Occurrence of pharmaceutically active compounds, parabens, and their main metabolites in soils amended with sludge and compost
, ... Abstract
The presence of pharmaceuticals and personal care products (PPCPs) in the environment has aroused considerable interest in the last few years. However, the occurrence and fate of parent compounds and their metabolites and/or transformation products in soils have been scarcely evaluated to date. In this work, the dissipation kinetics of PPCPs, which are most frequently detected in the environment, and their main metabolites, is evaluated in soil amended with digested and composted sludge. The studied compounds were monitored for 60 days after digested or compost application to the soil. Several kinetic models were used to evaluate the dissipation kinetics of the processes. Parabens and their metabolites were the compounds with the fastest dissipation rates, while the antiepileptic carbamazepine and its metabolites were the poorest degraded compounds studied. Most of the compounds showed a single first-order dissipation kinetics. The application of biphasic kinetic models can improve the knowledge about the dissipation behavior of some of them. For instance, whereas compounds such as carbamazepine showed a high persistence showing a lag phase in its dissipation, resulting in an approximately constant concentration for the first days of batch experiments, others, such as sulfamethoxazole and diclofenac, followed a dissipation kinetics in two phases: a fast dissipation attributed to the amount of compound associated to the soil-water solution and a slow dissipation that could be due to the amount of the compound adsorbed onto the soil particle. For most of the compounds, the dissipation was faster in sludge-amended soil than in soil without organic amendment. This fact could be due to the influence of microbial activity and organic matter on their dissipation.
Keywords
INTRODUCTION
Currently, one of the main destinations of sewage sludge is its valorization as organic amendment in agricultural soils. Such practice allows the improvement of biological activity and physical properties of the soil[1]. More than 50% of the approximately 15 million tons of sludge generated in the European Union in 2021[2] were applied to soils[3]. This practice is the main sludge disposal option in countries such as Cyprus, Denmark, France, Ireland, and Spain, and, to a lesser extent, in countries such as Finland or Sweden[4]. However, sewage sludge tends to concentrate persistent organic pollutants present in wastewater, which could be accumulated into soils by sludge amendment. As a result, inorganic and organic contaminants such as metals[5,6], plasticizers[7], surfactants, and other industrial pollutants[8,9], and pharmaceuticals and personal care products (PPCPs)[10] have been reported in soils at concentrations in the range from a few nanograms to milligrams per kilogram. Among them, PPCPs have arisen considerable interest not only because they are continuously being discharged into the environment but also because they have been designed to cause effects in living organisms.
Moreover, in the case of pharmaceutically active compounds, they have been designed to be nonbioaccumulative and easily removed after their administration[11]. As a result, these compounds can be accumulated in sewage sludge and on sludge-amended soils. They can be present not only as parent compounds but also as their metabolites that, in some cases, can be discharged into the environment at higher concentrations than their parent compounds[10]. However, although the presence of these compounds has been previously reported in soils[2,12,13], studies about their occurrence and distribution in the terrestrial environment are scarce[10,11,14,15]. The therapeutic groups of pharmaceutically active compounds most frequently detected, and the ones at the highest concentrations, are antibiotics, anti-inflammatories, and antiepileptics[2,14], whereas methylparaben (MeP) and propylparaben (PrP) are the most frequently detected personal care products. Only a few studies have evaluated the presence of metabolites of pharmaceutically active compounds and parabens in soils[14,16,17]. Among them, the most studied metabolites are N4-acetylsulfamethoxazole (Ac-SMZ), acetylsulfapiridine[16], carboxy-ibuprofen (CBX-IBU), 4-hydroxy-diclofenac (4-OH-DIC), and 10,11-dihydro-10-hydroxycarbamazepine (10-OH-CBZ). However, these studies were mainly focused on the evaluation of sorption-desorption processes[10,11,18-21].
The behavior and degradation of PPCPs and their metabolites in the soil depend not only on the physicochemical characteristics of the compound but also on the characteristics of the soil[10,22,23]. For example, Kodešová et al. reported that soil texture, organic matter content, and biological activity are key factors affecting the half-life of pharmaceutically active compounds[22]. Biel-Maeso et al. described a faster degradation of some pharmaceuticals such as ibuprofen, diclofenac, or gemfibrozil under aerobic conditions and pointed out the importance of microbial activity in their degradation process[18]. In 2021, Malvar et al. studied the distribution of PPCPs and their main metabolites, most frequently detected in the environment, in different types of Mediterranean soils. Their results revealed the influence of soil characteristics on the removal of the studied compounds and the high persistence of carbamazepine (CBZ) and its metabolites among the evaluated compounds[10]. However, although these studies have revealed the occurrence of some PPCPs in soil, only a few of them have addressed their dissipation kinetics in soil and, in most cases, a single first-order (SFO) kinetic model has been applied[10,11,15,18,22]. However, although SFO enables a first approximation to the kinetics of the process, the use of other kinetic models allows evaluation of other issues related to the heterogeneous nature of the soil, such as biphasic kinetics, or lag period, in which the compounds follow slow degradation for the first period followed by a fast kinetics. To the best of our knowledge, until now, only two studies have assessed different dissipation kinetic models of pharmaceuticals in soil using different models, and in these studies, only the kinetics of sulfamethoxazole[24] and metformin[25] were evaluated.
This work aimed to evaluate the influence of sludge or compost amendment on the dissipation of PPCPs in amended soils and to evaluate the kinetics of the process. The study was carried out using Mediterranean soil amended with anaerobically-digested and dehydrated sludge or with composted sludge. The studied compounds were selected considering their high consumption, frequency of detection and concentrations measured in environmental samples, their ecotoxicity effects and persistence. Moreover, the metabolites with the highest excretion rate were selected for each parent compound. According to these criteria, the selected compounds were the anti-inflammatories ibuprofen (IBU) and diclofenac (DIC) and their metabolites 1-hydroxyibuprofen (1-OH-IBU), 2-hydroxyibuprofen (2-OH-IBU), CBX-IBU and 4-OH-DIC; the antiepileptic carbamazepine (CBZ) and its metabolites 3-hydroxycarbamazepine (3-OH-CBZ), 10-OH-CBZ, and carbamazepine-10,11-epoxide (EP-CBZ); the antibiotic sulfamethoxazole (SMX) and its metabolite AcSMX; the parabens MeP and PrP and their metabolites methyl protocatechuate (OH-MeP), 3,4-dihydroxybenzoic acid (3,4-DHB), and 4-hydroxybenzoic acid (4-HB).
METHODS
Soil, sludge, and compost collection
The studies were carried out with Cambisol-type Mediterranean soil. It was selected from the results of a previous study in which the degradation rates of the target compounds in alluvial, terra rosa, and Cambisol soils were evaluated[10]. The poorest degradation rates were obtained in Cambisol soil, and because of that, it was selected for the present study. Cambisol type is a sandy loam soil (65.6%-18.4%-16.0%, sand-silt clay) widely extended in several European countries such as France, Germany, Greek, Italy, and Spain. Its physicochemical properties are shown in Supplementary Table 1. Five kilograms of soil were collected from the surface (approximately from 0 to 20 cm) of agricultural land sited in Seville (south of Spain).
Anaerobically-digested and dehydrated sludge was collected from a wastewater treatment plant sited in Seville City. Two kilograms of sludge were collected in an amber glass bottle after the centrifugation process.
Compost was collected from a composting plant sited in Seville, where anaerobically-digested and dehydrated sludge produced in Seville City is composted. The composting process consists of a natural process using dynamic batteries thermally controlled in which aeration is facilitated by turning. Two kilograms of compost were obtained by mixing several aliquots collected from different points of the compost pile, approximately 30 to 50 cm in depth.
All samples were homogenized, sieved (particle size < 2 mm), and stored (-18 ºC) until batch experiments were carried out.
Batch experiments
Two batch experiments were carried out by amending the soil with (i) anaerobically digested and dehydrated sludge and (ii) composted sludge. Although target compounds have been measured in digested and in composted sludge[26], both types of sludge were spiked with the compounds to ensure their presence in the amended soil. Sludge (or compost) was mixed with soil in a 1:99 proportion, according to the reported agricultural application rate in Andalucía Region (south of Spain) from 2020 to 2022 (rate from 0.01% to 0.61% (w/w) (Consejería de Agricultura, Pesca, Agua y Desarrollo Rural from the Andalucía government)[27].
Batch experiments were carried out according to OECD guideline 307[28] as described by Malvar et al.[10]. Sludge and compost were spiked at 50 mg·kg-1 of each compound to achieve an initial concentration of 500 µg·kg-1 in the sludge-amended soil. For this purpose, 50 mL of a standard solution containing the target compounds at 100 mg·L-1 was added to 100 g of sludge (or compost). The mixture was shaken for 2 h in a rotator shaker (LLG-uniLOOPMIX2) at 40 rpm. After that, they were stored at 4 ºC for 24 h. Then, the spiked soils were frozen, lyophilized (Cryodos-50 lyophilizer, Telstar, Terrasa, Spain), and sieved with a 2 mm mesh side. Fifteen grams of the sludge or compost were added to 1,485 g of the soil. The mixture was homogenized and distributed into five glass reactors (300 g of the mixture in each reactor). Then, deionized water was added to achieve a 20% v/w of soil humidity. Two reactors containing non-spiked sludge or compost were used as blanks. Batch experiments were done in a climatic chamber (Selecta, Barcelona, Spain) under the following climatic conditions: 33% of relative humidity, 20% of soil humidity, and an irradiation level of 1,091 W·m-2. The loss of humidity was corrected by adding the proper amount of deionized water.
Target compounds in each reactor were monitored for 60 days with the following sampling frequency: four weekly samples were collected during the first four weeks, three weekly samples were collected during the 5th and 6th weeks, and two weekly samples were collected during the 7th and 8th weeks. The last sample was collected on the 60th day. On each sampling day, 2 g of sample was collected from each reactor. Collected samples were frozen, lyophilized, and stored at -18 ºC until analysis.
Target compound analysis
Analysis of the target compounds was carried out according to a previously developed and validated analytical method[29]. The analytical method was based on selective pressurized liquid extraction and determination by liquid chromatography-tandem mass spectrometry. A brief description of the method can be found in the supplementary material. More details about the analytical method can be found in Malvar et al. and Supplementary Table 2[29].
Data analysis
Dissipation kinetic models
The evaluation of the dissipation kinetics was carried out following the FOCUS guidance document on estimating persistence and degradation kinetics from environmental fate studies on pesticides[30], and previously reported studies[24,31]. Several kinetic models were used to evaluate the kinetics of the dissipation processes and to estimate the endpoints for the studied compounds (DT50 and DT90, corresponding to the dissipation of the 50% and 90%, respectively, of the initial concentrations). Six kinetic models were used: a single first-order (SFO) kinetic model; three biphasic kinetic models: a bi-exponential model (BEM), a first-order two-compartment (FOTC) model, and a first-order double-exponential decay (FODED) model; and two models suitable to describe the dissipation pattern with a lag phase: a modified hockey-stick model (MHSM), and a logistic model (LM).
Single first-order kinetic model
This model assumes that the number of molecules in the soil is small in relation to the number of active sites in the soil. In this case, the dissipation of a compound is proportional to the remaining concentration. The equation for this model is:
where t corresponds to time (days), k1 is the first-order rate constant (day-1), C0 corresponds to the initial concentration (µg·kg-1), and C corresponds to the concentration at time t (day) (µg·kg-1).
Bi-exponential model
Bi-exponential model (BEM) considers that the degradation occurs in two compartments[24]. The first compartment corresponds to the soil-water phase, where the contaminants are available for microorganisms. In this compartment, a fast dissipation is expected. The second compartment corresponds to a phase where the contaminants are adsorbed to soil particles and the dissipation is controlled by sorption-desorption processes. In this model, the dissipation rate can be represented by the sum of first- and second-order differential equations (2):
This equation can be integrated as follow:
As a result:
where k1 is the first-order rate constant (day-1), k2 is the second-order rate constant (kg·µg-1·day-1), t is time, and C0 and C are the concentrations (µg·kg-1) at the beginning of the experiment and at time t, respectively. The equation is reduced to a single first-order equation when k2 = 0.
First-order double-exponential decay model
As in BEM model, First-order double-exponential decay (FODED) model considers the dissipation of the compound into two compartments, a solution phase, and a sorbed phase, where the compounds are adsorbed onto soil particles. However, FODED model considers first-order kinetics for both processes. For this model, the equation is (5):
where t is time (day); Csol and Csorb are the concentrations of the compounds initially distributed into the solution phase and adsorbed onto soil particles, respectively; k1 (day-1) and k2 (day-1) are the dissipation rate constants for each compartment. Usually, k1 is higher than k2, indicating a higher degradation rate of the compound in the solution compartment than adsorbed onto soil. Moreover, when k1 is equal to k2, or one of the concentrations (Csol or Csorb) is zero, the model is mathematically transformed into SFO model.
First-order two-compartment model
This model considers the dissipation of the compounds in a complex system as a heterogeneous soil. This dissipation takes place in a two-compartment system[32]. In one compartment, fast degradation occurs because of physical, photochemical and other processes. In the other compartment, the degradation is slow and mainly due to microbial, chemical or enzymatic processes[24,25,31]. First-order two-compartment (FOTC) model includes a component of transference between the two compartments not considered in BEM and FODED models. The equations can be written as follows:
where t is the time, kr (day-1) is the transfer rate constant of the compound between both compartments; k1 (day-1) and k2 (day-1) are the rate constants for the fast and slow degradation compartments, respectively; and C1 and C2, are the concentrations in each compartment. The total concentration at time t is the sum of C1 and C2. Therefore:
Modified hockey-stick model
The concentrations of organic compounds could be virtually constant (or with a very poor dissipation rate) for a period of time, followed by a dissipation process governed by SFO or biphasic kinetics[30]. The initial phase is referred to as the lag phase and could be caused by several facts such as a slow adaptation of degrading microorganisms, the inhibition of the degrading microflora at high concentrations of contaminants, or a high persistence of the compound, among others. Modified hockey-stick model (MHSM) considers a constant concentration of the compound (or a very slow decrease) up to the breakpoint (time = tb) followed by a SFO kinetic model. This model can be described by the equation (9):
where k (day-1) is the rate constant, tb is the time of lag phase, and C and C0 are the concentrations at time t and the initial concentration (µg·kg-1), respectively.
Logistic model
Logistic model (LM) considers that the degradation rate constant increases after the application of sludge or compost to a maximum value. Once the kinetic constant has reached its maximum value, the kinetic approach is first order[30]. The equation is as follows:
where a0 is the kinetic constant at the beginning of the batch experiment, amax is the maximum value of the kinetic constant, r is the microbial growth rate, t is time, and C and C0 are the concentrations at time t and at t = 0. For this model, the higher the r or a0 value, the lower the lag phase of the compound.
Evaluation of data fitting to the kinetic models
For each compound, the parameters implied in each of the studied models (C0, k1 in the case of SFO; C0, k1, and k2 for BEM; k1, k2, Csorb, and Csol for FODED; C0, k1, k2, and kr for FOTC; C0, k, and tb for MHSM, and C0, amax, a0, and r for the case of LM) were optimized using Microsoft Excel-Solver tool. The values of the model parameters providing the best fit to the experimental data were optimized by minimizing the mean of the square error, that is, the sum of the square of the residuals (SSR) divided by the number of experimental data.
Following the FOCUS guidance document[30], the fit of each model to the experimental data was first evaluated through visual assessment. For this evaluation, the experimental mean concentrations and the calculated curves were represented vs. time. In addition, experimental concentrations were plotted vs. the concentrations obtained by the models. The goodness fit of the model was evaluated considering the coefficient of determination (R2) and the slope (close to 1) of the linear curve.
DT50 and DT90 determination
For each applied model, DT50 and DT90 values were determined. In the case of the SFO model, DT50 and DT90 were determined by applying the following equations[30]:
There is no analytical equation for the estimation of DT50 and DT90 values in biphasic models. Consequently, these values should be determined using an interactive procedure that was carried out using the Goal Seek function from Excel.
RESULTS
Concentration of target compounds in spiked sludge and compost applied to the soil
Batch experiments were carried out using spiked anaerobically-digested and dehydrated sludge and compost. Previously to the spiking procedure, soil, sludge, and compost samples were analyzed
Concentrations and relative standard deviations (RSD) of the target compounds in spiked sludge-amended soil and compost-amended soil (n = 5)
| Spiked sludge-amended soil | Spiked compost-amended soil | |||||
| Range (µg·kg-1) | Mean (µg·kg-1) | RSD (%) | Range (µg·kg-1) | Mean (µg·kg-1) | RSD (%) | |
| MeP | 190-233 | 204 | 9.0 | 297-376 | 297 | 25 |
| OH-MeP | 264-305 | 282 | 6.3 | 133-386 | 281 | 34 |
| PrP | 278-337 | 304 | 7.9 | 266-515 | 352 | 27 |
| 3,4-DHB[1] | 114-135 | 123 | 6.6 | 89.1-123 | 107 | 44 |
| 4-HB[1] | 513-533 | 524 | 1.6 | 24.7-55.5 | 44.6 | 16 |
| CBZ | 379-449 | 422 | 6.8 | 342-728 | 575 | 25 |
| 3-OH-CBZ | 423-495 | 464 | 6.3 | 358-732 | 565 | 24 |
| 10-OH-CBZ | 381-484 | 438 | 8.9 | 315-647 | 516 | 24 |
| EP-CBZ | 291-347 | 325 | 9.2 | 274-473 | 374 | 26 |
| DIC | 250-275 | 260 | 3.9 | 221-419 | 311 | 24 |
| 4-OH-DIC | 209-296 | 244 | 13 | 122-356 | 194 | 14 |
| IBU | 250-292 | 270 | 5.6 | 199-356 | 308 | 21 |
| 1-OH-IBU | 373-453 | 424 | 7.6 | 294-640 | 537 | 26 |
| 2-OH-IBU | 303-364 | 341 | 6.8 | 248-523 | 443 | 25 |
| CBX-IBU | 318-376 | 356 | 6.6 | 267-540 | 462 | 24 |
| SMZ | 523-876 | 764 | 21 | 436-584 | 529 | 15 |
| AcSMZ | 373-452 | 416 | 7.3 | 310-637 | 496 | 24 |
Mean concentrations were in the ranges 123-764 µg·kg-1 and 44.5-575 µg·kg-1 in sludge-amended soil (SAS) and compost-amended soil (CAS), respectively. These concentrations correspond to a percentage of the spiked concentration from 24.6% to 153% in the case of SAS, and from 8.4% to 115% for CAS. For most of the target compounds, these percentages were in the range between 41% and 153% in the case of SAS, and from 56% to 115% for CAS. The concentrations of the metabolites of parabens and 4-OH-DIC in amended soils were even lower, from 9% to 39% of the spike concentration. This fact could be explained not only by their low adsorption on the solid, as was reported previously[20], but also by their high degradation in the soil[2,10]. However, the concentrations measured for all target compounds at the beginning of batch experiments were high enough to evaluate their degradation in sludge-amended soils.
Degradation of PPCPs and their metabolites in sludge-amended soil
Dissipation of SMZ and AcSMZ
The concentrations of SMZ and AcSMZ obtained in batch experiments are shown in Supplementary Figures 1 and 2, respectively. Optimized parameters of the kinetic models applied to the degradation of the studied compounds in SAS and CAS are shown in Supplementary Tables 4 and 5, respectively. The concentrations of both compounds decreased quickly during the first 7 days of the batch experiment to concentrations lower than 50% of the initial concentrations. Considering the dissipation kinetics of these compounds [Figure 1], all models applied were transformed to a SFO model. In the case of BEM model, because of the low k2 values; in the case of FODED model, because k1 = k2 or low value of k2; in the case of FOTC model, k1 = k2; in the case of MHSM model because of low tb value and; in the case of LM model
Figure 1. Mean concentrations of SMZ and AcSMZ as a function of time in (A) sludge-amended soil and (B) compost-amended soil with curves from kinetic models. AcSMZ: Acetylsulfamethoxazole; BEM: bi-exponential model; FODED: first-order double-exponential decay; FOTC: first-order two-compartment; SFO: a single first-order; SMZ: sulfamethoxazole.
Determination coefficients were calculated to evaluate the fit of the experimental concentrations to SFO model. They were in the range from 0.941 to 0.966 (slope from 0.88 to 1.0) for SMZ and from 0.964 to 0.996 for AcSMZ (slopes from 0.95 to 1.0). The SFO constants determined in SAS were in the range of 0.169 to 0.174 d-1 for SMZ and 0199 to 0.290 d-1 for AcSMZ. The SFO constants determined in CAS were slightly lower in the case of SMZ (from 0.121 to 0.159 d-1) and higher in the case of AcSMZ (0.305 d-1). These values were similar to those previously reported for SMZ. For example, Kodešová et al. evaluated the dissipation of SMZ and other pharmaceutical compounds in thirteen different soil conditions, obtaining kinetic constants for this compound in the range between 0.0045 and 0.149 d-1[22]. Furthermore, the application of FODED model to the dissipation of SMZ in CAS revealed that the compound could follow a biphasic kinetics with two compartments: in the first one, it follows a fast dissipation kinetics with a kinetic constant of 0.159 d-1, while in the second one, where it can be potentially adsorbed onto the soil, could follow a slow dissipation kinetics with a kinetic constant of 0.001 d-1. These results were similar to those observed by Srinivasan et al. (2014), who evaluated the dissipation of SMZ in topsoil and subsoil[24]. In this study, FODED and FOTC models were the best models for fitting the SMO dissipation. Moreover, this fact could explain the concentration of approximately 20 µg·kg-1 measured in SAS after 60 days of batch experiment.
To our knowledge, no data about the kinetic model of AcSMZ in soil has been reported yet. However, previous studies have reported that it suffers a quick dissipation in soil, higher than its parent compound SMZ[23,37,38]. This fact could be explained not only by its higher availability in soil, but also by its transformation by deacetylation into SMZ[37].
Endpoint values obtained for SMZ in SAS were slightly lower than those obtained in CAS. Moreover, these values were lower than in spiked soil without organic amended (SS) [Supplementary Table 6]. The opposite effect was observed for AcSMZ. Its DT50 and DT90 values decreased from SAS (from 3.48 to 5.62 days and from 11.2 to 16.1 days, respectively) to CAS (2.27-3.43 days and 7.55-8.18 days, respectively) and SS (1.20-1.31 days and 3.98-4.34 days, respectively). The behavior of SMZ could be due to the importance of biological processes in its dissipation, and the higher microbial biomass in SAS than in CAS or SS. The behavior of AcSMZ could be explained by greater sorption onto sludge than onto compost and soil, as previously reported[20], decreasing its bioavailability[36].
Degradation of DIC and 4-OH-DIC
The concentrations of DIC were reduced during the first 15 days of batch experiments from the initial concentrations (260 and 310 µg·kg-1 in SAS and CAS, respectively) to 60.0 and 19.9 µg·kg-1, respectively. In the following 45 days, the concentrations measured in SAS remained approximately constant (mean concentration: 40.1 µg·kg-1), while in CAS were reduced below 10 µg·kg-1 [Supplementary Figures 3 and 4]. The concentrations of 4-OH-DIC were reduced in the first 10 days from 244 µg·kg-1 to lower than 3.10 µg·kg-1 in SAS, and from 153 µg·kg-1 to lower than 5.48 µg·kg-1 in CAS.
Considering the dissipation kinetics of DIC [Figure 2], a similar behavior was observed in SAS and CAS. In both cases, BEM, MHSM and LM models were transformed to SFO model (kinetic constants of these models were 0.081 d-1 in SAS, and from 0.160 to 0.167 d-1 in CAS), while in the case of the FODED and FOTC models, the kinetics showed two components: a first component followed a SFO kinetics (kinetic constant of 0.132 and 0.178 d-1, for SAS and CAS) and a second component corresponding to a constant concentration of 30.3 and 5.84 µg·kg-1, respectively. These results could be explained by the distribution of DIC between the solution and the solid phase. In the solution phase, the concentration of DIC follows a SFO model. In the soil phase, its concentration remains constant, which can be explained by its adsorption onto soil particles making its degradation more difficult. This fact is consistent with the measured concentrations and the adsorption behavior of DIC in soil and sludge[39,40]. For example, Silva et al. showed the influence of pH on the adsorption of DIC onto soil[40]. They reported an increase in the ionic form of DIC at higher pH values which could explain its higher availability and, as a result, its higher microbial degradation in CAS (compost pH: 6.33) than in SAS (sludge pH: 5.85). In another study, Yan et al. reported a high correlation between the adsorption of DIC and the organic matter content of sludge[39]. This fact can explain the lower desorption and availability of DIC in sludge [OM: 48.5%, wt (by weight)] than in compost (OM: 27.3%, wt) [Supplementary Table 1].
Figure 2. Mean concentrations of DIC and 4-OH-DIC as a function of time in (A) sludge-amended soil and (B) compost-amended soil and curves of the applied kinetic models. BEM: Bi-exponential model; DIC: Diclofenac; FODED: first-order double-exponential decay; FOTC: first-order two-compartment; SFO: a single first-order; 4-OH-DIC: 4-hydroxy-diclofenac.
All kinetic models applied were well fitting to the dissipation of 4-OH-DIC. Determination coefficients were higher than 0.922 and slopes were higher than 0.91. FOTC, MHSM, and LM models revealed that dissipation corresponded to SFO model with kinetic constants of 0.646 and 0.653 d-1 in SAS and CAS, respectively. BEM and FODED models showed a biphasic kinetics in both matrices [Figure 2].
Regarding the endpoints, DT50 values of DIC were in the ranges 6.29-8.56 days, 3.98-4.33 days, and 5.89-6.66 days for SAS, CAS and SS, respectively. These results are in concordance with those previously reported. For example, Biell-Maeso et al. reported DT50 values of 3 and 4 days in two studied soils[18]. DT90 values were in the ranges from 25.6 to higher than 100 days in SAS, from 13.8 to 14.4 days in CAS, and from 22.1 to 25.1 days in SS. The highest DT90 value was obtained for SAS, which could be explained by the high retention of DIC in sludge resulting in the remaining amount in SAS (30.3 µg·kg-1). In the case of CAS, the measured concentrations and FODED and FOTC models showed the presence of DIC 60 days after batch experiments. These concentrations were lower than 10% of the initial concentrations. As a result, DT90 values were achieved before 15 days of batch experiments (25 days in the case of SS)
Dissipation of CBZ and its metabolites
CBZ and its metabolites showed similar behavior in SAS and CAS [Supplementary Figures 5 and 6]. Concentrations of CBZ and EP-CBZ were approximately constant in SAS during the first 27 and 23 days, respectively, of the batch experiments and during 27 days (both compounds) in CAS. After that, a fast decrease in the concentrations occurred, whereas the concentrations of the hydroxylated metabolites (3-OH-CBZ and 10-OH-CBZ) decreased since the beginning of the batch experiments.
No good fit to kinetic models was observed for the dissipation of CBZ and EP-CBZ [Figure 3], neither in SAS nor in CAS. Determination coefficients were in the ranges 0.530-0.538 and 0.676-718 for CBZ and EP-CBZ, respectively, in SAS and up to 0.685 and 0.648, respectively, in CAS for most of the models applied. This was in concordance with the poor fit to the SFO model reported by Al-Rajab et al.[41]. LM model was the best fitting one to the measured concentrations: determination coefficients were 0.683 and 0.853 for CBZ and EP-CBZ, respectively, (slopes 0.86 and 0.95) in SAS and 0.728 and 0.788 for CBZ and EP-CBZ, respectively, (slopes 1.0 and 1.1) for CAS. These results showed a lag phase for the dissipation of these compounds. In LM model, the rate constant increased from its initial value to a maximum value (0.067 d-1 for CBZ and 0.072 d-1 for EP-CBZ). Then, the diffusion behavior approached a first-order kinetics. The poor fit between measured concentrations and LM model could be explained by considering that, after the lag phase, the dissipation of CBZ and EP-CBZ occurs in two phases. In the first phase, a fast dissipation occurs, possibly in soil-water phase. In the second phase, the compound is poorly degraded, probably due to being adsorbed to soil particles. As a result, concentrations higher than 145 and 51.7 µg·kg-1 were measured for CBZ and EP-CBZ, respectively, in the last 25 days of batch experiments. These results are in concordance with the high persistence of CBZ previously noticed by other authors[15,42]. Despite the high persistence of these compounds in amended soil, the results obtained after the application of sludge and compost to the soil revealed a higher degradation of both compounds in amended soil than in SS[10], in which no dissipation, especially in the case of CBZ, was measured.
Figure 3. Mean concentrations of CBZ and its metabolites as a function of time in (A) sludge-amended soil and (B) compost-amended soil and curves from the applied kinetic models. BEM: Bi-exponential model; CBZ: carbamazepine; EP-CBZ: carbamazepine-10,11-epoxide; FODED: first-order double-exponential decay; FOTC: first-order two-compartment; LM: logistic model; MHSM: modified hockey-stick model; SFO: a single first-order; 3-OH-CBZ: 3-hydroxycarbamazepine; 10-OH-CBZ: 10-hydroxycarbamazepine.
The degradation of the hydroxylated metabolites was faster than that of CBZ and EP-CBZ. The analysis of kinetic models revealed a SFO kinetics for both compounds. All kinetic models were transformed to a SFO model. Only LM and FODED models showed different behavior. In the case of SAS, LM model gave the best fit with R2: 0.919 for 3-OH-CBZ and 0.933 for 10-OH-CBZ. These results show a slow degradation in SAS during the first days of the batch experiments, followed by a first-order kinetics. A faster dissipation was obtained for all compounds in CAS and SAS than in SS, except for 3-OH-CBZ. Mean DT50 and DT90 values in SAS were 22 and 67 days, respectively; 15 and 44 days, respectively, in CAS, and 74 and 281 days, respectively, in SS.
Dissipation of IBU and its metabolites
The concentrations of IBU and its metabolites obtained from batch experiments are shown in Supplementary Figures 7 and 8. The metabolite CBX-IBU showed a higher dissipation with concentrations lower than 5.71 µg·kg-1 ten days after the beginning of the experiments. However, its concentration increased from days 38 to 60. The compounds IBU and 1-OH-IBU showed higher persistence in SAS than in CAS. For example, whereas in SAS, the concentration of IBU decreased from 270 to 123 µg·kg-1 during the first 30 days, in CAS, the concentration of IBU decreased from 307 to 31.9 µg·kg-1 in the first 20 days. The same pattern was observed for 1-OH-IBU. The metabolites 2-OH-IBU and CBX-IBU had similar behavior in both matrices.
A high dispersion of results was obtained for all compounds, especially in SAS [Figure 4]. Because of that, a poor goodness-of-fit was achieved for all kinetic models applied. Their determination coefficients were worst in SAS [from 0.535 to 0.826 for most of the models and compounds, except for CBX-IBU (from 0.290 to 0.469)] than in CAS (from 0.641 to 0.980). The best goodness-of-fit was provided by SFO model (from 0.722 to 0.980). For all studied compounds, except for 2-OH-IBU, the obtained dissipation was slower in SAS (kinetic constant values from 0.048 to 0.119 d-1) than in CAS (0.143-0.229 d-1) and SS
Figure 4. Mean concentrations of IBU and its metabolites as a function of time in (A) sludge-amended soil and (B) compost-amended soil and curves of the applied kinetic models. BEM: Bi-exponential model; CBX-IBU: carboxy-ibuprofen; FODED: first-order double-exponential decay; FOTC: first-order two-compartment; IBU: Ibuprofen; LM: logistic model; MHSM: modified hockey-stick model; SFO: a single first-order; 1-OH-IBU: 1-hydroxyibuprofen; 2-OH-IBU: 2-hydroxyibuprofen.
The lowest endpoint values were obtained for CBX-IBU (mean DT50 and DT90 values up to 7.6 and 23 days, respectively), followed by 1-OH-IBU (up to 11.1 and 34.1 days, respectively), IBU (14.8 and 52 days), and 2-OH-IBU (19.8 and 63 days). The fast dissipation of CBX-IBU could be explained by its lower retention in the organic-amended soil[20] and, consequently, its higher availability. In addition, the greater mobility of the metabolite 1-OH-IBU could explain its faster dissipation in comparison to that of IBU. However, the metabolite 2-OH-IBU, in spite of its physicochemical properties and adsorption behavior onto the sludge and soil are similar to those of 1-OH-IBU[20], showed the slowest dissipation. This fact is in concordance with previous studies which reported a high persistence of 2-OH-IBU[10,43]. The high persistence of this compound could be explained not only by its poor degradation but also by the potential transformation of IBU into 2-OH-IBU under aerobic conditions, as has been previously described[44,45]. This fact could also explain the high variability of the concentrations measured for this compound in batch experiments.
Dissipation of parabens and their metabolites
The concentrations of parabens and their metabolites in SAS and CAS are shown in Supplementary Figures 9 and 10, respectively. These concentrations quickly decreased during the first ten days of batch experiments. In SAS, the concentration of MeP decreased in 2 days from the initial concentration of 204 µg·kg-1 to 2.16 µg·kg-1, whereas the concentration of PrP decreased from 304 µg·kg-1 to 3.45 µg·kg-1 in 3 days. The same behavior was observed in CAS, where the concentrations strongly decreased, except for the metabolite 4-HB, after the first days of batch experiments. These results are in concordance with the low persistence previously reported for parabens[46,47]. For example, Arachchige et al. reported that more than 90% of parabens were dissipated after three days of the beginning of the experiment[46]. The dissipation of the studied compounds was lower in SS than in SAS and CAS, especially in the case of MeP, PrP, and OH-MeP[10].
The results showed a good fit of the kinetics of dissipation to a SFO model for most of the compounds and matrices [Supplementary Figures 5 and 6]. Only the metabolite 4,3-DHB showed a low determination coefficient and poor slope (0.65-0.68) in sludge-amended soil, which could be due to the wide dispersion of concentrations measured in these samples. For OH-MeP in CAS, the best fit was obtained from FODED and FOCT models [Supplementary Table 5], which could show a degradation kinetics in two compartments for this compound in this experiment.
Figure 5. Mean concentrations of MeP and its metabolites measured as a function of time in (A) sludge-amended soil and (B) compost-amended soil and curves of the applied kinetic models. BEM: Bi-exponential model; FODED: first-order double-exponential decay; FOTC: first-order two-compartment; LM: logistic model; MeP: Methylparaben; MHSM: modified hockey-stick model; OH-MeP: methyl protocatechuate; SFO: a single first-order.
Figure 6. Mean concentrations of PrP and its metabolites as a function of time in (A) sludge-amended soil and (B) compost-amended soil and curves of the applied kinetic models BEM: Bi-exponential model; FODED: first-order double-exponential decay; FOTC: first-order two-compartment; LM: logistic model; MHSM: modified hockey-stick model; PrP: Propylparaben. SFO: a single first-order; 3,4-DHB: 3,4-dihydroxybenzoic acid; 4-HB: 4-hydroxybenzoic acid.
In general, after sludge or compost application onto soil, the parent compounds MeP and PrP showed the fastest dissipation (DT50 and DT90 values up to 0.37 and 1.24 days, respectively, in SAS, and up to 0.56 and 2.10 days, respectively, in CAS), while the metabolite OH-MeP showed the lowest dissipation rate (DT90 up to 18.0 days in SAS and 7.60 days in CAS). The dissipation of the parent compounds was considerably faster in amended soil than in SS [Supplementary Tables 4 to 6]. No relevant difference was obtained between the endpoint values of the metabolites. These results could be explained by the potential transformation of parent compounds into their metabolites under aerobic conditions. The degradation of MeP and PrP in amended soil could result in the formation of their metabolites, compensating for the decrease of the concentration of metabolites by dissipation. This compensation is most significant in amended soils than in soils where the dissipation of parent compounds is slower.
CONCLUSIONS
The occurrence and dissipation kinetics of most recurrent PPCPs and their metabolites in the environment have been evaluated in soil amended with anaerobically-digested and dehydrated sludge or with compost. Obtained results were compared with those reported for the same compounds in soil without organic amendment. Batch experiments were carried out using spiked sludge and compost. The adsorption of the studied compounds in sludge and compost (from 8.4% to 153%) was enough to evaluate their occurrence in soil after organic amendment.
In general, parabens and their metabolites were the compounds with the fastest dissipation in all studied matrices, followed by the antibiotic SMZ and its metabolite, DIC and its metabolite 4-OH-DIC, IBU and its metabolites, and CBZ and its metabolites, in that order. Most of them were well fitting to a SFO dissipation kinetic model. The application of FODED model showed that some of the studied compounds, such as DIC or SMZ, could follow a dissipation kinetics in two phases: a fast dissipation that could be attributed to the amount of the compound associated with the aqueous compartment and a slow dissipation that could be due related to the amount of the compound strongly adsorbed onto soil particles. This fact could result in a high persistence of the compound, similar to the case of DIC which remains in the soil 60 days after batch experiments.
The compounds CBZ and EP-CBZ showed a high persistence, with a lag phase on their dissipation, resulting in a concentration approximately constant during the first days of batch experiments.
For most of the compounds, the dissipation kinetics was faster in SAS and CAS than in SS, which could show the influence of microbial activity and organic matter content in their dissipation.
The use of several kinetic models, including SFO and biphasic models, to evaluate the dissipation of PPCPs and their metabolites can clarify the dissipation behavior of some of the studied compounds. For example, while compounds such as parabens and their metabolites followed a SFO model associated with a high degradation 60 days after batch experiments, compounds such as SMZ or DIC were still present in CAS and SAS after that period of time. This fact could be due to their higher adsorption onto the organic matter of sludge and compost.
DECLARATIONS
Authors’ contributionsInvolved in conceptualization: Santos JL, Alonso E
Involved in methodology and verification of results/experiments, application of statistical models to analyze or synthesize study data: Santos JL, Martín J
Involved in presentation and visualization of data: Martín J
Involved in writing - Original Draft: Santos JL
Involved in writing - Review & Editing: Aparicio I
Involved in project administration and funding acquisition: Alonso E
Availability of data and materialsData are available in the Supplementary Material.
Financial support and sponsorshipThis work was financially supported by the Ministerio de Economía y Competitividad and the Ministerio de Ciencia e Innovación from the Spanish Government (Project No. PID2020-117641RB-I00 and CTM2017-82778-R) and by Junta de Andalucía (Consejería de Economía y Conocimiento, Project I+D+i PAIDI Andalucía No. P20_00556).
Conflicts of interestAll authors declared that there is no conflict of interest.
Ethical approval and consent to participateNot applicable.
Consent for publicationNot applicable.
Copyright© The Author(s) 2023.
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Santos JL, Martín J, Aparicio I, Alonso E. Occurrence of pharmaceutically active compounds, parabens, and their main metabolites in soils amended with sludge and compost. Water Emerg Contam Nanoplastics 2023;2:11. http://dx.doi.org/10.20517/wecn.2023.14
AMA Style
Santos JL, Martín J, Aparicio I, Alonso E. Occurrence of pharmaceutically active compounds, parabens, and their main metabolites in soils amended with sludge and compost. Water Emerging Contaminants & Nanoplastics. 2023; 2(3): 11. http://dx.doi.org/10.20517/wecn.2023.14
Chicago/Turabian Style
Santos, Juan Luis, Julia Martín, Irene Aparicio, Esteban Alonso. 2023. "Occurrence of pharmaceutically active compounds, parabens, and their main metabolites in soils amended with sludge and compost" Water Emerging Contaminants & Nanoplastics. 2, no.3: 11. http://dx.doi.org/10.20517/wecn.2023.14
ACS Style
Santos, JL.; Martín J.; Aparicio I.; Alonso E. Occurrence of pharmaceutically active compounds, parabens, and their main metabolites in soils amended with sludge and compost. Water. Emerg. Contam. Nanoplastics. 2023, 2, 11. http://dx.doi.org/10.20517/wecn.2023.14
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