3.1. Fluoroquinolone Degradation Kinetics and Phototransformation Products
The oxidation process that occurs during the photocatalytic process involves five steps: (i) Diffusion of the reagent molecules from the bulk to the surface of the catalyst, (ii) adsorption of the reagent molecules on the catalyst surface, (iii) reaction on the surface and TPs formation, (iv) TPs desorption from the catalyst surface, and (v) TPs diffusion in bulk. The most common way to describe this process is the Langmuir–Hinshelwood (L–H) model (Equation (1)):
where
C is the substrate concentration in the liquid phase after equilibrium,
KHL is the LH adsorption coefficient,
kr is the reaction rate coefficient, and
r is the initial degradation rate.
For FQ concentrations ranging from tens of µg L
−1 to tens of mg L
−1,
KHL ×
C is much less than the unity, and the degradation rate can be simplified as follows:
All experimental data fit a pseudo-first-order equation:
where
C0 is the initial substrate concentration,
C is the substrate concentration at time
t, and
kdeg is the kinetic degradation constant.
Although many authors calculated the kinetic degradation constant (
kdeg) through the common linear correlation between ln [C]/[C
0] and the irradiation time, giving undue weight to the first points of the curve,
kdeg is a more convenient parameter to compare the photocatalytic performances of different systems, since it is independent of concentration. In
Table 1, the kinetic degradation constant values are reported.
As shown in
Table 1, TiO
2 Degussa P25 is used without modification [
49,
50]. Hombikat UV100 and PC500 were used by Paul et al. [
63] and Venacio et al. [
59] for comparison. Differing from P25, they displayed the sole crystallized anatase phase and a higher and microporous specific surface area of 330 and 290 m
2 g
−1, respectively [
50,
51,
58]. Venancio et al. [
59] synthesized highly crystalline TiO
2 nanoparticles with a high density through the sol-gel method. Despite the different morpho-structural characteristics, degradation rates were not significantly different from those obtained in the presence of TiO
2 P25.
The influence of different variables in the photo-oxidation process, such as the initial substrate concentration, amount of catalyst, solution pH, oxygen dissolved, temperature, oxidizing species, scavenger effects, and adsorption effects has been investigated by many research groups [
59,
60,
62,
65,
67,
68].
Concentrations of antibiotics in the range 10–50 mg L
−1 were appropriately used to facilitate the identification of photodegradation products [
60,
62,
63,
64,
67,
68,
70,
71] and the accurate quantification of dissolved organic carbon (DOC) removal [
60,
64,
67,
68,
71]—a few works were carried out at lower FQ concentrations (100–560 µg L
−1) [
59,
61,
65,
69] to better mimic actual conditions.
Ultrapure or deionized water was the most common solvent employed for the photochemical experiments [
61,
62,
63,
64,
70,
71]. It was chosen to avoid any possible interaction by a more complex matrix like surface water and wastewater. It is well known that the matrix constituents of both natural freshwaters and wastewaters may affect the photocatalytic process. Indeed, some inorganic ions and the fraction of natural dissolved organic matter (DOM) can act as photo-sensitizers able to absorb solar light directly [
72]. As hinted above, in several cases, transient species are produced in the photosensitized degradation of FQs which exhibit an absorption spectrum similar to that of the starting material, as is normal since it is mostly determined by the heteroaromatic moiety, and may make the process somewhat more complex. It can further be observed that Venancio et al. [
59] and Guzman et al. [
60] showed that fluoroquinolone degradation occurred slower in natural waters than in ultrapure water; more precisely, the rate of degradation followed the order ultrapure water < simulated water < bottle water < tap water (see
Figure 3).
Similar, but more marked results were obtained by Hapeshi et al. [
67] and Micheal et al. [
68] in secondary treated effluent samples taken from an urban WWTP. Around 80% of OFL degraded after 120 min of irradiation.
Additionally, the pH of the solution affects the degradation of these antibiotics. The pH is correlated to both the intrinsic properties of the substrate and the surface charge density of the catalyst. It is reported that the adsorption of the organic compound, and thus its degradation, is favored near the zero point charge (zpc) of the catalyst. A medium pH may affect both the position of vb and cb and the band gap energy [
65]. At pH 6, zpc of TiO
2 P25 is 6, and FQs are mainly in their cationic and zwitterion forms (half way between the
pKa of the acidic and the alkaline functional groups, see
Figure 4). Adsorption experiments demonstrated that FQs adsorption, ensured by an appropriate stirring time in the dark, onto the surface catalyst was favored at pH values close to neutral.
In neutral or slightly alkaline conditions, like those of natural freshwaters, hydroxyl radicals are more easily generated on the TiO
2 catalyst surface due to the oxidation of the adsorbed hydroxyl ions, and therefore the efficiency of the process is enhanced (see
Figure 5) [
62,
67].
As expected, the photocatalytic efficiency is affected by the catalyst loading. Indeed, by increasing the catalyst amount, the number of available active sites increases; consequently, the substrate degradation rate rises. On the other hand, concentrations higher than some grams per liter may cause light scattering, poor light penetration or agglomeration phenomena in the suspension, which leads to a sharp drop in the substrate degradation rate [
62,
65,
68].
Other experimental variables, in particular, temperature and oxygen percentage, seem to have minor relevance in the process efficiency. Specifically, Hapeshi et al. [
67] observed that the degradation rate was enhanced by bubbling oxygen through the solution, especially in the first steps of the process, while only in the presence of a saturated working solution, fluoroquinolones were quantitatively abated in half the time, as reported by Venancio et al. [
59].
Micheal et al. [
68] demonstrated that the effect of temperature had no significant impact on the conversion rate.
On the other hand, the addition of an oxidizing reagent, such as hydrogen peroxide, promotes both the oxidation process and total organic carbon (TOC) removal [
62,
67,
68]. A slight increase in the degradation rate was observed, increasing H
2O
2 from 0.14 to 5.5 mM (see
Table 1).
On the contrary, a decrease in the degradation rate was found by Biancullo et al. in the presence of a radical scavenger, such as MeOH (see
Table 1) [
65].
A comparison of the rates of degradation of CIP using different irradiation protocols revealed that UV-A/TiO
2 photocatalysis was more efficient than Vis/TiO
2 one, as shown by Paul et al. [
63].
A higher kinetic degradation constant value, and consequently, a higher percentage of dissolved organic carbon, was obtained by using UV-A sonophotocatalytic treatment. Ultrasonic radiation applied by Hapeshi et al. [
67] contributed to promoting the formation of reactive radicals, favored the substrate mass transfer from the bulk to the catalyst surface, and helped to avoid particle aggregation.
Biancullo et al. [
65] obtained the highest
kdeg value by using a UV-A irradiation LEDs system that was more efficient than the traditional lamps.
The identification and quantification of transformation products arising from photo-reactive molecules is a difficult task because no analytical standards are available for their identification and quantification; moreover, many different TPs may occur depending on the experimental and analytical setup [
73]. Despite this limitation, TPs’ evolution profiles are usually monitored, and chemical structures are proposed with an acceptable degree of certainty.
As for the photocatalytic degradation of FQs, many TPs were already present in the early steps of the process, with a lifetime usually in the same order as that of the starting molecule. On the contrary, the presence of secondary photoproducts was often confirmed by TOC measurements, which still indicated the presence of an organic load after the substrate degradation was complete en route to mineralization (see
Figure 6) [
60,
64,
67,
68,
71].
Generally, and in all the considered research papers, the chemical structures of TPs were based on high-resolution mass determination, and mechanistic pathways were quite accurately suggested. Unlike direct photolysis, in the presence of a photocatalyst, the oxidative degradation of the electron-rich amine side-chain results in the main chemical path followed by hydroxylation and decarboxylation, while photoproducts coming from direct photolysis, such as reductive dehalogenation and fluorine substitution, are not predominant as reported in most of the considered papers (see
Scheme 1 and
Table 1) [
45].
At the experimentally reported TiO
2 concentrations, most of the light is absorbed by the titania particles, which then interact through hydrogen or electron transfer with the adsorbed drug [
45,
74,
75]. Furthermore, FQ activation and degradation can also be initiated by energy transfer between photoactivated titania nanoparticles and FQs, producing a triplet of the latter, as reported in our previous works on Ofloxacin [
76]. This mechanism accounts for the identification of low quantities of secondary products from direct photolysis.
Direct photolysis initiated photodegradation suffers from the limited part of the solar spectrum the FQs can absorb (usually below 380 nm for most FQs), which does not cover the visible part of the spectrum, and from the competitive light absorption of many other environmentally present organic compounds. Furthermore, the main photo-reaction paths for FQs are accessible only through their long-living triplet excited state; this transient must be populated through Inter System Crossing (ISC) from the first excited singlet and, once formed, may suffer from different deactivation pathways which can compete with photoreaction (i.e., electron transfer and energy transfer) especially under real conditions where inorganic ions and organic compounds may easily foster these processes. When direct photolysis occurs, products from reductive dehalogenation and fluorine substitution dominate in the 6-FQs family, while the presence of a second fluorine atom at position 8, in addition, to increasing the photoreaction quantum yield, shifts the reactivity toward F elimination from C8. Unlike photolysis, during photocatalysis at the commonly employed concentrations, most of the light is absorbed by the semiconductor and not by the molecule used as the probe [
74,
75]. Furthermore, the TiO
2 UV-Vis absorption spectrum expands toward the visible region of light (400–410 nm, 3.0–3.2 eV), depending on the crystalline structure [
77]. This allows for a better exploitation of sunlight radiation. In this case, after light absorption and electron-hole pair generation in the semiconductor, the FQ molecule adsorbed onto the catalyst surface reacts mostly through either direct or OH radical-mediated hydrogen and electron transfer [
45]. This privileged reaction path in the presence of titanium dioxide photocatalyst massively increases the number of photoproducts originating from the oxidative degradation of the electron-rich FQs side-chains (
Scheme 1 and
Table 1). It ultimately initiates a cascade of reactions, eventually leading to complete mineralization and pollutant removal, with only water and carbon dioxide being released as by-products.
3.2. Antibacterial and Ecotoxicological Tests
As is apparent, TiO
2 is a strong photo-oxidant and has a bactericidal action towards a broad spectrum of harmful microorganisms [
14,
78,
79]. For this reason, microbiological and eco-toxicological tests performed on irradiated samples are useful tools to evaluate the efficiency and the efficacy of the photocatalytic process, both in terms of degradation (decomposition degree of the target molecules) and detoxification (inactivation of the active principles, evaluation of acute and chronic effects against microorganisms). Nevertheless, testing the residual toxicity of toxic substances, such as pharmaceuticals, is a critical task, as sometimes the degradation TPs are more toxic than the parent compound, and at the same time it is impossible to isolate TPs. Due to the different experimental set ups, a significant variability among the obtained data was observed, although each test was performed following the standard guidelines [
14,
71].
In general, the experimental measurements allowed us to calculate the growth inhibition (I%) according to the following equation:
where
Scontrol is the signal measured in the control sample, and
Ssample is the signal measured in the irradiated sample.
A dose–response curve may be obtained by plotting the I% calculated for each sample dilution against the log of the corresponding sample dilution. The log of the effective dose that causes 50% of the growth inhibition is reported as EC50.
An EC50 value against
P. subcapitata was reported by Van Doorslaer et al. [
71], who declared to have obtained the lowest value among those reported in the literature. It is apparent that by increasing the photocatalytic treatment, the inhibition decreases significantly, as shown by Paul et al. [
63] (see
Figure 7) and Venancio et al. [
59], despite the different experimental conditions.
Antibacterial and eco-toxicological tests used to evaluate changes in the residual antibacterial activity and the toxicity of FQs and their photoproducts are shown in
Table 1.
Several organisms were used as a probe in the experiments, such as the microorganisms as
Bacillus subtilis [
59,
62],
Staphylococcus aureus [
60] and
Escherichia coli [
59,
60,
63,
65,
70] collected from lab cultures, heterotrophic arthropods [
65] and rotifers such as
Daphnia magna [
67,
68] and
Brachionus calcyflorus [
69], algae (
Pseudokirchneriella subcapitata [
71],
Synechococcus. leopoliensis [
69]) and bacteria (
Enterococcaceae [
65],
Vibrio fischeri [
61,
64]).
Van Dooslar et al. chose
P. supcapitata as a model organism to evaluate the potential adverse effects occurring during MOX degradation [
71]. They observed a significant increase of EC50 in the first hour of irradiation when the antibiotic was still present in the solution, while for longer irradiation times (90–150 min, MOX < LOD), the growth inhibition decreased from 72% to 14%. The residual toxicity and the poor carbon mineralization compared with the parent compound conversion confirmed the presence of TPs. Their lower toxicity may be ascribable either to a decrease of biological activity or to a minor permeation through the cell membrane due to their larger hydrophilicity compared to MOX.
A UV-A sonophotocatalytic treatment of OFL aqueous solutions against
D. magna was carried out by Hapeshi et al. [
67]. OFL was quickly consumed after 30 min of treatment leading to the formation of 20 TPs, while
D. magna immobilization increased from 20% to 55% after 24 h and 48 h of exposure, respectively. A decrease in daphnid immobilization up to 5% was observed after 240 min of treatment, when TPs were removed.
Similar, but less marked results were obtained during the UV-A photocatalytic treatment of OFL, carried out under both the same experimental conditions [
67] and in other work [
68] in the presence of 3 g L
−1 TiO
2. In both cases, incomplete mineralization (around 10%) was observed after the antibiotic degradation confirmed the presence of toxic intermediates and daphnid immobilization up to 25% at a long-term exposure. The long-term toxicity was due to the formation of photoproducts that maintain the core quinolone structure.
Calza et al. [
64] reported a reduction in the natural emission of the luminescent bacteria
V. fisheri within the first 10 min of the OFL photolytic treatment, corresponding to a 55% inhibition after only 15 min of incubation. The simultaneous disappearance of the parent compound and the occurrence of TPs indicated that the toxicity was strictly correlated to TPs, especially when they retain the quinolone core structure. From 15 to 120 min of irradiation, the percentage inhibition decreased and TOC was almost completely abated (85%). An inhibition percentage below 1% and complete mineralization of the organic carbon was obtained after four hours of treatment (see
Figure 8).
The
V. fisheri bioassay was also adopted to investigate the eco-toxicological effects of CIP photodegradation [
61,
64]. For Calza et al. [
64], CIP aqueous solutions tested at different irradiation times (2–240 min) were non-toxic, indicating that CIP degradation proceeded through the formation of non-toxic TPs, even if more than four hours of irradiation were necessary to obtain a 100% TOC abatement (see
Figure 8).
In contrast, higher toxicity (70%) against
V. fisheri in more diluted CIP solutions and for long irradiation time (45 min) was observed by Silva et al. [
61], who did not exclude that irradiation in the presence of titanium dioxide may lead to toxic effects.
In good agreement with Calza’s results, Paul et al. [
63] demonstrated that for each mole of CIP degraded, the antimicrobial load
(E. coli) in the irradiated solution decreased by a “mole”, indicating that TPs retained a negligible antibacterial activity compared to the parent compound. The stoichiometric release of ammonium and fluoride ions confirmed quantitative CIP abatement, which was achieved faster under UV-A TiO
2 than Vis-TiO
2.
Comparable findings were reported by Guzman et al. [
60] both in distilled and mineral waters. CIP-irradiated samples (95% degradation, mineralization not exceeding 35%) were less harmful than non-irradiated samples, even if Gram-positive bacteria (
S. aureus) seemed more vulnerable than the Gram-negative bacteria (
E. coli) to the main TPs (by about two-fold), which conserved the quinolone ring but not the amino-side chain.
Only three papers among those examined investigated the biological activity of the sole parent compound [
59,
62,
70].
Venancio et al. [
59] assessed the residual CIP, LOM, and OFL antibacterial activity against Gram-positive (
B. subtilis) and Gram-negative (
E. coli) strains during TiO
2 P25 and PC500 photocatalytic treatment. TiO
2 P25 removed almost the total antibacterial activity in a shorter reaction time (120 min of irradiation) compared to PC500.
Similar results were obtained by Li et al. [
62]. An inhibition of 50% of the initial
E. coli activity was reached after 120 min of irradiation with concentrated solutions of CIP, ENR, NOR, OFL. Although mineralization required a longer irradiation time than FQs degradation, biological activity was removed within 180 min.
A total of 90 min of treatment on the LEV solution in the presence of a TiO
2 homemade catalyst was enough to inhibit
E. coli resistance completely [
70].
Only two studies analyzed the photocatalytic degradation of a mixture of pharmaceuticals in effluents from urban WWTPs [
65,
69]. Today, urban WWTPs are considered a hot-spot for antibiotic-resistant bacteria proliferation, and clear data on the adverse toxic effects due to pharmaceuticals at actual concentrations on simple living organisms are still lacking.
Preliminary results were proposed by Andreozzi et al. [
69]. They carried out toxicity tests exposing blue-green algae and a rotifer to an irradiated moderately hard synthetic aqueous sample containing OFL and other five pharmaceuticals at a concentration of hundredths of nanograms per liter.
The solutions were irradiated for 48h in the presence of either suspended or immobilized TiO2. A slight reduction (9%) in the initial bioactivity was observed in the catalyst suspension, while supported TiO2 nanoparticles on a membrane were less efficient than suspended nanoparticles. Based on the poor results obtained, the authors suppose that the presence of toxic TPs formed during the photocatalytic treatment.
Additional information was recently reported in a novel disinfection study by means of UV-A LEDs system applied to different WWTP samples collected on different days before and after spiking with a few hundredths of nanograms per liter of a mixture of antibiotics [
65]. Interestingly, the count of various bacterial groups (heterotrophs,
E. coli, and
Enterococcaceae) was carried out both after the photocatalytic treatment and after three days of storage in the dark at room temperature. Although one-hour UV-A LED-irradiation was enough to inhibit the bacterial activity (about two log-units), a significant regrowth was observed after three days of storage. Specifically, the antibiotic resistance after treatment was similar to that of non-treated WWTP samples for heterotrophs and
E coli and lower for
Enterococcaceae (see
Figure 9). This phenomenon, not yet completely investigated, may be attributed to organic matter produced during the oxidative process. As is evident, actual matrices can play an important role not only in the photodegradation of recalcitrant molecules, but also in the photo-detoxification process.
Lastly, Vasquez et al. [
66] described a 28 day biodegradability test where the microbiological conditions of the aqueous environment were simulated following the OECD guidelines. They evaluated the potential biodegradability of OFL and its transformation products formed during both the photocatalytic treatment and biotransformation processes. Although OFL was degraded quickly, mineralization remained low (6%), and the biodegradability value was negligible (around zero). These results indicated that primary TPs were not readily biodegradable, especially when fluorine was still present on the heteroaromatic ring.