1 Introduction

Alachlor, 2-chloro-2,6-diethyl-N-(methoxymethyl) acetanilide (ALC), is a herbicide used to control grasses and broadleaf weeds in crops such as corn, soybeans, and sorghum. In the USA, it is of the most widely used herbicides. However, currently, alachlor is included in the list of priority substances in European Union Water Framework Directive 2000/60/EC, because it is considered a carcinogen category 3 (Directive 2000/7/EC, 2000). In this way, the Environmental Protection Agency (EPA) has set as target for maximum contaminant level goal (MCLG) in water for human consumption at zero and a maximum contaminant level (MCL) value of 0.002 mg L−1 (EPA 2009).

Alachlor has been found in groundwater above the maximum concentration limit (MCL) of 0.002 mg L−1 (EPA 2009; Ritter 1990). It has a half-life in soil of 70 days and 30 days in water. More than ten transformation products of its degradation have been found in groundwater, in concentrations ranging from 4 to 570 ng L−1, exceeding the parent compound concentration by at least two-fold. Among those, 2′,6′-diethylaniline is of special interest. This compound is frequently detected as a promutagen and may be transformed by rat liver enzyme preparations to 2,6-diethylnitrosobenzene, which is strongly mutagenic (Potter and Carpenter 1995).

Use of advanced treatments for polluted water is becoming necessary to reduce alachlor concentration levels. In this way, advanced oxidation processes have been successfully used for this purpose. Different authors have evaluated the use of TiO2 heterogeneous photocatalysis, Fenton and photo-Fenton reactions, ultrasound, and some electrochemical techniques for pollutant removal, reaching relevant extents of ALC degradation. In particular, TiO2 heterogeneous photocatalysis and Fenton reactions have demonstrated their potential to remove a considerable number of organic pollutants from aqueous matrices (de Luna et al. 2015; Kumar et al. 2017; Potter and Carpenter 1995; Wang et al. 2016).

In the photocatalytic oxidation process with TiO2, photoexcitation of TiO2 particles promotes the generation of an electron (e−CB) from the conduction band, leaving a hole (h+) in the valence band. Hence, oxidative or reductive processes can occur at or near the surface of the catalyst particles and promote the degradation of organic matter. However, one of the disadvantages of this technology is the eventual recombination of the photo-generated electron-hole pairs, limiting the possibility of organic compound oxidation. In this way, it is known that by adding H2O2, it is possible to achieve favorable effects in photocatalytic degradation by means of further OH· radical generation. Additional OH· radicals are generated via interaction of H2O2 with conduction-band electrons as it is shown by Eq. 1, and with superoxide radicals present in the solution (Eq. 2) (Pichat et al. 1995).

$$ {\mathrm{H}}_2{\mathrm{O}}_2+{\mathrm{e}}_{\mathrm{CB}}^{-}\to {\mathrm{O}\mathrm{H}}^{\cdotp }+{\mathrm{O}\mathrm{H}}^{-} $$
(1)
$$ {\mathrm{H}}_2{\mathrm{O}}_2+{\mathrm{O}}_2^{\cdotp -}\to {\mathrm{O}\mathrm{H}}^{\cdotp }+{\mathrm{O}\mathrm{H}}^{-}+{\mathrm{O}}_2^{\cdotp } $$
(2)

In this sense, additional OH· radicals could promote a faster degradation and mineralization rates, but, in some cases, hydrogen peroxide may act as scavenger of the photogenerated holes and hydroxyl free radicals according to Eqs. 3 and 4 (Pichat et al. 1995).

$$ {\mathrm{H}}_2{\mathrm{O}}_2+2{\mathrm{h}}^{+}\to {\mathrm{O}}_2+2{\mathrm{H}}^{+} $$
(3)
$$ {\mathrm{H}}_2{\mathrm{O}}_2+{\mathrm{O}\mathrm{H}}^{\cdotp}\to {\mathrm{H}}_2\mathrm{O}+{\mathrm{H}\mathrm{O}}_2^{\cdotp } $$
(4)

Besides, hydrogen peroxide could also compete with organic matter for the absorption sites on the photocatalyst surface.

On the other hand, presence of persulfate ions in the solution could increase the photocatalytic degradation rates avoiding recombination of \( {\mathrm{e}}_{\mathrm{CB}}^{-}/{\mathrm{h}}^{+} \), leaving a larger number of holes able to produce OH·. Reactions for photocatalytic oxidation adding sodium persulfate as oxidant are expressed in Eqs. 5–8 (Pérez et al. 2006).

$$ {\mathrm{S}}_2{\mathrm{O}}_8^{2-}+{\mathrm{e}}_{\mathrm{CB}}^{-}\to \mathrm{S}{\mathrm{O}}_4^{2-}+{\mathrm{S}\mathrm{O}}_4^{\cdotp -} $$
(5)
$$ {\mathrm{SO}}_4^{\cdotp -}+{\mathrm{e}}_{\mathrm{CB}}^{-}\to \mathrm{S}{\mathrm{O}}_4^{2-} $$
(6)
$$ {\mathrm{SO}}_4^{\cdotp -}+{\mathrm{H}}_2\mathrm{O}\leftrightarrow \mathrm{S}{\mathrm{O}}_4^{2-}+\mathrm{oxidized}\ \mathrm{species} $$
(7)
$$ {\mathrm{S}}_2{\mathrm{O}}_8^{2-}+\mathrm{hv}\to 2{\mathrm{S}\mathrm{O}}_4^{\cdotp -} $$
(8)

Additionally, \( {\mathrm{SO}}_4^{\cdotp -} \) radicals are generated through persulfate UV radiation, and they could contribute to organic pollutant removal.

For its part, reactions for Fenton processes are shown in Eqs. 9–15, and for Fenton-like reactions in Eqs. 10 to 15 (Wu et al. 2012).

$$ {\mathrm{Fe}}^{2+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{Fe}}^{3+}+{\mathrm{O}\mathrm{H}}^{\cdotp }+{\mathrm{O}\mathrm{H}}^{-} $$
(9)
$$ {\mathrm{Fe}}^{3+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{Fe}}^{2+}+{\mathrm{H}\mathrm{O}}_2^{\cdotp }+{\mathrm{H}}^{+} $$
(10)
$$ {\mathrm{Fe}}^{3+}+{\mathrm{H}}_2{\mathrm{O}}_2\to \mathrm{Fe}-{\mathrm{O}}_2{\mathrm{H}}^{2+}+{\mathrm{H}}^{+} $$
(11)
$$ \mathrm{Fe}-{\mathrm{O}}_2{\mathrm{H}}^{2+}\to {\mathrm{Fe}}^{2+}+{\mathrm{H}\mathrm{O}}_2^{\cdotp } $$
(12)
$$ {\mathrm{Fe}}^{3+}+{\mathrm{H}}_2\mathrm{O}\to {\mathrm{Fe}\mathrm{OH}}^{2+}+{\mathrm{H}}^{+} $$
(13)
$$ {\mathrm{Fe}}^{2+}+{\mathrm{HO}}_2^{\cdotp}\to {\mathrm{Fe}}^{3+}+{\mathrm{HO}}_2^{-} $$
(14)
$$ {\mathrm{O}\mathrm{H}}^{\cdotp }+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{H}\mathrm{O}}_2^{\cdotp }+{\mathrm{H}}_2\mathrm{O} $$
(15)
$$ 2{\mathrm{O}\mathrm{H}}^{\cdotp}\to {\mathrm{H}}_2{\mathrm{O}}_2 $$
(16)

Additionally, in photo-Fenton and photo-Fenton-like processes, additional reactions take place,

$$ \mathrm{Fe}-{\mathrm{O}}_2{\mathrm{H}}^{2+}+\mathrm{hv}\to {\mathrm{Fe}}^{2+}+{\mathrm{O}\mathrm{H}}^{\cdotp } $$
(17)
$$ {\mathrm{H}}_2{\mathrm{O}}_2+\mathrm{hv}\to 2{\mathrm{O}\mathrm{H}}^{\cdotp } $$
(18)

Taking into account the above, in this study, ALC degradation using TiO2 heterogeneous photocatalysis including the effect of oxidizing agents H2O2 and Na2S2O8 was analyzed. In addition, pollutant removal under Fenton and photo-Fenton reactions was carried out, and the results were compared. Extent of organic carbon mineralization, ion generation rate, and sample toxicity reduction were also assessed.

2 Materials and Methods

2.1 Chemicals

Aqueous solutions were prepared using distilled purified water (conductivity: 90 μS, total organic carbon (TOC) < 0.5 mg L−1). Analytic grade alachlor (purity > 99%) was supplied by Chemservice and used as analytical standard in liquid and gas chromatography and in photochemical experiments. TiO2 Degussa P25, hydrogen peroxide 30 and 35% (weight/volume) were supplied by Panreac and Protoquímica. Sodium persulfate, iron sulfate heptahydrate, and ferric chloride hexahydrate analytical grade, supplied by Merck, were used as source of sulfate, ferrous, and ferric ions. Solution pH was adjusted with analytical grade sulfuric acid supplied by Merck.

2.2 Photochemical System

TiO2 photocatalysis, photo-Fenton, and photo-Fenton-like experiments were carried out in a tubular photo-reactor consisting of a quartz tube equipped with a concentric low pressure mercury lamp (Atlantic Ultraviolet, model MP36B, 38 W), which emits UV monochromatic radiation at 254 nm at a maximum power of 38 W. Reactor dimensions are 86 cm long and 1.5 cm diameter. Alachlor solution flowed through the space between quartz tube and an outer tube made of stainless steel. This reactor is connected to an acrylic circulation tank which has a submergible pump and stirring system for supplying oxygen to the solution. Figure 1 corresponds to a schematic representation of the used dispositive.

Fig. 1
figure 1

Schematic representation of the photo-reactor

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ALC initial concentration in the experiments was 50 mg L−1 (185.4 μmol L−1). Tests were carried out at 25 °C. Solution was recirculated at a rate of 300 mL min−1, until homogenization was achieved. The volume of solution in the photo-reactor was 10 L. The monitoring time for photolysis and photocatalysis started when the system was completely filled and mixed. Hydrolysis assays were made in three flasks filled with 500 mL of pesticide solution. pH was adjusted to 2.7, 7, and 9, and solutions were left in the dark for 5 days.

For photocatalysis tests, solution at desired initial concentration was transferred to the reactor. Mix was recycled through the system during 15 min to achieve homogenization. After homogenization, the lamp was turned on and samples were taken every 5 min at the beginning of irradiation, and subsequently every 10 min. No pH adjustment was made during the photo-degradation tests. Additional experiments at 50 mg L−1 of TiO2 in the presence of hydrogen peroxide (H2O2) at initial concentrations of 250, 500, and 750 mg L−1, and of sodium persulfate (Na2S2O8) at initial concentrations of 595, 1190, and 2381 mg L−1 were carried out.

In photo-Fenton and photo-Fenton-like processes, pH for all samples was adjusted to 2.8. Subsequently, iron ions were added as ferrous sulfate heptahydrate (FeSO4·7H2O) at concentrations of 2, 10, and 56 mg L−1 and recirculated every 15 min. H2O2 was added in a concentration of 500 mg L−1 and recirculated for 15 min. H2O2 was dosed during the experiment in order to obtain a concentration in excess during all the test. An additional Fenton-like test was made adding ferric chloride hexahydrate (FeCl3·6H2O) at a concentration of Fe+3 of 10 mg L−1.

2.3 Analytical Methods

For monitoring the concentration of alachlor, an HPLC system (Agilent Technologies 1100 Series) was used. It consists of a high pressure pump for four solvents, a diode array detector, and a system with sample injection valve, a 20-μm injection loop, and a thermostat column oven. The pesticide was analyzed using a C18 column (Merck, 125 mm × 4 mm; 5 μm). The wavelength for detection was 210 nm. Column temperature was 25 °C and injection volume was 10 μL. Eluent was ultrapure water/acetonitrile at a ratio of 40/60 and a flow rate of 1.0 mL min−1. Under these conditions, alachlor detection limit was 0.0047 mg L−1 and quantification limit was 0.1 mg L−1.

Degradation by-products were identified using a gas chromatograph (Agilent model 7890A) with autosampler coupled to a spectrometer 5975C. A column HP-5MS (30 m × 0.25 mm × 0.25 mm) and helium as carrier gas were used. Injector temperature was 250 °C, detector temperature 270 °C, and an oven temperature ramp from 100 to 210 °C at a rate of 15 °C min−1.

Chloride, nitrite, nitrate, and sulfate ions in solution were analyzed as mineralization products. For this purpose, a DIONEX-120 ion chromatograph, with an IonPac AS14 column (250 mm × 4 mm), an IonPac AG14 precolumn (4 mm), and an ionic conductivity suppressor membrane column ARS-Ultra, was used. The loop injection volume was 100 μL and the flow rate was 1.2 mL min−1. This device was connected to an electronic integrator (Hewlett Packard 3396 series II). The mobile phase was a solution of Na2CO3 3.5 mM and NaHCO3 1 mM.

Total organic carbon was evaluated with a TOC 1010 Analytical IO Total Organic Carbon Analyzer. TOC was determined by differences between total carbon (TC) and inorganic carbon (IC) contents.

H2O2 consumption was analyzed adding KI in excess to samples in acid medium for photo-Fenton and photocatalytic treatments assisted by H2O2. Starch was then added, and then, titration was made using a standard solution of sodium or potassium thiosulfate.

Na2S2O8 concentration was measured by precipitating the sulfate ion in acetic acid in presence of barium chloride.

Potentiometric determination of nitrates was made with an electrode of liquid ion exchange membranes, chlorides with a potentiometric titration system with a silver nitrate solution 0.0141 N in a system with glass electrode, and silver/silver chloride. Ammonia nitrogen and total nitrogen were also analyzed by the Kjeldahl method.

Finally, a potassium ferroxioxalate actinometer was used to determine the radiation intensity of the mercury lamp, via oxidation of oxalate K3Fe(C2O4)3·H2O, or reduction of Fe+3 based on the method described by Hatchard and Parker (1956).

2.4 Toxicity Analysis

Toxicity studies were performed with Daphnia pulex crustacean. It was acclimated to 21 ± 1 °C over a period of 16 h light and 8 of darkness and illumination of 1000 to 1500 lx, with dissolved oxygen above 60% saturation. They were fed daily with a mixture of alfalfa plus TetraMin. The specimens for toxicity test were fed 2 h before testing. Two hundred neonates were exposed for toxicity testing, with 10 individuals per test. Dilutions of samples from degradation treatments were prepared, 10 neonates transferred to each vessel for a period of 48 h. After incubation, test vessels were revised and the number of dead organisms (motionless) verified.

3 Results and Discussion

3.1 Hydrolysis Experiments

Control experiments with solutions of 50-mg L−1 alachlor initial concentration in the dark at pH 2.7, 7, and 9 (data not shown) indicated that hydrolysis in 5 days was negligible regardless of pH. This stability for alachlor in aqueous solution in the absence of light has been previously reported by Chu and Wong (2004) and Lizama and Silva (2006) (Lizama and Silva 2006; Chu and Wong 2004).

3.2 Photolysis Experiments

Photolysis results (data not shown) indicate that alachlor in aqueous solutions (initial concentration 50 mg L−1) could be completely degraded within 5 h of radiation (UV 254 nm) and a TOC reduction of 10%. These results are similar to those reported by other authors (Huang et al. 2017; Katsumata et al. 2006; Wong and Chu 2003). However, it is important to highlight that the low TOC reduction implies that some of the photo-generated by-products are recalcitrant and it is necessary to use other kind of treatment technologies in order to guarantee its removal from water bodies, in special because some of them could have more toxicity than ALC.

3.3 ALC Removal Using TiO2 Heterogeneous Photocatalysis

Results of ALC removal employing TiO2 photocatalysis are showed in Fig. 2. In this way, it can be seen that pollutant was removed totally in less than 80 min of photo-treatment, and a complete mineralization of the organic matter content in the solution was reached in around 400 min. In addition, Fig. 3 shows the influence of initial concentration of alachlor on degradation times. Effect of ALC initial concentration on process was evaluated considering values of 10, 25, 50, 83, and 124 mg L−1 obtaining total degradation times of less than 30, 50, 80, 90, and 120 min, respectively.

Fig. 2
figure 2

ALC removal using TiO2 heterogeneous photocatalysis (pollutant initial concentration: 50.0 mg L−1, catalyst initial concentration: 50.0 mg L−1, average temperature: 25 ± 2 °C, pH: 7.0)

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Fig. 3
figure 3

Effect of pollutant initial concentration on ALC removal using TiO2 heterogeneous photocatalysis (catalyst initial concentration: 50.0 mg L−1, average temperature: 25 ± 2 °C, pH: 7.0)

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Initial degradation rates (r0) and time necessary to achieve 90% mineralization (t90%,TOC) are presented in Table 1. From the table, it can be noted that as ALC initial concentration increases, initial degradation rate increases, but also does degradation and mineralization time.

Table 1 Time for 90% mineralization and initial alachlor degradation rates by photocatalysis (catalyst initial concentration: 50.0 mg L−1, average temperature: 25 ± 2 °C, pH: 7.0)
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3.4 ALC Removal Using TiO2 Heterogeneous Photocatalysis Under the Presence of H2O2 and Na2S2O8

Figure 4 shows the results for alachlor mineralization using TiO2 heterogeneous photocatalysis under the presence of H2O2.

Fig. 4
figure 4

Effect of H2O2 presence on ALC removal using TiO2 heterogeneous photocatalysis (pollutant initial concentration: 50.0 mg L−1, catalyst initial concentration: 50.0 mg L−1, average temperature: 25 ± 2 °C, pH: 7.0)

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According to Fig. 4 and Eqs. 3 and 4, addition of H2O2 into the solution has a positive effect on organic matter mineralization. However, if peroxide presence is very high, it acts as a radical scavenger and pollutant removal and mineralization are diminished. In this way, under the evaluated experimental conditions, the optimized H2O2/alachlor molar ratio corresponds to around 79 (500 mg L−1 H2O2/50 mg L−1 ALC), with a mineralization rate constant of 0.015 min−1, and mineralization time of around 300 min, shorter than that obtained using the conventional TiO2 heterogeneous photocatalysis. These results are consistent with those reported by other researches, who indicated that the optimal molar ratios for similar removal of similar organic compounds using H2O2/TiO2 photocatalysis are between 10 and 100 (Malato et al. 2000).

On the other hand, Fig. 5 shows the results obtained for alachlor mineralization using TiO2 photocatalysis under the presence of sodium persulfate. From the figure, it can be seen that increments on S2O82− presence in the solution contribute to organic matter oxidation. In particular, under 2380 mg L−1 concentration of that ion, total mineralization time was at least seven times shorter than the necessary under the treatment using only TiO2 and UV radiation. These results are consistent with the information presented by Eqs. 5–8, in which the positive effect of persulfate ions on organic pollutant removal is indicated.

Fig. 5
figure 5

Effect of S2O82− ion presence on ALC removal using TiO2 heterogeneous photocatalysis (pollutant initial concentration: 50.0 mg L−1, catalyst initial concentration: 50.0 mg L−1, average temperature: 25 ± 2 °C, pH: 7.0)

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3.5 ALC Removal Using Fenton and Photo-Fenton Technologies

Figure 6 shows the results about the alachlor degradation and mineralization using Fenton and photo-Fenton technologies under Fe+2 initial concentrations of 2.0, 10.0, and 56.0 mg L−1 and a H2O2 concentration of 500 mg L−1. Results indicated that pollutant concentration was significantly reduced during Fenton reaction in the dark for all the Fe+2-evaluated concentrations, being the most efficient 56 mg L−1. At this ferrous ion concentration, ALC was completely removed in less than 60 min without UV light irradiation, and total organic matter mineralization was reached after 120 min of irradiation. This aspect indicated that Fenton technology is able to remove ALC from aqueous matrices, but it is necessary to implement other kind of methodologies to get a total oxidation or the organic matter present in the evaluated samples.

Fig. 6
figure 6

ALC removal using Fenton and photo-Fenton (pollutant initial concentration: 50.0 mg L−1, H2O2 concentration: 500.0 mg L−1, average temperature: 25 ± 2 °C, pH: 3.0)

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Effectiveness on mineralization for photo-Fenton processes can be explained by reaction 17, in which radical Fe2+ is regenerated producing additional OH· radicals as consequence of the effect of UV light radiation. Besides, under UV radiation, H2O2 could generate additional hydroxyl radicals that would be available for organic compound elimination. In this way, photo-Fenton process would represent a more efficient way to achieve total pollutant mineralization, and implementation implies less H2O2 requirements and a faster ferrous ion regeneration.

On the other hand, the use of ferric salts for ALC degradation using Fenton technologies could represent an advantageous option in an economic point of view since they are cheaper than ferrous salts. However, Fe+3 ions react much slower with H2O2 than Fe+2 ions; hence, regeneration of Fe+2 (Eq. 12) is usually the limiting step of most Fenton and Fenton-like systems. In this way, in order to find out the effect of ferric ions in alachlor degradation, an experiment using a Fe+3 concentration of 10 mg L−1 at the same H2O2 concentration used in previous experiments with Fe+2 was carried out. Fe+3 source was ferric chloride. Figure 7 shows the obtained results and the pollutant removal under both photo-treatments.

Fig. 7
figure 7

ALC removal using ferric and ferrous ions in Fenton/photo-Fenton reactions (pollutant initial concentration: 50.0 mg L−1, H2O2 concentration: 500.0 mg L−1, average temperature: 25 ± 2 °C, pH: 3.0)

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From Fig. 7, it can be seen that without UV radiation, ALC removal with Fe2+ is faster than with Fe+3, a fact that could be explained by the slower reaction rate of Fe+3 ions with H2O2, but also by the OH· radicals scavenging by Cl− ions from the ferric salt. The reaction of OH· with Cl− leads to the formation of chlorine atoms (Cl·) and dichloride anion radicals (Cl−·2). These two species are strong oxidizing agents, being able to oxidize H2O2 and Fe2+, but are less reactive with organic solutes than OH· radicals (Pérez et al. 2006). In addition, chloride ion presence may have negative effects on the efficiency of the Fe+2/H2O2 and Fe+3/H2O2 systems as result of the formation of Fe+3 chlorocomplexes (Cui et al. 2015). However, once UV irradiation started, degradation rates for Fe+2 and Fe+3 systems were very similar. Then, it could be concluded that despite the negative effects of the Fe+3 and chloride ions in the system, FeCl3 would be a good ferric ion source to ALC removal using photo-Fenton-like processes.

Finally, Fig. 8 presents a comparison of pollutant and mineralization variation profiles using photocatalysis with 50 mg L−1 of TiO2, photocatalysis assisted by H2O2 (500 mg L−1) and Na2S2O8 (2381 mg L−1), and photo-Fenton with 56 mg L−1 of Fe+2 and 500 mg L−1 of H2O2.

Fig. 8
figure 8

ALC removal using heterogeneous photocatalysis and photo-Fenton technologies (pollutant initial concentration: 50.0 mg L−1, average temperature: 25 ± 2 °C)

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Times for alachlor total removal and mineralization followed the order photo-Fenton < photocatalysis with persulfate < photo-Fenton-like < photocatalysis with hydrogen peroxide < photocatalysis with TiO2. In this sense, organic matter mineralization under photo-Fenton process was around six times faster than using heterogeneous photocatalysis with TiO2.

3.6 Inorganic Ion Evaluation

In addition to sample TOC evolution, it might be of interest to compare inorganic release of heteroatoms, such as chlorine, ammonia, and/or nitrate. When total mineralization of alachlor is achieved, nitrogen could be released into the solution as a combination of ammonium and nitrate, but under a sufficient reaction time, ammonium could be totally oxidized to nitrate. In this way, Fig. 9 presents the ammoniacal and nitric nitrogen concentration evolution during sample treatment using heterogeneous photocatalysis and photo-Fenton. From results, it can be noted that at 400 min of photo-treatments, 100% conversion (stoichiometrically) of the alachlor nitrogen content was achieved. Besides, most of the N content in the solutions was in form of ammonia indicating that NO3− is susceptible to oxidation inside the reactive medium.

Fig. 9
figure 9

Nitrogen presence in treated solutions using a heterogeneous photocatalysis and b photo-Fenton technologies (pollutant initial concentration: 50.0 mg L−1, average temperature: 25 ± 2 °C)

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On the other hand, Fig. 10 allows to observe that chloride ion concentration in samples was quickly increased with increasing the reaction time, a situation that indicates that degradation/dechlorination stage occurs very fast during the pollutant photo-treatment. The total amount of chloride ion produced after 120 min of reaction was approximately 6.60 mg L−1 corresponding to a 100% conversion of the alachlor chloride content.

Fig. 10
figure 10

Chloride ion presence in treated solutions using heterogeneous photocatalysis and photo-Fenton technologies for ALC removal (pollutant initial concentration: 50.0 mg L−1, average temperature: 25 ± 2 °C)

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3.7 By-product Identification

Degradation by-products for ALC photocatalytic treatment were studied using a GC-MS mass spectrometer. In this way, hydroxyalachlor, N-(2,6-diethylphenyl) methylamine, and lactam (8-ethyl-methoxymethyl-4-methyl-2-oxo-1,2,3,4,tetraquinoline) were identified. In general, different authors have reported that pollutant removal under advanced oxidation processes proceeds via homolytic breakage of the C–Cl bond, followed by co-entrapment of radical co-formed by water or by an intramolecular hydrogen abstraction (Hapeman-Somich 1991; Kumar et al. 2017; Pérez et al. 2006). Finally, the results found in this study support that degradation of alachlor by photocatalysis with TiO2 comes from bond breaking or by oxidation of the N-methoxy methyl group and a subsequent mineralization of the organic matter. Table 2 presents the molecular structures, main fragments, and retention times from ALC and the identified by-products.

Table 2 Identified intermediates during photocatalytic degradation of ALC
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3.8 Toxicity Studies

Figure 11 shows the evolution of toxicity assays (expressed in mortality percentage) using Daphnia pulex crustacean during sample treatment under heterogeneous photocatalysis with TiO2 and photo-Fenton processes. From results, it can be inferred that during the photocatalysis process, a 50% mortality reduction was achieved after 450 min of treatment while during the photo-Fenton process, this reduction was reached after 180 min, demonstrating that photo-Fenton technology is faster for toxicity decreasing.

Fig. 11
figure 11

Toxicity evolution using heterogeneous photocatalysis and photo-Fenton technologies for ALC removal (pollutant initial concentration: 50.0 mg L−1, average temperature: 25 ± 2 °C)

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4 Conclusions

It was demonstrated that TiO2 heterogeneous photocatalysis and photo-Fenton processes with UV radiation at 254 nm are useful treatments for water contaminated with alachlor. Additionally, pollutant removal was faster using Fenton reaction than under the photocatalytic process with TiO2, even if oxidizing agents such as H2O2 and Na2S2O8 are added.

Fenton process was sufficient for a total alachlor degradation, but not for total mineralization, that was reached under irradiation conditions in all the evaluated systems. Without UV radiation, pollutant oxidation with Fe2+ was faster than oxidation with Fe3+, but once irradiation started, degradation rates were very similar, showing that FeCl3 could be a good ferric ion source to degrade alachlor under Fenton and photo-Fenton-like technologies.