Environmental fate of naproxen, carbamazepine and triclosan in wastewater, surface water and wastewater irrigated soil — Results of laboratory scale experiments

Science of the Total Environment 538 (2015) 350–362

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Environmental fate of naproxen, carbamazepine and triclosan in wastewater, surface water and wastewater irrigated soil — Results of laboratory scale experiments J.C. Durán-Álvarez a, B. Prado b, D. González c, Y. Sánchez c, B. Jiménez-Cisneros d,⁎ a

Centro de Ciencias Aplicadas y Desarrollo Tecnológico, Universidad Nacional Autónoma de México, Mexico Instituto de Geología, Universidad Nacional Autónoma de México, Mexico c Instituto de Ingeniería, Universidad Nacional Autónoma de México, Mexico d International Hydrological Programme (IHP), UNESCO, France b

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• Degradation and transport of pharmaceuticals studied in a wastewater irrigated area • Naproxen and triclosan readily photo and biodegraded, carbamazepine was recalcitrant • Biodegradation inhibited at high spiking concentrations for the tested matrices • Transport of triclosan and carbamazepine was delayed by soil chemical components • Naproxen can easily reach aquifer in case of low degradation in the soil

a r t i c l e

i n f o

Article history: Received 20 May 2015 Received in revised form 15 July 2015 Accepted 7 August 2015 Available online xxxx Editor: Kevin V Thomas Keywords: Adsorption Biodegradation

a b s t r a c t Lab-scale photolysis, biodegradation and transport experiments were carried out for naproxen, carbamazepine and triclosan in soil, wastewater and surface water from a region where untreated wastewater is used for agricultural irrigation. Results showed that both photolysis and biodegradation occurred for the three emerging pollutants in the tested matrices as follows: triclosan N naproxen N carbamazepine. The highest photolysis rate for the three pollutants was obtained in experiments using surface water, while biodegradation rates were higher in wastewater and soil than in surface water. Carbamazepine showed to be recalcitrant to biodegradation both in soil and water; although photolysis occurred at a higher level than biodegradation, this compound was poorly degraded by natural processes. Transport experiments showed that naproxen was the most mobile compound through the first 30 cm of the soil profile; conversely, the mobility of carbamazepine and triclosan through the soil was delayed. Biodegradation of target pollutants occurred within soil columns during transport experiments. Triclosan was not detected either in leachates or the soil in columns, suggesting its complete biodegradation.

⁎ Corresponding author at: 1 rue Miollis, 75732 Paris Cedex 15, France. E-mail address: [email protected] (B. Jiménez-Cisneros).

http://dx.doi.org/10.1016/j.scitotenv.2015.08.028 0048-9697/© 2015 Elsevier B.V. All rights reserved.

J.C. Durán-Álvarez et al. / Science of the Total Environment 538 (2015) 350–362 Transport Pharmaceutical compounds Photolysis

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Data of these experiments can be used to develop more reliable fate-on-the-field and environmental risk assessment studies. © 2015 Elsevier B.V. All rights reserved.

1. Introduction Reuse of treated and untreated wastewater for agricultural irrigation is a widespread practice in arid and semiarid areas (Raschid-Sally and Jayakody, 2008). Despite the benefits of reusing wastewater (Adrover et al., 2012; Raschid-Sally and Jayakody, 2008; Toze, 2006; Jiménez, 1995), the potential of spoiling soil, water sources and crops with a wide variety of contaminants, notably emerging pollutants, is a source of concern. Previous studies have reported the occurrence of several emerging pollutants in urban wastewater in developing countries, such as Mexico, South Africa, Brazil, Tunisia, Vietnam and other tropical Asian countries, at levels of ng/L–μg/L, and even at mg/L for some phthalate esters (Shimizu et al., 2013; Olujimi et al., 2012; Mnif et al., 2010; Moreira et al., 2009; Duong et al., 2008; Gibson et al., 2007). On the other hand, in rural areas or small towns, even when the load of most of emerging pollutants in wastewater is lower than that observed in big cities (Vystavna et al., 2012), the presence of antibiotics at levels of ng/L to μg/L has been systematically observed and attributed to human and livestock medications as well as to aquaculture (Zou et al., 2011; Cabello, 2006; Sarmah et al., 2006). Considering the aforementioned, the use of wastewater for agricultural irrigation in both urban and rural areas should be carried out with caution, and knowing as much as possible the effects and fate of the pollutants that are discharged into the environment. The harmful effects that emerging pollutants may cause to aquatic and soil organisms are as the matter of facts still poorly understood. Toxicity studies are mostly focused on acute damages to aquatic organisms. In the case of soil, some studies on the impact of emerging pollutants to plants, soil microfauna and accumulation in worms have been performed. Moreover, the information on the chronic effects of complex mixtures of emerging pollutants at trace levels, which are the most realistic conditions, is very scarce (see Cleuvers, 2004 as a good example of this). Extensive reviews on the effects that emerging pollutants cause to aquatic and soil organisms can be found in literature (Durán-Álvarez and Jiménez-Cisneros, 2014; Stuart et al., 2012; Pal et al., 2010; Farré et al., 2008). The study of the environmental fate of emerging pollutants has been performed mostly on wastewater and drinking water treatment systems (see reviews in Verlicchi et al., 2013 and Huerta-Fontela et al., 2011); thus the study of the potential degradation, mobility and accumulation of these chemicals in soil and natural waters, and assimilation by crops is necessary, particularly in regions where large volumes of wastewater are used for irrigation. In the environment, a succession of natural processes leading to the dissipation and/or the mobility of pollutants in soil and water are continuously occurring. Biodegradation, photolysis, hydrolysis, adsorption and leaching through soil have been previously reported for specific emerging pollutants in natural systems (Yamamoto et al., 2009; Farré et al., 2008). Some studies have conjunctively addressed the degradation and partition of specific pharmaceuticals in freshwater/sediments systems at laboratory scale (Yamamoto et al., 2009). From these studies, it is known that some pharmaceuticals, such as diclofenac, naproxen and sulfamethoxazole, are readily photodegradable in fresh water (Bahnmüller et al., 2014; Zhang et al., 2011; Isidori et al., 2005), whereas other compounds, such as anti-epileptic and β-blocker drugs, are recalcitrant to natural photolysis (Acuña et al., 2015; Dong et al., 2015; Kunkel and Radke, 2012). However, it is noteworthy that most of the reported studies addressing natural photolysis of emerging pollutants in water have been carried out in regions above 40° North latitude (Boreen et al., 2003), thus it is necessary to elucidate the potential of natural photolysis, not only in water but in soil, using sunlight

irradiation conditions that are found near the tropics, where many developing countries are found. Biodegradation of emerging pollutants has been extensively studied in wastewater treatment systems (Verlicchi et al., 2013; Salgado et al., 2012); although, biodegradation in agricultural soils and fresh water has been less explored. From what has been reported so far, it is known that some microorganisms capable of degrading highly recalcitrant chemicals, such as carbamazepine and polybrominated flame retardants, can be isolated from soils (Jelic et al., 2012; Rodríguez-Rodríguez et al., 2012). High biodegradation rates may be expected in wastewater and soils due to the high content of both organic matter and degrading microorganisms (Ma et al., 2015). Moreover, higher biodegradation rates may be expected in long-term wastewater irrigated soils than in rainfed or groundwater irrigated agricultural soils. This is because of the high and varied microorganism populations, the excess of nutrients in soil and the high acclimatization of the native microfauna to degrade toxic and recalcitrant chemicals (Ma et al., 2015; Müller et al., 2007). When wastewater is used to irrigate crop fields, it is naturally treated by infiltration through soil. Whether biodegradation occurs or not, the mobility of dissolved contaminants within the soil is influenced by the physical and chemical properties of the solid matrix (e.g., texture, porosity, pH, salinity and organic matter content) and the properties of the contaminants (such as pKa, log D, chemical structure and vapor pressure) (Wehrer and Totsche, 2008). Since soil properties are indeed modified by wastewater irrigation (Durán-Álvarez and JiménezCisneros, 2014), it is necessary to study the mobility of emerging pollutants in soils that have been irrigated using wastewater for long periods of time. The aim of this work was to determine the environmental fate of three emerging pollutants, namely naproxen, carbamazepine and triclosan, in wastewater, surface water and wastewater irrigated soils using samples coming from an agricultural irrigation district where untreated wastewater has been applied to soil for more than a century. These pollutants were selected for this study given that: a) they have been consistently found in different environmental compartments of the Tula Valley irrigation area (i.e., waste and surface waters and soil); and, b) these compounds display differences in their physical and chemical properties, which may result in differences in the environmental fate.

2. Materials and methods 2.1. Chemicals All reagents used in the experiments were analytical grade. The standards naproxen, carbamazepine and triclosan; the surrogate standards 3,4-dichlorophenoxyacetic acid (3,4-D), [2H4] 4-n-nonylphenol and [2H16] bisphenol-A; the internal standards clofibric acid, 4-nnonylphenol and 10,11-dihydrocarbamazepine; the reagents sodium bicarbonate, sodium sulfate, calcium chloride, diatomaceous earth and acetic acid; as well as the derivatizing agents N-tertbutyldimethylsilyl-N-methyltrifluoroacetamide (MTBSTFA) with 1% of tert-butyldimethylsilylchlorane and N,O-bis(trimethylsilyl) trifluoroacetamide (BSTFA) with 1% of trimethylsilylchlorane were obtained from Sigma-Aldrich (St. Louis, MO, USA). All the solvents and pure water were HPLC grade, purchased from Burdick and Jackson (Morristown, NJ, USA). Oasis HLB extraction cartridges (200 mg, 60 cm3) were bought from Waters (Milford, MA, USA). The relevant

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physical and chemical properties of the target emerging pollutants are shown in Table 1. 2.2. The study area For this study, wastewater, soil and surface water samples were taken from Tula Valley, Central Mexico (20° North latitude). This is a semiarid zone (climate classification BS1kw(i′)gw′) where annual precipitation is 550 mm in average (mainly between June and September), while evapotranspiration is higher than precipitation in a factor of three (1524 mm). In this area, the wastewater produced in Mexico City (around 5.2 million m3 per day) has been used for more than a century to irrigate 85,000 ha of agricultural land (Jiménez and Chávez., 2002). Wastewater is transported from Mexico City to the Tula Valley via 80 km of subterranean canals, taking nearly one day to travel from Mexico City to the first irrigation fields (Siemens et al., 2008). In the irrigation district, wastewater is distributed to small plots via a complex network of open unlined canals in which wastewater flows by gravity. The excess of wastewater is stored in reservoirs for further use during the dry season, from October to June. The hydraulic retention time in reservoirs is three months on average. Intensive production of forage such as alfalfa, corn and oats is practiced in the agricultural district. Crop yields in the irrigated area are usually 100–150% higher than those obtained in non-irrigated areas (Jiménez and Chávez, 2002; Jiménez, 1995); such production rates are due to the continuous supply of water and nutrients to soil via wastewater (Siebe, 1998). Wastewater irrigation is carried out by flooding the plots. In a typical irrigation event, a column of wastewater (190 mm in average) is applied to the soil, then water is allowed for infiltration through soil, while excess water runs off; each event takes from 2 to 5 h. Both evapotranspiration and infiltration of wastewater determine the fate of the pollutants entering to the irrigated area. The average evapotranspiration rate in the Tula Valley region is 1524 mm in spring and summer months, i.e., near 73% of the applied wastewater is lost by evaporation. In autumn and winter, evapotranspiration decays up to 300 mm. Infiltration of wastewater through soil has led to the incidental recharge of the aquifer. The recharge of the aquifer has resulted in upwelling springs with flow rates of 40–600 L/s. Both surface and groundwater are used to supply the entire population of the Tula Valley region (500,000 inhabitants), using only chlorination as drinking water treatment (Jiménez and Chávez, 2004). 2.3. Samples collection and characterization Grab samples of untreated wastewater, topsoil (corresponding to the first 5 cm layer) and surface water were obtained at specific locations of Tula Valley (Figure S1 in the Supplementary Information section). Wastewater samples were taken directly from irrigation canals

while soil samples were taken from a cropland that has been irrigated for 90 years. Surface water was sampled from a spring near the wastewater irrigated plot. All the samples were placed into glass amber bottles and stored at 4 °C overnight until characterization the next day. For transport experiments, undisturbed soil columns were collected alongside the surface soil sampling point. Acquisition of undisturbed soil columns was as follows (Figure S2): a 40 cm depth pit was dug; then, 25 × 15 cm cylindrical soil monoliths were carved. Monoliths were put into stainless steel cases and then liquid paraffin (at 60 °C) was poured in the free space between the soil monolith and the stainless steel case, in order to avoid the border effect during transport experiments. Three soil columns were obtained for transport experiments. In the laboratory, the surface soil sample was air-dried for 12 h; then it was gently grounded in a mortar and passed through a 2 mm sieve. Wastewater, soil and surface water samples were characterized in terms of the parameters shown in Tables 2 and 3, following standardized methods (Soon and Hendershot, 2008; Eaton et al., 1998). The background concentration of the target emerging pollutants was determined in the three matrices. 2.4. Photolysis tests Photolysis experiments were carried out under simulated solar irradiation using the Suntest CPS + equipment, Atlas Electronic Devices. The apparatus is equipped with a 1500 W arc Xenon lamp and UV filters to simulate the atmospheric attenuation of sunlight. The wavelengths tested ranged from 290 to 800 nm (UV-B, UV-A, and visible light spectrum). The irradiance of the lamp and the temperature within the Suntest device were kept constant at 250 W/m2 and 35 °C, respectively. Keeping such temperature in photolysis experiments reflects in a better way the concomitant effect of sunlight and high temperature that is typical of this semiarid area. In a typical experimental run, 100 mL of unfiltered surface or wastewater were put into 250 mL borosilicate glass flasks and fortified with the mixture of the target pollutants at concentration of 100 μg/L of each compound; initial concentration in experiments refers to spiking concentration plus background concentration. In the case of soil samples, the fortification was aimed to achieve the concentration of 100 μg/kg of each compound. To manage this, 4 mL of the concentrated solution of the target pollutants (25 mg/L) was diluted into 96 mL of pure water and then sprayed on 1000 g (dry mass) of soil. The fortified soil was thoroughly mixed with a stainless steel spatula and allowed for drying in the fume hood for 24 h. The concentration of each target pollutant was measured in liquid and soil samples after spiking. The gravimetric moisture and the concentration of each pollutant in the soil samples were 150 mg H2O/g soil and 98 μg/kg, respectively. For the photolysis assays, 30 g of fortified soil (dry mass) were put into Petri dishes (10 cm, inner diameter) as a

Table 1 Relevant chemical and physical properties of the target emerging pollutants (using information from the United States National Library of Medicine, 1986). Compound Naproxen

Carbamazepine

Triclosan

a

Chemical structure

pKaa

log Db

Water solubility (mg/L)c

Compound speciationd

4.1

−0.2

159

N99.99% as ionized form

13.9

2.3

17.7

8.4

4.75

9.6

N99.99% as nonionized form

94.9% as ionized form 5.1% as nonionized form

Acidic ionization constant. Partition coefficient at pH 7.5 calculated as: log D = log Kow + log(1 + 10pH–pKa)−1. At 25 °C. d The fraction of nonionized molecules was calculated considering a pH value of 7.5 by: log(1 + 10pH–pKa)−1, some of this information is available at United States National Library of Medicine Hazardous Substances Data Bank: available at: http://toxnet.nlm.nih.gov.pbidi.unam.mx:8080/cgi-bin/sis/htmlgen?HSDB. b c

J.C. Durán-Álvarez et al. / Science of the Total Environment 538 (2015) 350–362 Table 2 Relevant physical and chemical properties of water samples. Parameter

Wastewater

Surface water

pH Electrical conductivity (μS/cm) Total organic carbon (mg/L) Redox potential (mV) Turbidity (NTU) Total suspended solids (mg/L) Dissolved oxygen (mg/L) Nitrate (mg/L) Heterotrophic aerobic bacteria (CFU/100 mL)

7.4 1011 192 −152 68 117 0.2 b0.3 1.3 × 107

7.3 1319 0.6 63 0.2 2.3 4.7 53 2.1 × 103

Concentration of the target pollutants (ng/L) Naproxen Carbamazepine Triclosan

6359 193 1401

1.0 17.2 1.8

5 mm layer. Both water and soil samples were sterilized by autoclaving and UV irradiation before the photolysis tests. Both liquid and soil samples were exposed to simulated sunlight irradiation for 120 min in order to supply a dose of 3600 kJ/m2. Samples were taken after 5, 10, 20, 30, 60, 90 and 120 min of irradiation as well as at the beginning of the experiments. All experiments were carried out in triplicate and dark sterile controls were carried out in parallel in order to exclude the effect of other degradation mechanisms. 2.5. Biodegradation tests In biodegradation assays water and soil samples were initially fortified as described above. Two concentrations of each pollutant were tested in each matrix in order to assess the effect of the pollutants' concentration on the biodegradation rate. Fortification levels were determined based on those concentrations found in previous monitoring events (Gibson et al., 2007). For wastewater, fortification levels were set at 10 and 50 μg/L (considering the background concentration determined in wastewater and showed in Table 2), while for surface water concentrations of 20 and 100 ng/L were tested. Regarding soil, the initial concentrations of target pollutants were set at 10 and 50 μg/kg. The concentration of each target pollutant was measured in liquid and soil samples after spiking. For each matrix, sterile blanks were prepared and spiked. The sterilization of wastewater and soil samples was carried out by autoclaving at 121 °C and 15 psi over three 60 min events. On the other hand, surface water was sterilized by UV irradiation (λ = 254 nm) for 45 min. For all the liquid samples HgCl2 was added at 1 g/L to prevent the resurgence of microorganisms. Biodegradation assays were carried out at 25 °C under dark conditions; all the experiments were conducted in triplicate. For wastewater,

Table 3 Relevant physical and chemical properties of the tested soil. Parameter

Value

pH Total organic carbon (mg/g) Electrical conductivity (μS/cm) Texture Sand (%) Silt (%) Clay (%) Specific surface area (m2/g) Hydraulic conductivity (cm/h) Heterotrophic aerobic bacteria (CFU/g)

8.01 26 1792 Clay loam 13 38 49 66 1.3 4.1 × 106

Concentration of the target pollutants (ng/g) Naproxen Carbamazepine Triclosan

3.2 4.8 7.7

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100 mL samples were put into sterilized amber glass flasks (100 mL capacity), minimizing the headspace in order to prevent the increment of the dissolved oxygen concentration. The experiment was carried out in two stages, in an attempt to emulate conditions observed in the field. The first stage, which represents the passage of wastewater in the sewerage system and its transportation to the Tula Valley, lasted 3 days and was performed under anoxic conditions, maintaining minimum dissolved oxygen content in wastewater (near 0.3 mg/L). In the second stage, which represents the travel of wastewater in the irrigation zone through open canals, dissolved oxygen content was increased to ~ 2.8 mg/L. This phase lasted 7 days, and the dissolved oxygen content was maintained throughout the experiment by bubbling air each three days. Samples were taken each 24 h throughout the experiments in order to monitor the decay in the concentration of target pollutants. Biodegradation of target pollutants was expected to be lower in surface water and soil than in wastewater samples, thus experiments in such matrices were carried out for a longer time frame. Biodegradation assays using surface water were performed for 60 days under aerobic conditions. Surface water samples (50 mL, each) were put into sterilized amber bottles, keeping a head space of 50 mL, in order to maintain aerobic conditions. The dissolved oxygen content in water was maintained at ~ 4.7 mg/L throughout the experiment. Samples were taken at the beginning of the experiment and after 5, 10, 20, 30, 40, 50 and 60 days of incubation. For soil biodegradation tests, 30 g of the spiked moist soil (75% of field capacity; i.e., 270 mg H2O/g soil) were weighed into glass flasks and maintained under aerobic conditions for 70 days. Soil moisture was kept constant through the addition of sterile water each time the flasks' mass fell below of that recorded at the beginning of the experiment. Samples for monitoring the concentration of target pollutants were drawn at 5, 10, 20, 30, 40, 50, 60 and 70 days as well as at the beginning of the biodegradation test. 2.6. Transport experiments In transport experiments, irrigation events were simulated using undisturbed soil columns and 10 mM CaCl2 as irrigation water. Water was applied to the soil columns using transient flow conditions in order to emulate the infiltration of water through the soil after the irrigation events. The simulation of the irrigation events was as follows: 950 mL of CaCl2 solution, containing the conservative tracer bromide (0.1 M) and the mixture of the target pollutants (10 μg/L of each one), were applied as a 5 cm water column at the top of soil columns. Water was then allowed to infiltrate through the soil. Leachate was collected at the bottom of the columns and divided into 20 mL sub-samples. Bromide was determined in leachate samples using a Br- Ion Selective Electrode (ISE, Thermo Scientific). Subsequently, columns were allowed for drying over the next 4 days. After this period of time, 8 irrigation events, using only 10 mM CaCl2 solution were carried out as described above. At the end of the transport experiments, soil columns were weighed in order to calculate the volumetric water content and then cut into three 8 cm segments. The soil of each section was air-dried, slightly grounded and sieved, then analyzed to determine the residual concentration of the target emerging pollutants. 2.7. Analysis of the target emerging pollutants The concentration of the target emerging pollutants in liquid samples was determined following the procedure developed and validated by Gibson et al. (2007). In brief, the pH of samples was adjusted to 2.0; then, surrogate standards were added. Water samples were passed through Oasis HLB cartridges, previously conditioned twice with 5 mL of acetone and once with 5 mL of 5% acetic acid. After this, naproxen and

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carbamazepine were eluted from cartridges using 5.5 mL of a mixture 40:60 of acetone:0.1 M sodium bicarbonate (adjusted to pH 10); then, triclosan was eluted with 5 mL of pure acetone. Each eluate was concentrated using a gentle atmosphere of ultra-high purity nitrogen and dewatered by addition of sodium sulfate salt. Derivatization was performed prior to the chromatographic analysis by the production of methylsilyl derivates using the derivatizing agents MTBSTFA for naproxen and carbamazepine and BSTFA for triclosan. Analysis of soil samples was carried out using the method proposed and validated by Durán-Álvarez et al. (2009). The extraction of analytes from soil was carried out using the pressurized liquid extraction technique. For this, 5 g of soil were accurately weighed into 22 mL-ASE stainless steel cells and extracted using the hexane:acetone:acetic acid (49:49:2 v/v/v) solvent mixture. The extraction conditions were as follows: two cycles, extraction temperature of 100 °C, nitrogen pressure at 10.34 MPa, no pre-heat, 5 min of heating time, 5 min of static time and flush at 60%. After extraction, surrogate standards were added to samples; then extracts were evaporated to reach a volume of ~ 3 mL. After evaporation, 20 mL of HPLC-grade water were added to the concentrated extracts. The resulting solutions were passed through the Oasis HLB cartridges, previously conditioned as stated above. Naproxen and carbamazepine were eluted from the cartridges using the 40:60 acetone:0.1 M sodium bicarbonate mixture, while triclosan was eluted using 5 mL of 50:50 acetone:dichloromethane mixture. After elution, the sample preparation procedure was the same as described for liquid samples. Separation and quantification was carried out using an Agilent HP 6890 gas chromatograph in tandem with a HP 5397 mass spectrometer. The chromatographic column was a fused silica capillary column (30 m × 0.25 mm, 0.25 μm of film thickness). The carrier gas (helium) was supplied at a constant flow of 1 mL/min, and 1 μL of sample was injected in the splitless mode. The temperature of the injector was set at 250 °C, while the oven temperature program was as follows: 100 °C for 1 min, ramp of 20 °C/min to 280 °C, and 280 °C for 10 min. The detector was used in the single ion monitoring (SIM) mode. The temperature of the electron ionization source was 230 °C with electron energy of 70 eV. Calculation of the analytes' concentration was done through the internal standard method. Quality assurance was guaranteed by the use of surrogate standards. The average recoveries of the surrogate standards in liquid and soil samples were 98% for 3,4-D, 93% for [2H4] 4-nnonylphenol and 90% for [2H16] bisphenol-A. Information on the validation of the analytical technique, including the limit of detection of the analytes, can be found in Table S1 of the Supplementary Information section.

2.8. Data analysis UV-visible absorption spectrum of each target pollutant was obtained in HPLC-grade water using a Cary 5000 UV-vis-NIR spectrophotometer. Determinations were carried out at pH 7, since it was within the pH range of the study matrices. Fig. 1 depicts the absorption spectra of the target emerging pollutants, expressed as the molar extinction coefficient in the wavelength range from 250 to 350 nm. Extinction coefficients were calculated using Eq (1): ελ ¼

Aλ cl

ð1Þ

where Ɛλ is the molar extinction coefficient at a given wavelength (1/M cm); Aλ is the absorbance of each compound at a given wavelength (dimensionless); c refers to the concentration of each compound (i.e., 1 × 10− 4 M); and, ɭ is the path length (i.e., 1 cm). The results obtained in the photodegradation and biodegradation assays were fitted to the first order kinetic model described by Eq. (2): C t ¼ C 0  e–kt

ð2Þ

where C0 and Ct are the concentrations of the target pollutants at time 0 and t (in ng/L or μg/kg), respectively; k is the first order degradation rate constant (1/h for photolysis and 1/day for biodegradation tests), and t is the experimental time (in hours for photolysis and in days for biodegradation experiments). Since experiments were carried out in triplicate, average degradation rates and standard deviations were obtained. In the case of biodegradation tests, half-life time was determined for each pollutant in the tested conditions by using Eq. (3): t 1=2 ¼

ln 2 k

ð3Þ

where t1/2 is the half-life time for each compound under the experimental conditions (days) and k is the first order degradation rate constant (1/day). Regarding transport experiments, the breakthrough curves of both the bromide conservative tracer and the target pollutants were obtained from the analysis of leachates. The breakthrough curves express the relationship between the relative concentration of the target compounds and the pore volume; this latter was obtained through Eq. (4): VP ¼ VC  θ

ð4Þ

where Vp is the pore volume (dimensionless), Vc is the volume of soil monolith (cm3) and θ represents the volumetric moisture of soil at the end of the experiment (cm3/cm3). The experiment was performed in triplicate in order to assess the variation of the results due to the heterogeneity of the soil. Analysis of variance showed no significant differences between replicates; thus average breakthrough curves of both the target pollutants and the conservative tracer are presented in the Results and discussion section. 3. Results and discussion 3.1. Photolysis tests

Fig. 1. Absorption spectra of the target emerging pollutants, in pH 7 buffered solutions, and the solar irradiation spectrum of the Tula Valley region.

Fig. 1 shows the UV–vis absorption spectra of the target pollutants and the solar irradiation spectrum reported for the Tula Valley region, which is close to 20° North latitude. Since the absorption spectra of the target pollutants overlap with the solar irradiation spectrum, it is plausible to expect that compounds are susceptible to photodegrade (Peuravuori and Pihlaja, 2009).

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Fig. 2. Photodegradation kinetics of the target pollutants in (a) wastewater, (b) surface water and (c) soil.

Results obtained in dark controls showed no degradation of the target pollutants by other abiotic processes such as hydrolysis during photolysis experiments (see Figure S3 and Table S2 in the Supplementary Information section).

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3.1.1. Photolysis in wastewater Fig. 2a shows the photolysis kinetics of the target pollutants in wastewater. The total removal of triclosan was achieved upon 60 min of irradiation, whereas for naproxen and carbamazepine degradation rates of 92% ± 0.1 and 34% ± 3, respectively were observed upon 2 h of irradiation. Photolysis rates obtained for the three pollutants in wastewater are shown in Table 4 (in the case of triclosan, the initial photodegradation rate was obtained). The photodegradation rate of carbamazepine (the most recalcitrant compound) obtained in this study was higher than that reported in wastewater treatment plants in Europe (Andreozzi et al., 2003). Due to the high solar irradiation intensity in the Tula Valley region (20° N latitude), higher photodegradation rate values than those reported for regions at higher latitude are expected for the three contaminants. On the other hand, it is known that composition of wastewater is a key factor on the photodegradation performance. Given that wastewater used in Tula Valley is non-treated, it is possible to hypothesize that its composition hinders the photolysis of contaminants. For instance, suspended solids and turbidity indeed hamper the photolysis of the organic compounds since both the dissolved and the particulate organic matter can reflect the incident photons hence impeding direct photolysis. According to this, higher photodegradation of target pollutants would be expected if wastewater is treated. Indirect photolysis may occur by production of free radicals in wastewater following different paths. One of these paths is the photolysis of dissolved organic matter molecules. Two types of free radicals can be formed by this route (Grzybowski and Szydlowski, 2014; Zafiriou et al., 1984), on the one hand, oxygen-based radicals (e.g., •OH, H2O2, •O and •O− 2 ) which can increase the pH in wastewater, and on the other hand, excited dissolved organic matter molecules (•DOM), which are the transitional state achieved by the excitation of electrons in the dissolved organic matter molecules to achieve the triplet state. The background state of •DOM is rapidly recovered when excited electrons are transferred to other organic molecules or by dissipating the energy as heat. It is well known that oxygen-based radicals are more energetic than •DOM, thus they are more effective at degrading dissolved pollutants (Al Housari et al., 2010; Zafiriou et al., 1984). Both the high reactivity and the non-selectivity of free radicals are factors reducing the indirect photolysis of target pollutants in wastewater, since probabilistically these free radicals attack to matrix components more than organic molecules at trace levels. In addition, the production of free radicals may be also prevented by the light screening exerted by matrix components (Ryan et al., 2011; Boreen et al., 2003). Due to this, even when the generators of free radical are present in wastewater, limited indirect photolysis is expected in this matrix. 3.1.2. Photolysis in surface water Photodegradation of target pollutants in surface water was faster than that observed in wastewater (p b 0.05), as shown in Fig. 2b and

Table 4 Photolysis and biodegradation kinetic constants obtained for the three target emerging pollutants in the study matrices. Compound

Naproxen Carbamazepine Triclosan

Photolysis (k, 1/h)

Biodegradation (k, 1/d)

Wastewater

Surface water

Soil

2.4 × 10−2 (0.991) 2.8 × 10−3 (0.989) 5.9 × 10−2 (0.981)

4.1 × 10−2 (0.986) 6.9 × 10−3 (0.891) 1.3 × 10−1 (0.892)

3.6 × 10−3 (0.986) 4 × 10−4 (0.986) 4.7 × 10−3 (0.986)

Wastewater anoxic conditions ([O2] ~0.3 mg/L)

Wastewater oxic conditions ([O2] ~2.8 mg/L)

High level

Low level

High level

Low level

High level

Low level

High level

Low level

9.5 × 10−3 (0.961) 4.9 × 10−3 (0.879) 1.3 × 10−2 (0.998)

3.5 × 10−2 (0.959) 1.3 × 10−2 (0.968) 3.3 × 10−2 (0.904)

4.5 × 10−2 (0.982) 7.1 × 10−3 (0.931) 8.4 × 10−2 (0.971)

1.02 × 10−1 (0.924) 3.3 × 10−2 (0.923) 9.9 × 10−2 (0.97)

2.5 × 10−2 (0.978) 1.1 × 10−3 (0.965) 1.7 × 10−2 (0.945)

2.9 × 10−2 (0.957) 2.2 × 10−3 (0.962) 2.1 × 10−2 (0.93)

2.5 × 10−2 (0.994) 7 × 10−4 (0.918) 1.2 × 10−2 (0.969)

3.1 × 10−2 (0.971) 5 × 10−3 (0.984) 1.4 × 10−2 (0.956)

Surface water

Soil

Degradation rate constants obtained by fitting to the pseudo-first order model. R2 values are given in parentheses. Initial degradation rate was calculated in cases where the total degradation of the target compound was rapidly achieved. High concentrations refer to: 50 μg/L, 100 ng/L and 50 μg/kg for wastewater, surface water and soil, respectively. Low concentration levels are referred as: 10 μg/L, 20 ng/L and 10 μg/kg for wastewater, surface water and soil, respectively.

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Table 4. Removal of triclosan was complete upon 20 min of irradiation, while for naproxen the maximum removal registered was 98% ± 7 at the end of irradiation experiments. In contrast, the removal of carbamazepine at the end of the experiments was 60% ± 5. The high penetration of light, due to the low concentration of dissolved organic matter and suspended solids, in addition to the high content of dissolved salts in spring water (Table 2) may explain the higher photolysis rates of target pollutants in this matrix when compared to wastewater (Table 4). The penetration of sunlight through water indeed promotes the direct photolysis of the target pollutants. On the other hand, dissolved salts may either promote indirect photolysis by production of free radicals, as occurs for nitrate ions, or hinder this process by the occurrence of free radical scavengers, such as bicarbonate ions (Xu et al., 2011). It is well known that the occurrence of photosensitizers in water, such as nitrite and nitrate ions, promotes the production of oxygenbased radicals (Niu et al., 2013; Ji et al., 2012; Yamamoto et al., 2009) via Eqs. (5) and (6): NO3 þ H2 O þ λυ→• NO2 þ •OH þ OH

ð5Þ

•NO2 þ • OH þ OH→ NO2 þ 2• O þ 2H

ð6Þ

As might be noticed in Table 2, spring water presented a high content of nitrate. Conversely, bicarbonate ions may scavenge free radicals, competing with the target pollutants in the indirect photolysis process (Chowdhury et al., 2011; Zafiriou et al., 1984), as described in Eq. (7): HCO3 þ •OH→CO3 þ H2 O

ð7Þ

This phenomenon may also occur to free radicals formed by the excitation of dissolved organic matter. The pH value of spring water (Table 2) suggests the prevalence of bicarbonate ions in water. Moreover, previous characterization studies have found high concentration of bicarbonate ions in this surface water (401–536 mg/L as CaCO3). Due to this, low rates of photolysis may be expected in this surface water compared to those reported in literature (Boreen et al., 2003). 3.1.3. Photolysis in soil Fig. 2c shows the photolysis kinetics of the target pollutants in the wastewater irrigated soil. Removal rates were 47% ± 1 for triclosan, 32% ± 5 for naproxen and 8% ± 5 for carbamazepine upon 2 h of simulated sunlight irradiation. The photodegradation rates of target pollutants in soil were significantly lower than those obtained for water samples (p b 0.05) as can be seen in Table 4. This is explained by the low penetration of light through the solid matrix (photolysis occurs in the 0.5 mm depth layer). Additionally, the high content of carbonates in soil (due to the accumulation by wastewater irrigation and soil respiration) may result in the decreased photolysis rate of target pollutants even in the photic zone of soil (Mountacer et al., 2014). The extent of photolysis in soil depends on the structure of the soil, the quality and quantity of organic matter and the soil moisture. As occurs in fresh water, photolysis of the dissolved organic matter leads to the generation of free radicals (either oxygen-based radicals or •DOM). These free radicals are produced in the surface layer of the soil and can interact with emerging pollutants resulting in some extent of degradation. It is plausible to hypothesize that triplet state dissolved organic matter may migrate with water to underlying horizons as irrigation event occurs (Frank et al., 2002). 3.2. Biodegradation tests 3.2.1. Biodegradation in wastewater Fig. 3a and b show the biodegradation kinetics of the target pollutants in wastewater, while Table 4 displays the biodegradation rates

under anoxic and oxic conditions. Comparing the biodegradation kinetics obtained for the high and low initial concentration conditions, it is clear that, on the one hand, biodegradation rates decrease as the pollutants' concentration increases (p b 0.05); and on the other hand, biodegradation rate drops as dissolved oxygen concentration decreases. Both effects are clearly observable by comparing the half-life time values for the three emerging pollutants under the different tested conditions. When the initial concentration of the target pollutants was low (10 μg/L) the half-life times (t1/2) of naproxen and triclosan were around 7 ± 0.5 and 20 ± 1 days under oxic and anoxic conditions, respectively. On the other hand, when high initial concentration was used (50 μg/L), half-life time values increased to 8 ± 0.25 and 53 ± 2 for triclosan under oxic and anoxic conditions, respectively; while for naproxen half-life increased to 15 ± 1.2 and 74 ± 3 days under oxic and anoxic conditions, respectively. Regarding carbamazepine, halflife time was higher than 100 days when high initial concentration levels were used under oxic and anoxic conditions. The reduction in the biodegradation rate as the initial concentration of the target pollutants increases is indicative of the inhibitory effect that these molecules exert on the degrading microorganisms in wastewater. Fig. 4 shows the mass of pollutants that was degraded in wastewater using the two initial concentrations. The inhibitory effect was clear for the biodegradation of naproxen and carbamazepine (p b 0.05), while for triclosan no significant differences were observed. This suggests that, even when triclosan is an antibacterial agent, its presence in wastewater at high concentrations does not affect its biodegradation. The effect of triclosan on the degrading organisms is currently under debate; some authors assert that triclosan, along with other antimicrobial agents such as triclocarban, certainly inhibits biodegradation of other emerging pollutants in wastewater treatment systems (Svenningsen et al., 2011; Lawrence et al., 2009). On the other hand, other studies suggest that the occurrence of triclosan does not exert significant impact on the performance of degrading microorganisms in wastewater (Stasinakis et al., 2007). The biodegradation rate of target pollutants was low when dissolved oxygen concentration was ~0.3 mg/L and increased as dissolved oxygen concentration rose up to ~ 2.8 mg/L; this was especially true for naproxen and triclosan. According to our results, naproxen and triclosan can be classified as readily biodegradable in aerobic environments and poorly biodegradable under anaerobic/anoxic conditions, while carbamazepine can be categorized as recalcitrant under both schemes. This is consistent with that reported by Xue et al. (2010) and Suárez et al. (2010) for naproxen and triclosan in wastewater treatment systems under aerobic and anoxic conditions. In both treatment systems, biodegradation of emerging pollutants can be performed by heterotrophic nitrifying bacteria (Bagnall et al., 2012; Roh et al., 2009). Regarding carbamazepine, its recalcitrance has been widely observed in biological treatment systems under a variety of conditions (i.e., aerobic, anoxic, anaerobic, denitrifying and sulfate reducing conditions) (Lam et al., 2004). 3.2.2. Biodegradation in surface water Biodegradation kinetics of target pollutants in surface water are shown in Fig. 3c and d for initial concentrations of 20 and 100 ng/L, respectively. When low initial concentration was tested the half-life times of naproxen, carbamazepine and triclosan were 24 ± 2, 315 ± 4 and 33 ± 2.5 days, respectively. On the other hand, half-life times significantly higher were found when high initial concentration levels were used: 27 ± 1.5 days for naproxen, 630 ± 6 days for carbamazepine and 41 ± 3 days for triclosan. Similar to the observations made for wastewater, microbial inhibition occurred when the initial concentration of the pollutants was increased (Fig. 4). Nevertheless, low biodegradation rates may be also related to a limited biodegradation capacity of the degrading organisms. When the number of degrading cells is low compared to the concentration of the substrate, degradation might be limited in the initial phase even without any toxic or inhibitory effect

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Fig. 3. Biodegradation kinetics of the target pollutants in wastewater using high and low initial concentrations: a) 10 and b) 50 μg/L; surface water, using high and low initial concentrations c) 20 and d) 100 ng/L and soil, using high and low initial concentrations e) 10 and f) 50 μg/kg.

of the chemical. Biodegradation rates were lower in surface water than in wastewater (p b 0.05, Table 4). This may be explained by the lower content of degrading organisms and dissolved organic matter in surface

water compared to that observed in wastewater (Table 2). When inhibition of degrading microorganisms occurs, the biodegradation of target pollutants might be assisted by co-metabolism processes, as suggested

Fig. 4. Biodegraded mass of the target pollutants using two initial concentrations. Gray bars represent high initial concentration (50 μg/L, 100 ng/L and 50 μg/kg for wastewater, surface water and soil, respectively) and black bars are used for low initial concentration (10 μg/L, 20 ng/L and 10 μg/kg for wastewater, surface water and soil, respectively).

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elsewhere (Kim et al., 2011; Gauthier et al., 2010). Nevertheless, due to the low concentration of dissolved organic matter in surface water, co-metabolism might occur as slowly as metabolism. Further studies on the co-metabolic routes for degrading emerging pollutants in environmental matrices are necessary in order to clarify this point.

3.2.3. Biodegradation in soil Fig. 3e and f show the biodegradation kinetics obtained for target emerging pollutants in the wastewater irrigated soil. Biodegradation behavior in soil was similar to that observed for aqueous matrices, i.e., higher dissipation of naproxen and triclosan and the recalcitrance of carbamazepine (Table 4). Half-life times of naproxen and triclosan were 22 ± 1.5 and 52 ± 3 days, respectively, when initial concentration was low, and increased to 28 ± 3 and 58 ± 3 days, respectively, when initial concentration was high. Conversely, carbamazepine half-life time was as high as 138 ± 8 days for low initial concentration conditions, and increased to 990 ± 10 days when initial concentration was augmented. The inhibitory effect on the biodegradation was observed for naproxen and carbamazepine but not when initial concentration of triclosan was arisen (Fig. 4); reinforcing the hypothesis of the low toxicity of this antibacterial agent for the highly acclimatized degrading organisms in natural environments. Even when triclosan is a biocide, its antimicrobial activity is reduced when it is present at trace levels; moreover, considering that aerobic wastewater irrigated soil is an organic matter enriched system co-metabolic paths can be developed by acclimatized bacteria, as been reported for some antibiotics (Verlicchi et al., 2013; Xu et al., 2009). Similar to the observations made in wastewater, biodegradation of target pollutants resulted more efficient in aerobic conditions than in the anaerobic ones (data not shown), which is consistent with the previously reported in literature (Ying et al., 2007). Some studies have focused on finding the microorganisms responsible for the biodegradation of emerging pollutants in soil. These investigations have found that the fungus Trametes versicolor (Marco-Urrea et al., 2010) and the bacteria Pseudomonas putida can degrade naproxen in terrestrial systems, while Rhodococcus rhodochrous can efficiently biodegrade carbamazepine in soil (Gauthier et al., 2010). Future studies should aim to identify the microorganisms that contribute to rapid biodegradation of target pollutants in wastewater irrigated soils displaying differences in the physical and chemical properties. No degradation of target pollutants was detected in sterile controls, indicating that degradation by other abiotic processes did not occur during experiments (see Figure S4 and Table S2 in the Supplementary Information section).

Fig. 5. Breakthrough curves of the target pollutants naproxen, and carbamazepine and the conservative tracer bromide obtained in transport experiments.

3.3. Transport tests 3.3.1. Transport of the conservative tracer through soil columns Fig. 5 depicts the breakthrough curve of the conservative tracer bromide obtained in transport experiments. The asymmetric shape of the breakthrough curve indicates the non-uniform movement of water – and thus solutes – through soil (Melamed et al., 1994). This may be explained by two phenomena: a) the presence of preferential paths within the soil monolith (Simunek et al., 2003); and, b) the repulsion forces (anion exclusion) between bromide and the negatively charged soil particles (i.e., clay and organic matter) (Melamed et al., 1994). Preferential paths in the soil matrix are caused by large and highly interconnected pores in association with the occurrence of roots and stones, the activity of macrofauna (e.g., worms, ants, beetles) and tillage (Galantini and Rosell, 2006). These preferential paths cause the rapid vertical movement of water and solutes through the soil. Anion exclusion implies that bromide moves in the edge of the infiltration front rather than homogeneously dispersed in the infiltrating water, as represented in Fig. 6a (Melamed et al., 1994). The concomitant occurrence of both phenomena explains the high recovery of the conservative tracer at the beginning of the experiment and thus the shape of its breakthrough curve. Negatively charged organic compounds can also experience anion exclusion; for instance, the anti-inflammatory drug naproxen and the antibacterial agent triclosan are likely repelled by the negatively charged particles of the soil as water infiltrates. This is because these molecules are dissociated at some extent at the pH values of the irrigated soil (see log D values and speciation in Table 1). The mass balance of bromide showed no production, degradation or accumulation of the conservative tracer throughout experiments. 3.3.2. Transport of the target pollutants through soil columns Breakthrough curves of the target pollutants are shown in Fig. 5. These curves were more symmetrical than that of the conservative tracer, although the Gaussian form was not achieved, indicating physical non-equilibrium in the transport of the organic compounds. The transport parameters of the target pollutants and the conservative tracer are presented in Table 5. The three target pollutants leached out the soil column after the lixiviation of the conservative tracer, which reveals interactions between soil and organic molecules and thus the delay in their transport. According to the retardation factor (RF) values, the order at which pollutants left out the soil column was: naproxen followed by carbamazepine, whereas triclosan was barely detected in the leachate samples, suggesting its irreversible sorption onto the soil particles and/or its degradation within soil monolith. The latter is more consistent with the high biodegradation rate observed in batch biodegradation experiments (see Section 3.2.3). The most important phenomenon causing the retention of organic pollutants in soil is adsorption. Previous studies in the Tula Valley have reported the rapid adsorption of the target pollutants in the irrigated soil; which was notably stronger for carbamazepine and triclosan (Durán-Álvarez et al., 2012). Such behavior explains the slow movement of the target compounds observed in this column experiment. Hydrophobic compounds, such as triclosan and carbamazepine (see Table 1), can be retained onto soil particles by hydrophobic partition to soil organic matter. Chemical interactions between the aromatic groups within pollutants' molecules and the humified organic matter in soil can occur through the formation of π–π bonds (Chefetz et al., 2008; Benoit et al., 1996). On the other hand, the anionic form of naproxen is prevalent at the soil pH values (see Table 1) and thus, similar to bromide, naproxen can undergo anion exclusion. However, the transport of naproxen was found to be slower than that of the conservative tracer (Table 5). This may be explained by the hydrophobic nature of the organic molecule (as revealed by its chemical structure shown in Table 1), which can experience partition onto the soil organic matter. Moreover, the aromatic moieties within naproxen molecule can establish π–π bonds with the

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Fig. 6. (a) Schematic representation of anion exclusion within soil matrix, (b) mass balance of the target pollutants in the unaltered soil columns (units in μg).

humified organic matter in soil. However, π–π bonds could not be formed as effectively as may occur with carbamazepine, given that a part of the naproxen molecule (the carboxylate moiety) is at the same time repelled by negatively charged soil particles. Mass balance was performed at the end of the transport experiments (Fig. 6b); the missing mass of the pollutants both in soil and leachate samples was assumed to be biodegraded. According to mass balance, the biodegradation rate of carbamazepine was of 24%, while for naproxen it was 14%. The higher degradation of carbamazepine compared to that of naproxen was attributed to the shorter residence time of the anti-inflammatory agent in the soil column (i.e., naproxen leached out the soil column significantly earlier than carbamazepine and triclosan). Regarding triclosan, low recoveries were observed both in leachates and soil. This may be explained either by the formation of non-extractable residues in soil or by the biodegradation of the antibacterial agent. Due to the use of the exhaustive ASE extraction method, which showed recoveries levels higher than 90% in previous studies (Durán-Álvarez et al., 2012), the biodegradation of the compound is more likely to occur than the formation of non-extractable residues. Triclosan showed the highest biodegradation rate in the soil columns, since 93% of the initially added mass was not found in leachate samples or recovered by exhaustive extraction of soil. The results of batch biodegradation experiments showed disappearance of triclosan close to 80% after 70 days of incubation. Increment in the biodegradation rate (k) value in a factor of 2.7, 3.2 and 25 were found for naproxen, carbamazepine and triclosan, respectively, comparing column and batch tests. The increment in k values was directly related with the increase of log D value of the target compounds (see Table 1), i.e., with the increment in their retention within the soil column. Differences in the biodegradation rates for batch and column experiments can be explained by the differences in the experimental conditions of both tests. In the former, biodegradation occurred in stationary conditions, maintaining gravimetric soil moisture in 250 mg H2O/g soil, whereas in column experiments, biodegradation took place under dynamic flow conditions, in which the gravimetric soil moisture was 320 mg H2O/g soil. Additionally, the flow of water through the soil monolith might facilitate the transport of nutrients from the top of the soil column to deeper layers. Half-life time values were determined for naproxen, carbamazepine and triclosan in soil columns as 31, 43 and 86 days. According to these

Table 5 Transport parameters obtained for the target pollutants and the conservative tracer using the CXTFIT 2.1 code. Solute

RF

Degradation rate (1/d)

Model

Bromide Naproxen Carbamazepine Triclosan

b1 1.5 4.3 N8

0 2.2 × 10−2 6 × 10−3 3.6 × 10−1

Physical non-equilibrium

values, and considering that one pore volume leach out the soil columns in 27 h, it is possible to establish that triclosan is degraded within the soil column, while naproxen and carbamazepine are able to leave the soil monolith before complete biodegradation take place. In addition, it is important to consider the strong adsorption of the compounds as well as the formation of non-extractable residues of the compounds. Half-life times determined in soil columns reinforce the hypothesis of the complete biodegradation of triclosan rather than the formation of non-extractable residues in soil columns. 3.4. Environmental relevance In this study the degradation and transport parameters of target emerging pollutants were determined; the environmental relevance of such results is discussed below. Under the field conditions, degradation, adsorption and transport processes are occurring concomitantly, some of them can be somehow competitive (e.g., photolysis and adsorption in soil) or antagonistic (e.g., leaching and degradation). Environmental conditions in the field during and upon irrigation rule the environmental fate of emerging pollutants in wastewater, irrigated soils and surface water near the irrigation area. For instance, since Tula Valley is located in an arid zone, continuous sunlight irradiation and warm weather favor photo and biodegradation. The irradiation dose supplied to samples in photolysis experiments (3600 kJ/cm2) is equivalent to 48 h of direct summer-day-sunlight irradiation in the study area. Considering the photodegradation kinetics obtained in wastewater experiments (see Fig. 2), the complete removal of triclosan is achieved in less than 24 h of sunlight irradiation in summer season; this is less than the time wastewater takes to be transported from Mexico City to the croplands. However, given that triclosan is still present in wastewater of irrigation canals (data not shown), it is clear that photolysis in wastewater occurs more rapidly in laboratory than under field conditions. Photolysis of emerging pollutants in wastewater can be occurring more effectively during the storage in dams, and in a lower extent when wastewater is supplied to croplands upon irrigation. In order to better understand the impact of natural photolysis on the fate of emerging pollutants in wastewater, it is necessary to elucidate whether pollutants are photodegraded in wastewater before it infiltrates through soil or if photodegradation starts once pollutants are adsorbed onto the soil particles. The same components retarding the photodegradation of emerging pollutants in wastewater are actually favoring their biodegradation; high concentration of dissolved and particulate organic matter (the latter associated with the occurrence microorganisms) is a key factor in the biodegradation process. However, as is showed in Fig. 3, biodegradation is only occurring under aerobic conditions, thus removal of target pollutants is expected to occur only during transportation of wastewater within the irrigation area, since only in open canals turbulence and exchange of gases with atmosphere is achieved. According to this, even when high biodegradation rates

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were found in wastewater experiments, such degradation results are applicable only during a short time lapse. Photodegradation in soil occurred in a lower extent than that observed in wastewater. Adsorption of target pollutants onto the soil has a determinant role in photolysis. This is because, on the one hand, the organic molecules retained on the soil particles can be reached by sunlight, and conversely, some molecules can be adsorbed in sites inaccessible to irradiation, impeding their photolysis (Mountacer et al., 2014). Thus the impact of adsorption on the photodegradation of emerging pollutants in soil should be carefully studied. In this study, readily adsorbable compounds, such as triclosan and carbamazepine showed dissimilar photodegradation rates, indicating that some other factors are impacting in the photodegradation of these compounds. In the case of naproxen, leaching precedes photodegradation (as indicated the low retardation factor of the anti-inflammatory drug), which results in the nullification of photolysis. Additionally, it is important to take into account the effect of the soil structure in the photolysis; sunlight can easily penetrate through pores in structured soils, while photolysis may be lower in non-structured soils; thus degradation rates determined under laboratory conditions (in which soil was grounded and sieved prior to irradiation experiments) could underestimate the photolysis yields occurring in the field. Similar to that observed in wastewater, biodegradation in soil showed to be a more important degradation process than photolysis. In the field, air is displaced from soil pores as wastewater infiltrates, thus anaerobic conditions are temporarily achieved within the solid matrix; then aerobic conditions are gradually restored as wastewater infiltrates. Given that aerobic conditions prevail most of the time in the irrigated soil, it is plausible to elucidate the biodegradation in the field by using the degradation rates determined in this work. According to our lab-scale results, it is estimated that the complete biodegradation of naproxen and triclosan in soil is achieved 80 days after irrigation event. However, the occurrence of these emerging pollutants has been found as perennial in the agricultural soils of the zone in previous studies (Durán-Álvarez et al., 2009). This may be explained by two factors: a) wastewater irrigation is carried out each 20-30 days allowing the accumulation of pollutants in soil; and, b) soil moisture is dramatically reduced in the time frame between irrigation events, which decreases the microbial activity. In general terms, aerobic conditions in agricultural soils of the Tula Valley prevail up to 2 m depth; considering that soil retain water in lower horizons for longer periods of time, it is expected that aerobic biodegradation occurs even at such depth, especially for mobile compounds such as naproxen. Further studies should aim to determine the biodegradation and transport parameters of emerging pollutants in deeper layers of the soil profile, given that transport parameters may be indeed modified because of the changes in composition and aeration of deep soil. Target pollutants showed to be readily photodegraded in surface water. Considering that under field conditions surface water is exposed to sunlight for long periods and that photodegradation kinetics showed that naproxen and triclosan are photolyzed in less than 48 h of solar irradiation at field conditions, it is plausible to consider that both molecules are fully removed in surface water by photolysis more than biodegradation. In the case of carbamazepine, complete photodegradation occurs upon more prolonged time lapse than that observed for triclosan and naproxen thus its occurrence may be expected in this water source. Biodegradation behavior of naproxen and triclosan was quite similar in water matrices, while important differences were found when comparing results of both pollutants in soil. This may be attributable to the differences in the adsorption of both compounds, which indeed modify their biodegradation. 4. Conclusions Dissipation and transport experiments were conducted using three emerging contaminants as model molecules. The results of these tests showed that naproxen and triclosan are readily degraded by abiotic

and biotic processes in water and soil. Conversely, carbamazepine was observed to be recalcitrant both in photolysis and biodegradation tests, although it can be said that this compound was more photodegraded than biodegraded. The increment in the concentration of emerging pollutants in the study matrices positively decreased their biodegradation. Results of column tests indicate that naproxen leaches rapidly to the aquifer, although the compound could be readily biodegraded in soil and photodegraded during the transportation of wastewater in irrigation canals, thus the presence of this drug in the aquifer will depend on the rate of wastewater infiltration. Carbamazepine and triclosan, on the contrary, are retained by the soil where rapid biodegradation of the latter occurs while the former is accumulated due to its low degradation rate. Triclosan was found to be completely degraded in soil, thus further studies should focus on the environmental fate of its biodegradation by-products as well as the potential for them to reach the aquifer. Differences in biodegradation rates were found comparing batch and column experiments, such differences increased with the retention time of pollutants within the soil column. This work provides relevant information on the environmental fate of emerging pollutants in wastewater, surface water and wastewater irrigated soil. Such information may be used to optimize the conditions of wastewater reuse in agricultural districts of arid areas, taking advantage of factors such as: a) high solar irradiation; b) high load and acclimatization of degrading organisms in the soil; and, c) the changes in physical and chemical composition of the irrigated soils in order to achieve greater depuration of wastewater in terms of removal of emerging pollutants. In the case of wastewater, the same matrix components retarding the photodegradation favor biodegradation. Future experiments varying the organic matter content in water and soil matrices could give some insight on the role of this component in the biotic and abiotic degradation of emerging pollutants in soil and water. Acknowledgements The authors would like to express their gratitude to Professor René Alcalá from the Laboratorio de Fisica de Suelos del Instituto de Geología, UNAM for the support in the analysis of soil samples, as well as to Dirección General de Asuntos del Personal Académico de la UNAM (DGAPA) and the Instituto de Ciencia y tecnología del Distrito Federal for funding this work within the framework of projects IN101610 and ICyTDF/324/2011, respectively. Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2015.08.028. References Acuña, V., von Schiller, D., García-Galán, M.J., Rodríguez-Mozaz, S., Corominas, L., Petrovic, M., Poch, M., Barceló, D., Sabater, S., 2015. Occurrence and in-stream attenuation of wastewater-derived pharmaceuticals in Iberian rivers. Sci. Total Environ. 503, 133–141. Adrover, M., Farrús, E., Moyá, G., Vadell, J., 2012. Chemical properties and biological activity in soils of Mallorca following twenty years of treated wastewater irrigation. J. Environ. Manag. 95, S188–S192. Al Housari, F., Vione, D., Chiron, S., Barbati, S., 2010. Reactive photoinduced species in estuarine waters. Characterization of hydroxyl radical, singlet oxygen and dissolved organic matter triplet state in natural oxidation processes. Photochem. Photobiol. Sci. 9, 78–86. Andreozzi, R., Raffaele, M., Nicklas, P., 2003. Pharmaceuticals in STP effluents and their solar photodegradation in aquatic environment. Chemosphere 50, 1319–1330. Bagnall, J.P., Ito, A., McAdam, E.J., Soares, A., Lester, J.N., Cartmell, E., 2012. Resource dependent biodegradation of estrogens and the role of ammonia oxidizing and heterotrophic bacteria. J. Hazard. Mater. 239-240, 56–63. Bahnmüller, S., von Gunten, U., Canonica, S., 2014. Sunlight-induced transformation of sulfadiazine and sulfamethoxazole in surface waters and wastewater effluents. Water Res. 57, 183–192.

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