Photolysis occurs in the environment by exposure to sunlight and is commonly employed for disinfection in water treatment utilizing UV light. Upon absorption of light, energy is released from a molecule through physical and chemical processes which include the breakdown of a compound . Although ANTX readily degrades under sunlight in absence of photosensitizers (t1/2 = 1‑2 h at alkaline pH) , other cyanotoxins such as MCs and CYN are less susceptible to direct photodegradation by sunlight . Efficacy of photolytic treatment strongly depends on the wavelength, i.e. energy of the used light. For instance, ANTX has an absorption maximum in the range of 230‑240 nm, which determines the toxin’s resistance to UV-A irradiation (315‑400 nm), while it degrades by 70 % under UV-C irradiation at 254 nm . Similarly, NOD degradation also improved when UV light of a shorter wavelength, i.e. higher energy, was used . With vacuum-UV at 172 nm, water is directly photolyzed to form OH (eq. (9)), which further increased ANTX degradation and substantially reduced the UV dose required for complete removal. However, direct water photolysis is strongly limited to a light penetration depth in water of < 100 μm, which makes OH formation by vacuum-UV less attractive to drinking water treatment compared to other AOPs .
See formula 9 in the supplementary files.
Besides wavelength, light intensity is a crucial parameter as well. MC-LR degradation increased by about 30‑40 % when light intensity was tripled . Moreover, at 254 nm and a dose of 564 mJ cm‑2, approximately 66 % MC-LR degradation was achieved, while at 312 nm, a much higher dose of 11,304 mJ cm‑2 was required to yield similar results [55, 56]. In addition to irradiation, degradation also depends on the toxin structure as shown in a study on UV-photolytic treatment of four MCs, where degradation increased from MC-LR < ‑RR < ‑YR < ‑LA owing to the different amino acid structures (A = alanine, L = leucine, R = arginine, Y = tyrosine) .
UV-based treatments are so far the only methods for which MC-LR detoxification due isomerization of the 4(E),6(E)-Adda chain (Fig. 1) to 4(Z)- or 6(Z)-Adda was observed. Furthermore, degradation mechanisms include decarboxylation, which has only been reported for UV-based methods and sulfate radical-based AOPs (SR-AOPs; see the chapter on sulfate radical-based AOPs) [56, 57].
Cyanotoxins usually co-occur with NOM which can act as photosensitizer and improve the degradation. For instance, MC-RR photodegradation by sunlight substantially increased in presence of the cyanobacterial pigment phycocyanin . However, photosensitizer concentration is essential as it was shown for MC-LR degradation. At lower concentrations, pigment availability was the limiting factor, while at higher concentrations, light attenuation was significant . In a similar manner, ANTX photodegradation was more effective in the presence of NOM but the degradation decreased with increasing NOM concentration. Experiments with quenchers showed that besides excited NOM, 1O2 and OH also contributed to the toxin degradation and that 1O2 was more important than OH . In contrast, photosensitized CYN degradation was observed to be mainly driven by OH (about 65‑70 %), with 1O2 and excited NOM only playing minor roles . This disagreement may not only be related to the different toxins, but also to experimental conditions and using fulvic acid and solar light vs humic acid and UV‑C light, respectively. Although phycocyanin did not improve CYN photodegradation, other cyanobacterial compounds were observed to accelerate NOD and CYN degradation [53, 62, 63]. In fact, the presence of different pigment types was shown to affect MC-LR photodegradation effectivity in the following order: without pigment < chlorophyll a < β-carotene < water-extractable pigments < solvent-extractable pigments . Furthermore, higher light intensities led to pigment bleaching and degradation which adversely affected MC-LR degradation .
Turbidity is one of the most important water quality parameters for photodegradation. Light absorption by non-target water constituents not acting as photosensitizer attenuates light and reduces penetration depth. Therefore, photodegradation is usually efficient in relatively clear water, after most turbidity has been removed . Other water quality parameters may also affect the degradation as shown for ANTX degradation by UV-C radiation, where toxin removal was more effective at acidic pH with an optimum at pH = 6.4, most likely due to ANTX speciation under acidic conditions (pKa = 9.4) and possible inter- and intramolecular hydrogen bonding under alkaline pH. Also, higher temperatures led to increased ANTX degradation, but the changes became insignificant at T > 24 ℃. Lastly, as for most AOPs, alkalinity was observed to decrease ANTX degradation due to quenching of reactive species .
In order to achieve high degradation yields, UV doses substantially higher than those commonly used for disinfection in water treatment (10‑40 mJ cm‑2 ) are required. Consequently, to reduce energy demand and operating costs for large-scale water treatment, the combination of UV with oxidants or photocatalysts – as discussed in the following paragraphs – is inevitable.
Photolysis in combination with oxidants
The combination of UV radiation with H2O2 or O3 improves pollutant degradation due to the photolytic production of OH (eqs. (10) and (11)) . Moreover, OH and SO4- are produced from peroxymonosulfate (PMS) or peroxydisulfate (also persulfate, PS) upon UV activation (see the chapter on sulfate radical-based AOPs). In a UV/chlorine system, OH, Cl, OCl and other reactive species are formed following eqs. (12) to (16) .
See formulas 10-17 in the supplementary files.
UV in combination with oxidants has been studied for the removal of MCs, CYN, ANTX and BMAA. For all for toxins, UV-based treatment was substantially more effective when H2O2 was added [37, 39, 40, 54, 55, 60]. Increasing H2O2 concentration improved cyanotoxin degradation only up to a certain oxidant concentration. Once the optimal H2O2 level was exceeded, OH quenching by H2O2 (eq. (17)) outbalanced radical formation [54, 55, 60]. Different studies reported that MCs were degraded at higher rates compared to CYN, ANTX and BMAA because of their higher reactivity with OH. This is caused by MCs’ size and higher number of functional moieties that are partly more susceptible to radical attack [37, 41, 65]. The importance of the structure for the reactivity with OH is further affirmed when looking at different MCs. The major part of their structures is similar with the main difference being two amino acids (see Fig. 1). However, these minor differences suffice to yield different degradation rate constants: MC-YR (1.63 × 1010 M‑1 s‑1) > MC-RR (1.45 × 1010 M‑1 s‑1) > MC-LR (1.13 × 1010 M‑1 s‑1) > MC-LA (1.10 × 1010 M‑1 s‑1) .
When O3 was added to UV instead of H2O2, MC-LR degradation also became more effective compared to UV- and O3-only treatment. O3 decomposition to OH is accelerated under UV irradiation and as a consequence, both O3 and OH oxidize pollutants [56, 66]. Although O3, i.e. its production, may be more expensive compared to H2O2 and TiO2 (for UV/TiO2 see the chapter on photocatalysis), to achieve similar results, shorter reaction times and lower oxidant doses were required compared to UV/H2O2 treatment . Due to the UV irradiation, decarboxylation and isomerization of MC-LR was observed, which did not occur in O3-only treatment. Furthermore, compared to UV- and O3-only treatment, UV/O3 had a higher potential to degrade MC-LR and its degradation intermediates simultaneously under the same conditions .
As another, cheaper alternative to H2O2, the addition of chlorine has been studied in UV-based AOPs . UV/chlorine was shown to be more effective compared to UV/H2O2, UV- and chlorine-only MC-LR treatment. Besides producing a variety of reactive oxygen and chlorine species (eqs. (12) to (16)), Cl is more selective than OH and preferably reacts with electron-rich moieties . Similar to UV/H2O2, increasing the chlorine dose led to a more effective MC-LR degradation due to an increase in reactive chlorine species production and higher contribution to toxin degradation [64, 67]. However, the use of chlorine may lead to the formation of halogenated degradation products such as chloroform and dichloroacetic acid produced from MC-LR following a series of oxidation steps Duan et al. . Even though yields of these chlorinated byproducts increased with prolonged treatment time, residual cytotoxicity after UV/chlorine treatment was lower compared to untreated MC-LR .
Besides oxidant type and dose, the UV radiation itself is an important factor, as the peroxide bond in H2O2 is cleaved only upon irradiation with light of λ < 300 nm . Hence, MC-LR and ANTX degradation by UV‑A/H2O2 (λUV‑A ≈ 400‑315 nm) has been reported to be substantially less effective compared to UV‑B/H2O2 and UV‑C/H2O2 (λUV‑B ≈ 315‑280 nm, λUV‑C ≈ 280‑100 nm), respectively [39, 55, 68].
Similar to other AOPs, water quality parameters can influence UV/oxidant degradation efficacy. In the UV/oxidant setup NOM rather acts as oxidant and radical quencher than as photosensitizer, thus decreasing removal efficacy, which is in contrast with NOM action during photolysis without the addition of oxidants. NOM may compete with the oxidant for UV photons which consequently reduces reactive species formation [60, 65, 69]. The UV/O3 system was also affected by NOM but to lesser extent than O3-only treatment of MC-LR . In case of UV/chlorine degradation of MC-LR, NOM did not only decrease the degradation, but also resulted in a higher yield of chlorinated byproducts. This yield was observed to be dependent on NOM as well as chlorine dosage . In the presence of bromide, MC-LR degradation increased due to the formation of HOBr which is more reactive than HOCl toward phenolic and amine moieties. Furthermore, UV activation of HOBr formed reactive bromine species which may have contributed to MC-LR degradation . Alkalinity decreased UV/H2O2 and UV/chlorine degradation efficacy similar to NOM due to H2O2 and radical quenching [64, 69].
UV/oxidant removal efficacy is also affected by water pH. For ANTX removal by UV/H2O2, the highest efficacy was achieved at pH 6.7, while at lower pH the OH yield decreased due to reactions with H+ and at alkaline pH ANTX is deprotonated and exists as neutral amine (pKa = 9.4). In this form, inter- and intramolecular hydrogen bonds can form which affect ANTX reactivity with OH . In contrast, for BMAA removal, alkaline pH appeared to increase the degradation rate constant due to BMAA speciation at higher pH . In UV/O3 systems, the pH does not only determine toxin speciation, but also O3 stability, which decreases at alkaline pH and may affect toxin degradation. However, this effect seemed to be less influential for MC-LR degradation by UV/O3 compared to O3-only treatment . In UV/chlorine-based treatment, the oxidant itself is also strongly affected by the pH, when HOCl dissociates to OCl‑ at alkaline pH (pKa = 7.5). OCl‑ has a lower molar absorption and thus a lower radical yield. Furthermore, OCl‑ reacts at a higher rate with OH and Cl compared to HOCl. The optimum pH for MC-LR degradation by UV/chlorine was determined to be pH 7.4 . In contrast, in another study MC-LR degradation by UV/chlorine was shown to be most effective at pH 6 and decreased at pH 7 . Most of the experimental conditions seemed to be very similar, i.e. oxidant type and concentration, UV wavelength, MC‑LR concentration and pH‑buffer composition but notable differences were the UV intensity and pH‑buffer concentration, which could have affected the outcomes. Both studies also examined the contribution of different reactive species to MC-LR degradation and reported different findings. In the first study, at neutral pH, MC-LR degradation by UV/chlorine was dominated by OH (42.5 %), while Cl2 (25.4 %), ClO (13.3 %), Cl (11.1 %) and UV (8.5 %) contributions were lower . In the second study, at neutral pH, MC-LR degradation was driven by HOCl/OCl‑ (47.3 %), while reactive chlorine species (21.3 %), UV (21.1 %) and OH (10.3 %) were only partially responsible for MC-LR degradation. Also in the second study, the UV intensity was about twice as high compared to the first study, which may explain the difference in the higher UV contribution .