# Fate of sulfamethoxazole, its main metabolite N-ac-sulfamethoxazole and ciprofloxacin in agricultural soils amended or not by organic waste products

Aurore Andriamalala, Laure Vieublé, Valerie Dumeny, Philippe Cambier

► To cite this version:

Aurore Andriamalala, Laure Vieublé, Valerie Dumeny, Philippe Cambier. Fate of sulfamethoxazole, its main metabolite N-ac-sulfamethoxazole and ciprofloxacin in agricultural soils amended or not by organic waste products. Chemosphere, 2018, 191, pp.607-615. 10.1016/j.chemosphere.2017.10.093. hal-01629852

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1

Fate of Sulfamethoxazole, its main metabolite N-ac-Sulfamethoxazole and Ciprofloxacin
in agricultural soils amended or not by organic waste products

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Aurore Andriamalala<sup>a</sup>, Laure Vieublé-Gonod<sup>b</sup>, Valérie Dumeny<sup>a</sup>. Philippe Cambier<sup>a</sup>

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<sup>a</sup> UMR Ecosys, INRA, AgroParisTech, Université Paris-Saclay, 78850, Thiverval Grignon, France

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<sup>b</sup> UMR Ecosys, AgroParisTech, INRA, Université Paris-Saclay, 78850, Thiverval Grignon, France

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E-mail: laure.vieuble@agroparistech.fr

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Corresponding author:

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Philippe Cambier (Tel. 330130815988; E-mail:philippe.cambier@inra.fr)

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**Abstract:**

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Spreading organic waste products (OWP) issued from sewage sludge or manures into soil may disseminate antibiotics with unknown risks for human health and environment. Our objectives were to compare the fate of two sulfonamides, sulfamethoxazole (SMX) and its metabolite N-acetyl-sulfamethoxazole (N-ac-SMX), and one fluoroquinolone, ciprofloxacin (CIP), in an unamended soil, and two soils regularly amended since 1998 with a sewage sludge and green waste compost and with farmyard manure respectively. Incubations of soil spiked with <sup>14</sup>C labelled SMX or N-ac-SMX (0.02 mg.kg<sup>-1</sup>) or CIP (0.15 mg.kg<sup>-1</sup>) allowed a quantification of radiolabeled molecules in the mineralized, easily, hardly and non-extractable fractions after 3 and 156 days. Nature of <sup>14</sup>C molecules was also analyzed by HPLC in extractable fractions after 3 and 156 days.

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SMX and N-ac-SMX dissipation was fast and due to i) mineralization (~10% of recovered <sup>14</sup>C after 156 days) or incomplete degradation (production of metabolites), ii) adsorption, even if
1

Fate of Sulfamethoxazole, its main metabolite N-ac-Sulfamethoxazole and Ciprofloxacin
in agricultural soils amended or not by organic waste products

2

3

Aurore Andriamalala<sup>a</sup>, Laure Vieublé-Gonod<sup>b</sup>, Valérie Dumeny<sup>a</sup>. Philippe Cambier<sup>a</sup>

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<sup>a</sup> UMR Ecosys, INRA, AgroParisTech, Université Paris-Saclay, 78850, Thiverval Grignon, France

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<sup>b</sup> UMR Ecosys, AgroParisTech, INRA, Université Paris-Saclay, 78850, Thiverval Grignon, France

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E-mail: laure.vieuble@agroparistech.fr

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Corresponding author:

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Philippe Cambier (Tel. 330130815988; E-mail:philippe.cambier@inra.fr)

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**Abstract:**

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Spreading organic waste products (OWP) issued from sewage sludge or manures into soil may disseminate antibiotics with unknown risks for human health and environment. Our objectives were to compare the fate of two sulfonamides, sulfamethoxazole (SMX) and its metabolite N-acetyl-sulfamethoxazole (N-ac-SMX), and one fluoroquinolone, ciprofloxacin (CIP), in an unamended soil, and two soils regularly amended since 1998 with a sewage sludge and green waste compost and with farmyard manure respectively. Incubations of soil spiked with <sup>14</sup>C labelled SMX or N-ac-SMX (0.02 mg.kg<sup>-1</sup>) or CIP (0.15 mg.kg<sup>-1</sup>) allowed a quantification of radiolabeled molecules in the mineralized, easily, hardly and non-extractable fractions after 3 and 156 days. Nature of <sup>14</sup>C molecules was also analyzed by HPLC in extractable fractions after 3 and 156 days.

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SMX and N-ac-SMX dissipation was fast and due to i) mineralization (~10% of recovered <sup>14</sup>C after 156 days) or incomplete degradation (production of metabolites), ii) adsorption, even if
25 both sulfonamides present low Kd (<3 L.kg⁻¹) and iii) formation of non-extractable residues
26 (NER), representing more than 50% of recovered radioactivity.

27 N-ac-SMX was more mineralized than SMX, and formed more progressively NER, after a
28 step of deacetylation. Adsorption of CIP was fast and formed mainly NER (> 72%) whereas
29 its mineralization was negligible. Repeated applications of OWP tend to enhance adsorption
30 of antibiotics and lower their degradation, through the quantity and quality of the built up soil
31 organic matter. If applications of sewage sludge compost favor adsorption and inhibit
32 mineralization, applications of manure boost the formation of non-extractable residues.

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34 **Keywords:** Soil; Antibiotic; Metabolite; Organic waste products; Biodegradation; Adsorption

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36 **Highlights:**
37 N-ac-SMX slightly more mobile and more mineralized than SMX during soil incubations

38 Differences in fates of N-ac-SMX and SMX partly due to a step of deacetylation

39 SMX more available in soil than CIP but both mainly non-extractable after 156 days

40 Regular organic amendments enhance sorption and lower mineralization of antibiotics

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25 both sulfonamides present low Kd (<3 L.kg⁻¹) and iii) formation of non-extractable residues
26 (NER), representing more than 50% of recovered radioactivity.

27 N-ac-SMX was more mineralized than SMX, and formed more progressively NER, after a
28 step of deacetylation. Adsorption of CIP was fast and formed mainly NER (> 72%) whereas
29 its mineralization was negligible. Repeated applications of OWP tend to enhance adsorption
30 of antibiotics and lower their degradation, through the quantity and quality of the built up soil
31 organic matter. If applications of sewage sludge compost favor adsorption and inhibit
32 mineralization, applications of manure boost the formation of non-extractable residues.

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34 **Keywords:** Soil; Antibiotic; Metabolite; Organic waste products; Biodegradation; Adsorption

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36 **Highlights:**
37 N-ac-SMX slightly more mobile and more mineralized than SMX during soil incubations
38 Differences in fates of N-ac-SMX and SMX partly due to a step of deacetylation
39 SMX more available in soil than CIP but both mainly non-extractable after 156 days
40 Regular organic amendments enhance sorption and lower mineralization of antibiotics

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## 1. Introduction

The increasing occurrence of antibiotics (ATB) in soils, surface water and ground water becomes of more and more concern because of environmental and human health threat (Boxall et al., 2012).

Once in the soil, ATB may be degraded by biotic or abiotic path, retained on soil constituents and/or transferred to surface and groundwater through runoff and infiltration (Jechalke et al., 2014). Volatilization and photodegradation of antibiotics are generally negligible processes in soils (Ozaki et al., 2011; Jechalke et al., 2014). Indeed, the fate of antibiotics depends on antibiotics and soil properties. Generally detected in the µg-mg per kg range (Golet et al., 2002; Boxall et al., 2004), ATB may persist for years in soils (Du and Liu, 2012) while maintaining an activity even at a low dose (Boxall et al., 2004).

One of the entry routes of veterinary and human antibiotics into agricultural soils is the application of organic waste products (OWP) (Kümmerer, 2001), which is a worldwide increasingly common practice of recycling valuable nutrients and organic matter (Chang et al., 1995). However OWP may contain antibiotics. When administered to livestock, 10 to 90% of ATB are excreted as parent compounds or metabolites (Sarmah et al., 2006). Consequently ATB and metabolites can be present directly in livestock effluents (Boxall et al., 2004). ATB in human's urine and feces are not completely eliminated in waste water plants and can persist in effluents and/or sludge (Nieto et al., 2010).

OWP addition may also impact the fate of antibiotics in soils, through changing soil physical, chemical and biological properties (Li et al., 2015). Contradictory results were obtained under laboratory conditions. Some works highlighted a stimulation of ATB biodegradation after OWP addition (Accinelli et al., 2007; Zhang et al., 2017) while others showed an increase of ATB adsorption and inhibition of mineralization (Li et al., 2015).
## 1. Introduction

The increasing occurrence of antibiotics (ATB) in soils, surface water and ground water becomes of more and more concern because of environmental and human health threat (Boxall et al., 2012).

Once in the soil, ATB may be degraded by biotic or abiotic path, retained on soil constituents and/or transferred to surface and groundwater through runoff and infiltration (Jechalke et al., 2014). Volatilization and photodegradation of antibiotics are generally negligible processes in soils (Ozaki et al., 2011; Jechalke et al., 2014). Indeed, the fate of antibiotics depends on antibiotics and soil properties. Generally detected in the µg-mg per kg range (Golet et al., 2002; Boxall et al., 2004), ATB may persist for years in soils (Du and Liu, 2012) while maintaining an activity even at a low dose (Boxall et al., 2004).

One of the entry routes of veterinary and human antibiotics into agricultural soils is the application of organic waste products (OWP) (Kümmerer, 2001), which is a worldwide increasingly common practice of recycling valuable nutrients and organic matter (Chang et al., 1995). However OWP may contain antibiotics. When administered to livestock, 10 to 90% of ATB are excreted as parent compounds or metabolites (Sarmah et al., 2006). Consequently ATB and metabolites can be present directly in livestock effluents (Boxall et al., 2004). ATB in human's urine and feces are not completely eliminated in waste water plants and can persist in effluents and/or sludge (Nieto et al., 2010).

OWP addition may also impact the fate of antibiotics in soils, through changing soil physical, chemical and biological properties (Li et al., 2015). Contradictory results were obtained under laboratory conditions. Some works highlighted a stimulation of ATB biodegradation after OWP addition (Accinelli et al., 2007; Zhang et al., 2017) while others showed an increase of ATB adsorption and inhibition of mineralization (Li et al., 2015).
However studies based on addition of OWP into soils in the lab miss the long-term effects of OWP applications on soil and their indirect effect on the fate of ATB. In this study, we compared the fate of ATB in three soils from the same field experiment. Two have been amended every two years since 1998, with a farmyard manure (FYM) or a co-compost of sewage sludge and green waste (SGW), and the third one has never been amended (CONT). We focused on two ATB widely used in human and veterinary medicine, sulfamethoxazole (SMX) and ciprofloxacin (CIP), belonging respectively to sulfonamides and fluoroquinolones families. These molecules are among the most frequently detected ATB in the environment (Barnes et al., 2004; Kolpin et al., 2002), and present contrasting behavior in soils. SMX is relatively mobile and rapidly degraded with half-lives ranging from a few hours to a few days (Salvia et al., 2014; Srinivasan and Sarmah, 2014), while CIP is known to have a high affinity to soils (Nowara et al., 1997; Vasudevan et al., 2009) and may persist for several months to years (Hamscher et al., 2005; Rosendahl et al., 2012). However some studies suggested that sulfonamides may be more persistent than would be predicted from laboratory controlled studies (Bialk et al., 2005). The different characteristics of CIP and SMX make them good candidates to represent a range of antibiotics in the environment (Chen et al., 2011).

We also considered the main metabolite of SMX, the N-acetyl-sulfamethoxazole (N-ac-SMX). When excreted, this metabolite accounts for approximately 50% of the administered SMX. But N-ac-SMX may undergo deacetylation in the environment and SMX can be regenerated (Göbel et al., 2005).

Thus the objectives of this study were to determine i) the fate of antibiotics with contrasted characteristics in soils and ii) how long-term applications of different OWP to soils impacted the fate of these antibiotics.
However studies based on addition of OWP into soils in the lab miss the long-term effects of OWP applications on soil and their indirect effect on the fate of ATB. In this study, we compared the fate of ATB in three soils from the same field experiment. Two have been amended every two years since 1998, with a farmyard manure (FYM) or a co-compost of sewage sludge and green waste (SGW), and the third one has never been amended (CONT). We focused on two ATB widely used in human and veterinary medicine, sulfamethoxazole (SMX) and ciprofloxacin (CIP), belonging respectively to sulfonamides and fluoroquinolones families. These molecules are among the most frequently detected ATB in the environment (Barnes et al., 2004; Kolpin et al., 2002), and present contrasting behavior in soils. SMX is relatively mobile and rapidly degraded with half-lives ranging from a few hours to a few days (Salvia et al., 2014; Srinivasan and Sarmah, 2014), while CIP is known to have a high affinity to soils (Nowara et al., 1997; Vasudevan et al., 2009) and may persist for several months to years (Hamscher et al., 2005; Rosendahl et al., 2012). However some studies suggested that sulfonamides may be more persistent than would be predicted from laboratory controlled studies (Bialk et al., 2005). The different characteristics of CIP and SMX make them good candidates to represent a range of antibiotics in the environment (Chen et al., 2011).

We also considered the main metabolite of SMX, the N-acetyl-sulfamethoxazole (N-ac-SMX). When excreted, this metabolite accounts for approximately 50% of the administered SMX. But N-ac-SMX may undergo deacetylation in the environment and SMX can be regenerated (Göbel et al., 2005).

Thus the objectives of this study were to determine i) the fate of antibiotics with contrasted characteristics in soils and ii) how long-term applications of different OWP to soils impacted the fate of these antibiotics.
We monitored <sup>14</sup>C-labelled antibiotics during 6 months incubations by quantifying radioactivity in the mineralized, extractable and non-extractable fractions, and by qualifying radioactivity (parent molecules or metabolites) in the extractable fractions.

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## 2. Materials and methods

### 2.1. Antibiotics

<sup>14</sup>C-ring-labeled sulfamethoxazole (chemical purity 98%, specific activity 77 mCi.mmol<sup>-1</sup>) and <sup>14</sup>C-ring-labeled N-acetyl-sulfamethoxazole (chemical purity 93%, specific activity 77 mCi.mmol<sup>-1</sup>) were purchased from Isobio (ARC). <sup>14</sup>C-ring-labeled ciprofloxacin (chemical purity 98.4%, specific activity 15 mCi.mmol<sup>-1</sup>), was purchased from Moravek Biochemical. The structures, pKa, and labelling of molecules are presented in Fig. S1.

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### 2.2. Soil sampling

Soils were sampled in 2014 from the ploughed layers (0-25 cm) of the experimental field QualiAgro (48°53' N, 1°58' E) located at Feucherolles (Yvelines, France): a soil control that has never received OWP and 2 soils amended every 2 years since 1998 with a co-compost of green waste and sewage sludge (SGW) or farmyard manure (FYM) (Houot et al., 2002). At sampling, amended soils (SGW or FYM) have received 9 applications of OWP. The soil is a silt loam Glossic Luvisol and contains on average 19% clay, 75% silt and 6% sand in the ploughed layer. Clay minerals identified are kaolinite, illite and smectite, with traces of chlorite (Cornu, personal communication). The field is cropped with a biannual rotation winter wheat-maize. Selected properties of the soil samples and of the OWP applied in 2013 are listed in Table 1. The soils were homogenized, sieved through a 4-mm mesh and stored 3 weeks at 4°C in the dark before using.

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We monitored <sup>14</sup>C-labelled antibiotics during 6 months incubations by quantifying radioactivity in the mineralized, extractable and non-extractable fractions, and by qualifying radioactivity (parent molecules or metabolites) in the extractable fractions.

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## 2. Materials and methods

### 2.1. Antibiotics

<sup>14</sup>C-ring-labeled sulfamethoxazole (chemical purity 98%, specific activity 77 mCi.mmol<sup>-1</sup>) and <sup>14</sup>C-ring-labeled N-acetyl-sulfamethoxazole (chemical purity 93%, specific activity 77 mCi.mmol<sup>-1</sup>) were purchased from Isobio (ARC). <sup>14</sup>C-ring-labeled ciprofloxacin (chemical purity 98.4%, specific activity 15 mCi.mmol<sup>-1</sup>), was purchased from Moravek Biochemical. The structures, pKa, and labelling of molecules are presented in Fig. S1.

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### 2.2. Soil sampling

Soils were sampled in 2014 from the ploughed layers (0-25 cm) of the experimental field QualiAgro (48°53' N, 1°58' E) located at Feucherolles (Yvelines, France): a soil control that has never received OWP and 2 soils amended every 2 years since 1998 with a co-compost of green waste and sewage sludge (SGW) or farmyard manure (FYM) (Houot et al., 2002). At sampling, amended soils (SGW or FYM) have received 9 applications of OWP. The soil is a silt loam Glossic Luvisol and contains on average 19% clay, 75% silt and 6% sand in the ploughed layer. Clay minerals identified are kaolinite, illite and smectite, with traces of chlorite (Cornu, personal communication). The field is cropped with a biannual rotation winter wheat-maize. Selected properties of the soil samples and of the OWP applied in 2013 are listed in Table 1. The soils were homogenized, sieved through a 4-mm mesh and stored 3 weeks at 4°C in the dark before using.

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**Table 1**

Physicochemical properties of soils (CONT, SGW, FYM) and OWP used as amendment in 2013

<table><thead><tr><th rowspan="2"></th><th colspan="3">Soils</th><th colspan="2">OWP</th></tr><tr><th>CONT</th><th>SGW</th><th>FYM</th><th>Compost sewage sludge-green waste</th><th>Farmyard manure</th></tr></thead><tbody><tr><td>Organic carbon (g.kg<sup>-1</sup> dw)</td><td>10.2</td><td>17.6</td><td>15.5</td><td>257.3</td><td>354.7</td></tr><tr><td>Total nitrogen (g.kg<sup>-1</sup> dw)</td><td>1.0</td><td>1.7</td><td>1.4</td><td>25.5</td><td>20.1</td></tr><tr><td>pH</td><td>6.7</td><td>6.8</td><td>7.3</td><td>7.7</td><td>8.8</td></tr><tr><td>CEC (cmol+·kg<sup>-1</sup>)</td><td>7.6</td><td>9.8</td><td>9.9</td><td>-</td><td>-</td></tr><tr><td>C/N</td><td>10.2</td><td>10.4</td><td>11.1</td><td>9.9</td><td>17.6</td></tr><tr><td>I<sub>ROC</sub><sup>a</sup></td><td>-</td><td>-</td><td>-</td><td>78</td><td>67</td></tr><tr><td>[SMX] (µg.kg<sup>-1</sup> soil dw)<sup>b</sup></td><td>nd</td><td>nd</td><td>nd</td><td>nd</td><td>nd</td></tr><tr><td>[CIP] (µg.kg<sup>-1</sup> soil dw)<sup>b</sup></td><td>nd</td><td>7.3 ± 1.9</td><td>nd</td><td>309 ± 22</td><td>nd</td></tr></tbody></table>

a Proportion of total organic carbon potentially incorporated into soil organic matter (Lashermes et al., 2009).

b Concentrations of ATB were analysed in OWP of 2013 and soil sampled in November 2014 (Bourdat-Deschamps et al., 2017).

nd: not detected

## 2.3. Sorption Coefficient of SMX and N-ac-SMX

Sorption coefficients (Kd) of SMX and N-ac-SMX were determined on gamma-ray sterilized (40–70 kGy) samples of the 3 studied soils as follows: three replicates of 3 g equivalent dry soil were mixed with 10 ml of CaCl₂ 0.01 M solution sterilized by filtration at 0.2 µm and 7.96 µl of ¹⁴C labelled SMX or 10.14 µl of ¹⁴C labelled N-ac-SMX (0.02 µg.g⁻¹ soil), before shaking at 15 rotations.min⁻¹ for 24 h in the dark at 20°C. Samples were then centrifuged 10 min at 34155 g at 20°C and the ¹⁴C remaining in the supernatants was quantified by liquid scintillation counting.

Then Kd (L.kg⁻¹) was calculated as follows:

$$Kd = Cads/Ce$$
**Table 1**

Physicochemical properties of soils (CONT, SGW, FYM) and OWP used as amendment in 2013

<table><thead><tr><th rowspan="2"></th><th colspan="3">Soils</th><th colspan="2">OWP</th></tr><tr><th>CONT</th><th>SGW</th><th>FYM</th><th>Compost sewage sludge-green waste</th><th>Farmyard manure</th></tr></thead><tbody><tr><td>Organic carbon (g.kg<sup>-1</sup> dw)</td><td>10.2</td><td>17.6</td><td>15.5</td><td>257.3</td><td>354.7</td></tr><tr><td>Total nitrogen (g.kg<sup>-1</sup> dw)</td><td>1.0</td><td>1.7</td><td>1.4</td><td>25.5</td><td>20.1</td></tr><tr><td>pH</td><td>6.7</td><td>6.8</td><td>7.3</td><td>7.7</td><td>8.8</td></tr><tr><td>CEC (cmol+·kg<sup>-1</sup>)</td><td>7.6</td><td>9.8</td><td>9.9</td><td>-</td><td>-</td></tr><tr><td>C/N</td><td>10.2</td><td>10.4</td><td>11.1</td><td>9.9</td><td>17.6</td></tr><tr><td>I<sub>ROC</sub><sup>a</sup></td><td>-</td><td>-</td><td>-</td><td>78</td><td>67</td></tr><tr><td>[SMX] (µg.kg<sup>-1</sup> soil dw)<sup>b</sup></td><td>nd</td><td>nd</td><td>nd</td><td>nd</td><td>nd</td></tr><tr><td>[CIP] (µg.kg<sup>-1</sup> soil dw)<sup>b</sup></td><td>nd</td><td>7.3 ± 1.9</td><td>nd</td><td>309 ± 22</td><td>nd</td></tr></tbody></table>

a Proportion of total organic carbon potentially incorporated into soil organic matter (Lashermes et al., 2009).

b Concentrations of ATB were analysed in OWP of 2013 and soil sampled in November 2014 (Bourdat-Deschamps et al., 2017).

nd: not detected

## 2.3. Sorption Coefficient of SMX and N-ac-SMX

Sorption coefficients (Kd) of SMX and N-ac-SMX were determined on gamma-ray sterilized (40–70 kGy) samples of the 3 studied soils as follows: three replicates of 3 g equivalent dry soil were mixed with 10 ml of CaCl₂ 0.01 M solution sterilized by filtration at 0.2 µm and 7.96 µl of ¹⁴C labelled SMX or 10.14 µl of ¹⁴C labelled N-ac-SMX (0.02 µg.g⁻¹ soil), before shaking at 15 rotations.min⁻¹ for 24 h in the dark at 20°C. Samples were then centrifuged 10 min at 34155 g at 20°C and the ¹⁴C remaining in the supernatants was quantified by liquid scintillation counting.

Then Kd (L.kg⁻¹) was calculated as follows:

$$K_d = C_{ads}/C_e$$
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Cads was the concentration of ATB adsorbed on soil (mg.kg⁻¹) and Ce, the concentration of ATB in soil solution after equilibrium (mg.L⁻¹).

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## 2.4. Incubations

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Ten grams (equivalent dry soil) of CONT, SGW and FYM soils were amended with 1 ml of solutions of ¹⁴C-SMX, ¹⁴C-N-ac-SMX or ¹⁴C-CIP to provide final concentrations of 0.02, 0.02 or 0.15 µg.g⁻¹ soil respectively (1543 MBq.mM⁻¹ per sample) and a moisture content corresponding to 80% of maximum water holding capacity (0.22 g of water per g of soil). Six replicates per treatment were prepared. Spiked soils were incubated in 1 L airtight jars in presence of 2 vials containing respectively 10 ml of deionized water and 6 ml of sodium hydroxide 1N, at 28 °C for 156 days.

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At 3 days and then weekly, ¹⁴C-activity in the NaOH solutions was measured with a TriCarb liquid scintillation counter after mixing with 10 mL of UltimaGold liquid scintillation cocktail. Samples were kept for at least 24 h in the dark before counting to avoid potential chemiluminescence. The counting time was 10 min.

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The opening of the jars to replace the NaOH traps ensured aerobic conditions during all incubation time. The soil moisture was checked over time and readjusted when necessary to 80% of WHC by adding deionized water.

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## 2.5. Quantification of ¹⁴C residues

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At days 3 and 156, three successive extractions were performed to quantify the distribution of radioactivity from ¹⁴C-SMX, ¹⁴C-N-ac-SMX and ¹⁴C-CIP in the easily and hardly extractable and the non-extractable fractions. The extraction protocol was adapted from Ferhi et al. (2016).

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Cads was the concentration of ATB adsorbed on soil (mg.kg⁻¹) and Ce, the concentration of ATB in soil solution after equilibrium (mg.L⁻¹).

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## 2.4. Incubations

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Ten grams (equivalent dry soil) of CONT, SGW and FYM soils were amended with 1 ml of solutions of ¹⁴C-SMX, ¹⁴C-N-ac-SMX or ¹⁴C-CIP to provide final concentrations of 0.02, 0.02 or 0.15 µg.g⁻¹ soil respectively (1543 MBq.mM⁻¹ per sample) and a moisture content corresponding to 80% of maximum water holding capacity (0.22 g of water per g of soil). Six replicates per treatment were prepared. Spiked soils were incubated in 1 L airtight jars in presence of 2 vials containing respectively 10 ml of deionized water and 6 ml of sodium hydroxide 1N, at 28 °C for 156 days.

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At 3 days and then weekly, ¹⁴C-activity in the NaOH solutions was measured with a TriCarb liquid scintillation counter after mixing with 10 mL of UltimaGold liquid scintillation cocktail. Samples were kept for at least 24 h in the dark before counting to avoid potential chemiluminescence. The counting time was 10 min.

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The opening of the jars to replace the NaOH traps ensured aerobic conditions during all incubation time. The soil moisture was checked over time and readjusted when necessary to 80% of WHC by adding deionized water.

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## 2.5. Quantification of ¹⁴C residues

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At days 3 and 156, three successive extractions were performed to quantify the distribution of radioactivity from ¹⁴C-SMX, ¹⁴C-N-ac-SMX and ¹⁴C-CIP in the easily and hardly extractable and the non-extractable fractions. The extraction protocol was adapted from Ferhi et al. (2016).

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To measure the easily extractable fractions, 30 ml of cyclodextrin HPCD (0.05 mol.L⁻¹) were added to soil samples in 250 ml PPCO tubes. After shaking during 16 hours at 15 rotations.min⁻¹, samples were centrifuged (20 min, 20°C, 34155 g). Supernatants were weighted and mixed with 10 ml of Ultima Gold XR scintillation liquid (Perkin Elmer) for counting. The next extraction was performed by adding acetonitrile MeCN (99.9% Acetonitrile Carlo Erba Reagents Dasit Group, 27.5 ml) combined with a Mc Ilvaine pH9 buffer solution (9.3 mL of 1/29 (v/v) citric acid/phosphate). pH9 was reached with addition of NaOH 10 N. Samples were ultra-sonicated twice 10 min before centrifugation (20 min, 20°C, 34155 g). Radioactivity was quantified in the supernatants by liquid scintillation. The last extraction was performed by addition of MeCN (99.9% Acetonitrile Carlo Erba Reagents Dasit Group, 27.5 ml) combined with a Mc Ilvaine pH2 buffer solution (11 mL of 27/3 (v/v) citric acid/phosphate). pH2 was reached by adding HCl 37%. Samples were ultra-sonicated twice for 10 min, before centrifugation (20 min, 20°C, 34155 g). Because of a salting out effect due to the high quantity of EDTA salt, two phases were obtained. Radioactivity was quantified in the lower and upper phases by liquid scintillation. The hardly extractable fraction was defined as the sum of pH9 and pH2 extracts.

After extractions, soil samples containing non-extractable ¹⁴C-residues (NER) were air dried at 35°C. Radioactivity from NER was measured by combustion (Biological Oxidizer OX 700, Zinsser Analytic, Frankfurt, Germany) and trapping of the ¹⁴CO₂ in Carbosorb and liquid scintillation counting (Packard Instruments, Meriden, CT, USA).

## 2.6. Qualitative analysis of ¹⁴C residues in extractable fractions

A qualitative analysis of ¹⁴C in HPCD and MeCN extracts was carried out by high-performance liquid chromatography (HPLC) with a Waters (Milford, MA, USA) chromatography appliance (Systemcontroller 610, Autosampler 717) coupled with a
To measure the easily extractable fractions, 30 ml of cyclodextrin HPCD (0.05 mol.L⁻¹) were added to soil samples in 250 ml PPCO tubes. After shaking during 16 hours at 15 rotations.min⁻¹, samples were centrifuged (20 min, 20°C, 34155 g). Supernatants were weighted and mixed with 10 ml of Ultima Gold XR scintillation liquid (Perkin Elmer) for counting. The next extraction was performed by adding acetonitrile MeCN (99.9% Acetonitrile Carlo Erba Reagents Dasit Group, 27.5 ml) combined with a Mc Ilvaine pH9 buffer solution (9.3 mL of 1/29 (v/v) citric acid/phosphate). pH9 was reached with addition of NaOH 10 N. Samples were ultra-sonicated twice 10 min before centrifugation (20 min, 20°C, 34155 g). Radioactivity was quantified in the supernatants by liquid scintillation. The last extraction was performed by addition of MeCN (99.9% Acetonitrile Carlo Erba Reagents Dasit Group, 27.5 ml) combined with a Mc Ilvaine pH2 buffer solution (11 mL of 27/3 (v/v) citric acid/phosphate). pH2 was reached by adding HCl 37%. Samples were ultra-sonicated twice for 10 min, before centrifugation (20 min, 20°C, 34155 g). Because of a salting out effect due to the high quantity of EDTA salt, two phases were obtained. Radioactivity was quantified in the lower and upper phases by liquid scintillation. The hardly extractable fraction was defined as the sum of pH9 and pH2 extracts.

After extractions, soil samples containing non-extractable ¹⁴C-residues (NER) were air dried at 35°C. Radioactivity from NER was measured by combustion (Biological Oxidizer OX 700, Zinsser Analytic, Frankfurt, Germany) and trapping of the ¹⁴CO₂ in Carbosorb and liquid scintillation counting (Packard Instruments, Meriden, CT, USA).

## 2.6. Qualitative analysis of ¹⁴C residues in extractable fractions

A qualitative analysis of ¹⁴C in HPCD and MeCN extracts was carried out by high-performance liquid chromatography (HPLC) with a Waters (Milford, MA, USA) chromatography appliance (Systemcontroller 610, Autosampler 717) coupled with a
radioactive flow detector (Flo-one A-500, Packard-Radiomatic) when radioactivity in samples was greater than 250 Bq.mL⁻¹. In certain cases, the 3 repetitions per extracts were pooled to obtain sufficient radioactivity for analysis (Table S1). HPCD extracts were concentrated on SPE cartridges (Strata X-A 33 µm Polymeric Stron Anion, Phenomenex) and then with nitrogen before HPLC analyses. Radioactivity in HPCD extracts was too low after 156 days of incubation to be analyzed. MeCN-McIlvaine extracts were evaporated before addition of 2 mL methanol. After centrifugation, samples were analyzed by HPLC. Radioactivity in the lower phase after MeCN-McIlvaine extraction was insufficient to be analyzed. Samples analyzed by HPLC (Perkin Elmer: Flexar System Quaternary Pump, Automatic Sample, PDA detector, Chromera® CDS piloting and acquisition Software; Column: Waters X-Bridge C18 5µ / 4.6x250 mm; detector: Radio-HPLC FLO ONE/Beta/Gamma C150TR; elution method: ACN + 0.01% TFA and H₂O + 0.01% TFA) are synthetized in Table S1.

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2.7. Statistical analyses

Statistical analyses, using R software, were performed via ANOVA and Tukey tests when values were parametric, Kruskal and Dunn tests when values were non-parametric, to compare the fate of ¹⁴C SMX, N-ac-SMX and CIP in the 3 studied soils at d3 and d156. P-values less than 0.05 were considered as statistically significant.

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radioactive flow detector (Flo-one A-500, Packard-Radiomatic) when radioactivity in samples was greater than 250 Bq.mL⁻¹. In certain cases, the 3 repetitions per extracts were pooled to obtain sufficient radioactivity for analysis (Table S1). HPCD extracts were concentrated on SPE cartridges (Strata X-A 33 µm Polymeric Stron Anion, Phenomenex) and then with nitrogen before HPLC analyses. Radioactivity in HPCD extracts was too low after 156 days of incubation to be analyzed. MeCN-McIlvaine extracts were evaporated before addition of 2 mL methanol. After centrifugation, samples were analyzed by HPLC. Radioactivity in the lower phase after MeCN-McIlvaine extraction was insufficient to be analyzed. Samples analyzed by HPLC (Perkin Elmer: Flexar System Quaternary Pump, Automatic Sample, PDA detector, Chromera® CDS piloting and acquisition Software; Column: Waters X-Bridge C18 5µ / 4.6x250 mm; detector: Radio-HPLC FLO ONE/Beta/Gamma C150TR; elution method: ACN + 0.01% TFA and H₂O + 0.01% TFA) are synthetized in Table S1.

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## 2.7. Statistical analyses

Statistical analyses, using R software, were performed via ANOVA and Tukey tests when values were parametric, Kruskal and Dunn tests when values were non-parametric, to compare the fate of ¹⁴C SMX, N-ac-SMX and CIP in the 3 studied soils at d3 and d156. P-values less than 0.05 were considered as statistically significant.

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## 3. Results

### 3.1. Fate of SMX and N-ac-SMX into soils amended or not with OWP

#### 3.1.1. Kd of SMX and N-ac-SMX

Kd of SMX was lower in CONT than in SGW and intermediate in FYM (Table 2). Kd of N-ac-SMX was significantly different in the three soils, in the following order: SGW>CONT>FYM (Table 2).

<caption>**Table 2**</caption>

Sorption coefficients (Kd) of SMX and N-ac-SMX (L.kg⁻¹) added at a concentration of 0.02 µg.g⁻¹ soil in sterilized CONT, SGW and FYM soils (mean ± standard-deviation from 3 replicates). Different letters indicate statistical differences (P<0.05) between the three soils.

<table><thead><tr><th>Kd (L.kg<sup>-1</sup>)</th><th>SMX</th><th>N-ac-SMX</th><th>SMX/N-ac-SMX</th></tr></thead><tbody><tr><td>CONT</td><td>2.58 ± 0.11</td><td>b</td><td>1.21 ± 0.03</td><td>b</td></tr><tr><td>SGW</td><td>3.20 ± 0.10</td><td>a</td><td>1.33 ± 0.00</td><td>a</td></tr><tr><td>FYM</td><td>2.95 ± 0.07</td><td>ab</td><td>1.02 ± 0.02</td><td>c</td></tr></tbody></table>

Kd of SMX appeared highly correlated to the C content of soils (R²=0.94) in contrast to N-ac-SMX (R²=0.008).

In the three soils, Kd of SMX was two to three folds higher than Kd of N-ac-SMX. In all cases, at a concentration of 0.02 µg.g⁻¹, Kd values were low, whatever the molecules and whatever the soils.

#### 3.1.2. Fate of SMX

SMX mineralization started without latency in the 3 soils and reached 6.6, 8.9 and 10.5% of added ¹⁴C-SMX in the SGW, FYM and CONT soils respectively after 156 days (Fig. 1).

Although low in all cases, SMX mineralization was significantly greater in CONT than in FYM and SGW.
## 3. Results

### 3.1. Fate of SMX and N-ac-SMX into soils amended or not with OWP

#### 3.1.1. Kd of SMX and N-ac-SMX

Kd of SMX was lower in CONT than in SGW and intermediate in FYM (Table 2). Kd of N-ac-SMX was significantly different in the three soils, in the following order: SGW>CONT>FYM (Table 2).

<caption>**Table 2**</caption>

Sorption coefficients (Kd) of SMX and N-ac-SMX (L.kg⁻¹) added at a concentration of 0.02 µg.g⁻¹ soil in sterilized CONT, SGW and FYM soils (mean ± standard-deviation from 3 replicates). Different letters indicate statistical differences (P<0.05) between the three soils.

<table><thead><tr><th>Kd (L.kg<sup>-1</sup>)</th><th>SMX</th><th>N-ac-SMX</th><th>SMX/N-ac-SMX</th></tr></thead><tbody><tr><td>CONT</td><td>2.58 ± 0.11</td><td>b</td><td>1.21 ± 0.03</td><td>b</td></tr><tr><td>SGW</td><td>3.20 ± 0.10</td><td>a</td><td>1.33 ± 0.00</td><td>a</td></tr><tr><td>FYM</td><td>2.95 ± 0.07</td><td>ab</td><td>1.02 ± 0.02</td><td>c</td></tr></tbody></table>

Kd of SMX appeared highly correlated to the C content of soils (R²=0.94) in contrast to N-ac-SMX (R²=0.008).

In the three soils, Kd of SMX was two to three folds higher than Kd of N-ac-SMX. In all cases, at a concentration of 0.02 µg.g⁻¹, Kd values were low, whatever the molecules and whatever the soils.

#### 3.1.2. Fate of SMX

SMX mineralization started without latency in the 3 soils and reached 6.6, 8.9 and 10.5% of added ¹⁴C-SMX in the SGW, FYM and CONT soils respectively after 156 days (Fig. 1).

Although low in all cases, SMX mineralization was significantly greater in CONT than in FYM and SGW.
<figure><img src="image_1.png" /><figcaption>**Fig. 1.** Cumulative $^{14}$C-SMX mineralization in the control soil and in the soils that have received 9 applications of a co-compost of sewage sludge and green waste (SGW) or farmyard manure (FYM). Error bars indicate standard deviations of 3 replicates.</figcaption></figure>

The total recovery of $^{14}$C in the mineralized, extractable (HPCD and MeCN) and non-extractable $^{14}$C fractions reached 87.4-88.8% of added $^{14}$C-SMX after 3 days and 81.7-91% after 156 days. For interpretation, the results were expressed as percentages of recovered $^{14}$C.

The sum of all fractions corresponds therefore to 100% (Fig. 2).

<figure><img src="image_2.png" /></figure>
<figure><img src="image_2.png" /><figcaption>**Fig. 1.** Cumulative $^{14}$C-SMX mineralization in the control soil and in the soils that have received 9 applications of a co-compost of sewage sludge and green waste (SGW) or farmyard manure (FYM). Error bars indicate standard deviations of 3 replicates.</figcaption></figure>

The total recovery of $^{14}$C in the mineralized, extractable (HPCD and MeCN) and non-extractable $^{14}$C fractions reached 87.4-88.8% of added $^{14}$C-SMX after 3 days and 81.7-91% after 156 days. For interpretation, the results were expressed as percentages of recovered $^{14}$C.

The sum of all fractions corresponds therefore to 100% (Fig. 2).

<figure><img src="image_3.png" /></figure>
239 **Fig. 2.** Distribution of ¹⁴C-SMX in the different soil fractions (mineralized, HPCD and MeCN extractable fractions, non-extractable residues), at (A) day 3, and (B) day 156, in the control soil (CONT) and the soils amended with a sewage sludge-green waste compost (SGW) or farmyard manure (FYM).

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244 At d3 (Fig. 2A), ¹⁴C was mainly distributed in the extractable and non-extractable fractions.
245 SMX mineralization was negligible and varied between 1.1% (FYM) and 1.7% (CONT and
246 SGW). Radioactivity in easily extractable (HPCD) fraction ranged between 17.9% (FYM) and
247 20.9% (SGW) of recovered ¹⁴C. In hardly extractable (MeCN) fraction, ¹⁴C ranged between
248 21.3% (FYM) and 27.3% (SGW). HPCD and MeCN extractable fractions were significantly
249 lower in FYM. NER represented more than half of the recovered radioactivity as early as 3
250 days (Fig. 2A). They were significantly higher in FYM than in CONT and SGW.

251 Between 3 and 156 days, we observed mainly a decrease of the HPCD fraction (from about
252 20% to 6.1-7.7%) and an increase of the mineralized fraction. During the same period,
253 mineralization was multiplied by a factor of about 6.5 and reached 7.4% (SGW), 11.0%
254 (FYM) and 11.5% (CONT) of recovered ¹⁴C (Fig. 2B).

255 The MeCN fraction slightly but significantly increased to reach 24.1-28.7%, whereas NER
256 did not significantly vary between d3 and d156.

257 After 156 days, CONT presented the highest SMX mineralization and SGW the lowest
258 mineralization and the highest HPCD extractable fraction. At last, FYM presented the lowest
259 MeCN ¹⁴C fraction.

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239 **Fig. 2.** Distribution of ¹⁴C-SMX in the different soil fractions (mineralized, HPCD and MeCN extractable fractions, non-extractable residues), at (A) day 3, and (B) day 156, in the control soil (CONT) and the soils amended with a sewage sludge-green waste compost (SGW) or farmyard manure (FYM).
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244 At d3 (Fig. 2A), ¹⁴C was mainly distributed in the extractable and non-extractable fractions.
245 SMX mineralization was negligible and varied between 1.1% (FYM) and 1.7% (CONT and
246 SGW). Radioactivity in easily extractable (HPCD) fraction ranged between 17.9% (FYM) and
247 20.9% (SGW) of recovered ¹⁴C. In hardly extractable (MeCN) fraction, ¹⁴C ranged between
248 21.3% (FYM) and 27.3% (SGW). HPCD and MeCN extractable fractions were significantly
249 lower in FYM. NER represented more than half of the recovered radioactivity as early as 3
250 days (Fig. 2A). They were significantly higher in FYM than in CONT and SGW.
251

251 Between 3 and 156 days, we observed mainly a decrease of the HPCD fraction (from about
252 20% to 6.1-7.7%) and an increase of the mineralized fraction. During the same period,
253 mineralization was multiplied by a factor of about 6.5 and reached 7.4% (SGW), 11.0%
254 (FYM) and 11.5% (CONT) of recovered ¹⁴C (Fig. 2B).
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255 The MeCN fraction slightly but significantly increased to reach 24.1-28.7%, whereas NER
256 did not significantly vary between d3 and d156.
257

257 After 156 days, CONT presented the highest SMX mineralization and SGW the lowest
258 mineralization and the highest HPCD extractable fraction. At last, FYM presented the lowest
259 MeCN ¹⁴C fraction.
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260
261 3.1.3. Fate of N-ac-SMX

262 As for SMX, N-ac-SMX was immediately mineralized and cumulative ¹⁴C-CO₂ ranged
263 between 8.5 and 11.6% at d156 (Fig. 3). As for SMX, N-ac-SMX mineralization was higher
264 in CONT than in amended soils and higher in FYM compared to SGW.

<figure><img src="image_3.png" /></figure>

265

266 **Fig. 3.** Cumulative ¹⁴C N-ac-SMX mineralization in the control soil (CONT) and the soils
267 amended with a sewage sludge-green waste compost (SGW) or farmyard manure (FYM).
268 Error bars indicate standard deviations of 3 replicates.

269

270 The total recovery of ¹⁴C in the different fractions ranged from 71.7 to 82.2% of added ¹⁴C N-
271 ac-SMX after 3 days and from 82.1 to 86.8% after 156 days. For interpretation, the results
272 were expressed as percentages of recovered ¹⁴C (Fig. 4).

273

<figure><img src="image_4.png" /></figure>
261 3.1.3. Fate of N-ac-SMX

262 As for SMX, N-ac-SMX was immediately mineralized and cumulative ¹⁴C-CO₂ ranged
263 between 8.5 and 11.6% at d156 (Fig. 3). As for SMX, N-ac-SMX mineralization was higher
264 in CONT than in amended soils and higher in FYM compared to SGW.

<figure><img src="image_4.png" /></figure>

265

266 **Fig. 3.** Cumulative ¹⁴C N-ac-SMX mineralization in the control soil (CONT) and the soils
267 amended with a sewage sludge-green waste compost (SGW) or farmyard manure (FYM).
268 Error bars indicate standard deviations of 3 replicates.

269

270 The total recovery of ¹⁴C in the different fractions ranged from 71.7 to 82.2% of added ¹⁴C N-
271 ac-SMX after 3 days and from 82.1 to 86.8% after 156 days. For interpretation, the results
272 were expressed as percentages of recovered ¹⁴C (Fig. 4).

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<figure><img src="image_5.png" /></figure>
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**Fig. 4.** Distribution of <sup>14</sup>C-N-ac-SMX in the different soil fractions (mineralized, HPCD and MeCN extractable fractions, non-extractable residues), at (A) day 3, and (B) day 156, in the control soil (CONT) and the soils amended with a sewage sludge-green waste compost (SGW) or farmyard manure (FYM).

After 3 days, N-ac-SMX mineralization was low (2.1-3.0%) and followed the order CONT≥FYM>SGW. The HPCD fractions represented 22.9-25.6% of recovered <sup>14</sup>C and were significantly higher in SGW than in CONT, FYM presenting intermediate values. Hardly (MeCN) extractable fractions were not significantly different in the 3 soils (34.4-37.3%). NER varied between 35% (SGW) and 37.7% (FYM).

Between 3 and 156 days, mineralization increased by a 4.4-4.8 factor. NER increased to reach 49.7% (CONT), 55.9% (SGW) and 60.1% (FYM) of recovered <sup>14</sup>C. On the contrary, the HPCD and MeCN <sup>14</sup>C fractions decreased and ranged between 6.3-7.3% and 21.2-30.6% respectively. At d156, the HPCD fractions were not significantly different between the 3 soils.

CONT presented the highest mineralized and MeCN fractions. FYM was characterized by a lower MeCN fraction and higher NER. The majority of <sup>14</sup>C was concentrated in the NER for the 3 soils.

### 3.1.4. Comparison between SMX and N-ac-SMX at d3 and d156

At d3, radioactivity in the mineralized, HPCD and MeCN extractable fractions was significantly lower for SMX than for N-ac-SMX (Table S2). The mineralization from <sup>14</sup>C-SMX was about half of that of N-ac-SMX. So, at d3, N-Ac-SMX was more mineralized and remained more extractable than SMX. At the opposite, addition of SMX produced more NER than N-ac-SMX. After 156 days, the HPCD fraction was significantly higher for SMX than N-ac-SMX except in CONT where it was similar. The MeCN fraction was similar for SMX
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**Fig. 4.** Distribution of <sup>14</sup>C-N-ac-SMX in the different soil fractions (mineralized, HPCD and MeCN extractable fractions, non-extractable residues), at (A) day 3, and (B) day 156, in the control soil (CONT) and the soils amended with a sewage sludge-green waste compost (SGW) or farmyard manure (FYM).

After 3 days, N-ac-SMX mineralization was low (2.1-3.0%) and followed the order CONT≥FYM>SGW. The HPCD fractions represented 22.9-25.6% of recovered <sup>14</sup>C and were significantly higher in SGW than in CONT, FYM presenting intermediate values. Hardly (MeCN) extractable fractions were not significantly different in the 3 soils (34.4-37.3%). NER varied between 35% (SGW) and 37.7% (FYM).

Between 3 and 156 days, mineralization increased by a 4.4-4.8 factor. NER increased to reach 49.7% (CONT), 55.9% (SGW) and 60.1% (FYM) of recovered <sup>14</sup>C. On the contrary, the HPCD and MeCN <sup>14</sup>C fractions decreased and ranged between 6.3-7.3% and 21.2-30.6% respectively. At d156, the HPCD fractions were not significantly different between the 3 soils.

CONT presented the highest mineralized and MeCN fractions. FYM was characterized by a lower MeCN fraction and higher NER. The majority of <sup>14</sup>C was concentrated in the NER for the 3 soils.

### 3.1.4. Comparison between SMX and N-ac-SMX at d3 and d156

At d3, radioactivity in the mineralized, HPCD and MeCN extractable fractions was significantly lower for SMX than for N-ac-SMX (Table S2). The mineralization from <sup>14</sup>C-SMX was about half of that of N-ac-SMX. So, at d3, N-Ac-SMX was more mineralized and remained more extractable than SMX. At the opposite, addition of SMX produced more NER than N-ac-SMX. After 156 days, the HPCD fraction was significantly higher for SMX than N-ac-SMX except in CONT where it was similar. The MeCN fraction was similar for SMX
299

and N-ac-SMX in the three soils. NER were significantly higher in CONT after addition of SMX while they were similar for SMX and N-ac-SMX in SGW and FYM. When comparing fraction by fraction, differences between SMX and N-ac-SMX were similar for the 3 soils.

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### 3.1.5. Qualitative analysis of ¹⁴C in extractable fractions

304

Three days after SMX addition, HPCD fractions that represented around 20% of recovered ¹⁴C, were mainly composed of ¹⁴C-SMX (71.3-80.5% of extractable ¹⁴C). Four other metabolites including N-ac-SMX were detected but in smaller quantities. Metabolites differed among the 3 soils (Table 3).

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In the MeCN pH9 fraction, we quantified no SMX but polar metabolites (M1, M2, M3) and an apolar metabolite (M9) at d3. Results were similar for the 3 soils. Between 3 and 156 days, metabolites evolved differently in the MeCN pH9 extracts. M2 and M3 disappeared while M9 and especially M1 increased. M1 (66.7-83.4% of extractable ¹⁴C and 10.1-11.9% of total ¹⁴C), the main metabolite and M9 were present in the 3 soils. M5 seemed specifically appear in the amended soils after 156 days.

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SMX was present in the upper phase of MeCN pH2 extracts and represented 46.1 to 100% of extractable ¹⁴C (3.9 and 7.7% of total ¹⁴C) in CONT and SGW respectively. SMX was predominant in this fraction for the 3 soils but metabolites (M6, M8, M9) were detected at smaller doses in one or another soil. CONT contained a greater diversity of metabolites than amended soils. In the MeCN pH2 extracts, SMX decreased over time but was still present at d156 in CONT and SGW. M8 and M9 disappeared and M4 and M6 remained the only quantifiable metabolites, representing 22.8-66% of extractable ¹⁴C (2.4-7% of total ¹⁴C).

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Three days after its addition, N-ac-SMX has almost completely disappeared in the HPCD fraction, subsisting only in CONT at about 12% of recovered ¹⁴C, and was converted into SMX and other metabolites (M6 et M7). SMX was present in the 3 soils but in larger

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and N-ac-SMX in the three soils. NER were significantly higher in CONT after addition of SMX while they were similar for SMX and N-ac-SMX in SGW and FYM. When comparing fraction by fraction, differences between SMX and N-ac-SMX were similar for the 3 soils.

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### 3.1.5. Qualitative analysis of ¹⁴C in extractable fractions

304

Three days after SMX addition, HPCD fractions that represented around 20% of recovered ¹⁴C, were mainly composed of ¹⁴C-SMX (71.3-80.5% of extractable ¹⁴C). Four other metabolites including N-ac-SMX were detected but in smaller quantities. Metabolites differed among the 3 soils (Table 3).

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In the MeCN pH9 fraction, we quantified no SMX but polar metabolites (M1, M2, M3) and an apolar metabolite (M9) at d3. Results were similar for the 3 soils. Between 3 and 156 days, metabolites evolved differently in the MeCN pH9 extracts. M2 and M3 disappeared while M9 and especially M1 increased. M1 (66.7-83.4% of extractable ¹⁴C and 10.1-11.9% of total ¹⁴C), the main metabolite and M9 were present in the 3 soils. M5 seemed specifically appear in the amended soils after 156 days.

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SMX was present in the upper phase of MeCN pH2 extracts and represented 46.1 to 100% of extractable ¹⁴C (3.9 and 7.7% of total ¹⁴C) in CONT and SGW respectively. SMX was predominant in this fraction for the 3 soils but metabolites (M6, M8, M9) were detected at smaller doses in one or another soil. CONT contained a greater diversity of metabolites than amended soils. In the MeCN pH2 extracts, SMX decreased over time but was still present at d156 in CONT and SGW. M8 and M9 disappeared and M4 and M6 remained the only quantifiable metabolites, representing 22.8-66% of extractable ¹⁴C (2.4-7% of total ¹⁴C).

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Three days after its addition, N-ac-SMX has almost completely disappeared in the HPCD fraction, subsisting only in CONT at about 12% of recovered ¹⁴C, and was converted into SMX and other metabolites (M6 et M7). SMX was present in the 3 soils but in larger

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quantities in the amended soils compared to CONT. CONT also presented a higher diversity
of metabolites.

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The metabolites present in the MeCN pH9 fraction 3 days after N-ac-SMX addition were the
same in the 3 soils (M1, M2, M3, M9) with similar distributions to those obtained after SMX
addition.

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329

In the MeCN pH2 extracts at d3, N-ac-SMX subsisted around 12% of extractable ¹⁴C in the 3
soils and was mostly transformed into SMX (70%-87%).

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At d156, only the MeCN pH9 extracts contained a sufficient radioactivity and could be
analyzed by HPLC. Between d3 and d156, the polar metabolite M1 increased by a factor of
2.3 to 3.5 and represented 59-71% of extractable ¹⁴C in the 3 soils. M2 and M3 decreased or
disappeared and M9 increased to reach around 20% of extractable ¹⁴C in CONT and SGW. As
for SMX, M5 appeared only in amended soils. The evolution of the MeCN pH9 metabolites
was the same for SMX and N-ac-SMX (Table 3).

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quantities in the amended soils compared to CONT. CONT also presented a higher diversity
of metabolites.

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326

The metabolites present in the MeCN pH9 fraction 3 days after N-ac-SMX addition were the
same in the 3 soils (M1, M2, M3, M9) with similar distributions to those obtained after SMX
addition.

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329

In the MeCN pH2 extracts at d3, N-ac-SMX subsisted around 12% of extractable ¹⁴C in the 3
soils and was mostly transformed into SMX (70%-87%).

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At d156, only the MeCN pH9 extracts contained a sufficient radioactivity and could be
analyzed by HPLC. Between d3 and d156, the polar metabolite M1 increased by a factor of
2.3 to 3.5 and represented 59-71% of extractable ¹⁴C in the 3 soils. M2 and M3 decreased or
disappeared and M9 increased to reach around 20% of extractable ¹⁴C in CONT and SGW. As
for SMX, M5 appeared only in amended soils. The evolution of the MeCN pH9 metabolites
was the same for SMX and N-ac-SMX (Table 3).

332

333

334

335

336

337
338 Table 3

Qualitative analysis of ¹⁴C by HPLC after addition of SMX or N-ac-SMX in the easily extractable fraction (HPCD), hardly extractable fraction at pH9 (MeCN pH9) and hardly extractable fraction (upper phase) at pH2 (MeCN pH2) of CONT, SGW and FYM respectively after 3 (d3) and 156 days (d156) of incubation: detection of SMX, N-ac-SMX and 9 other metabolites (M1-M9)

Results are expressed as % of extractable ¹⁴C or % of total ¹⁴C.

343 - Extracts non analyzed by HPLC because of a too low radioactivity

<table><thead><tr><th rowspan="3"></th><th rowspan="3"></th><th colspan="8">SMX</th><th colspan="8">N-ac-SMX</th></tr><tr><th colspan="2">CONT</th><th colspan="2">SGW</th><th colspan="2">FYM</th><th colspan="2">CONT</th><th colspan="2">SGW</th><th colspan="2">FYM</th></tr><tr><th>Extractable ¹⁴C (%)</th><th>Total ¹⁴C (%)</th><th>Extractable ¹⁴C (%)</th><th>Total ¹⁴C (%)</th><th>Extractable ¹⁴C (%)</th><th>Total ¹⁴C (%)</th><th>Extractable ¹⁴C (%)</th><th>Total ¹⁴C (%)</th><th>Extractable ¹⁴C (%)</th><th>Total ¹⁴C (%)</th><th>Extractable ¹⁴C (%)</th><th>Total ¹⁴C (%)</th></tr></thead><tbody><tr><td rowspan="6">HPCD</td><td>M6 (24.4')</td><td>0</td><td>0</td><td>8.7</td><td>1.8</td><td>10.3</td><td>1.8</td><td>27.8</td><td>6.4</td><td>0</td><td>0</td><td>0</td></tr><tr><td>SMX (25.4')</td><td>74.7</td><td>15.6</td><td>71.3</td><td>14.9</td><td>80.5</td><td>14.4</td><td>36.1</td><td>8.3</td><td>67.9</td><td>17.3</td><td>100</td></tr><tr><td>M7 (26.3')</td><td>18.6</td><td>3.9</td><td>13.2</td><td>2.8</td><td>0</td><td>0</td><td>24.2</td><td>5.5</td><td>32.1</td><td>8.2</td><td>0</td></tr><tr><td>N-ac-SMX (27.8')</td><td>0</td><td>0</td><td>6.9</td><td>1.4</td><td>0</td><td>0</td><td>11.9</td><td>2.7</td><td>0</td><td>0</td><td>0</td></tr><tr><td>M9 (37.3')</td><td>6.7</td><td>1.4</td><td>0</td><td>0</td><td>9.2</td><td>1.7</td><td>0</td><td>0</td><td>0</td><td>0</td><td>0</td></tr><tr><td>Total</td><td>100</td><td>20.9</td><td>100</td><td>20.9</td><td>100</td><td>17.9</td><td>100</td><td>22.9</td><td>100</td><td>25.5</td><td>100</td></tr><tr><td rowspan="6">d3</td><td>M1 (3.7')</td><td>23.6</td><td>4.1</td><td>24.0</td><td>4.5</td><td>27.3</td><td>3.4</td><td>30.6</td><td>7.3</td><td>21.0</td><td>5.6</td><td>28.7</td></tr><tr><td>M2 (6.5')</td><td>37.8</td><td>6.6</td><td>27.1</td><td>5.0</td><td>20.9</td><td>2.6</td><td>20.1</td><td>4.8</td><td>20.0</td><td>5.4</td><td>17.6</td></tr><tr><td>M3 (9.8')</td><td>31.0</td><td>5.4</td><td>42.0</td><td>7.8</td><td>44.7</td><td>5.6</td><td>43.1</td><td>10.3</td><td>50.0</td><td>13.5</td><td>42.8</td></tr><tr><td>M9 (37.3')</td><td>7.7</td><td>1.3</td><td>7.0</td><td>1.3</td><td>7.2</td><td>0.9</td><td>6.2</td><td>1.5</td><td>8.8</td><td>2.4</td><td>11.0</td></tr><tr><td>Total</td><td>100</td><td>17.4</td><td>100</td><td>18.6</td><td>100</td><td>12.5</td><td>100</td><td>23.8</td><td>100</td><td>26.9</td><td>100</td></tr><tr><td>M6 (24.4')</td><td>10.3</td><td>0.9</td><td>0</td><td>0</td><td>0</td><td>0</td><td>2.8</td><td>0.3</td><td>0</td><td>0</td><td>0</td></tr><tr><td rowspan="4">MeCN pH9</td><td>SMX (25.4')</td><td>46.1</td><td>3.9</td><td>100</td><td>7.7</td><td>76.7</td><td>5.2</td><td>74.7</td><td>6.9</td><td>87.4</td><td>8.2</td><td>69.9</td></tr><tr><td>N-ac-SMX (27.8')</td><td>0</td><td>0</td><td>0</td><td>0</td><td>0</td><td>0</td><td>12.5</td><td>1.2</td><td>12.6</td><td>1.2</td><td>10.1</td></tr><tr><td>M8 (35.0)</td><td>16.2</td><td>1.4</td><td>0</td><td>0</td><td>23.3</td><td>1.6</td><td>0</td><td>0</td><td>0</td><td>0</td><td>0</td></tr><tr><td>M9 (37.3')</td><td>27.4</td><td>2.3</td><td>0</td><td>0</td><td>0</td><td>0</td><td>10.1</td><td>0.9</td><td>0</td><td>0</td><td>20.0</td></tr><tr><td></td><td>Total</td><td>100</td><td>8.4</td><td>100</td><td>7.7</td><td>100</td><td>6.8</td><td>100</td><td>9.2</td><td>100</td><td>9.4</td><td>100</td></tr><tr><td>d156</td><td>MeCN</td><td>M1 (3.7')</td><td>66.7</td><td>10.1</td><td>67.1</td><td>11.9</td><td>69.6</td><td>11.6</td><td>71.2</td><td>13.1</td><td>61.0</td><td>9.4</td><td>58.8</td></tr></tbody></table>
338 Table 3

Qualitative analysis of ¹⁴C by HPLC after addition of SMX or N-ac-SMX in the easily extractable fraction (HPCD), hardly extractable fraction at pH9 (MeCN pH9) and hardly extractable fraction (upper phase) at pH2 (MeCN pH2) of CONT, SGW and FYM respectively after 3 (d3) and 156 days (d156) of incubation: detection of SMX, N-ac-SMX and 9 other metabolites (M1-M9)

Results are expressed as % of extractable ¹⁴C or % of total ¹⁴C.

343 - Extracts non analyzed by HPLC because of a too low radioactivity

<table><thead><tr><th rowspan="3"></th><th rowspan="3"></th><th colspan="6">SMX</th><th colspan="6">N-ac-SMX</th></tr><tr><th colspan="2">CONT</th><th colspan="2">SGW</th><th colspan="2">FYM</th><th colspan="2">CONT</th><th colspan="2">SGW</th><th colspan="2">FYM</th></tr><tr><th>Extractable ¹⁴C (%)</th><th>Total ¹⁴C (%)</th><th>Extractable ¹⁴C (%)</th><th>Total ¹⁴C (%)</th><th>Extractable ¹⁴C (%)</th><th>Total ¹⁴C (%)</th><th>Extractable ¹⁴C (%)</th><th>Total ¹⁴C (%)</th><th>Extractable ¹⁴C (%)</th><th>Total ¹⁴C (%)</th><th>Extractable ¹⁴C (%)</th><th>Total ¹⁴C (%)</th></tr></thead><tbody><tr><td rowspan="6">HPCD</td><td>M6 (24.4')</td><td>0</td><td>0</td><td>8.7</td><td>1.8</td><td>10.3</td><td>1.8</td><td>27.8</td><td>6.4</td><td>0</td><td>0</td><td>0</td></tr><tr><td>SMX (25.4')</td><td>74.7</td><td>15.6</td><td>71.3</td><td>14.9</td><td>80.5</td><td>14.4</td><td>36.1</td><td>8.3</td><td>67.9</td><td>17.3</td><td>100</td></tr><tr><td>M7 (26.3')</td><td>18.6</td><td>3.9</td><td>13.2</td><td>2.8</td><td>0</td><td>0</td><td>24.2</td><td>5.5</td><td>32.1</td><td>8.2</td><td>0</td></tr><tr><td>N-ac-SMX (27.8')</td><td>0</td><td>0</td><td>6.9</td><td>1.4</td><td>0</td><td>0</td><td>11.9</td><td>2.7</td><td>0</td><td>0</td><td>0</td></tr><tr><td>M9 (37.3')</td><td>6.7</td><td>1.4</td><td>0</td><td>0</td><td>9.2</td><td>1.7</td><td>0</td><td>0</td><td>0</td><td>0</td><td>0</td></tr><tr><td>Total</td><td>100</td><td>20.9</td><td>100</td><td>20.9</td><td>100</td><td>17.9</td><td>100</td><td>22.9</td><td>100</td><td>25.5</td><td>100</td></tr><tr><td rowspan="6">d3</td><td>M1 (3.7')</td><td>23.6</td><td>4.1</td><td>24.0</td><td>4.5</td><td>27.3</td><td>3.4</td><td>30.6</td><td>7.3</td><td>21.0</td><td>5.6</td><td>28.7</td></tr><tr><td>M2 (6.5')</td><td>37.8</td><td>6.6</td><td>27.1</td><td>5.0</td><td>20.9</td><td>2.6</td><td>20.1</td><td>4.8</td><td>20.0</td><td>5.4</td><td>17.6</td></tr><tr><td>M3 (9.8')</td><td>31.0</td><td>5.4</td><td>42.0</td><td>7.8</td><td>44.7</td><td>5.6</td><td>43.1</td><td>10.3</td><td>50.0</td><td>13.5</td><td>42.8</td></tr><tr><td>M9 (37.3')</td><td>7.7</td><td>1.3</td><td>7.0</td><td>1.3</td><td>7.2</td><td>0.9</td><td>6.2</td><td>1.5</td><td>8.8</td><td>2.4</td><td>11.0</td></tr><tr><td>Total</td><td>100</td><td>17.4</td><td>100</td><td>18.6</td><td>100</td><td>12.5</td><td>100</td><td>23.8</td><td>100</td><td>26.9</td><td>100</td></tr><tr><td>M6 (24.4')</td><td>10.3</td><td>0.9</td><td>0</td><td>0</td><td>0</td><td>0</td><td>2.8</td><td>0.3</td><td>0</td><td>0</td><td>0</td></tr><tr><td rowspan="4">MeCN pH9</td><td>SMX (25.4')</td><td>46.1</td><td>3.9</td><td>100</td><td>7.7</td><td>76.7</td><td>5.2</td><td>74.7</td><td>6.9</td><td>87.4</td><td>8.2</td><td>69.9</td></tr><tr><td>N-ac-SMX (27.8')</td><td>0</td><td>0</td><td>0</td><td>0</td><td>0</td><td>0</td><td>12.5</td><td>1.2</td><td>12.6</td><td>1.2</td><td>10.1</td></tr><tr><td>M8 (35.0)</td><td>16.2</td><td>1.4</td><td>0</td><td>0</td><td>23.3</td><td>1.6</td><td>0</td><td>0</td><td>0</td><td>0</td><td>0</td></tr><tr><td>M9 (37.3')</td><td>27.4</td><td>2.3</td><td>0</td><td>0</td><td>0</td><td>0</td><td>10.1</td><td>0.9</td><td>0</td><td>0</td><td>20.0</td></tr><tr><td></td><td>Total</td><td>100</td><td>8.4</td><td>100</td><td>7.7</td><td>100</td><td>6.8</td><td>100</td><td>9.2</td><td>100</td><td>9.4</td><td>100</td></tr><tr><td>d156</td><td>MeCN</td><td>M1 (3.7')</td><td>66.7</td><td>10.1</td><td>67.1</td><td>11.9</td><td>69.6</td><td>11.6</td><td>71.2</td><td>13.1</td><td>61.0</td><td>9.4</td><td>58.8</td></tr></tbody></table>
<table><tbody><tr><td rowspan="4">pH9</td><td>M2 (6.5')</td><td>12.4</td><td>1.9</td><td>0</td><td>0</td><td>0</td><td>0</td><td>7.8</td><td>1.4</td><td>0</td><td>0</td><td>12.9</td></tr><tr><td>M5 (21.6')</td><td>0</td><td>0</td><td>16.2</td><td>2.9</td><td>17.0</td><td>2.8</td><td>0</td><td>0</td><td>19.5</td><td>3.0</td><td>22.3</td></tr><tr><td>M9 (37.3')</td><td>20.9</td><td>3.2</td><td>16.7</td><td>3.0</td><td>13.5</td><td>2.3</td><td>21.0</td><td>3.9</td><td>19.6</td><td>3.0</td><td>6.0</td></tr><tr><td>Total</td><td>100</td><td>15.2</td><td>100</td><td>17.8</td><td>100</td><td>16.7</td><td>100</td><td>18.4</td><td>100</td><td>15.4</td><td>100</td></tr><tr><td rowspan="3">MeCN<br>pH2</td><td>M4 (16.7')</td><td>22.8</td><td>2.4</td><td>0</td><td>0</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>M6 (24.4')</td><td>66.0</td><td>7.0</td><td>43.8</td><td>3.5</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td></tr><tr><td>SMX (25.4')</td><td>11.2</td><td>1.2</td><td>56.2</td><td>4.4</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td></tr><tr><td></td><td>Total</td><td>100</td><td>10.6</td><td>100</td><td>7.9</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td></tr></tbody></table>

344

345
<table><tbody><tr><td rowspan="4">pH9</td><td>M2 (6.5')</td><td>12.4</td><td>1.9</td><td>0</td><td>0</td><td>0</td><td>0</td><td>7.8</td><td>1.4</td><td>0</td><td>0</td><td>12.9</td></tr><tr><td>M5 (21.6')</td><td>0</td><td>0</td><td>16.2</td><td>2.9</td><td>17.0</td><td>2.8</td><td>0</td><td>0</td><td>19.5</td><td>3.0</td><td>22.3</td></tr><tr><td>M9 (37.3')</td><td>20.9</td><td>3.2</td><td>16.7</td><td>3.0</td><td>13.5</td><td>2.3</td><td>21.0</td><td>3.9</td><td>19.6</td><td>3.0</td><td>6.0</td></tr><tr><td>Total</td><td>100</td><td>15.2</td><td>100</td><td>17.8</td><td>100</td><td>16.7</td><td>100</td><td>18.4</td><td>100</td><td>15.4</td><td>100</td></tr><tr><td rowspan="3">MeCN<br>pH2</td><td>M4 (16.7')</td><td>22.8</td><td>2.4</td><td>0</td><td>0</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>M6 (24.4')</td><td>66.0</td><td>7.0</td><td>43.8</td><td>3.5</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td></tr><tr><td>SMX (25.4')</td><td>11.2</td><td>1.2</td><td>56.2</td><td>4.4</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td></tr><tr><td></td><td>Total</td><td>100</td><td>10.6</td><td>100</td><td>7.9</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td></tr></tbody></table>

344

345
346

## 3.2. Fate of CIP into soils amended or not with OWP

347

The ¹⁴C balances ranged from 76 to 91.7% of added ¹⁴C-CIP after 3 days and from 73.8 to 83.2% after 156 days. Presented results are expressed as percentages of recovered ¹⁴C (Fig. 5).

349

350

<figure><img src="image_5.png" /></figure>

351

352

**Fig. 5.** Distribution of ¹⁴C-CIP in the different soil fractions (mineralized, HPCD and MeCN extractable fractions, non-extractable residues), at (A) day 3, and (B) day 156, in the control soil (CONT) and the soils amended with a sewage sludge-green waste compost (SGW) or farmyard manure (FYM).

353

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356

357

We observed few evolutions between d3 and d156 in all of the ¹⁴C fractions. CIP mineralization was considered as negligible since % of mineralization (0.5-0.6%) were below the impurity threshold (1.6%) in the 3 soils after 156 days. The radioactivity present in the HPCD fraction represented only a few percent of recovered ¹⁴C. At d156, radioactivity was concentrated in MeCN fractions (20.4-25.9%) and mainly in NER (72.2-77.6%). CIP became rapidly strongly adsorbed or non-extractable. When comparing the 3 soils, NER were significantly higher in FYM while MeCN residues were higher in CONT. SGW presented intermediate values.

358

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360

361

362

363

364
346

## 3.2. Fate of CIP into soils amended or not with OWP

347

The ¹⁴C balances ranged from 76 to 91.7% of added ¹⁴C-CIP after 3 days and from 73.8 to 83.2% after 156 days. Presented results are expressed as percentages of recovered ¹⁴C (Fig. 5).

349

350

<figure><img src="image_6.png" /></figure>

351

**Fig. 5.** Distribution of ¹⁴C-CIP in the different soil fractions (mineralized, HPCD and MeCN extractable fractions, non-extractable residues), at (A) day 3, and (B) day 156, in the control soil (CONT) and the soils amended with a sewage sludge-green waste compost (SGW) or farmyard manure (FYM).

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We observed few evolutions between d3 and d156 in all of the ¹⁴C fractions. CIP mineralization was considered as negligible since % of mineralization (0.5-0.6%) were below the impurity threshold (1.6%) in the 3 soils after 156 days. The radioactivity present in the HPCD fraction represented only a few percent of recovered ¹⁴C. At d156, radioactivity was concentrated in MeCN fractions (20.4-25.9%) and mainly in NER (72.2-77.6%). CIP became rapidly strongly adsorbed or non-extractable. When comparing the 3 soils, NER were significantly higher in FYM while MeCN residues were higher in CONT. SGW presented intermediate values.

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364
## 4. Discussion

### 4.1. Fate of SMX and N-ac-SMX

After addition of ¹⁴C-SMX and ¹⁴C N-ac-SMX as a solution, the fraction of easily extractable ¹⁴C dropped in all soils till d3 and then decreased more slowly between 3 and 156 days. The decrease with time of SMX and N-ac-SMX in the HPCD fraction can be due to i) sorption on soil constituents, ii) physical sequestration in soil and/or iii) degradation. Then, radioactivity was quickly transferred to the hardly extractable fraction and NER. SMX and N-ac-SMX sorption affinity to soils was relatively low as shown by the Kd values in the 3 soils (1-3.2 L.kg⁻¹) that are in the same range of other studies (Drillia et al., 2005; Höltge and Kreuzig, 2007; Kodešová et al., 2015). Different interactions may be involved in sorption of sulfonamides in soils, including surface complexation, ionic and hydrophobic interactions, some of them being pH-dependent (Leal et al., 2013; Tolls, 2001; Gao et al., 2010). In this study, soil pH ranged between 6.7 and 7.3, above the pKas of SMX and N-ac-SMX (Fig. S1), so that these molecules mainly existed in their anionic form, limiting ionic interactions with clay minerals (Gao et al., 2005). But sorption also depends on hydrophobicity, and SMX and its metabolites may thus interact with soil OM (Gao et al., 2010; Leal et al., 2013; Li et al., 2015). Thiele-Bruhn et al. (2004) highlighted the strong affinity between soil organic carbon and the sulfonamides. Furthermore, according to Martinez-Hernandez et al. (2016), sorption is fast and can play a key role during the first 48h of soil incubations. Gulkowska et al. (2014) attributed the fast formation of NER to chemical bonds between humic substances and the sulfonamide sulfamethazine. Thus sorption and chemical reactions of SMX and its metabolites could partially explain the high increase of ¹⁴C in the MeCN and non-extractable fractions at d3, leading to unavailability of these compounds to soil microorganisms. This hypothesis was also confirmed by HPLC analyses showing that after 3 days, SMX, N-ac-
**4. Discussion**

**4.1. Fate of SMX and N-ac-SMX**

After addition of ¹⁴C-SMX and ¹⁴C N-ac-SMX as a solution, the fraction of easily extractable ¹⁴C dropped in all soils till d3 and then decreased more slowly between 3 and 156 days. The decrease with time of SMX and N-ac-SMX in the HPCD fraction can be due to i) sorption on soil constituents, ii) physical sequestration in soil and/or iii) degradation. Then, radioactivity was quickly transferred to the hardly extractable fraction and NER. SMX and N-ac-SMX sorption affinity to soils was relatively low as shown by the Kd values in the 3 soils (1-3.2 L.kg⁻¹) that are in the same range of other studies (Drillia et al., 2005; Höltge and Kreuzig, 2007; Kodešová et al., 2015). Different interactions may be involved in sorption of sulfonamides in soils, including surface complexation, ionic and hydrophobic interactions, some of them being pH-dependent (Leal et al., 2013; Tolls, 2001; Gao et al., 2010). In this study, soil pH ranged between 6.7 and 7.3, above the pKas of SMX and N-ac-SMX (Fig. S1), so that these molecules mainly existed in their anionic form, limiting ionic interactions with clay minerals (Gao et al., 2005). But sorption also depends on hydrophobicity, and SMX and its metabolites may thus interact with soil OM (Gao et al., 2010; Leal et al., 2013; Li et al., 2015). Thiele-Bruhn et al. (2004) highlighted the strong affinity between soil organic carbon and the sulfonamides. Furthermore, according to Martinez-Hernandez et al. (2016), sorption is fast and can play a key role during the first 48h of soil incubations. Gulkowska et al. (2014) attributed the fast formation of NER to chemical bonds between humic substances and the sulfonamide sulfamethazine. Thus sorption and chemical reactions of SMX and its metabolites could partially explain the high increase of ¹⁴C in the MeCN and non-extractable fractions at d3, leading to unavailability of these compounds to soil microorganisms. This hypothesis was also confirmed by HPLC analyses showing that after 3 days, SMX, N-ac-
390

SMX and other metabolites were partially transferred to the MeCN fraction. Then sulfonamides may desorb with time from this fraction to be transferred to readily extractable or non-extractable fractions (Drillia et al., 2005; Sukul et al., 2008). At the end of incubation, NER represented more than 50% of recovered ¹⁴C for SMX and N-ac-SMX. NER can have a chemical origin with binding of ATB and their metabolites to soil constituents (Gulkowska et al., 2014) but can also result from a physical sequestration of compounds or a microbial transformation of ATB and subsequent incorporation into microbial biomass (Kästner et al., 2014).

391

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Previous work showed that SMX and N-ac-SMX are mainly degraded by soil microorganisms and that abiotic degradation is negligible (Acinnelli et al., 2007; Alvarino et al., 2016). Our results showed that SMX and N-ac-SMX biodegradation could be complete (~ 10% of mineralization at d156) or incomplete with metabolite production as previously showed by Liu et al. (2009) and Martinez-Hernandez et al. (2016). The production of metabolites, mainly more polar than parent compounds, started quickly after spiking soils with SMX and N-ac-SMX and confirmed results previously observed by Richardson and Ternes (2011) and Koba et al. (2017). Few metabolites were identified but the degradation pathway of SMX and N-ac-SMX path remained unclear. The main metabolite of sulfonamides is their acetylated form (Baran et al., 2011), whose formation can be reversible (Höltge & Kreuzig, 2007). Metabolites may also vary according to microbial species (Larcher and Yargeau, 2012) and environmental conditions (Müller et al., 2013).

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401

402

403

404

405

406

407

408

409

410

Mineralization was significantly greater for N-ac-SMX than for SMX in the first days of incubation. These results are consistent with the lower adsorption and the higher extractability of N-ac-SMX in soil, leading to a higher availability of N-ac-SMX to soil microorganisms. At the opposite, the higher Kd of SMX can partially explain the higher production of abiotic NER. Höltge and Kreuzig (2007) previously observed higher Kd for SMX compared to N-ac-SMX.

411

412

413

414
SMX and other metabolites were partially transferred to the MeCN fraction. Then sulfonamides may desorb with time from this fraction to be transferred to readily extractable or non-extractable fractions (Drillia et al., 2005; Sukul et al., 2008). At the end of incubation, NER represented more than 50% of recovered ¹⁴C for SMX and N-ac-SMX. NER can have a chemical origin with binding of ATB and their metabolites to soil constituents (Gulkowska et al., 2014) but can also result from a physical sequestration of compounds or a microbial transformation of ATB and subsequent incorporation into microbial biomass (Kästner et al., 2014).

Previous work showed that SMX and N-ac-SMX are mainly degraded by soil microorganisms and that abiotic degradation is negligible (Acinnelli et al., 2007; Alvarino et al., 2016). Our results showed that SMX and N-ac-SMX biodegradation could be complete (~ 10% of mineralization at d156) or incomplete with metabolite production as previously showed by Liu et al. (2009) and Martinez-Hernandez et al. (2016). The production of metabolites, mainly more polar than parent compounds, started quickly after spiking soils with SMX and N-ac-SMX and confirmed results previously observed by Richardson and Ternes (2011) and Koba et al. (2017). Few metabolites were identified but the degradation pathway of SMX and N-ac-SMX path remained unclear. The main metabolite of sulfonamides is their acetylated form (Baran et al., 2011), whose formation can be reversible (Höltge & Kreuzig, 2007). Metabolites may also vary according to microbial species (Larcher and Yargeau, 2012) and environmental conditions (Müller et al., 2013).

Mineralization was significantly greater for N-ac-SMX than for SMX in the first days of incubation. These results are consistent with the lower adsorption and the higher extractability of N-ac-SMX in soil, leading to a higher availability of N-ac-SMX to soil microorganisms. At the opposite, the higher Kd of SMX can partially explain the higher production of abiotic NER. Höltge and Kreuzig (2007) previously observed higher Kd for SMX compared to N-ac-
SMX in most of soils and OWP they studied. Considering HPLC analyses, N-ac-SMX has almost completely disappeared from the HPCD extracts after 3 days and was present only in the MeCN pH2 fractions. The quantification of SMX as the most abundant molecule in the HPCD and MeCN pH2 extracts after N-ac-SMX addition in the 3 soils confirmed the reversible transformation of N-Ac-SMX to SMX (Höltge and Kreuzig, 2007; Zarfl et al., 2009).

The fate of SMX and N-ac-SMX also differs in the second part of incubation. For SMX, the production of MeCN and non-extractable residues was very fast up to 3 days and then evolved little between d3 and d156. For N-ac-SMX, hardly extracted residues increased till 3 days before decreasing. The production of NER was more progressive with N-ac-SMX than for SMX and was explained by a decrease of easily and hardly extractable fractions. Zarfl et al. (2009) suggested by simulation that NER were formed directly from easily extractable residues after sulfonamide addition. This would mean that some of the hardly extractable residues would go back to soil solution before contributing to the formation of NER.

Even if evolution with time was different for the 2 compounds, the HPCD, MeCN and non-extractable fractions were very similar at d156. For SMX and N-ac-SMX, radioactivity was mainly concentrated in NER and represented more than 50% of recovered ¹⁴C. Besides, Zarfl et al. (2009) also showed that the formation of NER from an acetylated metabolite of sulfadiazine was negligible. Therefore, deacetylation of N-ac-SMX to reform the parent compound should be necessary before its transfer to NER. Thus the transformation of N-SMX into SMX leads at long term to a similar distribution of radioactivity in extractable and non-extractable fractions for N-ac-SMX and SMX. On the other hand, the mineralization of the 2 molecules started from the first days of incubation and explains why in this case, differences persisted over time.
415

SMX in most of soils and OWP they studied. Considering HPLC analyses, N-ac-SMX has almost completely disappeared from the HPCD extracts after 3 days and was present only in the MeCN pH2 fractions. The quantification of SMX as the most abundant molecule in the HPCD and MeCN pH2 extracts after N-ac-SMX addition in the 3 soils confirmed the reversible transformation of N-Ac-SMX to SMX (Höltge and Kreuzig, 2007; Zarfl et al., 2009).

416

417

418

419

420

421

The fate of SMX and N-ac-SMX also differs in the second part of incubation. For SMX, the production of MeCN and non-extractable residues was very fast up to 3 days and then evolved little between d3 and d156. For N-ac-SMX, hardly extracted residues increased till 3 days before decreasing. The production of NER was more progressive with N-ac-SMX than for SMX and was explained by a decrease of easily and hardly extractable fractions. Zarfl et al. (2009) suggested by simulation that NER were formed directly from easily extractable residues after sulfonamide addition. This would mean that some of the hardly extractable residues would go back to soil solution before contributing to the formation of NER.

422

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426

427

428

429

Even if evolution with time was different for the 2 compounds, the HPCD, MeCN and non-extractable fractions were very similar at d156. For SMX and N-ac-SMX, radioactivity was mainly concentrated in NER and represented more than 50% of recovered ¹⁴C. Besides, Zarfl et al. (2009) also showed that the formation of NER from an acetylated metabolite of sulfadiazine was negligible. Therefore, deacetylation of N-ac-SMX to reform the parent compound should be necessary before its transfer to NER. Thus the transformation of N-SMX into SMX leads at long term to a similar distribution of radioactivity in extractable and non-extractable fractions for N-ac-SMX and SMX. On the other hand, the mineralization of the 2 molecules started from the first days of incubation and explains why in this case, differences persisted over time.

430

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435

436

437

438

439
440 **4.2. Fate of CIP**

441 The fate of the fluoroquinolone CIP is very different from sulfonamides. No significant
442 mineralization (less than 1%) was detected after 156 days. Except in artificial conditions, e.g.,
443 pure cultures with optimal growth conditions, liquid media, bioreactor (Dorival-Garcia et al.,
444 2013) or specific conditions with microorganisms from drinking water biofilter and adapted to
445 high concentrations in CIP (Liao et al., 2016), most of the studies concluded that
446 ciprofloxacin was recalcitrant to biodegradation (Girardi et al., 2011, Jechalke et al., 2014;
447 Goulas et al., 2016).

448 The low CIP mineralization is due to its strong sorption and its resulting low bioavailability.
449 Kd of CIP was not measured in this study, but the literature reports very high values of Kd, >
450 500 kg.L⁻¹ (Leal et al., 2013; Vasudevan et al., 2009). Different mechanisms such as cation
451 exchange, cation bridging, complexation, and H-binding may be responsible for sorption of
452 fluoroquinolones in soils (Tolls, 2001; Vasudevan et al., 2009). At soil pH, zwitterionic
453 species predominate in the soil solution. Hence a strong attraction between positive charges of
454 CIP and negative charges of soil constituents can explain fast and strong CIP adsorption in
455 soil. Drillia et al. (2005) measured very high desorption coefficients of fluoroquinolones,
456 meaning that they mainly remained sorbed in the soil. Furthermore CIP addition in soil
457 induces a high production of NER (Girardi et al., 2011). The decline of extractable CIP is thus
458 mainly governed by sorption and formation of NER. In our study, from 3 days and throughout
459 the incubation time, the radioactivity was concentrated in the hardly extractable fraction
460 (20.4-25.9% at d156) and even more in NER (72.2-77.6% at d156). Girardi et al. (2011)
461 comparing CIP degradation in biotic and abiotic conditions concluded that NER were mainly
462 from abiotic origin.

463 But no mineralization does not mean no biodegradation. Due to the low radioactivity in the
464 extracts, HPLC analyses could not be performed in this work. But other studies highlighted a
440 **4.2. Fate of CIP**

441 The fate of the fluoroquinolone CIP is very different from sulfonamides. No significant
442 mineralization (less than 1%) was detected after 156 days. Except in artificial conditions, e.g.,
443 pure cultures with optimal growth conditions, liquid media, bioreactor (Dorival-Garcia et al.,
444 2013) or specific conditions with microorganisms from drinking water biofilter and adapted to
445 high concentrations in CIP (Liao et al., 2016), most of the studies concluded that
446 ciprofloxacin was recalcitrant to biodegradation (Girardi et al., 2011, Jechalke et al., 2014;
447 Goulas et al., 2016).

448 The low CIP mineralization is due to its strong sorption and its resulting low bioavailability.
449 Kd of CIP was not measured in this study, but the literature reports very high values of Kd, >
450 500 kg.L⁻¹ (Leal et al., 2013; Vasudevan et al., 2009). Different mechanisms such as cation
451 exchange, cation bridging, complexation, and H-binding may be responsible for sorption of
452 fluoroquinolones in soils (Tolls, 2001; Vasudevan et al., 2009). At soil pH, zwitterionic
453 species predominate in the soil solution. Hence a strong attraction between positive charges of
454 CIP and negative charges of soil constituents can explain fast and strong CIP adsorption in
455 soil. Drillia et al. (2005) measured very high desorption coefficients of fluoroquinolones,
456 meaning that they mainly remained sorbed in the soil. Furthermore CIP addition in soil
457 induces a high production of NER (Girardi et al., 2011). The decline of extractable CIP is thus
458 mainly governed by sorption and formation of NER. In our study, from 3 days and throughout
459 the incubation time, the radioactivity was concentrated in the hardly extractable fraction
460 (20.4-25.9% at d156) and even more in NER (72.2-77.6% at d156). Girardi et al. (2011)
461 comparing CIP degradation in biotic and abiotic conditions concluded that NER were mainly
462 from abiotic origin.

463 But no mineralization does not mean no biodegradation. Due to the low radioactivity in the
464 extracts, HPLC analyses could not be performed in this work. But other studies highlighted a
465

possible incomplete degradation and the appearance of metabolites (Girardi et al., 2011), especially in presence of fungi (Parshikov et al., 1999; Wetzstein et al., 1999).

466

467

468

### 4.3. Influence of soil properties on the fate of antibiotics

469

The use of soils amended or not with OWP had for objective to study how long term applications of different OWP to soils impacted the fate of ATB. Two hypotheses were tested:

470

471

- OWP repeated application may increase ATB biodegradation (Accinelli et al., 2007)

472

stimulating the soil indigenous microorganisms, introducing exogenous degrading

473

microorganisms in soils or promoting the development of an adapted microflora due to the

474

presence of antibiotics in OWP (Topp et al., 2013)

475

- OWP repeated application may enhance antibiotics sorption limiting their bioavailability and decreasing their mineralization (Li et al., 2015).

476

477

We found SMX and N-ac-SMX mineralization lower in OWP amended soils compared to CONT. Li et al (2015) previously observed that the addition of biosolids and compost to soil decreased SMX mineralization but the level of inhibition depended on the quantity of OM.

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480

When comparing results of HPLC analyses between amended and unamended soils after SMX addition, the differences mainly appeared in the MeCN pH2 fraction, with less SMX and more metabolites present in CONT, as if the biodegradation was more advanced while remaining incomplete. These effects appeared at d3 and persisted over time. In soils incubated with N-ac-SMX, the main effect of amendments appeared in HPCD fractions, where organic matter seemed to have accelerated the de-acetylation into SMX and inhibited the formation of one metabolite (M6). The M5 metabolite present at d156 after addition of SMX and N-ac-SMX seemed specifically produced in amended soils.

481

482

483

484

485

486

487

488

On another side, differences were observed between SGW and FYM, i.e., SMX and N-ac-SMX mineralization was more inhibited in SGW than in FYM. HPLC analyses also showed

489
465

possible incomplete degradation and the appearance of metabolites (Girardi et al., 2011), especially in presence of fungi (Parshikov et al., 1999; Wetzstein et al., 1999).

466

467

468

### 4.3. Influence of soil properties on the fate of antibiotics

469

The use of soils amended or not with OWP had for objective to study how long term applications of different OWP to soils impacted the fate of ATB. Two hypotheses were tested:

470

471

- OWP repeated application may increase ATB biodegradation (Accinelli et al., 2007)

472

stimulating the soil indigenous microorganisms, introducing exogenous degrading microorganisms in soils or promoting the development of an adapted microflora due to the

473

474

presence of antibiotics in OWP (Topp et al., 2013)

475

- OWP repeated application may enhance antibiotics sorption limiting their bioavailability and decreasing their mineralization (Li et al., 2015).

476

477

We found SMX and N-ac-SMX mineralization lower in OWP amended soils compared to CONT. Li et al (2015) previously observed that the addition of biosolids and compost to soil decreased SMX mineralization but the level of inhibition depended on the quantity of OM.

478

479

480

When comparing results of HPLC analyses between amended and unamended soils after SMX addition, the differences mainly appeared in the MeCN pH2 fraction, with less SMX and more metabolites present in CONT, as if the biodegradation was more advanced while remaining incomplete. These effects appeared at d3 and persisted over time. In soils incubated with N-ac-SMX, the main effect of amendments appeared in HPCD fractions, where organic matter seemed to have accelerated the de-acetylation into SMX and inhibited the formation of one metabolite (M6). The M5 metabolite present at d156 after addition of SMX and N-ac-SMX seemed specifically produced in amended soils.

481

482

483

484

485

486

487

488

On another side, differences were observed between SGW and FYM, i.e., SMX and N-ac-SMX mineralization was more inhibited in SGW than in FYM. HPLC analyses also showed

489
that the proportion of easily extractable SMX was higher in FYM while strongly adsorbed
SMX was higher in SGW. These results are consistent with soil organic carbon contents
higher in SGW than in FYM (17.6 vs 15.5 g.kg⁻¹) and values of Kd higher in SGW than in
FYM for both SMX and N-ac-SMX. Furthermore, Racke and Lichtenstein (1985) reported
that the quality of OM and its degree of decomposition were also key factors. OM is more
stable and less degradable in compost of sewage sludge and green waste than in manure.
Finally organic amendments stimulated the production of NER whatever the studied antibiotic
(SMX, N-ac-SMX, CIP). The increase of NER occurred very quickly with SMX and CIP and
later with N-ac-SMX. In addition, this effect was more pronounced in FYM than in SGW. Li
et al. (2015) also showed that OWP amendment increased the formation and accumulation of
NER in a soil spiked with SMX from the first days of incubation. They suggested that OWP
provide additional sites facilitating binding or conjugation interactions between contaminants
and soil constituents and/or promoted microbial transformation leading to the formation of
intermediates that were readily bound to the matrix.

504

## 5. Conclusion

The fate of two sulfonamides, SMX and N-ac-SMX, and one fluoroquinolone, CIP, in soils amended or not with manure or a co-compost of sludge and green waste depends on molecules and soil characteristics. Antibiotic mineralization was nil or low after 156 days (CIP << SMX < N-ac-SMX) while the production of NER accounted for more than 50% of recovered ¹⁴C.

Organic amendments decrease antibiotics degradation and mineralization (SMX and N-ac-
SMX) and increase the production of NER (SMX, N-ac-SMX and CIP). The effects also
depend on the quantity and quality of added OM. The SGW compost further inhibits
mineralization while FYM enhances the production of NER. Land application of OWP is an
that the proportion of easily extractable SMX was higher in FYM while strongly adsorbed
SMX was higher in SGW. These results are consistent with soil organic carbon contents
higher in SGW than in FYM (17.6 vs 15.5 g.kg⁻¹) and values of Kd higher in SGW than in
FYM for both SMX and N-ac-SMX. Furthermore, Racke and Lichtenstein (1985) reported
that the quality of OM and its degree of decomposition were also key factors. OM is more
stable and less degradable in compost of sewage sludge and green waste than in manure.
Finally organic amendments stimulated the production of NER whatever the studied antibiotic
(SMX, N-ac-SMX, CIP). The increase of NER occurred very quickly with SMX and CIP and
later with N-ac-SMX. In addition, this effect was more pronounced in FYM than in SGW. Li
et al. (2015) also showed that OWP amendment increased the formation and accumulation of
NER in a soil spiked with SMX from the first days of incubation. They suggested that OWP
provide additional sites facilitating binding or conjugation interactions between contaminants
and soil constituents and/or promoted microbial transformation leading to the formation of
intermediates that were readily bound to the matrix.

504

## 5. Conclusion

The fate of two sulfonamides, SMX and N-ac-SMX, and one fluoroquinolone, CIP, in soils amended or not with manure or a co-compost of sludge and green waste depends on molecules and soil characteristics. Antibiotic mineralization was nil or low after 156 days (CIP << SMX < N-ac-SMX) while the production of NER accounted for more than 50% of recovered ¹⁴C.

Organic amendments decrease antibiotics degradation and mineralization (SMX and N-ac-
SMX) and increase the production of NER (SMX, N-ac-SMX and CIP). The effects also
depend on the quantity and quality of added OM. The SGW compost further inhibits
mineralization while FYM enhances the production of NER. Land application of OWP is an

490

491

492

493

494

495

496

497

498

499

500

501

502

503

504

505

506

507

508

509

510

511

512

513

514
increasing practice in agriculture and urban gardening and so the impact of OWP on the fate
of ATB and its associated health and environmental risks are of great interest to be studied.
But OWP often contain multiple contaminants (other xenobiotics, metals ...), as well as a
complex organic matrix that interacts with the contaminants during storage or treatment
(composting, methanisation, liming, drying...), and then after soil spreading. Consequently it
would be interesting to evaluate whether the concomitant addition of contaminants (cocktail
effects) in soils and/or the evolution of OM-ATB interactions over time have an impact on the
fate of antibiotics.

523

524 6. Acknowledgements

Acknowledgements are due to ADEME and INRA, Department of Environment and
Agronomy, for the PhD grant of the first author, and to the French National Research Agency
who financed the project CEMABS (ANR-13-CESA-0016-01). Soil samples came from the
QualiAgro field experiment supported by Veolia Research and Innovation and by AnaEE-
France. We also thank the staff of ECOSYS who helped us during sampling, experiments and
data treatments.

531

532 7. Bibliographie

533 Accinelli, C., Koskinen, W.C., Becker, J.M., Sadowsky, M.J. (2007). Environmental Fate of
Two Sulfonamide Antimicrobial Agents in Soil. J. Agric. Food. Chem., 55, 2677-2682.
534 http://doi.org/10.1021/jf063709j
535

536 Alvarino, T., Nastold, P., Suarez, S., Omil, F., Corvini, P.F.X., Bouju, H. (2016). Role of
biotransformation, sorption and mineralization of 14 C-labelled sulfamethoxazole under
different redox conditions. *Science of the Total Environment*, 542, 706-715.
537
538
http://doi.org/10.1016/j.scitotenv.2015.10.140
539
increasing practice in agriculture and urban gardening and so the impact of OWP on the fate
of ATB and its associated health and environmental risks are of great interest to be studied.
But OWP often contain multiple contaminants (other xenobiotics, metals ...), as well as a
complex organic matrix that interacts with the contaminants during storage or treatment
(composting, methanisation, liming, drying...), and then after soil spreading. Consequently it
would be interesting to evaluate whether the concomitant addition of contaminants (cocktail
effects) in soils and/or the evolution of OM-ATB interactions over time have an impact on the
fate of antibiotics.

523

524 6. Acknowledgements

Acknowledgements are due to ADEME and INRA, Department of Environment and
Agronomy, for the PhD grant of the first author, and to the French National Research Agency
who financed the project CEMABS (ANR-13-CESA-0016-01). Soil samples came from the
QualiAgro field experiment supported by Veolia Research and Innovation and by AnaEE-
France. We also thank the staff of ECOSYS who helped us during sampling, experiments and
data treatments.

531

532 7. Bibliographie

533 Accinelli, C., Koskinen, W.C., Becker, J.M., Sadowsky, M.J. (2007). Environmental Fate of
Two Sulfonamide Antimicrobial Agents in Soil. J. Agric. Food. Chem., 55, 2677-2682.
534 http://doi.org/10.1021/jf063709j
535

536 Alvarino, T., Nastold, P., Suarez, S., Omil, F., Corvini, P.F.X., Bouju, H. (2016). Role of
biotransformation, sorption and mineralization of 14 C-labelled sulfamethoxazole under
different redox conditions. *Science of the Total Environment*, 542, 706-715.
537
538
539 http://doi.org/10.1016/j.scitotenv.2015.10.140
540

Baran, W., Adamek, E., Ziemiańska, J., Sobczak, A. (2011). Effects of the presence of

541

sulfonamides in the environment and their influence on human health. J. Hazard. Mater.,

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196, 1-15.

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Barnes, K.K., Christenson, S.C., Kolpin D.W., Focazio, M., Furlong, E.T., Zaugg, S.D.,

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Meyer, M.T., Barber, L.B. (2004). Pharmaceuticals and other organic waste water

545

contaminants within a leachate plume downgradient of a municipal landfill. Ground

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Water Monit. Rem., 24 (2), 119-126.

547

Bialk, H.M., Simpson A.J., Pedersen, J.A. (2005). Cross-Coupling of Sulfonamide

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Antimicrobial Agents with Model Humic Constituents. Env. Sci. Technol., 39(12),

549

4463-4473. doi 10.1021/es0500916

550

Bourdat-Deschamps, M., Ferhi, S., Bernet, N., Feder, F., Crouzet, O., Patureau, P.,

551

Montenach, D., Moussard, G.D., Mercier, V., Benoit, P., Houot, S. (2017). Fate and

552

impacts of pharmaceuticals and personal care products after repeated applications of

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organic waste products in long-term field experiments. Science of the Total Environment

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607-608, 271-280.

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Boxall, A.B.A., Kay, P., Blackwell, P.A., Fogg, L.A., (2004). Fate of veterinary medicines

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applied to soils. In: Kümmerer, K. (Ed.), Pharmaceuticals in the Environment: Sources,

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Fate, Effects and Risks. Springer-Verlag, Berlin, pp. 165-180.

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Boxall, A.B.A., Rudd, M.A., Brooks, B.W., Caldwell, D.J., Choi, K., Hickmann, S., Innes, E.,

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Ostapyk, K., Staveley, J., P., Beazley, K.F. (2012). Review Pharmaceuticals and

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Personal Care Products in the Environment: What Are the Big Questions?. Environ.

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Health Perspect., 120(9), 1221-1229. doi 10.1289/ehp.1104477

562

Chang, A.C., Pan, G., Page, A.L., Asano, T. (1995). Developing Human Health-related

563

Chemical Guidelines for Reclaimed Waster and Sewage Sludge Applications in

564

Agriculture. World Health Organization, Geneva, 1995., WHO/EOS/95.20.
540 Baran, W., Adamek, E., Ziemiańska, J., Sobczak, A. (2011). Effects of the presence of

541 sulfonamides in the environment and their influence on human health. J. Hazard. Mater.,

542 196, 1-15.

543 Barnes, K.K., Christenson, S.C., Kolpin D.W., Focazio, M., Furlong, E.T., Zaugg, S.D.,

544 Meyer, M.T., Barber, L.B. (2004). Pharmaceuticals and other organic waste water

545 contaminants within a leachate plume downgradient of a municipal landfill. Ground

546 Water Monit. Rem., 24 (2), 119-126.

547 Bialk, H.M., Simpson A.J., Pedersen, J.A. (2005). Cross-Coupling of Sulfonamide

548 Antimicrobial Agents with Model Humic Constituents. Env. Sci. Technol., 39(12),

549 4463-4473. doi 10.1021/es0500916

550 Bourdat-Deschamps, M., Ferhi, S., Bernet, N., Feder, F., Crouzet, O., Patureau, P.,

551 Montenach, D., Moussard, G.D., Mercier, V., Benoit, P., Houot, S. (2017). Fate and

552 impacts of pharmaceuticals and personal care products after repeated applications of

553 organic waste products in long-term field experiments. Science of the Total Environment

554 607-608, 271-280.

555 Boxall, A.B.A., Kay, P., Blackwell, P.A., Fogg, L.A., (2004). Fate of veterinary medicines

556 applied to soils. In: Kümmerer, K. (Ed.), Pharmaceuticals in the Environment: Sources,

557 Fate, Effects and Risks. Springer-Verlag, Berlin, pp. 165-180.

558 Boxall, A.B.A., Rudd, M.A., Brooks, B.W., Caldwell, D.J., Choi, K., Hickmann, S., Innes, E.,

559 Ostapyk, K., Staveley, J., P., Beazley, K.F. (2012). Review Pharmaceuticals and

560 Personal Care Products in the Environment: What Are the Big Questions?. Environ.

561 Health Perspect., 120(9), 1221-1229. doi 10.1289/ehp.1104477

562 Chang, A.C., Pan, G., Page, A.L., Asano, T. (1995). Developing Human Health-related

563 Chemical Guidelines for Reclaimed Waster and Sewage Sludge Applications in

564 Agriculture. World Health Organization, Geneva, 1995., WHO/EOS/95.20.
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Dorival-García, N., Zafra-Gómez, A., Navalón, A., González, J., Hontoria, E., Vílchez, J.L. (2013). Removal of quinolone antibiotics from wastewaters by sorption and biological degradation in laboratory-scale membrane bioreactors. Sci. Total Environ., 442, 317-328. http://doi.org/10.1016/j.scitotenv.2012.10.026

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Du, L., Liu, W. (2012). Occurrence, fate, and ecotoxicity of antibiotics in agro-ecosystems. A review. Agron. Sustain. Dev., 32, 309-327. http://doi.org/10.1007/s13593-011-0062-9

Ferhi, S., Bourdat-Deschamps, M., Daudin, J.J., Houot, S., Nélieu, S. (2016). Factors influencing the extraction of pharmaceuticals from sewage sludge and soil: an experimental design approach. Anal Bioanal Chem., 408, 6153-6168.

Gao, J., Pedersen, J. (2005). Adsorption of Sulfonamide Antimicrobial Agents to Clay Minerals. Environ. Sci. Technol., 39, 9509-9516.

Gao, J., Pedersen, J. (2010). Sorption of Sulfonamide Antimicrobial Agents to Humic Acid – Clay Complexes. J. Environ. Qual., 39, 228-235. doi:10.2134/jeq2008.0274

Girardi, C., Greve, J., Lamshöft, M., Fetzer, I., Miltner, A., Schäffer, A., Kästner, M. (2011). Biodegradation of ciprofloxacin in water and soil and its effects on the microbial communities. J. Hazard. Mater., 198, 22-30. http://doi.org/10.1016/j.jhazmat.2011.10.004

Göbel, A., Thomsen, A., McArdell, C.S., Joss, A., Giger, W. (2005). Occurrence and sorption behavior of sulfonamides, macrolides, and trimethoprim in activated sludge treatment.
Chen, H., Gao, B., Li, H., Ma, L.Q. (2011). Effects of pH and ionic strength on sulfamethoxazole and ciprofloxacin transport in saturated porous media. *J. Contam. Hydrol.*, 126(1–2), 29–36. http://doi.org/10.1016/j.jconhyd.2011.06.002

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Du, L., Liu, W. (2012). Occurrence, fate, and ecotoxicity of antibiotics in agro-ecosystems. A review. *Agron. Sustain. Dev.*, 32, 309–327. http://doi.org/10.1007/s13593-011-0062-9

Ferhi, S., Bourdat-Deschamps, M., Daudin, J.J., Houot, S., Nélieu, S. (2016). Factors influencing the extraction of pharmaceuticals from sewage sludge and soil: an experimental design approach. *Anal Bioanal Chem.*, 408, 6153–6168.

Gao, J., Pedersen, J. (2005). Adsorption of Sulfonamide Antimicrobial Agents to Clay Minerals. *Environ. Sci. Technol.*, 39, 9509–9516.

Gao, J., Pedersen, J. (2010). Sorption of Sulfonamide Antimicrobial Agents to Humic Acid – Clay Complexes. *J. Environ. Qual.*, 39, 228–235. doi:10.2134/jeq2008.0274

Girardi, C., Greve, J., Lamshöft, M., Fetzer, I., Miltner, A., Schäffer, A., Kästner, M. (2011). Biodegradation of ciprofloxacin in water and soil and its effects on the microbial communities. *J. Hazard. Mater.*, 198, 22–30. http://doi.org/10.1016/j.jhazmat.2011.10.004

Göbel, A., Thomsen, A., McArdell, C.S., Joss, A., Giger, W. (2005). Occurrence and sorption behavior of sulfonamides, macrolides, and trimethoprim in activated sludge treatment.
590 Env. Sci. Technol., 39, 3981-3989. http://doi.org/10.1021/es048550a

591 Golet, E.M., Strehler, A., Alder, A.C., Giger, W. (2002). Determination of fluoroquinolone
592 antibacterial agents in sewage sludge and sludge-treated soil using accelerated solvent
593 extraction followed by solid-phase extraction. Anal. Chem., 74(21), 5455-5462.

594 Goulas, A., Haudin, C., Bourdat-Deschamps, M., Benoit, P., Ferhi, S. (2016). A new
595 extraction method to assess the environmental availability of ciprofloxacin in agricultural
596 soils amended with exogenous organic matter. Chemosphere, 165, 460-469.
597 doi.org/10.1016/j.chemosphere.2016.09.040

598 Gulkowska, A., Thalmann, B., Hollender, J., Krauss, M. (2014). Nonextractable residue
599 formation of sulfonamide antimicrobials: New insights from soil incubation experiments.
600 Chemosphere 107, 366-372.

601 Hamscher, G., Pawelzick, H.T., Höper, H., Nau, H. (2005). Different behaviour of
602 tetracyclines and sulfonamides in sandy soils after repeated fertilization with liquid
603 manure. Environ. Toxicol. Chem., 24, 861-8.

604 Höltge, S., Kreuzig, R. (2007). Laboratory testing of sulfamethoxazole and its metabolite
605 acetyl-sulfamethoxazole in soil. Clean Soil Air Water, 35(1), 104-110.
606 doi.org/10.1002/clen.200600019

607 Houot, S., Clergeot, D., Michelin, J., Francou, C., Bourgeois, S., Caria, G., Ciesielski, H.
608 (2002). Agronomic value and environmental impacts of urban composts used in
609 agriculture. In: Insam, H., Klammer, S. (Eds.), Microbiology of Composting. Springer-
610 Verlag, Berlin Heidelberg, pp. 457-472

611 Jechalke, S., Heuer, H., Siemens, J., Amelung, W., Smalla, K. (2014). Fate and effects of
612 veterinary antibiotics in soil. Trends Microbiol., 22(9), 536-545.
613 http://doi.org/10.1016/j.tim.2014.05.005.

614 Kästner, M., Nowak, K.M.; Miltner, A., Trapp, S., Schaffer, A. (2014) Classification and
590

Env. Sci. Technol., 39, 3981-3989. http://doi.org/10.1021/es048550a

591

Golet, E.M., Strehler, A., Alder, A.C., Giger, W. (2002). Determination of fluoroquinolone antibacterial agents in sewage sludge and sludge-treated soil using accelerated solvent extraction followed by solid-phase extraction. *Anal. Chem.*, **74**(21), 5455–5462.

592

593

594

Goulas, A., Haudin, C., Bourdat-Deschamps, M., Benoit, P., Ferhi, S. (2016). A new extraction method to assess the environmental availability of ciprofloxacin in agricultural soils amended with exogenous organic matter. *Chemosphere*, **165**, 460–469.
doi.org/10.1016/j.chemosphere.2016.09.040

595

596

597

598

Gulkowska, A., Thalmann, B., Hollender, J., Krauss, M. (2014). Nonextractable residue formation of sulfonamide antimicrobials: New insights from soil incubation experiments. *Chemosphere* **107**, 366–372.

599

600

601

Hamscher, G., Pawelzick, H.T., Höper, H., Nau, H. (2005). Different behaviour of tetracyclines and sulfonamides in sandy soils after repeated fertilization with liquid manure. *Environ. Toxicol. Chem.*, **24**, 861–8.

602

603

604

Höltge, S., Kreuzig, R. (2007). Laboratory testing of sulfamethoxazole and its metabolite acetyl-sulfamethoxazole in soil. *Clean Soil Air Water*, **35**(1), 104–110.
doi.org/10.1002/clen.200600019

605

606

607

Houot, S., Clergeot, D., Michelin, J., Francou, C., Bourgeois, S., Caria, G., Ciesielski, H. (2002). Agronomic value and environmental impacts of urban composts used in agriculture. In: Insam, H., Klammer, S. (Eds.), *Microbiology of Composting*. Springer-Verlag, Berlin Heidelberg, pp. 457-472

608

609

610

611

Jechalke, S., Heuer, H., Siemens, J., Amelung, W., Smalla, K. (2014). Fate and effects of veterinary antibiotics in soil. *Trends Microbiol.*, **22**(9), 536–545.
doi://doi.org/10.1016/j.tim.2014.05.005.

612

613

614

Kästner, M., Nowak, K.M.; Miltner, A., Trapp, S., Schaffer, A. (2014) Classification and
615

Modelling of Non extractable Residue (NER) Formation of Xenobiotics in Soil - A
Synthesis. Crit. Rev. Env. Sci. Technol., 44(19), 2107-2171.

616

617

Koba, O., Golovko, O., Kodešová, R., Fér, M., Grabic, R. (2017). Antibiotics degradation in soil: A case of clindamycin, trimethoprim, sulfamethoxazole and their transformation products. Environ. Pollut., 220, 1251-1263.

618

619

620

Kodešová, R., Grabic, R., Kočárek, M., Klement, A., Golovko, O., Fér, M., Nikodem, A., Jakšík. O. (2015). Pharmaceuticals' sorptions relative to properties of thirteen different soils. Sci. Total Environ., 511, 435-443. doi: 10.1016/j.scitotenv.2014.12.088

621

622

623

Kolpin, D.W., Furlong, E.T., Meyer, M.T., Thurman, E.M., Zaugg., S.D., Barber, L.B., Buxton, H.T. (2002). Pharmaceuticals, Hormones, and Other Organic Wastewater Contaminants in U.S. Streams, 1999 - 2000: A National Reconnaissance. Environ. Sci. Technol., 36(6), 1202-1211.

624

625

626

627

Kümmerer, K. (2001). Drugs in the environment: emission of drugs, diagnostic aids and disinfectants into wastewater by hospitals in relation to other sources - a review. Chemosphere, 45,957-969.

628

629

630

Larcher, S and Yargeau, V. (2012). Biodegradation of sulfamethoxazole: current knowledge and perspectives. Appl Microbiol Biotechnol., 96:309-318.

631

632

Lashermes, G., Nicolardot, B., Parnaudeau, V., Thuriès, L., Chaussod, R., Guillotin, M.L., Linères, M., Mary, B., Metzger, L., Morvan, T., Tricaud, A., Vilette, C., Houot, S. (2009). Indicator of Potential Residual Carbon in Soils after Exogenous Organic Matter Application. European Journal of Soil Science., 60 (2), 297-310.

633

634

635

636

Leal, R.M.P., Alleoni, L.R.F., Tornisielo, V.L., Regitano, J.B. (2013). Sorption of fluoroquinolones and sulfonamides in 13 Brazilian soils. Chemosphere, 92(8), 979-985.
http://doi.org/10.1016/j.chemosphere.2013.03.018

637

638

639

Li, J., Ye, Q., Gan, J. (2015). Influence of organic amendment on fate of acetaminophen and
615

Modelling of Non extractable Residue (NER) Formation of Xenobiotics in Soil - A
Synthesis. Crit. Rev. Env. Sci. Technol., 44(19), 2107-2171.

616

617

Koba, O., Golovko, O., Kodešová, R., Fér, M., Grabic, R. (2017). Antibiotics degradation in soil: A case of clindamycin, trimethoprim, sulfamethoxazole and their transformation products. Environ. Pollut., 220, 1251-1263.

618

619

620

Kodešová, R., Grabic, R., Kočárek, M., Klement, A., Golovko, O., Fér, M., Nikodem, A., Jakšík. O. (2015). Pharmaceuticals' sorptions relative to properties of thirteen different soils. Sci. Total Environ., 511, 435-443. doi: 10.1016/j.scitotenv.2014.12.088

621

622

623

Kolpin, D.W., Furlong, E.T., Meyer, M.T., Thurman, E.M., Zaugg., S.D., Barber, L.B., Buxton, H.T. (2002). Pharmaceuticals, Hormones, and Other Organic Wastewater Contaminants in U.S. Streams, 1999 - 2000: A National Reconnaissance. Environ. Sci. Technol., 36(6), 1202-1211.

624

625

626

627

Kümmerer, K. (2001). Drugs in the environment: emission of drugs, diagnostic aids and disinfectants into wastewater by hospitals in relation to other sources - a review. Chemosphere, 45,957-969.

628

629

630

Larcher, S and Yargeau, V. (2012). Biodegradation of sulfamethoxazole: current knowledge and perspectives. Appl Microbiol Biotechnol., 96:309-318.

631

632

Lashermes, G., Nicolardot, B., Parnaudeau, V., Thuriès, L., Chaussod, R., Guillotin, M.L., Linères, M., Mary, B., Metzger, L., Morvan, T., Tricaud, A., Vilette, C., Houot, S. (2009). Indicator of Potential Residual Carbon in Soils after Exogenous Organic Matter Application. European Journal of Soil Science., 60 (2), 297-310.

633

634

635

636

Leal, R.M.P., Alleoni, L.R.F., Tornisielo, V.L., Regitano, J.B. (2013). Sorption of fluoroquinolones and sulfonamides in 13 Brazilian soils. Chemosphere, 92(8), 979-985.
http://doi.org/10.1016/j.chemosphere.2013.03.018

637

638

639

Li, J., Ye, Q., Gan, J. (2015). Influence of organic amendment on fate of acetaminophen and
640

sulfamethoxazole in soil. Environ. Pollut., 206, 543-550.

641

http://doi.org/10.1016/j.envpol.2015.08.019

642

Liao, X., Li, B., Zou, R., Dai, Y., Xie, S., Yuan, B. (2016). Biodegradation of antibiotic

643

ciprofloxacin: pathways, influential factors, and bacterial community structure. Environ.

644

Sci. Pollut. Res. Int., 23(8), 7911-7918. http://doi.org/10.1007/s11356-016-6054-1

645

Liu, F., Ying, G.G., Tao, R., Zhao, J.L., Yang, J.F., Zhao, L.F. (2009). Effects of six selected

646

antibiotics on plant growth and soil microbial and enzymatic activities. Environ. Pollut.,

647

157(5), 1636-1642. http://doi.org/10.1016/j.envpol.2008.12.021

648

Martínez-Hernández, V., Meffe, R., Herrera, S., Bustamante, I.D. (2016). The role of sorption

649

and biodegradation in the removal of acetaminophen, carbamazepine, caffeine, naproxen

650

and sulfamethoxazole during soil contact: A kinetics study. Sci. Total Envir., 559, 232-

651

241. http://doi.org/10.1016/j.scitotenv.2016.03.131

652

Müller, M., Schüssler, W., Horn H., Lemmer, H. (2013). Aerobic biodegradation of the

653

sulfonamide antibiotic sulfamethoxazole by activated sludge applied as co-substrate and

654

sole carbon and nitrogen source. Chemosphere 92, 969-978.

655

Nieto A., Borruli F., Pocurull E., Marce R.M., (2010). Ocurence of pharmaceuticals and

656

hormones in sewage sludge. Environ Toxicol Chem., 29, 1484-1489.

657

Nowara, A., Burhenne, J., Spiteller, M. (1997). Binding of Fluoroquinolone Carboxylic Acid

658

Derivatives to Clay Minerals. J. Agric. Food. Chem., 45, 1459-1463.

659

http://doi.org/10.1021/jf960215l

660

Ozaki, N., Bester, K., Moldrup, P., Henriksen, K., Komatsu, T. (2011). Photodegradation of

661

the synthetic fragrance OTNE and the bactericide triclosan adsorbed on dried loamy sand

662

Results from models and experiments. Chemosphere, 83(11), 1475-1479.

663

http://doi.org/10.1016/j.chemosphere.2011.03.006

664

Parshikov, I.A., Freeman, J.P., Lay, J.O., Beger, R.D., Williams, A.J., Sutherland, J.B. (1999).
640

sulfamethoxazole in soil. Environ. Pollut., 206, 543-550.

641

http://doi.org/10.1016/j.envpol.2015.08.019

642

Liao, X., Li, B., Zou, R., Dai, Y., Xie, S., Yuan, B. (2016). Biodegradation of antibiotic

643

ciprofloxacin: pathways, influential factors, and bacterial community structure. Environ.

644

Sci. Pollut. Res. Int., 23(8), 7911-7918. http://doi.org/10.1007/s11356-016-6054-1

645

Liu, F., Ying, G.G., Tao, R., Zhao, J.L., Yang, J.F., Zhao, L.F. (2009). Effects of six selected

646

antibiotics on plant growth and soil microbial and enzymatic activities. Environ. Pollut.,

647

157(5), 1636-1642. http://doi.org/10.1016/j.envpol.2008.12.021

648

Martínez-Hernández, V., Meffe, R., Herrera, S., Bustamante, I.D. (2016). The role of sorption

649

and biodegradation in the removal of acetaminophen, carbamazepine, caffeine, naproxen

650

and sulfamethoxazole during soil contact: A kinetics study. Sci. Total Envir., 559, 232-

651

241. http://doi.org/10.1016/j.scitotenv.2016.03.131

652

Müller, M., Schüssler, W., Horn H., Lemmer, H. (2013). Aerobic biodegradation of the

653

sulfonamide antibiotic sulfamethoxazole by activated sludge applied as co-substrate and

654

sole carbon and nitrogen source. Chemosphere 92, 969-978.

655

Nieto A., Borruli F., Pocurull E., Marce R.M., (2010). Ocurence of pharmaceuticals and

656

hormones in sewage sludge. Environ Toxicol Chem., 29, 1484-1489.

657

Nowara, A., Burhenne, J., Spiteller, M. (1997). Binding of Fluoroquinolone Carboxylic Acid

658

Derivatives to Clay Minerals. J. Agric. Food. Chem., 45, 1459-1463.

659

http://doi.org/10.1021/jf960215l

660

Ozaki, N., Bester, K., Moldrup, P., Henriksen, K., Komatsu, T. (2011). Photodegradation of

661

the synthetic fragrance OTNE and the bactericide triclosan adsorbed on dried loamy sand

662

Results from models and experiments. Chemosphere, 83(11), 1475-1479.

663

http://doi.org/10.1016/j.chemosphere.2011.03.006

664

Parshikov, I.A., Freeman, J.P., Lay, J.O., Beger, R.D., Williams, A.J., Sutherland, J.B. (1999).
665 Regioselective transformation of ciprofloxacin to N -acetylciprofloxacin by the fungus
666 *Mucor ramannianus*. FEMS Microbiol. Lett., 177, 131–135.

667 Racke, K.D., Lichtenstein, E.P. (1985). Effects of soil microorganisms on the release of
668 bound 14C residues from soils previously treated with [14C] parathion. J. Agric. Food
669 Chem., 33, 938-943.

670 Richardson, S.D., Ternes, T.A. (2011). Water analysis: emerging contaminants and current
671 issues. Anal. Chem., 83, 4614-4648.

672 Rosendahl, I., Siemens, J., Kindler, R., Groeneweg, J., Zimmermann, J., Czerwinski, S.,
673 Lamshöft, M., Wilke, B.M., Vereeken, H., Amelung, W. (2012). Persistence of the
674 Fluoroquinolone Antibiotic Difloxacin in Soil and Lacking Effects on Nitrogen
675 Turnover. J. Environ. Qual., 41, 1275-1283. http://doi.org/10.2134/jeq2011.0459

676 Salvia, M.-V., Experton, J., Geandel, C., Cren-Olivé, C., Vulliet, E. (2014). Fate of
677 pharmaceutical compounds and steroid hormones in soil: study of transfer and
678 degradation in soil columns, Environ. Sci. Pollut. Res., 21, 10525-10535.
679 http://doi.org/10.1007/s11356-014-3038-x

680 Sarmah, A. K., Meyer, M. T., Boxall, A.B. (2006). A global perspective on the use, sales,
681 exposure pathways, occurrence, fate and effects of veterinary antibiotics (VAs ) in the
682 environment. Chemosphere, 65, 725-759.
683 http://doi.org/10.1016/j.chemosphere.2006.03.026

684 Srinivasan, P., Sarmah, A.K. (2014). Assessing the sorption and leaching behaviour of three
685 sulfonamides in pasture soils through batch and column studies. Sci. Total Environ., 493,
686 535-543. http://doi.org/10.1016/j.scitotenv.2014.06.034

687 Sukul, P., Lamshöft, M., Zühlke, S., Spiteller, M. (2008). Sorption and desorption of
688 sulfadiazine in soil and soil-manure systems. Chemosphere, 73(8), 1344-1350.
689 http://doi.org/10.1016/j.chemosphere.2008.06.066
665

Regioselective transformation of ciprofloxacin to N -acetylciprofloxacin by the fungus

666

*Mucor ramannianus*. FEMS Microbiol. Lett., 177, 131–135.

667

Racke, K.D., Lichtenstein, E.P. (1985). Effects of soil microorganisms on the release of

668

bound 14C residues from soils previously treated with [14C] parathion. J. Agric. Food

669

Chem., 33, 938-943.

670

Richardson, S.D., Ternes, T.A. (2011). Water analysis: emerging contaminants and current

671

issues. Anal. Chem., 83, 4614-4648.

672

Rosendahl, I., Siemens, J., Kindler, R., Groeneweg, J., Zimmermann, J., Czerwinski, S.,

673

Lamshöft, M., Wilke, B.M., Vereeken, H., Amelung, W. (2012). Persistence of the

674

Fluoroquinolone Antibiotic Difloxacin in Soil and Lacking Effects on Nitrogen

675

Turnover. J. Environ. Qual., 41, 1275-1283. http://doi.org/10.2134/jeq2011.0459

676

Salvia, M.-V., Experton, J., Geandel, C., Cren-Olivé, C., Vulliet, E. (2014). Fate of

677

pharmaceutical compounds and steroid hormones in soil: study of transfer and

678

degradation in soil columns, Environ. Sci. Pollut. Res., 21, 10525–10535.

679

http://doi.org/10.1007/s11356-014-3038-x

680

Sarmah, A. K., Meyer, M. T., Boxall, A.B. (2006). A global perspective on the use, sales,

681

exposure pathways, occurrence, fate and effects of veterinary antibiotics (VAs ) in the

682

environment. Chemosphere, 65, 725–759.

683

http://doi.org/10.1016/j.chemosphere.2006.03.026

684

Srinivasan, P., Sarmah, A.K. (2014). Assessing the sorption and leaching behaviour of three

685

sulfonamides in pasture soils through batch and column studies. Sci. Total Environ., 493,

686

535–543. http://doi.org/10.1016/j.scitotenv.2014.06.034

687

Sukul, P., Lamshöft, M., Zühlke, S., Spiteller, M. (2008). Sorption and desorption of

688

sulfadiazine in soil and soil-manure systems. Chemosphere, 73(8), 1344–1350.

689

http://doi.org/10.1016/j.chemosphere.2008.06.066
690 Thiele-Bruhn, S., Seibicke, T., Schulten, H.R., Leinweber P. (2004). Sorption of sulfonamide pharmaceutical antibiotics on whole soils and particle-size fractions. J. Environ. Qual., 33, 1331-1342.

693 Tolls, J. (2001). Sorption of Veterinary Pharmaceuticals in Soils: A Review. Envir. Sci. Technol., 35(17), 3397-3406.

695 Topp, E., Chapman, R., Devers-Lamrani, M., Hartmann, A., Marti, R., Martin-Laurent, F., Sabourin, L., Scott, A., Sumarah, M. (2013). Accelerated Biodegradation of Veterinary Antibiotics in Agricultural Soil following Long-Term Exposure, and Isolation of a Sulfamethazine-degrading sp. J. Environ. Qual., 42(1), 173-178.
699 http://doi.org/10.2134/jeq2012.0162

700 Vasudevan, D., Bruland, G.L., Torrance, B.S., Upchurch, V.G., MacKay, A.A. (2009). pH-dependent ciprofloxacin sorption to soils: Interaction mechanisms and soil factors influencing sorption. Geoderma, 151, 68-76.
703 http://doi.org/10.1016/j.geoderma.2009.03.007

704 Wetzstein, H.G., Stadler, M., Tichy, H.V., Dalhoff, A., Karl, W. (1999). Degradation of Ciprofloxacin by Basidiomycetes and Identification of Metabolites Generated by the Brown Rot Fungus *Gloeophyllum striatum*. Appl. Environ. Microbiol., 65(4), 1556-1563.

708 Zarfl, C., Klasmeier, J., Matthies, M. (2009). A conceptual model describing the fate of sulfadiazine and its metabolites observed in manure-amended soils. Chemosphere, 77(6), 720-726. http://doi.org/10.1016/j.chemosphere.2009.08.035

711 Zhang, Y., Hu, S., Zhang, H., Shen, G., Yuan, Z., & Zhang, W. (2017). Degradation kinetics and mechanism of sulfadiazine and sulfamethoxazole in an agricultural soil system with manure application. Science of the Total Environment., 607-608, 1348-1356.
714 http://doi.org/10.1016/j.scitotenv.2017.07.08
690 Thiele-Bruhn, S., Seibicke, T., Schulten, H.R., Leinweber P. (2004). Sorption of sulfonamide pharmaceutical antibiotics on whole soils and particle-size fractions. J. Environ. Qual., 33, 1331-1342.

693 Tolls, J. (2001). Sorption of Veterinary Pharmaceuticals in Soils: A Review. Envir. Sci. Technol., 35(17), 3397-3406.

695 Topp, E., Chapman, R., Devers-Lamrani, M., Hartmann, A., Marti, R., Martin-Laurent, F., Sabourin, L., Scott, A., Sumarah, M. (2013). Accelerated Biodegradation of Veterinary Antibiotics in Agricultural Soil following Long-Term Exposure, and Isolation of a Sulfamethazine-degrading sp. J. Environ. Qual., 42(1), 173-178.
699 http://doi.org/10.2134/jeq2012.0162

700 Vasudevan, D., Bruland, G.L., Torrance, B.S., Upchurch, V.G., MacKay, A.A. (2009). pH-dependent ciprofloxacin sorption to soils: Interaction mechanisms and soil factors influencing sorption. Geoderma, 151, 68-76.
703 http://doi.org/10.1016/j.geoderma.2009.03.007

704 Wetzstein, H.G., Stadler, M., Tichy, H.V., Dalhoff, A., Karl, W. (1999). Degradation of Ciprofloxacin by Basidiomycetes and Identification of Metabolites Generated by the Brown Rot Fungus *Gloeophyllum striatum*. Appl. Environ. Microbiol., 65(4), 1556-1563.

708 Zarfl, C., Klasmeier, J., Matthies, M. (2009). A conceptual model describing the fate of sulfadiazine and its metabolites observed in manure-amended soils. Chemosphere, 77(6), 720-726. http://doi.org/10.1016/j.chemosphere.2009.08.035

711 Zhang, Y., Hu, S., Zhang, H., Shen, G., Yuan, Z., & Zhang, W. (2017). Degradation kinetics and mechanism of sulfadiazine and sulfamethoxazole in an agricultural soil system with manure application. Science of the Total Environment., 607-608, 1348-1356.
714 http://doi.org/10.1016/j.scitotenv.2017.07.08
715

716

717
715

716

717
## Appendix A. Supplementary data

<table><caption>**Fig. S1.** Molecular structure, pKa and labelling of sulfamethoxazole, N-acetyl-sulfamethoxazole and ciprofloxacin (* $^{14}$C labelling)</caption><thead><tr><th>Compounds</th><th>Molecular structure</th><th>Molecular speciation</th></tr></thead><tbody><tr><td>Sulfamethoxazole (SMX)</td><td><img src="sulfamethoxazole_molecular_structure.png" alt="Sulfamethoxazole molecular structure"></td><td>pKa 5.6<br><img src="sulfamethoxazole_molecular_speciation.png" alt="Sulfamethoxazole molecular speciation"></td></tr><tr><td>N-acetyl-Sulfamethoxazole (N-ac-SMX)</td><td><img src="n-acetyl_sulfamethoxazole_molecular_structure.png" alt="N-acetyl-Sulfamethoxazole molecular structure"></td><td>pKa 5.6<br><img src="n-acetyl_sulfamethoxazole_molecular_speciation.png" alt="N-acetyl-Sulfamethoxazole molecular speciation"></td></tr><tr><td>Ciprofloxacin (CIP)</td><td><img src="ciprofloxacin_molecular_structure.png" alt="Ciprofloxacin molecular structure"></td><td>pKa 6.2<br><img src="ciprofloxacin_molecular_speciation.png" alt="Ciprofloxacin molecular speciation"></td></tr></tbody></table>
## Appendix A. Supplementary data

<table><caption>**Fig. S1.** Molecular structure, pKa and labelling of sulfamethoxazole, N-acetyl-sulfamethoxazole and ciprofloxacin (* $^{14}$C labelling)</caption><thead><tr><th>Compounds</th><th>Molecular structure</th><th>Molecular speciation</th></tr></thead><tbody><tr><td>Sulfamethoxazole (SMX)</td><td><img src="smx_molecular_structure.png" alt="Sulfamethoxazole (SMX) molecular structure"></td><td>pKa 5.6<br><img src="smx_molecular_speciation.png" alt="Sulfamethoxazole (SMX) molecular speciation"></td></tr><tr><td>N-acetyl-Sulfamethoxazole (N-ac-SMX)</td><td><img src="nacSMX_molecular_structure.png" alt="N-acetyl-Sulfamethoxazole (N-ac-SMX) molecular structure"></td><td>pKa 5.6<br><img src="nacSMX_molecular_speciation.png" alt="N-acetyl-Sulfamethoxazole (N-ac-SMX) molecular speciation"></td></tr><tr><td>Ciprofloxacin (CIP)</td><td><img src="cip_molecular_structure.png" alt="Ciprofloxacin (CIP) molecular structure"></td><td>pKa 6.2<br><img src="cip_molecular_speciation.png" alt="Ciprofloxacin (CIP) molecular speciation"></td></tr></tbody></table>
<table><caption>**Table S1**</caption><thead><tr><th rowspan="2">Molecule</th><th rowspan="2">Soil</th><th rowspan="2">Day</th><th colspan="2">HPCD extractions</th><th colspan="2">MeCN pH9 extractions</th><th colspan="2">MeCN pH2 upper phase extractions</th></tr><tr><th>Concentration</th><th>Pool</th><th>Concentration</th><th>Pool</th><th>Concentration</th><th>Pool</th></tr></thead><tbody><tr><td rowspan="2">N-ac-SMX</td><td>CONT</td><td rowspan="2">3</td><td rowspan="2">SPE cartridge strata XA type</td><td rowspan="2">no</td><td rowspan="2">Evaporation</td><td rowspan="2">no</td><td rowspan="2">Evaporation</td><td rowspan="2">yes</td></tr><tr><td>SGW FYM</td></tr><tr><td rowspan="2">N-ac-SMX</td><td>CONT</td><td rowspan="2">156</td><td rowspan="2"></td><td rowspan="2"></td><td rowspan="2">Evaporation</td><td rowspan="2">no</td><td rowspan="2">Evaporation</td><td rowspan="2">yes</td></tr><tr><td>SGW FYM</td></tr><tr><td rowspan="2">SMX</td><td>CONT</td><td rowspan="2">3</td><td rowspan="2">SPE cartridge strata XA type</td><td rowspan="2">no</td><td rowspan="2">Evaporation</td><td rowspan="2">no</td><td rowspan="2">Evaporation</td><td rowspan="2">yes</td></tr><tr><td>SGW FYM</td></tr><tr><td rowspan="2">SMX</td><td>CONT</td><td rowspan="2">156</td><td rowspan="2"></td><td rowspan="2"></td><td rowspan="2">Evaporation</td><td rowspan="2">no</td><td rowspan="2">Evaporation</td><td rowspan="2">yes</td></tr><tr><td>SGW FYM</td></tr></tbody></table>

**Table S2.** Ratio between percentages of ¹⁴C-SMX and ¹⁴C- N-ac-SMX in the mineralized, HPCD, MeCN and non-extractable fractions at day 3 and day 156.

* means that the percentages of radioactivity in the considered fraction are significantly different for SMX and N-ac-SMX.

<table><thead><tr><th rowspan="3"></th><th colspan="8">SMX/N-ac-SMX</th></tr><tr><th colspan="2">Mineralized fraction</th><th colspan="2">HPCD extractable fraction</th><th colspan="2">MeCN extractable fraction</th><th colspan="2">Non extractable fraction</th></tr><tr><th>d3</th><th>d156</th><th>d3</th><th>d156</th><th>d3</th><th>d156</th><th>d3</th><th>d156</th></tr></thead><tbody><tr><td>CONT</td><td>0.6 *</td><td>0.8 *</td><td>0.9 *</td><td>0.9</td><td>0.7 *</td><td>0.9</td><td>1.4 *</td><td>1.1 *</td></tr><tr><td>SGW</td><td>0.5 *</td><td>0.7 *</td><td>0.8 *</td><td>1.1 *</td><td>0.7 *</td><td>1.0</td><td>1.4 *</td><td>1.0</td></tr><tr><td>FYM</td><td>0.6 *</td><td>0.9</td><td>0.7 *</td><td>1.1 *</td><td>0.6 *</td><td>1.1</td><td>1.7 *</td><td>1.0</td></tr></tbody></table>

Selected extracts for qualitative analyses of radioactivity by HPLC
<table><caption>**Table S1**</caption><thead><tr><th rowspan="2">Molecule</th><th rowspan="2">Soil</th><th rowspan="2">Day</th><th colspan="2">HPCD extractions</th><th colspan="2">MeCN pH9 extractions</th><th colspan="2">MeCN pH2 upper phase extractions</th></tr><tr><th>Concentration</th><th>Pool</th><th>Concentration</th><th>Pool</th><th>Concentration</th><th>Pool</th></tr></thead><tbody><tr><td rowspan="2">N-ac-SMX</td><td>CONT</td><td rowspan="2">3</td><td rowspan="2">SPE cartridge strata XA type</td><td rowspan="2">no</td><td rowspan="2">Evaporation</td><td rowspan="2">no</td><td rowspan="2">Evaporation</td><td rowspan="2">yes</td></tr><tr><td>SGW FYM</td></tr><tr><td rowspan="2">N-ac-SMX</td><td>CONT</td><td rowspan="2">156</td><td rowspan="2"></td><td rowspan="2"></td><td rowspan="2">Evaporation</td><td rowspan="2">no</td><td rowspan="2">Evaporation</td><td rowspan="2">yes</td></tr><tr><td>SGW FYM</td></tr><tr><td rowspan="2">SMX</td><td>CONT</td><td rowspan="2">3</td><td rowspan="2">SPE cartridge strata XA type</td><td rowspan="2">no</td><td rowspan="2">Evaporation</td><td rowspan="2">no</td><td rowspan="2">Evaporation</td><td rowspan="2">yes</td></tr><tr><td>SGW FYM</td></tr><tr><td rowspan="2">SMX</td><td>CONT</td><td rowspan="2">156</td><td rowspan="2"></td><td rowspan="2"></td><td rowspan="2">Evaporation</td><td rowspan="2">no</td><td rowspan="2">Evaporation</td><td rowspan="2">yes</td></tr><tr><td>SGW FYM</td></tr></tbody></table>

**Table S2.** Ratio between percentages of <sup>14</sup>C-SMX and <sup>14</sup>C- N-ac-SMX in the mineralized, HPCD, MeCN and non-extractable fractions at day 3 and day 156.

* means that the percentages of radioactivity in the considered fraction are significantly different for SMX and N-ac-SMX.

<table><thead><tr><th rowspan="3"></th><th colspan="8">SMX/N-ac-SMX</th></tr><tr><th colspan="2">Mineralized fraction</th><th colspan="2">HPCD extractable fraction</th><th colspan="2">MeCN extractable fraction</th><th colspan="2">Non extractable fraction</th></tr><tr><th>d3</th><th>d156</th><th>d3</th><th>d156</th><th>d3</th><th>d156</th><th>d3</th><th>d156</th></tr></thead><tbody><tr><td>CONT</td><td>0.6 *</td><td>0.8 *</td><td>0.9 *</td><td>0.9</td><td>0.7 *</td><td>0.9</td><td>1.4 *</td><td>1.1 *</td></tr><tr><td>SGW</td><td>0.5 *</td><td>0.7 *</td><td>0.8 *</td><td>1.1 *</td><td>0.7 *</td><td>1.0</td><td>1.4 *</td><td>1.0</td></tr><tr><td>FYM</td><td>0.6 *</td><td>0.9</td><td>0.7 *</td><td>1.1 *</td><td>0.6 *</td><td>1.1</td><td>1.7 *</td><td>1.0</td></tr></tbody></table>

Selected extracts for qualitative analyses of radioactivity by HPLC