Degradation kinetics and mechanism of sulfadiazine and sulfamethoxazole in an agricultural soil system with manure application

Science of the Total Environment 607–608 (2017) 1348–1356

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Degradation kinetics and mechanism of sulfadiazine and sulfamethoxazole in an agricultural soil system with manure application Yu Zhang a,b, Shuangqing Hu b, Hongchang Zhang b, Genxiang Shen b, Zhejun Yuan c, Wei Zhang a,⁎ a State Environmental Protection Key Laboratory of Environmental Risk Assessment and Control on Chemical Process, School of Resource and Environmental Engineering, East China University of Science and Technology, Shanghai 200237, China b Shanghai Academy of Environmental Sciences, Shanghai 200233, China c College of Environmental Science and Engineering, Donghua University, Shanghai 201620, China

H I G H L I G H T S

G R A P H I C A L

• Accelerated removal of SDZ and SMX in an agricultural soil system was achieved by the addition of manure. • Degradation kinetics and metabolites of SDZ and SMX were revealed. • Degradation pathways and mechanisms of SDZ and SMX were proposed.

The removal rates of the two antibiotics under different incubation conditions after 49 days (non-sterile for a non-sterile soil, sterile for a sterile soil, Single-application for manure added 8 g on 0 d, and Repeated-application for manure added 2 g on 0, 7, 14 and 21 d, respectively).

a r t i c l e

a b s t r a c t

i n f o

Article history: Received 25 May 2017 Received in revised form 10 July 2017 Accepted 10 July 2017 Available online xxxx Editor: Jay Gan Keywords: Sulfadiazine Sulfamethoxazole Soil Manure Degradation pathways

⁎ Corresponding author. E-mail address: [email protected] (W. Zhang).

http://dx.doi.org/10.1016/j.scitotenv.2017.07.083 0048-9697/© 2017 Elsevier B.V. All rights reserved.

A B S T R A C T

Recently, under the application of waste-water, manure and biosolids, antibiotics have been used massively in agriculture resulted in antibiotic resistance and potential environmental risks. In the present study, the removal of sulfadiazine (SDZ) and sulfamethoxazole (SMX) in an agricultural soil system was explored. All the experiments were conducted under different incubation conditions for 49 days. The experimental results indicated that all the degradation processes could effectively follow a first-order kinetic model. Based on the analyses of these two antibiotics, SDZ had a higher reaction rate and a shorter DT50 value. Additionally, there were no marked differences in DT50 values at varying initial concentrations under the same conditions (p N 0.05). Compared with the non-sterile soil, the degradation rates of SMX and SDZ were slower (b 70%), and the associated DT50 values (N21 days) were higher in the sterile soil. Because the biodegradation played a major role, it may be effective for the removal of these contaminants from the soils. The processes of SDZ and SMX degradations were slightly accelerated by applying manure (b20%). There were different accelerating effects on the removal of SDZ and SMX in soils by manure Single- and Repeated-application, which may be related to the amount of manure during the degradation processes, and different methods of adding manure could only affect the degradation rate. The major intermediate products were derived from the hydroxylation, sulfonamide S\\N bond cleavage and aniline moiety oxidation. Therefore, the present study inferred that possible degradation pathways

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of SDZ and SMX were hydroxylation of the benzene ring, oxidation of the amine group at the benzene ring, ring open and S\\N bond cleavage. Results revealed that more attention should be paid to the transformation products because they could be more toxic than the parent compounds. © 2017 Elsevier B.V. All rights reserved.

1. Introduction Antibiotics are a large group of pharmaceuticals that have been used worldwide to prevent bacterial infections and treat diseases (Fedorova et al., 2014; Grabicova et al., 2015; Johnson et al., 2015). The consumption of antibiotics has been gradually increasing with industrial, agricultural and medical development in recent years. In the United States, the utilization of antibiotics for livestock production annually exceeded 11,200 t (Mellon et al., 2001). In China, it was reported that nearly 1.50 Mt of antibiotics were annually employed in the livestock industry by wastewater effluent (Zhang et al., 2015). Among the different classes of antibiotics, sulfonamide antibiotics (SAs) were of particular interests (Hu et al., 2010). Sulfadiazine (SDZ) and sulfamethoxazole (SMX) were regarded as representatives of the universal antibiotic types with higher concentrations in the environment due to their heavy usages in human and animal medicines (Yan et al., 2013). There were substantial transformation products excreted by animals that would be released into the environment, and could be subsequently converted to more toxic products than their parent compounds (Kühne et al., 2000; Søeborg et al., 2004). It has been reported that antibiotic metabolites may maintain their pharmacological activities (Johne et al., 2005). Metabolites had longer persistence in soils despite those target pollutants had been degraded completely, and they would transport to ground water (Koba et al., 2016). Investigating their degradation contributes to the understanding of their persistence and potential mobility. Furthermore, the antibiotics usages for veterinary medicine introduced constant field contamination through animal or land application of manure. Therefore, it was essential to investigate the degradation kinetics and mechanism of antibiotics in a soil system. Most reports have implemented the degradation kinetics of antibiotics in aqueous and soil environments (Dorival-García et al., 2013; Hektoen et al., 1995). Several literatures have reported that antibiotics persistence in soils was mostly dependent on soil types (Koba et al., 2016; Kodešová et al., 2016), and dissipation could be affected by many environmental factors (Feng et al., 2010; Srinivasan and Sarmah, 2014). Additionally, some scholars investigated land application of biosolids or compost on mineralization (Li et al., 2015), sorption (Kodešová et al., 2015) and transport in soil columns (Siemens et al., 2010). However, the information about SAs by adding organic manure into soil after the fate of system characteristics was still scarce. Due to deficient information related to antibiotics degradation in soils, the transformation products, and further degradation processes, as well as different removal mechanisms can't be better understood. The aim of this study was (1) to investigate the degradation behaviors of two select antibiotics (SDZ and SMX) in the sterile and nonsterile soils under aerobic conditions; (2) to compensate the shortage of investigations of sulfonamide antibiotics by adding organic manure into soil after the fate of system characteristics; (3) to reveal the fate of the selected antibiotics and their possible metabolites in soils. This paper proposed the reaction pathways related to the transformation of SDZ or SMX on the conclusions of previously corresponding studies. 2. Materials and methods 2.1. Chemicals, soil and manure Sulfadiazine (SDZ; purity N 98%), Sulfamethoxazole (SMX; purity N 98%) and Sulfamethoxazole-13C6 ([13C6]-SMX; purity N 99.5%) were

obtained from Dr. Ehrenstorfer (Augsburg, Germany). All solvents (acetonitrile, methanol and formic acid) used were of high-performance liquid chromatography (HPLC) grade from Tedia Company (Fairfield, USA) and HPLC grade H2O was used for dilutions. Stock solutions of 1000 mg·L− 1 were made by dissolving each standard in methanol and storing at −20 °C in a brown glass bottle. Working solutions and internal standard (5 mg·L−1) were made freshly by diluting the stock solutions with methanol and stored at 4 °C. The physiochemical properties of these two antibiotics were presented in Table 1. The surface soil (0–20 cm) used in these experiments were collected from an agricultural field in Shanghai, southern China. SDZ and SMX were not detectable at this site. Soil samples were sieved (b 2 mm) to remove stones, gravel and plants after air drying, then stored at 4 °C until use. The cattle manure used in this research was collected from a cattle farm, where both SDZ and SMX were not detected. The manure was stored at 4 °C for repeated applications (the structure of dominant microbial communities and physiochemical properties were barely changed). The physiochemical properties of the soil and the manure were listed in Table 2.

2.2. Degradation experiment The degradations of SDZ and SMX in the sterile and non-sterile treatments were examined under an aerobic condition. Exactly 200 g of dried agricultural soil was weighed into a 250 mL conical flask. The soils were sterilized by autoclaving at 121 °C for 20 min for three times before spiking with the antibiotics and being pre-incubated at 20 °C for 2 days (autoclaving was a less harmful method for physicochemical properties of the soil). The soil moisture was held at 60% of maximum water holding capacity (MWHC) by adding sterilized deionized water. Three different initial concentrations (4, 10 and 20 mg·kg−1, SDZ and SMX were mixed into the soils) of each antibiotic were selected considering their instrumental detection limits, and then the methanol was evaporated in a fume cupboard. The conical flasks were weighed every day and deionized water was added to make up for a water loss. Each opened conical flask was hand-shaken to let air in to keep an aerobic condition when necessary. The conical flasks were wrapped with aluminum foils to avoid photolysis. The samples were protected from light and incubated in an airy incubator at room temperature (20 ± 1 °C). The same experimental procedure was carried out in the non-sterile treatments except autoclaving. SDZ and SMX with manure mixed into the soil were explored in another experiment. The manure containing these two antibiotics was mixed with 200 g dried soil to attain a typical manure addition level of 4% (w/w) (Heuer et al., 2011), three antibiotic concentrations of 4, 10, 20 mg·kg− 1 with dry soil, and then stirred well by using a glass rod. The mixture was placed in a fume cupboard to evaporate the methanol. Two ways of applying manure were as follows: Single-application (treatment A, added 8 g on 0 d) and Repeated-application (treatment B, added 2 g on 0, 7, 14 and 21 d, respectively). The experimental treatments included Manure (Single-application) + SDZ + SMX and Manure (Repeated-application) + SDZ + SMX. The following steps were same as that described above for the sterile and non-sterile groups. Three conical flasks were prepared at each concentration for per treatment, and the samples (5 g dried soil) were collected on 0, 1, 3, 7, 14, 21, 28, 35, 42 and 49 d from every conical flask. All experiments were conducted in triplicates.

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Table 1 The physiochemical properties of these two antibiotics. Molecular formula

Molecular weight

pKa

Koc

logKow

Sulfadiazine

C10H10N4O2S

250.28

pKa1 = 2.0 pKa2 = 6.5a

114.4a

0.8a

Sulfamethoxazole

C10H11N3O3S

253.27

pKa1 = 1.8 pKa2 = 6.0a

94.9a

0.9a

Antibiotic

Structure

pKa, negative log of acid dissociation constant. logKow, the logarithm of the octanol/water partition coefficient. a Stoob et al. (2007).

d) was calculated by the equation (DT50 = ln2/k). The removal ratio of each antibiotic was expressed by the following equation:

2.3. Extraction and sample clean-up Exactly 5 g of dried soil was weighted into the tube, and then 50 ng [13C6]-SMZ was added. 20 mL of McIlvaine extraction buffer at pH 4 containing 0.1 M EDTA was added to the samples. McIlvaine buffer (500 mL) was prepared by mixing 243 mL of 0.1 M citric acid (CA) and 257 mL of 0.2 M Na2HPO4 (Mcilvaine, 1921). The samples were mixed on a vortex for 1 min to homogenize the solution, and then the mixture was subsequently placed into an ultrasonic bath for 30 min at 25 °C. After cooling to room temperature, they were centrifuged at an approximate 3000 rpm for 20 min. Finally, 10 mL of the supernatants were collected into 50 mL centrifuge tubes. The above operations were repeated for three times, and the extracts were then passed through HLB cartridges for purification. The SPE cartridges were conditioned with 5 mL of methanol (MeOH) and 5 mL of H2O. Samples were loaded on the cartridges by gravity. The SPE cartridges were washed with 5 mL of H2O:MeOH (95:5, v/v). The analyses were eluted with 10 mL of MeOH into conical-bottom centrifuge tubes, and then were evaporated under a gentle nitrogen stream at room temperature before HPLC analysis. 2.4. Method validation Identification and analysis of compounds and their transformation products were determined by HPLC-MS/MS and UHPLC-Q-TOF-MS (Text S2). All samples were analyzed in triplicate. The recoveries of these two antibiotics ranged from 79.92 to 100.67% on average (Table S1). The LOQs of SDZ and SMX ranged from 0.56 to 2.76 μg·kg−1. Eight concentrations (0.1–12.8 mg·L−1) of individual antibiotics were used to construct a calibration curves (R2 N 0.99). Masslynx 3.4 work station software was used for data processing. 2.5. Data analysis The data from the degradation experiments were analyzed using a first-order kinetic model, and they were fitted into the following kinetic formula: Ct ¼ C0 e−kt

ð1Þ

Where C0 is the initial concentration of these two antibiotics added to soil (mg·kg−1), Ct is the time-varying concentration of the antibiotic (mg·kg−1), e is the base of the natural logarithm, t is the time point in days, and k is the degradation rate constant (d−1). The half-life (DT50,

Removal ratio ¼ ðC0 −Ct Þ C0 −1  100%:

ð2Þ

Degradation rate equation, which was calculated using the derivative of first-order reaction kinetic Eq. (1) with respect to time, was defined as follows: dCt η ¼ dt

ð3Þ

where η denotes the degradation rate for a specific time (mg·kg−1·d−1). All statistical analyses were carried out using Origin 8.0 for Windows. The presented results were mean values ± standard deviation from three independent samples with an experimental error below 10%. 3. Results and discussion 3.1. Degradations of these two antibiotics in the sterile and non-sterile soils Concentration variations of SDZ and SMX in the test soils during the 49 d incubation period were showed in Fig. 1a and b. For SDZ at 4 mg·kg−1, a rapid degradation in the non-sterile soil was observed in the beginning of the 7 days with a removal ratio N 60%, followed by a very slow dissipation (Fig. 1a). After the experiment, the concentrations of SDZ in the non-sterile and sterile soils were 0.38 and 1.54 mg·kg−1, respectively. Their degradation fractions were 90.5 and 61.5%. The data in Fig. 1c and d revealed that degradation rates were slower in the sterile soil compared with the non-sterile soil. Similarly, for SDZ at 4 mg·kg− 1, the degradation rates on days 0, 1, 3 and 7 in the nonsterile soil were 0.261, 0.241, 0.205 and 0.148 mg·kg−1·d−1, respectively. Additionally, in the sterile soil they were 0.081, 0.079, 0.075 and 0.069 mg·kg−1·d−1, respectively. Similar results were observed between these two antibiotics under the same incubation conditions (p N 0.05). Degradation rates at varying initial concentrations (4, 10 and 20 mg·kg− 1) for these two chemicals at the same time point were obviously different (p b 0.05), indicating that degradation rate was relied on organic matter and primary substrate activity. A faster degradation rate during the first 7 days might be explained by the existence of a great deal of organisms, and an antibiotic resistance development during the exposure period. It had been shown earlier that the faster degradation was observed in the non-sterile clay loam under an aerobic condition, with SAs having DT50 of 2 days (Feng et al., 2010).

Table 2 The physiochemical properties of the soil and the manure. Materials

Total organic carbon (g·kg−1)

Moisture content

Cation exchange capacity (cmol(+)·kg−1)

pH

Clay

Silt

Sand

Soil Manure

4.71 362.5

60% 40%

7.0 –

7.6 8.1

4.28% –

58.76% –

36.96% –

Y. Zhang et al. / Science of the Total Environment 607–608 (2017) 1348–1356

(a)

21

(c)

20 Non-sterile 20 Sterile

14

0.4

Degradation rate (mg kg d )

-1

21

10 Non-sterile 10 Sterile

·

14

·

7 0 6

4 Non-sterile 4 Sterile

3

10 Non-sterile 10 Sterile

0.8 0.4 0.0

4 Non-sterile 4 Sterile

0.8 0.4 0.0

0 0

10

20 30 Time (d)

40

50

0

10

20 30 Time (d)

40

50

(d)

21

20 Non-sterile 20 Sterile

14

20 Non-sterile 20 Sterile

0.8 0.4

Degradation rate (mg kg d )

7

-1

-1

0 -1

0.0

·

-1

-1

Residual (mg kg )

0

Residual (mg kg )

20 Non-sterile 20 Sterile

0.8

7

(b)

1351

21

10 Non-sterile 10 Sterile

·

14

·

7 0

6

4 Non-sterile 4 Sterile

3

0.0

·

10 Non-sterile 10 Sterile

0.8 0.4 0.0

4 Non-sterile 4 Sterile

0.8 0.4 0.0

0 0

10

20

30

40

50

0

Time (d)

10

20 30 Time (d)

40

50

Fig. 1. Dissipation and degradation rates of SDZ and SMX in sterile and non-sterile soils at various initial concentrations (4 for 4 mg·kg−1, 10 for 10 mg·kg−1 and 20 for 20 mg·kg−1): (a) dissipation of SDZ, (b) dissipation of SMX, (c) degradation rate of SDZ, (d) degradation rate of SMX. The points and error bars represent the mean and standard deviation of replicates, respectively (n = 3).

The degradation kinetic data set for all the treatments between the experimental conditions during the incubation period was presented in Table 3. The coefficients of determination (R2 ) ranged from 0.861 to 0.933 (p N 0.05). As shown in Table 3, DT50 values for these two antibiotics in the non-sterile soil were lower than those in the sterile soil. Additionally, SDZ degraded fastest in the non-sterile soil at varying initial concentrations, and SMX had the highest DT50 value in the sterile soil. It was suggested that the biodegradation could play a major role in the degradations of SAs in

soils, which was in agreement with previous reports (Feng et al., 2010; Lin and Gan, 2011). The results indicated that the residuals of SDZ and SMX within 49 days would decrease slightly based on abiotic processes in the sterile treatments. Photolysis could play no effects on the abiotic degradations of these two chemicals because the samples were incubated in the dark. It could be considered that hydrolysis was a main abiotic degradation pathway, because it had structural features relating to be hydrolysed. Irreversible binding with the soil components was attributed to the

Table 3 First-order reaction kinetic modeling data for the degradations of SDZ and SMX in sterile and non-sterile soils. Antibiotic

Initial concentration (mg·kg−1)

Treatments

k (mg·kg−1·day−1)

Equation

R2

DT50 (days)

SDZ

4

Non-sterile Sterile Non-sterile Sterile Non-sterile Sterile Non-sterile Sterile Non-sterile Sterile Non-sterile Sterile

0.0817 0.0230 0.0772 0.0261 0.0678 0.0327 0.0507 0.0302 0.0674 0.0209 0.0641 0.0304

Ct = 3.20e−0.0817t Ct = 3.50e−0.0230t Ct = 8.32e−0.0772t Ct = 9.22e−0.0261t Ct = 15.87e−0.0678t Ct = 17.78e−0.0327t Ct = 3.26e−0.0507t Ct = 3.65e−0.0302t Ct = 7.94e−0.0674t Ct = 9.08e−0.0209t Ct = 16.69e−0.0641t Ct = 18.28e−0.0304t

0.906 0.878 0.901 0.906 0.915 0.933 0.905 0.936 0.861 0.863 0.905 0.928

8.48 30.09 8.97 26.53 10.22 21.21 13.68 22.99 10.28 33.24 10.81 22.79

10 20 SMX

4 10 20

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abiotic losses of SDZ and SMX in a soil system. The sulfonamide antibiotics tend to bind with soil and manure organic matter via the stable covalent binding mechanisms (Bialk et al., 2005; Kahle and Stamm, 2007). Previous reports revealed that the biotransformation processes of SAs were discovered in water and sediment (Radke et al., 2009), although they were resistant to the biodegradation in other reports (Lai and Hou, 2008; Radke et al., 2009). The present study explicitly illustrated that biological processes were a dominating factor in the degradations of SAs in soils (Al-Ahmad et al., 1999; Lindberg et al., 2005). The results also demonstrated that there was no marked differences in the concentration dissipations and DT50 values (p N 0.05) for both SDZ and SMX at varying initial concentrations under the same conditions. Though there was a 15–17% increase in the rate constants of these two antibiotics at high concentrations, it was inconclusive whether the persistence of SAs was affected by initial chemical levels. In contrast to our findings, Wang et al. (2006) signaled that the degradation rate declined with increasing sulfamethoxine concentration in manure, suggesting that the activities of degrading microorganisms were inhibited at high concentrations. Yang et al. (2009) reported that the DT50 values for 1, 10 and 25 mg·kg−1 initial levels of SDZ were 2, 18 and 34 days, respectively. An examination of literature data suggested that degradation kinetics of SAs were dependent on the initial concentrations and removal rates would be slower at a higher concentration. A plausible explanation was that the higher initial concentration of antibiotics often involved larger volumes of methanolic stock solutions, which could indirectly inhibit microbial activities. Another reason was

(a)

21

the high dosage of antibiotics could be lethal to soil microbes (Ma et al., 2001), thus lowering microbial activities of the soils. 3.2. Degradations of these two antibiotics with manure application under various conditions The degradations of SDZ and SMX in the test soils with different manure application methods were presented in Fig. 2. The results displayed that the concentrations of SDZ and SMX declined over time for all treatments. It could be observed from the degradation rates (Fig. 2c and d), signifying that a rapid degradation occurred in the first 7 days. Thereafter, the antibiotics degraded much more slowly and their residuals became balanced after 49 d incubation. During a 7 d incubation period, the removal percentages of SDZ (4 mg·kg− 1) for treatment A and B were 66.25 and 60.75%, respectively. For both of SDZ and SMX, a relatively higher degradation rate was probed from day 0 to day 14 for group A (Fig. 2c and d). Although the total amount of manure added to these two groups after 21 days was equal, the degradation rate in treatment B was much slower. It indicated that the amount of manure added in the beginning stage may increase microbial biomass and adsorption capacity. The rapid degradations of SDZ and SMX may have been caused by soil and manure adsorption of antibiotics through stable covalent binding to soil and manure organic matter (Förster et al., 2009; Schmidt et al., 2008). After the experiment, the concentrations of SDZ for treatment A and treatment B were 0.276 and 0.526 mg·kg− 1 , and

(c)

20 Single 20 Repeated

14

0.8

-1

-1 -1

0

Residual (mg kg )

Degradation rate (mg kg d )

7

21

10 Single 10 Repeated

·

14 7 0 6

4 Single 4 Repeated

0.0

·

1.6 0.8 0.0

0.8 0.0 0

10

20 30 Time (d)

40

50

0

20 Single 20 Repeated

20 30 Time (d)

40

50

20 Single 20 Repeated

1.6

7

-1

-1

Degradation rate (mg kg d )

0.8

0 -1

10

(d)

21 14

Residual (mg kg )

4 Single 4 Repeated

1.6

0

·

10 Single 10 Repeated

·

3

(b)

20 Single 20 Repeated

1.6

21

10 Single 10 Repeated

14 7 0 6

4 Single 4 Repeated

3

·

0.0 10 Single 10 Repeated

·

1.6 0.8 0.0

4 Single 4 Repeated

1.6 0.8

0

0.0 0

10

20 30 Time (d)

40

50

0

10

20 30 Time (d)

40

50

Fig. 2. Dissipation and degradation rates of SDZ and SMX in single and repeated manure application soils at various initial concentrations (4 for 4 mg·kg−1, 10 for 10 mg·kg−1 and 20 for 20 mg·kg−1): (a) dissipation of SDZ, (b) dissipation of SMX, (c) degradation rate of SDZ, (d) degradation rate of SMX. The points and error bars represent the mean and standard deviation of replicates, respectively (n = 3).

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Table 4 First-order reaction kinetic modeling data for the degradations of SDZ and SMX with manure application under various conditions. Antibiotic

Initial concentration (mg·kg−1)

Treatments

k (mg·kg−1·day−1)

Equation

R2

DT50 (days)

SDZ

4

Single Repeated Single Repeated Single Repeated Single Repeated Single Repeated Single Repeated

0.1077 0.0615 0.1771 0.0760 0.1252 0.0665 0.0737 0.0558 0.1042 0.0617 0.1318 0.0847

Ct = 3.33e−0.1077t Ct = 3.20e−0.0615t Ct = 8.26e−0.1771t Ct = 7.71e−0.0760t Ct = 16.23e−0.1252t Ct = 16.66e−0.0665t Ct = 3.04e−0.0737t Ct = 3.16e−0.0558t Ct = 8.09e−0.1042t Ct = 8.08e−0.0617t Ct = 16.78e−0.1318t Ct = 16.44e−0.0847t

0.912 0.860 0.887 0.911 0.891 0.899 0.868 0.857 0.857 0.955 0.872 0.911

6.44 11.27 3.92 9.12 5.54 10.41 9.41 12.42 6.65 11.22 5.26 8.19

10 20 SMX

4 10 20

their removal ratios were 93.10 and 86.85%, respectively. There were no significant differences (p N 0.05). Therefore, the degradation may be related to the amount of manure, whereas different manure additions would only affect the degradation rates, not the removal ratios. Throughout the experiment, the degradation rates on days 0, 1, 3 and 7 in group A were 0.358, 0.322, 0.259 and 0.169 mg·kg− 1·d− 1, respectively. In group B they were 0.196, 0.185, 0.163 and 0.128 mg·kg− 1·d− 1, respectively. For both SDZ and SMX, degradation rates during the same incubation period were significantly different (p b 0.05), while there were similar removal ratios (p N 0.05) at varying initial concentrations under the same conditions.

As shown in Table 4, the coefficients of determination (R2) ranged from 0.857 to 0.955 (p N 0.05). SDZ dissipated faster and had lower DT50 values for treatment A at varying initial concentrations, while SMX had higher DT50 values for treatment B. The manure application had little effects on the degradations of these two antibiotics (b 20%). The results showed that the removal of SDZ and SMX in soils was slightly promoted by the presence of manure, which was also proved by Carlson and Mabury (2006). Manure addition accelerated the degradation, which can be explained that the addition of organic supplements could facilitate the contaminant removal, while an excessive supplementation would hinder the degradation. Inhibiting microbial degradations of the

Fig. 3. Proposed degradation pathways of SDZ in soil.

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aimed contaminants was most probably responsible for extrinsic carbon sources which would be preferentially degraded towards target compounds (Chang et al., 2002). As the contents of SDZ or SMX in the sterile soil (Fig. 1) were still high on day 49, manure application treatments could reduce final residuals (removal ratios N 90%). It can be concluded that a biodegradation should be considered as a major pathway for antibiotics removal in soil. It has been exhibited above that sulfamethoxazole and trimethoprim were degraded faster in the non-sterile soil (Feng et al., 2010; Srinivasan and Sarmah, 2014), suggesting that the biodegradation may accelerate the removal of antibiotics from the soil. Since microbial degradation processes attached most importance to the antibiotics and their metabolites behaviors in soil, this issue, investigated by substantially previous studies, should be performed. Degradation can also be influenced by many other factors such as physiochemical properties of the chemical, soil composition and environmental factors. Antibiotics persistence in soils was mostly dependent on soil-type conditions. Lower organic carbon contents and soil pH values associated with lower microbial activities resulted in higher DT50 values (Kodešová et al., 2016). Previous reports showed that SAs had relatively low sorption coefficients (Koc b 100 L·kg− 1), as their sorption onto soil was very weak (Höltge and Kreuzig, 2010; Kodešová et al., 2015; Stoob et al., 2007; Wu et al., 2009). Based on the data, it can be concluded that the lower octanol-water partitioning coefficient (Kow) value of SAs and lower sorption of the compounds may lead to their higher mobility and available amount of removal. It would make sense that SAs can be rapidly degraded in soils.

3.3. Degradation pathways and mechanisms of these two antibiotics Experiments were conducted by UHPLC-Q-TOF-MS techniques to identify SDZ and SMX degradation products in soils (the samples on days 3, 7 and 21 were selected for the determination). Figs. S1 and S2 in Supplementary material showed ESI (+) MS spectrums of purified and concentrated samples under a full scan mode, in which several transformation products (TPs) with a different m/z were determined. Generally, five reaction types were discussed as follows: S\\C bond cleavage, C\\N bond cleavage, S\\N bond hydrolysis, ring open and hydroxylation/oxidation, which proved that some types of reactions could be generalized across these two SAs in the present study. Aniline moiety in a SDZ molecule was validated to be a primary reactive site (Fig. 3). The reactive route might be due to the delocalization of electrophilic nature, resulting in carbons adjacent to the sulfur atom liable to a nucleophilic attack (Zhao et al., 1999). The groups were at the para and/or ortho locations on the aromatic or heteroaromatic rings. The group attached to the sulfonyl was an aromatic ring with an amino at the para location, so as to cause a cleavage of the S\\C bond. An alternative route of coming into the cleavage could be the production of a sulfamic acid derivative and aniline (product c). TP (c) could further react with the conversion of phenol to produce (d), and (e) was generated by the substitution reaction, and then (f) was formed by hydroxylation. This further proved that the degradation was not only dependent on biotic factors, but also associated with SAs tending to bind with soil and manure organic matter via the stable covalent binding mechanisms (Bialk et al., 2005; Kahle and Stamm, 2007). SDZ has

Fig. 4. Proposed degradation pathways of SMX in soil.

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been reported to undergo desulfonation reactions following a biodegradation (Schwarz et al., 2010). An alternative structure could be presented with the hydroxylation occurring in the amino group of the aniline (Gassman and Campbell, 1971). A hypothesis was that the parent compound firstly suffered an S\\N hydrolysis reaction, and then the products were combined to form desulfonation products (LekkerkerkerTeunissen et al., 2012). Based on the predominant oxidizing species and the given TPs' structure, the transformation pathways of SMX were illustrated in Fig. 4. Six TPs were identified in total, including hydroxylation, ring open, sulfonamide S\\N bond cleavage and aniline moiety oxidation (Supplementary material, Table S3). Results also demonstrated that the aniline group was a prioritized reactive moiety of SMX, but it was deactivated by the sulfonyl. Due to the electrical effect of \\NH2 and electrophilic effect of \\SO2\\NH\\, the monohydroxylation was confirmed to occur at an ortho location of \\NH2 (Dirany et al., 2012). The hydroxylation could activate the aromatic ring because of its increased electron density derived from the hydroxyl. Additionally, the hydrolysis of the S\\N bond would produce a free primary amine and sulfanilic acid. It has been reported that SMX followed different kinds of oxidation reactions (Bonvin et al., 2013; Gao et al., 2012). The sulfonamide S\\N bond was easily cleaved by the attack of oxidizing species, resulting in the formation of (c). Products (d) and (e) were rapidly produced after the reaction was initialized (Fig. 4). The processes of oxazole ring open and gradual oxidation were observed during the degradation of SMX (product a). The reasons could be attributed to the unstable N\\O bond. It was essential to indicate that the toxicity remained increased despite the target pollutants being removed completely. Results clearly suggested that the measured toxicity came from the oxidation products during the degradation period. Phenolic compounds were produced during the removal of SDZ and SMX (Supplementary material, Figs. S1 and S2). The increases of toxicity based on toxic TPs formed could generate more serious hazards than the parent compounds, which was identical to the conclusions from Olmez-Hanci et al. (2013) and Rosal et al. (2009). Previous reports revealed that the derivative TPs of SMX maintained a bacteriostatic mechanism towards bacteria, and characterized a toxicophore-like moiety (Majewsky et al., 2014). However, the transformation products like sulfanilic acid, barely displayed a toxicophore, thus losing their antibacterial activity. 4. Conclusions This study explored the degradation kinetics and mechanisms of SDZ and SMX in an agricultural soil system with manure application. The removal of these two antibiotics was well fit with first-order kinetics. Compared with the non-sterile soil, the degradation rates were slower and the associated DT50 values were higher in the sterile soil, suggesting that the biological processes may be a major contributing factor. Additionally, there were higher degradation rates in the treatments containing manure due to a higher microbial biomass carbon and nitrogen in soils. During the process, Single-application would accelerate microbial degradation and adsorption capacity in comparison with Repeated-application. However, different manure additions could only affect the degradation rate, not the removal ratio. The degradation mechanisms of SDZ and SMX in soils were proposed: (1) hydroxylation of the benzene ring; (2) oxidation of the amine group at the benzene ring; (3) cleavage of the sulfonamide bond; and (4) oxazole ring open. It can be concluded a toxicity increase in the treated samples because the toxic TPs formed could generate more serious toxicity than the parent compounds. Acknowledgements This research was supported by projects of the National Water Pollution Control and Treatment Science and Technology Major Project

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(2017ZX07207002), the Shanghai Science and Technology Committee (16DZ1204700), the National Natural Science Foundation of China (41371467), the Shanghai Pujiang Program (15PJD013), and the National Key Research and Development Program (2016YFD0800405). Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2017.07.083.

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