The selectivity of the processes toward the pollutant degradation in the urine matrix was established. Firstly, special attention was paid to the action routes of the processes involved in the elimination of losartan. Besides, computational analyses using DFT/Fukui functionals were performed to determine the most regions on losartan reactive to hydroxyl radicals, and these theoretical results were related to primary degradation products coming from the processes action. Additionally, considering the possible reuse of treated urine for water irrigation extra analyses such as mineralization and phytotoxicity were carried out.

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

#### *2.1. Reagents*

Losartan tablets (50 mg each) were purchased from La Santé S.A. Acetonitrile (HPLC grade), ammonium heptamolybdate (>99.3%), methanol (HPLC grade), potassium iodide (>99.5%), potassium perchlorate (>99.5%), sodium acetate (>99%), sodium chloride (99.9%), sodium dihydrogen phosphate (>99.0%), sodium hydroxide (>99.0%), sodium sulfate (>99.0%), sulfuric acid (95–97%), and urea (>99.0%) were provided by Merck. Ammonium chloride (>99.8%), calcium chloride dihydrate (>99.0%), ferrous sulfate heptahydrate (>99.0%), formic acid (99.0%), hydrogen peroxide (30% w/v), and magnesium chloride hexahydrate (>99.0%) were provided by PanReac. All the reagents were used as received.

The solutions were prepared using distilled water. In all cases, the initial losartan concentration was 43.38 μM (i.e., 20 mg L<sup>−</sup>1, which is a plausible amount of the antihypertensive excreted in human urine [21]). The fresh urine used for the tests was prepared according to Table 1. The fresh urine was used immediately after its preparation and the pH was adjusted to 6.1.


**Table 1.** Composition of fresh urine 1.

<sup>1</sup> Composition taken from Amstutz et al. [22].

#### *2.2. Reaction Systems*

For the UVC/H2O2 process, a homemade aluminum reflective reactor box equipped with UVC lamps (OSRAM HNS®, with the main emission peak at 254 nm, 60 W) was used (Figure 1a). Losartan solutions (50 mL) were placed in beakers under constant stirring. Meanwhile, the sonochemical treatments were performed in a Meinhardt cylindrical glass reactor containing 250 mL of losartan solution. Ultrasonic waves of 375 kHz and 106.3 W L−<sup>1</sup> (actual ultrasound power density determined by the calorimetric method) were emitted from a transducer at the bottom of the reactor (Figure 1b). For both processes, the experimental conditions (i.e., reagents concentrations, ultrasonic frequency, light power) were selected based on previous works [23,24].

#### *2.3. Analyses*

## 2.3.1. Chromatographic Analyses

Losartan evolution was followed by using a UHPLC Thermo Scientific Dionex UltiMate 3000 instrument equipped with an Acclaim™ 120 RP C18 column (5 μm, 4.6 × 150 mm) and a diode array detector (operated at 230 and 254 nm). The mobile phase was methanol (10% v/v), acetonitrile (44% v/v), and formic acid (46% v/v, 10 mM, and pH 3.0) at a flow of 0.6 mL min−1. Primary transformation products were elucidated by HPLC–MS analyses in our previous work [24]. For the chromatographic analyses, samples of 0.5 mL were periodically taken from the reaction systems (the total taken volume was always lower than 10% of the initial volume in each system). All experiments were performed at least in duplicate.

#### 2.3.2. Oxidizing Species Accumulation

Accumulation of sonogenerated hydrogen peroxide was estimated by iodometry [25]. An aliquot of 600 μL from the reactors was added to a quartz cell containing 1350 μL of potassium iodide (0.1 M) and 50 μL of ammonium heptamolybdate (0.01 M). After 5 min, the absorbance at 350 nm was measured using a Mettler Toledo UV5 spectrophotometer.

#### 2.3.3. Mineralization Determinations

Mineralization degree was established as removal of total organic carbon (TOC). TOC content of the samples was measured using a Shimadzu LCSH TOC analyzer (previously calibrated), according to Standard Methods 5310B (high-temperature combustion method), in which the water sample is homogenized and injected into a heated reaction chamber packed with an oxidative catalyst (platinum spheres). The water is vaporized, and the organic carbon is oxidized to CO2. The CO2 from oxidation is transported by a carrier gas stream and is then measured using an IR detector. The TOC analyzer performed the catalytic combustion at 680 ◦C using high-purity oxygen gas at a flow rate of 190 mL min<sup>−</sup>1. The apparatus had a nondispersive infrared detector. For the TOC analyses, samples of 7.0 mL were taken from the reaction systems, and for the TOC analyses, the experiments were carried out independently from the initial tests of degradation (to avoid retire amounts higher than 10% of the initial volume in each system).

**Figure 1.** Reactors used in the degradation of losartan. (**a**) UVC/H2O2 process; (**b**) sonochemical treatment.

#### 2.3.4. Phytotoxicity Tests

Toxicity against radish seeds (*Raphanus sativus*) was considered. For such purpose, the ratio of seeds germinated (RSG, Equation (6)) and the ratio of root length (RRG, Equation (7)) were determined. As a phytotoxicity parameter, the germination index (GI, Equation (8)) was assessed according to N.J. Hoekstra et al. [26]. For the phytotoxicity tests, samples of 5.0 mL were taken.

$$\text{RSG } (\%) = \frac{\text{Number of seconds gemimated in sample}}{\text{Number of seconds gemimated in control}} \times 100\tag{6}$$

$$\text{RRG } (\%) = \frac{\text{mean root length in sample}}{\text{mean root length in control}} \times 100\tag{7}$$

$$\text{GI}(\%) = \frac{\text{RSG} \times \text{RRG}}{100} \tag{8}$$

#### 2.3.5. Computational Analyses

For the determination of regions on losartan most susceptible to the attack of radical species and electrophilic oxidants, computational analyses were performed by applying the framework of functional density theory (DFT). The antihypertensive structure was optimized with the B3LYP hybrid functional density [27], 6-311++G(2d,2p) method [28] using the dielectric constant for water to simulate the aqueous environment. Thus, f<sup>+</sup> and f- (i.e., nucleophilic and electrophilic Fukui function indices) values and the average between such values (**fave**) were calculated.

#### **3. Results**

#### *3.1. Treatment of Fresh Urine Loaded with Losartan*

The two processes were individually applied to degrade losartan in the simulated fresh urine (FU, whose composition is presented in Table 1). In addition to degradation in urine, losartan was also treated in distilled water (DW). The degradations followed pseudo-first-order kinetics, and their respective rate constants (k) in both matrices were established (see Figure S1 in Supplementary material). Then, the ratio between the degradation rate constants (Rk:kFU/kDW) was calculated. This Rk parameter is an indicator of both the selectivity of processes toward the antihypertensive degradation in the complex matrix and the inhibitory effect of losartan elimination caused by the fresh urine components. Table 2 contains the k and Rk values for each process.

**Table 2.** Kinetic constants (in min<sup>−</sup>1) determined in the degradation of losartan in fresh urine (kFU) and distilled water (kDW) for each advanced oxidation processes 1.


<sup>1</sup> Experimental conditions: [Pollutant] = 43.38 μM, pH: 6.1. Sonochemistry: 106.3 W L−<sup>1</sup> (375 kHz). UVC/H2O2: [H2O2] = 500 μM, 60 W.

#### *3.2. Degradation Routes of Losartan (LOS) in Di*ff*erent AOPs*

To elucidate the routes of the processes action, some specific experiments and measures in distilled water were carried out and results were compared to those obtained in the urine to understand the effect of the matrix components. Results for each treatment in distilled water are detailed in the following subsections.

## 3.2.1. Action Routes of the UVC/H2O2 Process

The UVC/H2O2 process may include the action of light of 254 nm, hydrogen peroxide, and radicals. To identify the routes involved in the process, control tests for the individual effects of UVC and H2O2 were carried out. Figure 2 compares the degrading effect of individual components of the process in distilled water, plus the losartan elimination in both distilled water and fresh urine (FU) by UVC/H2O2.

**Figure 2.** Determination of action routes of the UVC/H2O2 process on losartan degradation in distilled water (DW) and fresh urine (FU). Conditions: [LOS] = 43.38 μM, [H2O2] = 500 μM, UVC light: 60 W, pH: 6.1.

#### 3.2.2. Degradation Routes Involved in the Sonochemical Treatment

Figure 3 depicts the degradation of LOS in both distilled water and fresh urine (FU) by ultrasound. To determine the degradation route in the sonochemical process, the accumulation of sonogenerated hydrogen peroxide in the presence and absence of the pollutant was also measured (results also presented in Figure 3).

**Figure 3.** Determination of action routes of the sonochemical treatment on losartan in distilled water (DW) and fresh urine (FU). Conditions: [LOS] = 43.38 μM, 106.3 W L−<sup>1</sup> (375 kHz), pH = 6.1.

#### *3.3. Analysis of Losartan Susceptibility to Attacks by Radical Species*

To establish electron-rich regions on losartan susceptible to attacks of radicals, computational analyses were performed [29–31], and the results from the theoretical calculations were used to better understand the formation of the degradation intermediaries. Table 3 depicts the moieties on losartan having more electron density according to Fukui function indices. In addition to these indices, other related quantities such as local softness and global hardness were determined, the values of which were 17.513 and 0.0571 eV, respectively. Moreover, a donor–acceptor diagram (DAM), to show the donor capability of the pharmaceutical concerning hydroxyl radical (HO·), hydroperoxyl radical (HOO·), and superoxide anion radical (·O2 −), was elaborated (Figure S2).


**Table 3.** Results of computational analysis for losartan 1.

<sup>1</sup> Boxes in gray color contains atoms having high values for the Fukui function indices. It should be mentioned that the computational calculations were done for LOS in water.

#### *3.4. Mineralization and Toxicity Evolution in Distilled Water*

The ability of the two processes to mineralize losartan was analyzed. The experiments were carried out in distilled water to avoid interfering effects of matrix and understand the fundamental aspects of the mineralizing action of the processes. We can mention that if mineralization is carried out in the fresh urine matrix, the urea that has a higher concentration masks the contribution of losartan, making it difficult to evaluate the mineralization of the contaminant under the oxidation processes. The TOC removal, at different treatment times normalized concerning the time necessary to completely degrade losartan in distilled water, was evaluated. Two different treatment times were considered: T (when losartan is 100% degraded) and 2T (the double of time required to 100% remove the antihypertensive). Results for mineralization are presented in Figure 4A.

On the other hand, toxicity modifications exerted by treatment with ultrasound and UVC/H2O2 to the distilled water loaded with losartan were tested. Radish seeds (*Raphanus sativus*) were used as probe organisms. The growth index (GI) was used as the toxicity measure (phytotoxicity). Phytotoxicity was established at 2T of treatment for both processes (Figure 4B).

**Figure 4.** Extent of advanced oxidation treatments in distilled water. (**A**) Mineralization of losartan during the application of different processes; (**B**) evolution of the toxicity of losartan treated solutions against radish seeds. Note: the time was normalized concerning the time necessary to completely degrade losartan. Then, T is the time when losartan is 100% degraded, and 2T means the double of time required to 100% remove the antihypertensive. Experimental conditions as described in Figures 2 and 3.

#### **4. Discussion**

#### *4.1. Treatment of Fresh Urine Loaded with Losartan*

The Rk values for the ultrasound and UVC/H2O2 were 0.79 and 0.46, respectively (Table 2). It can be noted that ultrasound had the highest value for Rk; indicating that the losartan degradation through such process is affected at a low extent (21%) by urine matrix components. Meanwhile, for UVC/H2O2, the urine matrix presented a moderate inhibition (54%) of the antihypertensive elimination. These results suggest that the matrix components decreased the efficiency of the processes, which can be related to modifications of degradation routes. The explanations are presented in detail in the next subsections.

#### *4.2. Degradation Routes of Losartan (LOS) in the Di*ff*erent AOPs*

## 4.2.1. UVC/H2O2 Process

After the application of the individual components of the UVC/H2O2 system to LOS, it was found that hydrogen peroxide (even at 500 μM) did not induce significant removal of losartan (less than 5% elimination after 20 min of treatment). On the contrary, the treatment with the UVC light degraded ≈33.5% of the antihypertensive at 20 min of irradiation. The ultraviolet spectrum of losartan shows light absorption at 254 nm (Figure S3), which suggests that this molecule can be transformed by the UVC light. This is corroborated with the relative high photodegradation coefficient for losartan at UVC light (Cp, 123–190 L Einstein−<sup>1</sup> cm−<sup>1</sup> [32]). In fact, organic compounds having Cp values higher than 40 L Einstein−<sup>1</sup> cm−<sup>1</sup> can experience direct photolysis [33], which is currently related to the presence of aromatics rings, π-conjugated systems, and heteroatoms [34], as contained in the losartan structure (e.g., biphenyl, imidazole, and tetrazole). These aspects explain the losartan degradation by the UVC light. When losartan was treated by the complete UVC/H2O2 system, 65.7% of removal after 20 min was observed (Figure 2). The significant improvement of losartan elimination with the combination of hydrogen peroxide and UVC suggests the participation of radical species in the pollutant degradation. Indeed, as indicated earlier, the UVC/H2O2 process generates hydroxyl radical by homolytic rupture of peroxide by UVC light (Equation (1)). Hence, it can be indicated that in this process, the main action routes are the UVC photolysis and the attacks of hydroxyl radicals.

#### 4.2.2. Ultrasound Process

The sonochemical system has three reaction zones: the inner part of cavitation bubbles, where volatile molecules are pyrolyzed by high temperatures and pressures [35–37]; the interfacial region, where hydrophobic substances can react with the sonogenerated hydroxyl radical [38]; the solution bulk, where a small number of hydroxyl radical can react with hydrophilic compounds [39].

When losartan in distilled water was treated by high-frequency ultrasound (375 kHz and 106.3 W L<sup>−</sup>1), this process led to 97% of pollutant concentration reduction after 60 min of treatment (Figure 3). Since losartan is a nonvolatile compound, degradation by pyrolysis is negligible. Thus, the antihypertensive elimination would be associated with the attack of hydroxyl radicals. On the other hand, it is well-known that during the sonochemical process, hydrogen peroxide is formed by the combination of hydroxyl radicals (Equation (9)). Indeed, H2O2 production is an indicator of pollutant interaction with sonochemically formed HO• [40]. Thereby, to prove the participation of sonogenerated HO• in the pollutant degradation, the accumulation of H2O2 was determined. Figure 3 shows that the accumulations of H2O2 after 60 min of sonication in the absence and the presence of losartan were ≈180 and ≈110 μM, respectively.

$$\rm{HO}^{\bullet} + \rm{HO}^{\bullet} \rightarrow \rm{H}\_{2}\rm{O}\_{2} \tag{9}$$

The oxidation of losartan by the accumulated hydrogen peroxide was discarded because pollutant removal by H2O2 even at 500 μM was not observed (Figure 2). Then, the lower accumulation of hydrogen peroxide in the presence of losartan is an indicator of the reaction between the HO• and losartan. Moreover, due to the hydrophobic character of losartan (denoted by its high Log KOW value, which is >4.0) [24]), its degradation is expected to occur in the interfacial zone of the system [41] by the sonogenerated hydroxyl radicals.

#### 4.2.3. Understanding the Interference of Urine Matrix

Based on the degradation routes previously established, the interference of the urine matrix on the pharmaceutical degradation by the considered processes can be rationalized. During the application of UVC/H2O2 (which has both radical and photolytic routes), the antihypertensive removal was inhibited (by ≈55%) by the urine matrix components (see Rk value in Table 2). This was related to two aspects: the shielding of UVC light and scavenging of hydroxyl radicals. The shielding effect of the urine matrix was demonstrated through the evaluation of the only action of UVC light on losartan in both matrices (i.e., urine and distilled water), which showed a Rk value of 0.8 (Figure S4).

In turn, it is recognized that the inorganic anions such as chloride or bicarbonate, and organic substances like urea and acetate, present in the fresh urine, have relatively high rate constants with hydroxyl radicals (see Table 4), and as a result, they also affect the losartan degradation. It can be remarked the significant contribution of UVC photolysis to the degradation of losartan, as well as to the relative low interference of urine components for the light absorption (see Rk value for photolysis in Figure S4). Considering these findings, a scheme of losartan degradation by UVC/H2O2 was proposed in Figure 5a. It can be mentioned that the action of the photogenerated hydroxyl radicals induces transformations to losartan (such topic discussed below in Section 4.3), which is also schematized in this figure.

**Figure 5.** Scheme of degradation routes and interfering action of the urine components on the tested processes and generation of primary degradation products. (**a**) UVC/H2O2; (**b**) sonochemical treatment. Note: black arrows mean degradation routes and red arrows represent interfering action of the urine components.

**Table 4.** Rate constants of the reactions between hydroxyl radical and the diverse components of fresh urine.


In the case of the sonochemical process, for the rationalization of the low inhibitory effect of the urine matrix for the degradation of losartan (Figure 3), we must consider both the degradation route of losartan and the hydrophobic/hydrophilic nature of the substances in the matrix. The urine components are very hydrophilic, as evidenced by their Log Kow values (which are close to zero or negative, see Table S1). Thus, such components are mainly placed in the bulk of the solution and losartan is in the interfacial zone (where there is a high concentration of the sonogenerated HO•). Consequently, this pharmaceutical is slightly affected by the ions and/or organic compounds of the urine matrix (as schematized in Figure 5b). It should be indicated that the action of the sonogenerated hydroxyl radicals modifies the structure of losartan (such topic discussed below in Section 4.3), which is also schematized in this figure.

#### *4.3. Analysis of Losartan Susceptibility to Attacks by Radical Species*

The values of the hardness and softness for losartan indicate its high donor capacity. This is advantageous for attacks of the radicals to the pharmaceutical. Such behavior was also observed in the DAM (Figure S2), which shows that the losartan molecule has a better donor capacity concerning hydroxyl radical, hydroperoxyl radical, and superoxide anion radical. Besides, the computational analyses revealed that atoms on the imidazole moiety (15C, 18C, and 6N), aromatic rings (3C, 5C, 10C, and 12C), tetrazole (1C, 1N, 3N, and 4N), alcohol (O), and alkyl chain (21C) on losartan have the highest values for **fave** (this suggests that such regions on losartan are the most susceptible to transformations by radicals such as HO•). Indeed, we can mention that the atom with the highest Fukui function indices is more reactive to hydroxyl radical (the main degrading radical species in the tested AOPs). In the case of losartan, its C18 atom presents a **f**ave of 1.480, the highest value concerning all the atoms in the entire molecule. This behavior can be associated with the stabilization by resonance among the imidazole ring for the radical generated (Figure S5). In contrast, the attack of hydroxyl radical on the C1 atom in the tetrazole ring for the hydroxyl radical does not lead to such stabilization (Figure S6). In fact, the Fukui function indices for the tetrazole system are smaller than for the imidazole ring. Additionally, in a previous work from our research team, it was reported that for losartan molecule, the HOMO is located in the imidazole ring, whereas LUMO is on the tetrazole ring [46].

The primary products of losartan degradation in distilled water present a good agreement to the computational analysis on reactive regions of losartan (see Table S2 and Table 3). In the sonochemical treatment, three transformation products coming from imidazole ring rupture (TP1, TP2, and TP3), several isomers of biphenyl hydroxylation (TP4a-f), and one product of alcohol moiety oxidation (TP5) have been observed. Additionally, products of hydroxylation/oxidation of the alkyl chain on the antihypertensive have been found (TP6 and TP7, Table S3). Furthermore, analogous primary transformations of losartan induced by UVC/H2O2 and photo-Fenton were recently reported. Kaur and Dulova also found the formation of TP2 TP3, TP4, TP5, TP6, and TP7, in addition to TP8 (product of hydroxylation at the imidazole ring) and TP9 (transformation coming from a chlorine removal of the imidazole structure, see Table S3) [4]. In this sense, the region attackable by the hydroxyl radical, indicated by theoretical results correlates with the reported primary transformation products. This highlights the usefulness of computational analysis as a tool to establish the regions on losartan susceptible to degradation by the radicals from the AOPs.

#### *4.4. Mineralization and Toxicity Evolution*

The ability of the two processes to mineralize losartan in distilled water was tested, showing that none of these processes transformed losartan into carbon dioxide, water, and inorganic ions even at longer treatment times (2T) (Figure 4A). These results can be understood based on the degradation routes involved in each process. In the case of ultrasound, the attack of sonogenerated radicals in the interfacial zone (main route above described) led to hydroxylations/oxidations and rupture of pollutant molecules (see Table S3), which typically generates products more hydrophilic than the parent compound [47]. Hence, due to the hydrophilic nature of losartan degradation products, they are placed far away from the cavitation bubble, and consequently far away from the sonogenerated HO•. Thereby, the mineralization of losartan by ultrasound is not observed.

In the case of losartan elimination by the UVC/H2O2 process, it was noted the high participation of light (Section 4.2.1). Although UVC has a strong degrading ability through isomerizations or carbon-heteroatoms bond cleavages, its mineralizing power is very low [48]. On the other hand, although the mineralizing ability of HO• is widely recognized, under the tested conditions (moderate H2O2 concentration; i.e., 500 μM), the formed amount of such species seems to be not enough to reach some mineralization of losartan. Due to the nonmineralizing ability of ultrasound and UVC/H2O2 toward LOS, it was necessary to test the toxicity. To establish the potential reuse of the treated urine for irrigating crops; toxicity tests against radish seeds (*Raphanus sativus*) were performed (Figure 4B). It should be noted that the UVC/H2O2 process inhibits the germination of the seeds, this is associated with noxious substances generated in this system. In fact, recent research on losartan degradation by UVC/H2O2 process also evidenced that toxicity of solutions against *Daphnia magna* and *Desmodesmus subspicatus* augmented after the treatment [49].

Unlike UVC/H2O2, in the sonochemical process, the growth of the radish seeds increased with treatment (see 2T in Figure 4B). This suggests that the losartan by-products generated at large treatment periods of the sono-treatment are beneficial/less toxic for the indicator organism than the parent compound. Such results are coincident with several studies, which reported that the treatment of polluting substances using ultrasonic irradiation reduces the toxicity of solutions [50]. It must be indicated that although both UVC/H2O2 and sonochemistry can generate similar primary transformation products by hydroxyl radical attacks to losartan in distilled water (Section 4.3) at long treatment periods they may differ. Additionally, it must be considered that in the sonochemical process mainly acts hydroxyl radicals, whereas in the UVC/H2O2 both the radicals and UVC light are responsible for pollutant degradation (Section 4.2.2). Then, the observed differences in toxicity between both processes would be associated with their degradation mode. In the UVC/H2O2, the noxious substances could come from the action of UVC light on losartan or its primary degradation products (indeed, a previous work about the treatment of other emerging concern pollutants by UVC also reported the generation of toxic products for some organisms produced by this irradiation [32]).

#### **5. Conclusions**

It can be concluded that this research provides relevant information to understand the elimination of a representative pharmaceutical in fresh urine by two advanced oxidation processes having different nature (a photochemical treatment and other sonochemical system). The application of ultrasound and UVC/H2O2 individually, for the removal of the model pharmaceutical (antihypertensive losartan) in simulated fresh urine, showed that the sonochemical process was little affected by the urine matrix, exhibiting a high selectivity (Rk = 0.79) for the removal of losartan, which was related to degradation of the pharmaceutical at the interface of the cavitation bubble by the action of HO•. Meanwhile, the UVC/H2O2 process experienced moderate impacts of the matrix (Rk = 0.46) on the removal of losartan, because their degradation routes involved both photolysis and radical attacks. In turn, both ultrasound and UVC/H2O2 processes showed no mineralization of the pollutant in distilled water. Nevertheless, differently to UVC/H2O2, the sonochemical system transformed losartan into nonphytotoxic products (evidencing the potential reuse of sono-treated urine to irrigate crops). This illustrates the positive potentiality of ultrasound for the treatment of pharmaceuticals with hydrophobic characteristics in the simulated fresh urine. On the other hand, the computational analyses indicated that atoms on imidazole moiety on losartan were the most susceptible to transformations by the radical species. Such analysis was in good agreement with primary degradation products coming from UVC/H2O2 and sonochemical treatments, evidencing that theoretical methods are a useful tool to predict and rationalize the attacks of degrading species in the considered AOPs. Finally, it must be mentioned that losartan degradation was carried out at a pH value of 6.1; however, urine ranges from 4.5 to 8, and the modification of such parameter may change the results about the degradation of

pharmaceuticals by the AOPs. Thus, the effect of the urine pH should be evaluated in more detail in future studies.

**Supplementary Materials:** The following are available online at http://www.mdpi.com/2073-4441/12/12/3398/s1, Text S1: Determination of pseudo-first-order kinetic constants (k), Figure S1: Determination of the kinetic constants, Figure S2: Donor–acceptor diagram (DAM), Figure S3: Absorption spectra of losartan, Figure S4: Comparison of Rk for UVC/H2O2 and UVC alone, Figure S5: Resonance hybrid, Figure S6: Hydroxyl radical attack to the tetrazole ring, Table S1: Log KOW of losartan and the components of urine, Table S2: Primary transformation products of losartan during sonochemical treatment, Table S3: Additional products of losartan transformation by UVC/H2O2 and photo-Fenton.

**Author Contributions:** Conceptualization: R.A.T.-P. and E.A.S.-G.; methodology: J.F.G.-L. and E.A.S.-G.; software: Y.Á.-T.; formal analysis: J.F.G.-L. and E.A.S.-G.; investigation: J.F.G.-L.; resources: R.A.T.-P. and Y.Á.-T.; writing—original draft preparation: J.F.G.-L.; writing—review and editing: E.A.S.-G., R.A.T.-P. and Y.Á.-T.; supervision: Y.Á.-T. and R.A.T.-P.; project administration: Y.Á.-T. and R.A.T.-P.; funding acquisition: Y.Á.-T. and R.A.T.-P. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded by UNIVERSIDAD SANTIAGO DE CALI, project DGI No. 934 621120-G04; Universidad de Antioquia UdeA; MINCIENCIAS COLOMBIA (before named COLCIENCIAS), grant "Jóvenes investigadores por la Paz 2018- J19-18-1, 5381" and project No. 111577757323.

**Acknowledgments:** Researchers from Grupo de Investigación en Remediación Ambiental y Biocatálisis (GIRAB) thanks Universidad de Antioquia UdeA for the support provided through "PROGRAMA DE SOSTENIBILIDAD"; E. A. Serna-Galvis thanks MINCIENCIAS COLOMBIA for his PhD fellowship during July 2015–June 2019 (Convocation 647 de 2014).

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

## **References**


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## **Towards the Removal of Antibiotics Detected in Wastewaters in the POCTEFA Territory: Occurrence and TiO2 Photocatalytic Pilot-Scale Plant Performance**

**Samuel Moles 1,\*, Rosa Mosteo 1, Jairo Gómez 2, Joanna Szpunar 3, Sebastiano Gozzo 3, Juan R. Castillo <sup>4</sup> and María P. Ormad <sup>1</sup>**


Received: 25 April 2020; Accepted: 15 May 2020; Published: 20 May 2020

**Abstract:** This research aims to assess the presence of four antibiotic compounds detected in the influent and effluent of wastewater treatment plants (WWTPs) in the POCTEFA territory (north of Spain and south of France) during the period of 2018–2019, and to relate the removal of antibiotic compounds with the processes used in the WWTPs. The performance of a photocatalytic TiO2/UV-VIS pilot-scale plant was then evaluated for the degradation of selected antibiotics previously detected in urban treated effluent. The main results reflect that azithromycin had the highest mass loadings (11.3 g/day per 1000 inhabitants) in the influent of one of the selected WWTPs. The results also show considerable differences in the extent of antibiotics removal in WWTPs ranging from 100% for sulfadiazine to practically 0% for trimethoprim. Finally, the photocatalytic TiO2/UV-VIS pilot-scale plant achieved the removal of the four antibiotics after 240 min of treatment from 78%–80% for trimethoprim and enrofloxacin, up to 100% for amoxicillin, sulfadiazine and azithromycin. The catalyst recovery via mechanical coagulation–flocculation–decantation was almost total. The Ti concentration in the effluent of the TiO2/UV-VIS pilot-scale plant was lower than 0.1% (w/w), and its release into the environment was subsequently minimized.

**Keywords:** antibiotics; wastewater; removal efficiency; photocatalysis; slurry reactor

## **Highlights:**


## **1. Introduction**

The problem of the presence of pharmaceutical compounds in wastewater has recently become a matter for concern, not only in terms of human health, but also for the preservation of the environment [1]. Antibiotics are an important group of medicines suitable for the treatment of human infections and in

veterinary medicine. Many of them are not completely metabolized by the body so between 30%–90% are excreted and, as a result, they end up in wastewater [2,3]. The main difference with other organic pollutants is that antibiotics represent a potential risk if they are released into the environment because they a direct biological action on microorganisms, generating antimicrobial-resistant bacteria (ARB). As suggested by other authors [4], ARB of animal origin can also be transmitted to humans.

Several studies have pointed out that conventional wastewater treatment plants (WWTPs) are not designed to remove pharmaceuticals, metabolites or drugs [3–5]. Besides urban plants and hospitals, slaughterhouses also generate wastewaters which are not usually incorporated into sewage systems. As a result, they represent a significant source of antibiotics released into the environment. The European Surveillance of Veterinary Antimicrobial Consumption (ESVAC) collects information on how antimicrobial medicines are used in animals across the European Union (EU). According to their latest report [6], Spain is known to be one of the main consumers of veterinary drugs in the EU. As a result, many studies have monitored the occurrence of the most commonly administered pharmaceuticals in urban wastewater, groundwaters and surface water in Spain. The literature informs that concentrations of antibiotics from ng/L to μg/L [7–10]. sulfonamides [11,12], trimethoprim [13,14], β-lactams [15,16], fluoroquinolones [17,18] and macrolides [19–21] all represent a potential risk for the environment. Consequently, a representative antibiotic from each one these groups was analyzed in this research work: sulfadiazine (veterinary use, sulfonamide), trimethoprim (human and veterinary use, trimethoprim), amoxicillin (human and veterinary use, β-lactam), enrofloxacin (veterinary use, fluoroquinolone) and azithromycin (human use, macrolide).

Among the various water treatment techniques used to eliminate these drugs, advanced oxidation processes (AOPs) are suitable for antibiotic degradation [22–24]. Other techniques, such as activated carbon or reverse osmosis, only transfer the contaminants from one phase to another without degrading them. Nevertheless, photocatalysis has been demonstrated to be effective for wastewater treatment as it is cost-effective and simultaneously oxidizes various organic contaminants into inorganic compounds, water and carbon dioxide, and pathogenic microorganisms [24]. Several semiconductors are used in photocatalysis, such as TiO2, ZnO, and CdS. Among these, TiO2 has been widely used because of its strong oxidizing power, availability, nontoxicity and price. The catalyst can be employed either in a colloidal or in an immobilized form. Although immobilizing the catalyst might improve the catalyst recovery, immobilized systems show lower degradation efficiencies compared to the suspended counterpart because of a reduction in the surface area [25,26]. Whenever the nanoparticles are dispersed in an aqueous medium, the depth of penetration of the radiation is limited because of absorption/scattering by the catalyst nanoparticles and the dissolved organic species. These systems also require an additional separation process to prevent Ti emission to the environment, and this stage induces further costs [27].

Pilot-scale plants represent the previous step to industrial scale plant. Literature suggest how to operate at lab-scale photocatalytic systems [28–30]. However, design and operation with a pilot-scale plant are necessary to determine how to deal with possible operational problems and establish the optimal operational parameters for real scale operation. Pilot-scale plants also allow one to determine if the real scale process would be economically feasible. Some studies about the application of TiO2 photocatalysis in wastewater have been reported [13,14,28–31]. However, these research works do not focus on the simultaneous antibiotic removal by a TiO2 photocatalysis pilot-scale plant applied to real wastewater.

The aim of this research work is to evaluate the presence of selected antibiotic compounds in the inlet and outlet of four WWTPs for the period of 2018–2019. Another objective is to treat selected antibiotics present in real wastewater in a photocatalytic plant by applying TiO2 in suspension. Finally, the Ti concentration in the final effluent was controlled, to prevent Ti emission to the environment.

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

#### *2.1. Site Description and Sample Collection*

This research is focused on four WWTPs located in the POCTEFA territory (north of Spain and south of France). These WWTPs are designed to treat urban wastewater of domestic and industrial origins. Table 1 shows the main characteristics of selected WWTPs.

Samples were collected in four sampling campaigns for two years (in the spring and autumn of 2018 and 2019). The inlet and outlet of each WWTP were selected as sampling points to estimate the current removal performance of selected antibiotics in the four WWTPs, aiming to compare the different treatment lines.


**Table 1.** Main characteristics of each wastewater treatment plants (WWTP).

The sampling was carried out following the EPA method 1694 for the analysis of pharmaceuticals and personal care products in water, soil, sediment, and biosolids by liquid chromatography tandem mass spectrometry (LC/MS/MS) [32]. Amber glass bottles were used to collect 1000 mL samples which were stored under refrigeration at 4 ◦C. The bottles were fully filled to avoid the presence of air and properly sealed by means of a PTFE seal. According to EPA 1694, the filtration of the samples is necessary in order to remove suspension solids. Two filtration steps were carried out prior to analysis using glass fiber filters (1.6 μm, supplied by GVS) for the first filtration stage and nylon filters (0.45 μm, supplied by GVS) for the second stage, as suggested in other research works [33–35].

Mass loadings of the antibiotics were calculated in each sampling period as the product of the individual concentration of each antibiotic in the samples and the daily flow rate of each WWTP.

Removal efficiencies of the target compounds were determined as the difference between the inlet mass loading and the outlet mass loading divided by the inlet mass loading and expressed as a percentage (Equation (1)).

$$\text{Removal} \, f \,\text{icincy } \left( \% \right) : \frac{\left( m\_{\inf} - m\_{\inf} \right)}{m\_{\inf}} \times 100 \tag{1}$$

#### *2.2. Antibiotic Characterization*

The quantification of the concentration of antibiotics was carried out via HPLC/MS/MS. Samples were centrifuged for 10 min at 13,000 rpm in Eppendorf tubes and then diluted 40-fold with 0.1% formic acid/MeOH/ACN (80%/10%/10%) before LC-MS/MS analysis. Chromatographic separations were carried out on an Ultra Performance Liquid Chromatography (UPLC) Ultimate 3000 RSLC system (Thermo Fisher Scientific). The column used was an Accucore C18 100 × 2.1 mm, 2.6 μm (Thermo Fisher Scientific). The mobile phases were A H2O 0.4% formic acid + 5 mM ammonium formate and B MeOH/ACN 1:1 (v/v). A 20μL sample aliquot was injected. Detection was performed on a Q Exactive Plus (Thermo Fisher Scientific) mass spectrometer operated in the targeted single ion monitoring

(SIM) positive mode with a resolution of 70,000. External calibration was used for quantification and validated by standard additions for selected samples; the samples were prepared and analyzed in triplicate. The limits of detection and quantification of each antibiotic are featured in Table 2.


**Table 2.** Limits of detection and quantification of selected antibiotics.

The four antibiotics investigated in this research work were selected according to their potential risk for the environment and reported occurrence [11–21]. All of them are representative human-use and veterinary-use antibiotics belonging to the main antibiotic groups. The standards were supplied by Sigma-Aldrich. Some characteristics of the selected antibiotic compounds are included in Table 3.

**Table 3.** Selected antibiotics and chemical information.


#### *2.3. Total Ti Assessment in the E*ffl*uent*

The effluent from the pilot-scale plant might contain some Ti which would then be emitted to the environment. The Ti concentration was quantified by an Inductively Coupled Plasma Mass Spectrometry (ICP/MS) ELAN DRC-e, PerkinElmer, Toronto, Canada. A discrete volume sample (100 μL) was injected through a six-way valve, and the carrier was delivered directly to the nebulizer of the spectrometer. A glass concentric slurry nebulizer with a cyclonic spray chamber (Glass Expansion, Melbourne, Australia) was used. Default values were used for the rest of the instrumental parameters. The quantification of TiO2 was based on monitoring the ICP-MS signal of the isotope 49Ti, using 74Ge as an internal standard. From an on-line calibration with an ionic titanium standard diluted in nitric acid (1%), intensity signals from the ICP-MS for samples were transformed into mass values by integrating the area of the transient signals obtained. All samples were injected in triplicate. The limit of detection of the method was established at 0.81 μg/L and the limit of quantification at 2.70 μg/L.

#### *2.4. Photocatalytic Oxidation Experiment*

The oxidation assays were carried out in the facility detailed in Figures 1 and 2. First, a 1 m<sup>3</sup> storage tank provided with a stirrer was filled with the water sample. The solution was then pumped (8–16 L/min) to a 0.1 m<sup>3</sup> mixer decanter where the catalyst was stored. The mixer decanter was provided with a stirrer to mix the influent with the catalyst. Subsequently, the mixture was placed in four identical slurry reactors. These reactors are made of aluminium because this material is known to have a high degree of light reflection. The reactors had a volume of 17 L and were provided with a UVA lamp (330–390 nm) of 40W. When the reactors were completely full, the UV/vis lamp in each reactor was turned on and stirring by means of compressed air took place. After treatment in the reactors, the treated water was pumped again to the decanter where a mechanical coagulation–flocculation–decantation treatment (CFD) was applied. Two steps take place in the separation process, coagulation 200 rpm

during 5 min, flocculation 40 rpm during 25 min. and 90 min of decantation, resulting in a 120 min total process. Coagulant was not added to the mechanical CFD separation process.

The effluent (clarified phase) was generated and the catalyst remained in the decanter for the next cycle.

**Figure 2.** Pilot-scale plant process images.

The sample was prepared by the addition of individual concentrations of 1 mg/L of each antibiotic (amoxicillin, azithromycin, enrofloxacin, trimethoprim and sulfadiazine) simultaneously in the effluent from WWTP2. The physicochemical characteristics of this wastewater were pH = 7.6, DQO = 90 mg/L, Turbidity =11 NTU. The catalyst was applied in suspension in a concentration of 1 g/L of TiO2 FN2 (supplied by Levenger S.L.). Radiation per unity of volume was 0.3 W/L in each reactor. The temperature

ranged from 14 ◦C to 20 ◦C during the experiment. The experiments were conducted twice, and their average is represented in the results. The individual antibiotic concentration was quantified via HPLC/MS/MS during 240 min of photocatalytic treatment following the procedure described in Section 2.2.

#### **3. Results**

#### *3.1. Occurrence of Target Antibiotics in Urban Wastewaters*

Figure 3 shows the mass loading of the selected antibiotics at the four WWTPs during 2018 and 2019. It should be noted that atypical points are not represented. However, they are all available in Tables S1 and S2 of the Supplementary Material. The tables show that the highest mass loading corresponds to azithromycin. This human-use antibiotic presented an average load of around 925 mg/day per 1000 inhabitants and reached a maximum load of 11,332 mg/day per 1000 inhabitants in WWTP4. The azithromycin concentration increased over the four sampling campaigns in all the WWTPs. This fact might be attributed to a major increase in the use of azithromycin [36]. Enrofloxacin (fluoroquinolone group) also followed the same trend: higher loads were detected over the campaigns analyzed in this study. This veterinary-use antibiotic showed median mass loadings of around 200 mg/day per 1000 inhabitants and a maximum of 4329 mg/day per 1000 inhabitants. The enrofloxacin mass load was higher in the spring of both years, while it was rarely found in autumn, and always in lower loads, reflecting a seasonal use of this fluoroquinolone. The seasonal appearance of enrofloxacin has also been recently reported in another research work [37].

**Figure 3.** Boxplots of each antibiotic indicating total mass loading values: (**a**) azithromycin, (**b**) enrofloxacin, (**c**) sulfadiazine (**d**) trimethoprim.

By contrast, the average mass loadings of the other two antibiotics (trimethoprim and sulfadiazine) ranged from 35 mg/day per 1000 inhabitants for trimethoprim, to undetectable levels for sulfadiazine. Comparing the median mass loadings of trimethoprim and sulfadiazine with enrofloxacin and azithromycin reveals a difference greater than one order of magnitude.

Amoxicillin was not found in any sample (influent and effluent of the WWTPs). This fact could be attributed to the low stability of amoxicillin and the subsequent generation of degradation products such as amoxicillin penicilloic acid or amoxicillin-diketopiperazine-2', 5', as is suggested in the literature [38–40]. These degradation products were found in subsequent campaigns carried out in the same sampling points.

Finally, Tables S1 and S2 show that the reported load of each antibiotic varies significantly, by more than one order of magnitude, between the years and seasons. This trend could be related to the differences in rainfall patterns between the two years: high flows in particular were reported in spring 2018 and in autumn 2019 in both the Ebro River basin and the Cantabrico Occidental River basin. However, the total mass loading of the four antibiotics was relatively higher in the spring: SPRING = 69.4 g/day per 1000 inhabitants versus AUTUMN = 13.6 g/day per 1000 inhabitants.

#### *3.2. Removal E*ffi*ciency of Selected WWTPs*

The results of the removal efficiency of the antibiotics in each WWTP is shown in Figure 2. These results are also fully detailed in Table S3. Figure 4 shows that the removal efficiencies range from 2%–100%, demonstrating the fact that WWTPs can partially or almost totally remove the target antibiotics.

**Figure 4.** Wastewater treatment plants' removal efficiency for each antibiotic in 2018–2019 (AZI = azithromycin, ENR = enrofloxacin, SDZ = sulfadiazine, TMT = trimethoprim).

The results also suggest a significant variation for each antibiotic in each WWTP, indicating that the removal efficiency strongly depends not only on the specific matrix but also on the season and associated flow. This variation might also be attributed to the different physicochemical characteristics of the antibiotics, such as the degradation rates in the water, organic carbon–water partition coefficients or acid dissociation constants and water solubilities. More precisely, the results suggest that sulfadiazine is the antibiotic with the highest removal efficiency in every case. This trend may be due to the low mass loadings of this antibiotic, which make its removal easier [41]. Enrofloxacin also presents a high removal efficiency after the wastewater treatment, reaching 100% in several samples. By contrast, azithromycin and trimethoprim showed lower removal efficiencies in all the WWTPs. It should be noted that azithromycin had the highest mass loadings, more than one order of magnitude greater than the other antibiotics. It might be also attributed to the fact that azithromycin and trimethoprim

have similar carbon–water partition coefficients [42]. Some studies suggest that biological processes, which are present in the four WWTPs, can remove fluoroquinolones effectively, while trimethoprim is more difficult to remove by means of biological treatments [43].

Comparing the different treatments of each WWTP, it can be observed that the facilities provided with a trickling filter (WWTP2 and WWTP3) showed higher removal efficiencies of the target antibiotics. This is consistent with some studies which demonstrate that trickling filters can remove antibiotics and other pharmaceuticals as well as personal care products [44–46]. WWTP4 showed the highest average antibiotic removal in most samples. However, it is important to note that this WWTP is in the Cantabrico River basin where there is higher rainfall than in the Ebro River Basin. As a result, the selected antibiotics had lower mass loadings in WWTP4.

#### *3.3. Photocatalytic Oxidation of Antibiotics*

Figure 5 shows the performance of the photocatalytic assays for the simultaneous oxidation of amoxicillin, enrofloxacin, sulfadiazine, trimethoprim and azithromycin during 4 h of treatment. The results show that in only 30 min of treatment, azithromycin and amoxicillin reached a degradation rate of 85% and 75%, respectively. Moreover, after 120 min of treatment, both antibiotics were completely removed from the wastewater. Amoxicillin was previously reported to be easily removed from waters by TiO2 photocatalysis applied to the isolated compound at lab scale [47]. However, sulfadiazine shows a slower degradation rate, achieving degradation yields of 25% and 100% in 30 min and 240 min, respectively. In contrast, enrofloxacin and trimethoprim were not completely removed from the wastewater after the treatment. The degradation rate of enrofloxacin was relatively high at the beginning of the process (degradation yield of 50% in 30 min of treatment), but complete degradation was not achieved by the end of the treatment. A similar trend has been reported in other research work at lab scale [31]. Finally, trimethoprim showed the slowest initial degradation rate during the first 120 min, as reported in other studies at lab scale [45], while its degradation yield at the end of the treatment was close to 70%.

**Figure 5.** TiO2 photocatalysis degradation yields of each antibiotic in WWTPE during 240 min of treatment (individual antibiotic initial concentration = 1 mg/L, TiO2 initial concentration = 1 g/L, radiation per unity of volume = 0.3 W/L). AMX =amoxicillin, AZI = azithromycin, ENR = enrofloxacin, SDZ = sulfadiazine, TMT = trimethoprim.

#### *3.4. Ti Assessment in the E*ffl*uent of the Photocatalytic Treatment Plant*

In order to determine whether Ti is released into the environment, a quantification of the Ti concentration in the effluent was carried out applying a coagulation–flocculation–decantation treatment. TiO2 has been demonstrated to be effective after several cycles of photocatalytic treatment for the degradation of pharmaceuticals [48], so it is important to recover it effectively. Fortunately, the results of

the Ti assessment suggest that no more than 0.1% of the initial Ti concentration remained in the effluent. This reflects the fact that recovery by means of mechanical coagulation–flocculation–decantation treatment is quite efficient for recovering TiO2 when used in suspension [48].

## **4. Conclusions**

This research work evaluates the behavior of four antibiotic compounds in four different WWTPs located in the north of Spain. The mass loadings of amoxicillin, enrofloxacin, sulfadiazine, trimethoprim and azithromycin were analyzed in the influent and effluent of the WWTPs. The performance of a TiO2 photocatalytic treatment plant applied to the simultaneous removal of the antibiotics from real urban treated water was evaluated. This showed that the technology can be used to totally remove some of the selected antibiotics at slightly higher concentrations than those commonly found in wastewaters. The conclusions can be summarized as follows:


**Supplementary Materials:** The following are available online at http://www.mdpi.com/2073-4441/12/5/1453/s1, Table S1: Spring and autumn 2018 Load selected antibiotics in mg/day/1000 inhabitants. Table S2: Spring and autumn 2019 Load selected antibiotics in mg/day/1000 inhabitants. Table S3: Removal efficiencies of the different WWTPs for selected antibiotics.

**Author Contributions:** Planning photocatalysis experimentation, R.M. and S.M.; general experimental work coordinator, M.P.O.; LC/MS/MS analysis, S.G. (technician) and J.S. (responsible); design and construction of pilot-scale plant, J.G. and S.M.; responsible of nanoparticles detection in the effluent, J.R.C.; experimental photocatalysis technician and principal author, S.M. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded by DGA\_FSE Research Team "Water and Environmental Health" Ref: B43-20R and cofounded by EFA 183/16/OUTBIOTICS.

**Acknowledgments:** This work was financed by DGA\_FSE Research Team "Water and Environmental Health" Ref: B43-20R in the framework of the project EFA 183/16/OUTBIOTICS, Program Interreg-POCTEFA 2014-2020, funded by FEDER.

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

## **References**


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