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Advanced oxidation processes
AOPs were first proposed in the 1980s for potable water treatment and later were widely applied for the treatment of different wastewaters [32,33]. AOPs mainly rely on the highly reactive oxygen species (ROS) generated from photocatalysis, electrochemistry, or peroxides, which can initiate the oxidation reaction in water to degrade micropollutants and damage microorganisms [34]. For chemical oxidation based-AOPs, ozone (O3), hydrogen peroxide (H2O2), and persulfate (PS) are the most commonly used oxidant precursors [27,33]. Various oxidants activation methods, such as heat, ultraviolet, alkaline, transition metals, and carbonaceous materials, have been applied to generate ROS more efficiently [33,35,36]. A possible classification of AOPs includes two groups: homogeneous AOPs and heterogeneous AOPs.
Homogeneous AOPs
Here, the homogeneous AOPs employ O3, H2O2, and persulfate, including peroxydisulfate (PDS, S2O82-) and peroxymonosulfate (PMS, HSO5-)) are mainly discussed.
Ozone-based AOPs
Ozone is a strong oxidant with a high redox potential (E0 = 2.1 V) and can oxidize inorganic and organic substances, inactivate pathogens and ARB&Gs from water [22,29,37–39]. Ozonation of contaminants are processed via two ways: direct reaction by O3 molecule, which preferentially reacts with the ionized and dissociated form of organic compounds, and indirect oxidation by hydroxy radical (•OH), which was generated from O3 (Eq. (1)) to initiate the indiscriminate oxidation [33,38].
Ozone act as a disinfectant and as an oxidant has been applied at full-scale as tertiary treatment of urban wastewater to remove micropollutants and microorganisms in many industrialized countries before discharge into the environment [38]. Switzerland is among the first to implement ozonation in full-scale and plans to upgrade about 100 out of 700 wastewater treatment plants [40]. Generally, an ozonation system is implemented after the activated sludge treatment where dissolved organic carbon (DOC) is lowest to minimize the ozone consumption, and followed by at least a biological filtration step (e.g., sand, biological activated carbon filtration) to remove biodegradable organic carbons (BDOC) generated as a consequence of the reaction of ozone and dissolved organic matters (DOM) and to prevent the regrowth of pathogenic bacteria after ozonation [38].
Micropollutant abatement is one important objective of ozone in wastewater treatment. Rizzo et al. recently reviewed the abatement kinetics of CECs at commonly applied specific ozone doses (0.4–0.6 g O3/g DOC) and classified them into three categories, as shown in Table 2 [41]. CECs of group A, including azithromycin, bisphenol-A, carbamazepine, ciprofloxacin, clarithromycin, diclofenac, erythromycin, metoprolol, sulfamethoxazole, and the hormones 17‑alpha‑ethinylestradiol and 17‑beta‑estradiol, that predominantly react with ozone because their electron-rich moieties highly react with ozone [41]. The selectivity of the ozone reactivity is still decisive for the abatement of CECs of group B, including benzotriazole, bezafibrate, mecoprop and methylenbenzotriazole [41].
However, ozonation may result in the oxidation by-products deriving from the wastewater matrix, such as N‑nitrosodimethylamine (NDMA) and bromate, and the formation of transformation products (TPs) resulting from the oxidation of CECs, and these TPs maybe not susceptible to ozone and can be even more toxic than the initial compounds [29,41]. The subsequent biological filter treatment step can be used to reduce the effluent toxicity.
Overall, the ozonation combined with a biological filter can be a good option for large UWTPs to simultaneously effectively degrade a wide range of CECs as well as limit the toxicity of its TPs and remove the abundance of pathogens as well as the ARB&Gs. For example, for a large UWTP (10,000 to 500,000 person equivalents) and DOC contents (6 to 20 g DOC m−3), considering the capital investment and annual operation cost including post-filtration step, the total costs of ozonation were estimated to range between 0.05 and 0.15 € m−3, depending on plant size and secondary effluent quality [38]. However, ozonation is generally only used at medium to large-sized plants after at least secondary treatment and is not economical for wastewater with high levels of
organic carbon [27,44]. However, it should be noted that for wastewater reuse in irrigation, in order to save transportation costs, the decentralized and small-sized UWTPs near farmland in the suburbs or rural areas are the priority, thus ozonation is not considered in this study.
H2O2, PMS, and PDS-based homogeneous AOPs have been studied for water and wastewater treatment over past decades and exhibited effective degradation of a broad range of micropollutants, inactivation of pathogens, and reduction of ARB&Gs from urban wastewater [5,19,27,45–51]. Reactive species, such as free •OH, sulfate (SO4•−), and superoxide ion radicals (•O2−) are usually believed as the primary ROSs and can be produced by different activation methods, such as UV, transition metal ions, complexing agents.
UV/ H2O2 and UV/PS
Conventional UV lamps use of mercury have widely applied for urban wastewater disinfection, which can effectively inhibit cell replication by damaging their DNA [17]. The application of light emitting diodes (LED) make UV-C LED very promising for water disinfection since they are much cheaper, smaller, lighter, and less fragile than traditional mercury vapor lamps [5,52]. Single UV can effectively degrade some organic contaminants but at high dose (> 10, 000 mJ/cm2), such as high abatement efficiency (>85%) of sulfonamides, quinolones, NDMA, herbicides and pesticides from water [5,53,54], however, it exhibited low removal capacity for most CECs [54,55].
In combination with peroxides (e.g., H2O2, PS), UV-based AOPs leading to the formation of and •OH and SO4•− through the fission of O-O bond by energy input of UV as presented in Eq. (5) – (7), or through photolysis of water (Eq. (8)) at a UV less than 242 nm [33,55], can significantly enhance the micropollutants degradation and pathogens inactivation efficiency in water [56–58].
However, for wastewater application, since turbidity and TSS can drastically decrease UV efficiency, UV irradiation requires extensive secondary treatment and conventional depth filtration (e.g., sand filtration) to reach a recommended water transmittance value over 50% and probably it is not adapted to open treatment systems (e.g., lagoons or constructed wetlands) where green colored microscopic algae can grow [5,15]. More importantly, the efficiency of those technologies is decreased in wastewater since the radicals are easily scavenged by DOM, and the toxic and sensory-unpleasant TPs are commonly generated during UV-AOPs of organic micropollutants and DOM [59,60]. These are all potential barriers to their practical applications.
Photo Fenton and Fenton-like AOPs
Compared with energy input activation methods (e.g., UV), transition metal ions, such as ferrous ions (Fe2+), exhibited the advantages of low energy consumption and strong activity. The classical Fenton reaction describes the activation H2O2 of by dissolving Fe2+ in solution to generate •OH (Eq. (9)) [61]. Fe2+ is also efficient for PMS and PDS activation (Eq. (10)-(12)) [62].
The formed ferric ion (Fe3+) is a major limitation of homogeneous Fe-based AOPs because it starts to precipitate in the form of ferric hydroxides when pH is higher than 3 [62]. Even though the regeneration of Fe2+ occurred in the system, noted as Eq. (13), but its reaction rate is much slower than the oxidation of Fe2+, and the steady accumulation of Fe3+ can cause heavy iron sludge production.
Therefore, homogeneous Fe-based AOPs are limited for practical application because of the strict low pH operation condition and heavy iron sludge production [51,61]. To overcome the disadvantages of the conventional Fenton process, apart from using the appropriate ratio of iron ions to oxidants, adding Fe solution sequentially, photo-Fenton processes were widely studied, which can enhance the reduction of Fe3+ to Fe2+ and improve the generation of generate •OH, as presented in Eq. (5), (9), and (14) [33,62– 65].
With the addition of irradiation, less H2O2 and Fe2+ consumption leading less iron sludge production compared with the conventional Fenton process [63]. Besides, photo-Fenton can use sunlight instead of UV light to save energy costs and to make the process sustainable [30,63]. Nevertheless, the strict acidic working pH problem of photo-Fenton processes still exists. Under these circumstances, to take advantage of the good performance of the photo-Fenton process, one possible way to improve this process to work at wider pH is to introduce chelating agents (L), such as oxalate, ethylenediamine-N, N’-disuccinic acid (EDDS), to the solar photo-Fenton system to solubilize Fe3+ to form Fe(III) complex, as noted in Eq. (15), which can absorb UV–vis light and undergo photochemical reductions leading to Fe2+ ions to realize the Fe redox cycle, as presented in Eq. (16) [64].
For example, Maniakova et al., recently reported efficient simultaneous removal CECs and pathogens from urban wastewater by solar driven photo-Fenton with the addition of EDDS [30]. However, the cost of the chelator, the biodegradability and ecotoxicity of the complex formed, and the strong weather variation effects of solar photo-Fenton processes should be considered. In addition, one limitation is the low volume able to be treated by this photo-Fenton even when using raceway open reactors [30,65].
Another method to overcome the processing and economic constraints associated with homogeneous Fenton and Fenton-like oxidation processes, such as recycling difficulty, acidic working pH, and iron sludge production, is to make use of recyclable solid catalysts instead of iron metal ions [63], which defined as heterogeneous Fenton and Fenton-like oxidation processes (HFOPs).
Heterogeneous Fenton-like oxidation processes
HFOPs using transition metal oxides, such as iron oxides, copper oxides (CuO), and manganese (MnO2), as catalysts have drawn increasing attention recently to degrade organic contaminants and inactivate pathogens effectively due to their low cost, relatively low toxicity levels, high catalytic activity and easy methods for recovery and scale-up [27,66–69]. However, there has been a long-term debate on the peroxide’s activation pathways and the produced intermediate oxidants (e.g., radicals, singlet oxygen, catalysts – peroxide complexes, high valent metal ions) in HFOPs [66,70,71].
Radical-based systems
•OH (E0 (•OH) = 1.9−2.7 V) and SO4•− (E0 (SO4•−) = 2.5−3.1 V) with high redox potential are long recognized as the responsible reactive species in HFOPs to degrade organic pollutants and inactivate pathogens [27,66,67,72]. Using iron oxides as examples, the possible surface bound radicals-based activation pathways (Figure 3) in HFOPs are similar to homogeneous peroxide activation (Eq. (9)-(11)) [73–75].
•OH is known as non-selective species that can degrade a wide range of organic contaminants, while SO4•− has a higher redox potential of key radical species (E0 (SO4•−) = 2.5−3.1 V > E0 (•OH) = 1.9−2.7 V), a longer half time (3-4 ×10-5 s > 2 ×10-8 s), a wider working pH range (pH = 2.0-8.0), and higher selectivity to the electron-rich compounds via a single electron transfer pathway [36,76]. However, both radicals can be quenched by organic matter, moreover, the oxidation by SO4•− favors the direct electron abstraction and can transform some anions (e.g., Cl- and NO2-) into their corresponding radicals potentially leading to troublesome chlorinated and nitrated TPs [48,77]. It should be noted that the HFOPs with the purpose to remove micropollutants, pathogens, and antibiotic resistance are widely applied as the last barrier in the wastewater reclamation train after carbon and nutrients removals. However, in irrigation application, what is needed is a technology that can work in presence of high contents of dissolved organic matter and nutrients because both are valuable agronomic parameters to be preserved [4]. Thus, HFOPs relying on radical-based mechanisms undergo efficiency losses in organic-rich wastewater such as urban wastewater due to the competing organic/inorganic constituents of water [48] and due to the difficulty in the regeneration of active sites on the surface of the catalyst due to the energetically unfavorable metal ion redox cycle [69].
In contrast to radical-based processes, the non-radical oxidative pathway is much more selective than radicals, accounting for the only removal of electron-rich organic contaminants (e.g., phenols and anilines) [78,79]. In addition, the mild non-radical based processes consume less amount of oxidant and are less influenced by the competing organic/inorganic constituent in urban wastewater, thereby limiting the formation of hazardous TPs.
Non-radical-based systems
The non-radical pathway including (1) oxidant surface metastable complexes over the catalyst, (2) singlet oxygen (1O2), and (3) the high valent metal-oxo oxidant species have been proposed for HFOPs [36,69,70,80].
(1) Surface activated complexes
Surface activated complexes result from the interaction between catalyst and oxidant through an outer-sphere surface complexation [78,79]. Surface activated complexes have not been reported on single iron oxides yet but on FeMn composite oxide and layered CuFe oxide [81,82]. As shown in Figure 4a, the activated PDS acted as a two-electrons transfer oxidant that directly accepts electrons from organic pollutants through an electron shuttle material leading to sulfate anions (SO42-) release [78,79]. To the best of our knowledge, the investigation of surface activated complexes has only been reported in the field of organic contaminants degradation.
(2) Single oxygen
Certain recent studies have reported the occurrence of 1O2 as another type of non-radical process as previously reported in the HFOPs [83–86]. •O2− might function as a precursor for the generation of 1O2 through •O2−oxidation or disproportionation reactions, as shown in Eq. (17) and Figure 4b.
The 1O2-based systems are very known for their resistance to background substances in the water matrix and for inactivating a wide range of bacteria and viruses as a potential disinfectant due to its oxidation capacity of proteins and DNA [87–89]. However, 1O2 based systems have been regarded to be poorly efficient because 1O2 is rapidly quenched by water [90,91].
(3) High valent metal ions
High valent metal ions, such as ferryl ions (Fe(IV)) and cupryl ions (Cu(III)) have been also recently suggested and proved to be the working oxidant in homogeneous (Figure 5a and b) and heterogeneous (Figure 5c and d) Fenton and Fenton-like processes [92– 96].
As shown in Figure 5a, the generated ferry oxo (FeIVO2+) is a powerful oxidant that can react with inorganic ions and organic compounds (e.g., aromatic substrates and aliphatic alcohols) through electron or atom (hydride, hydrogen, or oxygen) transfer, it can also oxidize sulfoxide to sulfone, which is the specific reaction associated to Fe(IV) and Cu(III) [96]. The generation of Cu(III) in a highly electron deficient state offers the highest redox potential for PMS activation (E0Cu(III)/Cu(II) = 2.3 V) that can oxidize electron-rich organic contaminants [94,99] or decompose into •OH, depending on the water pH value [100]. Fe(IV) and Cu(III) have also been identified as the working species in HFOPs during PS activation by using Fe(III) doped g-C3N4 and CuO-Fe3O4 [97,98]. However, it has not been reported yet on single iron oxide or copper oxide. In addition to the electron-rich compounds degradation capacity, Cu(III) also exhibited excellent antimicrobial activity, especially against viruses [101,102].
Therefore, this study aimed to contemplate the possibility to implement a HFOP relying on more selective reactive species than sulfate and hydroxyl radicals, which might work in organic-rich wastewater for urban wastewater treatment in terms of micro-pollutant degradation, pathogen inactivation, and reduction of antibiotic resistance. CuO was selected as the heterogeneous catalyst because it has a similar redox behavior to iron oxide, but has been less studied and has more diverse activation pathways [69]. In addition, CuO in form of nanoparticles exhibits effective antimicrobial activity [103– 106]. PDS was selected as the working oxidant taking into consideration its lower costs of storage and transportation due to the availability and stability of peroxydisulfate salts (0.74 $/kg of PDS vs. 1.5 $/kg of H2O2 vs. 2.2 $/kg of PMS) [48].
Table of contents :
Chapter 1. Introduction and context
1. Water scarcity
2. Wastewater reuse in irrigation
2.1. Guidelines and regulations
2.2. Urban wastewater treatment technologies
3. Advanced oxidation processes
3.1. Homogeneous AOPs
3.1.1. Ozone-based AOPs
3.1.2. UV/H2O2 and UV/PS
3.1.3. Photo Fenton and Fenton-like AOPs
3.2. Heterogeneous Fenton-like oxidation processes
3.2.1. Radical-based systems
3.2.2. Non-radical-based systems
4. Fixed-bed column
5. Aims
Chapter 2. Peroxydisulfate activation process on copper oxide: Cu(III) as the predominant selective intermediate oxidant for phenol and waterborne antibiotics removal
1. Introduction
2. Material and Methods
2.1. Chemicals
2.2. Characterization of CuO
2.3. Experimental procedures
2.4. Analytical methods
3. Results and discussion
3.1. Characterization of CuO
3.2. PDS activation by CuO-batch experiments
3.3. Identification of reactive species
3.4. Phenol transformation pathways and mechanisms
3.5. Influence of operating parameters
3.6. Applicability of CuO/PDS system
4. Conclusions
Chapter 3. Copper oxide / peroxydisulfate system for urban wastewater disinfection: Performances, reactive species, and antibiotic resistance genes removal
1. Introduction
2. Material and Methods
2.1. Chemicals
2.2. Pathogens inactivation
2.3. Identification of reactive species
2.4. Antibiotic-resistant-bacteria and antibiotic resistance genes
3. Results and discussion
3.1. Inactivation performances of pathogens
3.2. Identification of reactive species
3.3. Antibiotic-resistant-bacteria and antibiotic resistance genes removal
4. Conclusions
Chapter 4. Peroxydisulfate activation by CuO pellet in a fixed-bed column for antibiotics degradation and urban wastewater disinfection
1. Introduction
2. Experimental section
2.1. Chemicals
2.2. Characterization of CuO pellet
2.3. Fixed-bed column set up
2.4. Experimental procedures
2.5. Analytical methods
3. Results and discussion
3.1. Working mode of CuO fixed-bed column
3.2. Degradation of antibiotics
3.3. Pathogens inactivation
3.4. Stability of FBC-CuO
4. Conclusions
Chapter 5. Conclusions and perspectives
