In the current scenario of increasing world population and environmental degradation, the transition to a Circular Economy model requires, among many other aspects, the reuse and sustainable management of wastes and by-products. The application of sewage sludge (SS) to soil is a common agricultural practice that can certainly lead to agronomic improvements such as, for instance: (i) increased soil organic matter (OM) and nutrient content (Lashermes et al., 2009); (ii) enhanced soil porosity and bulk density (Singh and Agrawal, 2007; Annabi et al., 2011); (iii) greater water holding capacity (Bulluck et al., 2002); (iv) higher soil microbial activity (García et al., 1992); and (v) improved overall soil quality (Powlson et al., 2011). The application of organic amendments (e.g., SS) into soil can also result in structural and functional changes in soil bacterial communities (Innerebner et al., 2006; Epelde et al., 2014).

However, the application of SS to agricultural soil as organic amendment can likewise lead to potential risks for human and environmental health, in particular, owing to the presence in SS of: (i) elevated concentrations of potentially toxic metals (Singh and Agrawal, 2007); (ii) organic contaminants (Alvarenga et al., 2015); (iii) nanoparticles (Fijalkowski et al., 2017); (iv) microplastics; and (v) pharmaceutical compounds (Martín et al., 2015), including antibiotics and their transformation products. These later chemical emerging contaminants (antibiotics and their transformation products) are often accompanied in SS by several biological contaminants of great concern, i.e., antibiotic resistance genes (ARGs), antibiotic resistant bacteria (ARB) and mobile genetic elements (MGEs).

Most worryingly, the use, overuse and misuse of antibiotics for medical and veterinary use have promoted the emergence and spread of antibiotic resistance in the environment, including the soil ecosystem (Cytryn, 2013). Antibiotic resistant bacteria can transfer to other bacteria (including potential human pathogens) the ARGs they harbor through horizontal gene transfer (HGT) via a variety of MGEs such as plasmids, transposons, integrative conjugative elements, phages, integrons, genomic islands, etc. (Pruden et al., 2006; Zhu et al., 2013).

In the European Union, in 2014, the consumption of antimicrobials for medical and veterinary use, including animal production, reached a staggering value of 12,720 tons of active substance (European Centre for Disease Prevention and Control [ECDC], European Food Safety Authority [EFSA], and European Medicines Agency [EMA], 2017). It has been reported (Van Boeckel et al., 2017) that, worldwide, more than 73% of all antimicrobials are administered to animals for veterinary or food-producing purposes. Relevantly, a considerable amount (between 30 and 90%) of the antibiotics administered for human or veterinary purposes are excreted in the urine and feces, essentially unchanged or as active metabolites (Sarmah et al., 2006). Regrettably, wastewater treatment plants (WWTPs) are not designed to efficiently remove these emerging contaminants and, then, they are unsurprisingly regarded as hotspot for the emergence and dissemination of antibiotic resistance (Rizzo et al., 2013).

In SS, the concentration of widely use antibiotics ranges from μg kg^–1 to mg kg^–1 (Lillenberg et al., 2009; Nieto et al., 2010). Although there are certainly many differences between countries and specific WWTPs, resulting in the impossibility to describe a common pattern, several authors (Zhang and Li, 2011; Rutgersson et al., 2020) detected lower levels of sulfonamides, macrolides and tetracyclines, compared to quinolones, in SS samples.

In the European Union, the most common method of SS disposal remains its application to agricultural soil, followed by thermal disposal and landfill (Christodoulou and Stamatelatou, 2016). The improper disposal of SS can lead to serious environmental risks such as, for instance, the contamination of aquifers and soils with a variety of potentially toxic inorganic and organic compounds (Sharma et al., 2017). The majority of legislations on SS management, such as Directive 86/278/ECC dealing with soil protection when SS is used in agriculture (European Commision, 1986), do not consider the abovementioned emerging contaminants. Actually, these legislations have traditionally been focused on total metal concentrations, both in the SS itself and in the amended agricultural soil, as well as on the presence of potential human pathogens.

The aim of this study was to assess the impact of the application of thermally-dried anaerobically-digested SS on: (i) soil physicochemical and microbial properties, including soil prokaryotic diversity and composition; and (ii) the presence and relative abundance of ARGs and MGE-genes in soil. In particular, regarding the application of SS, three factors were studied: dose of application, dosage of application, and elapsed time after the last application. The term “dose” refers here to a specific amount of SS applied at one time. Instead, the term “dosage” refers to the total amount of SS applied to those fields (i.e., the sum of all the individual applications). The novelty of the study is supported by the following facts: (1) the study was carried out with samples taken from a high number of real agricultural farms, i.e., 20 farms; (2) the studied agricultural fields differed in the abovementioned three factors, thus allowing the assessment of their individual and combined influence on the main topic under study, i.e., antibiotic resistance; and (3) the SS had been thermally-dried and anaerobically-digested prior to their application which, a priori, could have considerably reduced the risk of antibiotic resistance spread. We hypothesized that the application of SS would enhance soil physicochemical and microbial properties, alter the composition of soil prokaryotic communities and, finally, increase the relative abundance of ARGs and MGE-genes in soil. Moreover, we expected these changes to be dependent upon the three factors mentioned above. This study was carried out with SS from a single WWTP. Nonetheless, SS composition and management (e.g., storage, treatment, dose, and mode of application, etc.) can significantly differ between WWTPs. This fact must be taken into consideration when comparing our results with those from other studies.

The application of organic amendments to agricultural soil is a common practice because of its potential positive effects on soil quality (Epelde et al., 2018; Urra et al., 2019b). Furthermore, this practice reduces the need for energetically costly synthetic fertilizers, while reusing a material that would otherwise be treated as a waste. Relevantly, in the current scenario of climate change, the application of SS to agricultural soils can help increase their soil OM content (a pertinent objective for many agricultural soils with low OM content), thus enhancing their capacity to sequester carbon. Regrettably, the application of organic amendments to agricultural soil, such as SS, can likewise lead to potential risks for human and environmental health, in particular, due to the presence of traditional and emerging contaminants (Singh and Agrawal, 2007; Alvarenga et al., 2015; Martín et al., 2015). In particular, nowadays, emerging contaminants (e.g., pesticides, pharmaceuticals, nanomaterials, personal and house care products, etc.) are a matter of much concern (Epelde et al., 2018). The possible dissemination of ARGs is of special concern owing to the risk of transfer of ARGs to human pathogens (Marshall and Levy, 2011; Epelde et al., 2018).

In this study, SS-amended soils presented higher Olsen P contents than unamended soils, as previously described (Börjesson and Kätterer, 2018; Urra et al., 2019a). The rate and amount of SS application to agricultural soil is frequently based on the N requirement of crops. Then, other nutrients, such as P and K⁺, can be over-applied (Antoniadis et al., 2015). It is a well-known fact that SS from WWTPs can present elevated levels of P (Alvarenga et al., 2015), a fact that must be taken into consideration as this key nutrient is being depleted at an alarming rate due chiefly to its use in fertilizers. In accordance with other studies (Kirchmann et al., 2017; Bogusz and Oleszczuk, 2018; Urra et al., 2019a), in our study, the application of SS also resulted in significantly higher soil Zn concentrations. Sewage sludge application to agricultural soil can lead to accumulation of potentially toxic metals, with concomitant adverse effects for crop quality (Muchuweti et al., 2006). On the other hand, in accordance with previous studies (Lloret et al., 2016; Urra et al., 2019a), the application of SS resulted in reduced soil microbial diversity, with potential negative consequence for soil functioning and resilience (Delgado-Baquerizo et al., 2016). On the contrary, some authors found an increase in diversity in SS-amended soils (Chen et al., 2016; Bai et al., 2019).

The application of SS to soil can result in an increase of microbial biomass and activity, associated with the supply of easily available carbon sources for the soil microbial communities (García-Gil et al., 2004; Sánchez-Monedero et al., 2004; Fernández et al., 2009). The application of organic amendments has been frequently reported to enhance soil microbial activity and biomass (Dinesh et al., 2010; Zhen et al., 2014; Hernández et al., 2016; Reardon and Wuest, 2016; Das et al., 2017; Urra et al., 2019a). In our study, unexpectedly, no differences in microbial biomass and activity values were found between SS-amended and unamended soils. At soil sampling times, our agricultural fields were planted with different crops (barley, beans, rapeseed, wheat). Taking into account that our soil samples were collected from the topsoil (0–10 cm depth), it is possible that the expected stimulation of soil microbial biomass and activity, derived from the rhizodeposition of readily available carbon sources, might have veiled the positive effect of SS on soil microbial biomass and activity. It is also feasible that, at sampling times, most of the easily available C substrates initially present in SS were long gone (elapsed times after the last SS application range from 1 to 4 years).

In our study, all the sampled soils and the SS itself met the EU Directive 86/278/EEC (European Commision, 1986) regarding total metal concentration values. However, the bioavailable fraction of soil metals is considered more relevant from an environmental risk point of view (Olaniran et al., 2013). Bioavailable metal concentrations in soil usually show an inverse relationship with soil pH (Rieuwerts et al., 1998). The agricultural soils sampled for this study have an alkaline pH (mean value: 8.4 ± 0.1), which could explain the fact that bioavailable metal levels were below the quantification limit, suggesting a lack of metal-induced ecotoxicity. Similarly, Urra et al. (2019a) found that the application of SS increased the total concentration of Cu and Zn in amended soils, but without affecting their bioavailability, possibly due to the high values of soil pH and OM content.

It is a well-known fact that metals might promote the spread of antibiotic resistance via co-selection mechanisms. In consequence, due to the current growing concern over the emergence and dissemination of antibiotic resistance, an increasing interest exists in the study of the generalities and particulars related to the role of metal contamination as selective agent in the proliferation of environmental antibiotic resistance. This co-selection depends on a variety of factors, including the nature and concentration of metal contaminants, as well as the type of antibiotic and its mechanism of action. In any case, the co-selection of antibiotic and metal resistance rests on two main mechanisms: (i) co-resistance, when different resistance determinants are present in the same genetic element; and (ii) cross-resistance, when the same genetic determinant is responsible for resistance to both antibiotics and metals (Baker-Austin et al., 2006). Due to the magnitude of the problem of metal contamination, as well as to the exceptionally long residence times of metals in the environment, metals represent a widespread and recalcitrant selection pressure contributing to the emergence, maintenance and spread of antibiotic resistance factors. In our study, total Cu concentration in soil was positively correlated with the relative abundance of both ARGs and MGE-genes. Other authors have reported that the selective pressure caused by the presence of metals in SS can result in increased abundance of ARGs (Gullberg et al., 2014). In a previous study on the long-term impact of SS on agricultural soil (Urra et al., 2019a), we found a positive correlation between soil Cu and Zn concentrations and the abundance of ARGs and MGE-genes in soils amended with thermally-dried anaerobically-digested SS. Similarly, we observed that all metals were positively correlated with, at least, one ARG (soil metal concentrations also showed positive correlations with MGE-genes, particularly with the gene tnpA-07) (Urra et al., 2018).

The content of Olsen P explained a considerable percentage of the variation in the relative abundances of ARGs and MGE-genes in our soils, which could possibly be related to the abovementioned high P content of SS (Daneshgar et al., 2018). Soil CEC was also positively correlated with the relative abundance of ARGs, which could be, at least partly, due to the interaction (e.g., sorption mechanisms) between antibiotics and soil surfaces in SS-amended soils (Tolls, 2001). The distribution and fate of antibiotics in soil hinge on their sorption coefficient value (K_d) which, in turn, depends on a variety of soil properties, such as soil pH, OM, and CEC (Lützhøft et al., 2000; Chee-Sanford et al., 2009; Wegst-Uhrich et al., 2014). In any event, although environmental parameters such as soil pH and OM content are the most significant variables that affect antibiotic sorption in soil, it is important to take into consideration the concentrations used, the analytical method employed, and the transformations that can occur when determining K_d values (Wegst-Uhrich et al., 2014). Antibiotics with low K_d values are highly mobile (they are not strongly bound to soil particles) and, hence, more bioavailable and degradable. The principal route for antibiotic degradation in soil is aerobic biodegradation. Antibiotic degradation rates, usually expressed as half-lives, can range from days to years (Boxall et al., 2004). Macrolides and β-lactams have been reported to be less persistent in soil than tetracyclines and quinolones (Boxall et al., 2004).

As stated above, the application of organic wastes of anthropogenic origin, such as SS, can increase the risk of dissemination of ARGs into agricultural soil (Munir and Xagoraraki, 2011). Values of ARG relative abundance in our SS were higher compared to those observed by other authors: in anaerobically digested sludge from South Korea, Yoo et al. (2020) reported relative abundance values of 8.93 × 10^–2, 5.74 × 10^–2, and 5.71 × 10^–2 for MLSB, sulfonamide and tetracycline genes, respectively. Xu et al. (2020) found ARG relative abundance values of 3.3–5.1 × 10^–1 in SS samples from two Chinese WWTPs operated via anaerobic/anoxic/oxic and anoxic zone nitrification techniques. Han and Yoo (2020) reported ARG relative abundance values from 5.80 × 10^–5 to 1.20 × 10^–1 and from 4.62 × 10^–6 to 4.48 × 10^–1 in activated and dewatered sludge, respectively. Differences in antibiotic consumption patterns, as well as in SS composition, treatment and management, are possibly responsible for the observed differences in ARG relative abundance values. In any case, it is strongly recommended that SS is properly treated prior to its application in order to minimize or prevent the spread of ARB and ARGs from WWTPs to agricultural soils. This same recommendation was reported by Urra et al. (2019c) in their study on the impact of six different manure-derived amendments on agricultural soil quality. According to Sun et al. (2016), anaerobic digestion generally reduces integron abundance, eliminating the aerobic hosts of such integrons. The presence of the class 1 integron has been reported to be associated with the proliferation of ARGs in soil (Binh et al., 2008). In fact, the intl1 gene was proposed as a proxy for anthropogenic contamination, due to its relationship with ARGs, disinfectants and metals (Gillings et al., 2015).

In any event, when interpreting data on antibiotic resistance in soil, it must not be forgotten that natural soil microbial communities are potentially a large environmental reservoir of antibiotic resistance (Allen et al., 2009). Although, in this study, thermally-dried and anaerobically-digested SS was used as amendment, out of the 85 ARGs measured here, 77 and 74 were detected in soil samples and SS itself, respectively. Furthermore, the 10 MGE-genes determined in our study were detected in both soil samples and SS.

In our study, the relative abundance of both ARGs and MGE-genes was higher in SS-amended soils, compared to non-amended soils, particularly in those with a more recent SS application, suggesting that the application of thermally-dried anaerobically-digested SS to agricultural soil can increase the risk of environmental antibiotic resistance, as previously described by Urra et al. (2019a). Rahube et al. (2014) reported the impact of fertilizing with anaerobically-digested vs. raw SS on the abundance of antibiotic-resistant coliforms, antibiotic resistance genes, and pathogenic bacteria in soil and on vegetables at harvest. Instead, Rutgersson et al. (2020) did not find evidence of antibiotic accumulation or enrichment of ARGs or ARB in soil amended with digested and stored SS at doses of up to 12 tons per hectare every 4 years.

The use (above all, the overuse and misuse) of antibiotics for animal, human and agricultural purposes has led to the emergence and dissemination, by HGT, of antibiotic resistance among bacteria. The spread of antibiotic resistance by HGT is an ancient phenomenon. Nonetheless, the use of antibiotics has raised the corresponding selective pressure, thus increasing HGT (von Wintersdorff et al., 2016). As abovementioned, our results showed a stimulatory effect of SS on the relative abundance of both ARGs and MGE-genes, especially when the elapsed time after the last application was only 1 year, in agreement with previous studies (Chen et al., 2016; Xie et al., 2016). A priori, the increased relative abundances of ARGs and MGE-genes in SS-amended soils could be the consequence of: (i) an increase microbial growth induced by the input of organic carbon sources present in the SS itself; and/or (ii) the dissemination of ARGs from SS to soil microorganisms through HGT. Since the intensity of these processes will decrease over time after the application of SS, it is expected that the resistome risk will also decrease over time after the last application of SS. In a much shorter microcosm experiment, Chen C. et al. (2019) found that risk scores for amended soils decreased to levels comparable to those of unamended soil after just 120 days. However, as stated before, the persistent accumulation of heavy metals in SS-amended soils provide a long-term selective pressure which can result in co-selection for antibiotic resistance (Song et al., 2017) and, thus, a persistent resistome risk.

Urra et al. (2019c) observed that genes encoding MGEs (tnpA, intI1) were positively correlated with ARGs, suggesting a risk of dissemination of antibiotic resistance via HGT in agricultural soils, as a result of the application of livestock manure-derived amendments. Garbisu et al. (2018) screened metal contaminated soil treated with organic wastes for the presence of MGEs, confirming the occurrence of conjugative IncP-1 and mobilizable IncQ plasmids, as well as of class 1 integrons, suggesting that bacteria from those soils had gene-mobilizing capacity with implications for the dissemination of resistance factors. Moreover, their data pointed out the role of spontaneous mutations in achieving low-level antibiotic resistance in a short time, which was associated with a trade-off in the capability to metabolize specific carbon sources (Garbisu et al., 2018).

Regarding prokaryotic taxa at phylum level, the two dominant phyla in SS were Proteobacteria and Bacteroidetes. Both phyla have previously been reported as dominant taxa in SS (Liao et al., 2018; Zhao et al., 2019) and can harbor ARGs (Sun et al., 2016). In accordance with previous studies (Zhang et al., 2019), in our soils, Xanthomonadaceae was linked with the relative abundance of ARGs and MGE-genes. Interestingly, this family was the fifth most abundant family in the SS itself. In their study on the long-term impact of SS application of agricultural soil, Urra et al. (2018) found that some ARGs correlated positively with particular prokaryotic taxa, being Gemmatimonadetes the taxon with the greatest number of positive correlations at phylum level. However, no positive correlation was detected between prokaryotic taxa and genes encoding resistance to sulfonamides and FCA (Urra et al., 2018).

Interestingly, Shannon’s and Simpson’s index values correlated negatively with ARG and MGE-gene relative abundances, as previously reported (Chen Q. L. et al., 2019), suggesting that a decrease in prokaryotic diversity could facilitate the proliferation and spread of ARGs and MGE-genes in soil. Since microbial diversity is essential for ecosystem functioning and resilience (Delgado-Baquerizo et al., 2016), the impact of organic amendments on such diversity must be carefully studied.

According to data from our Kendall’s correlation analysis, 30 families showed correlation with at least two genes that are known to confer resistance against several antibiotic families. Among these 30 multi-resistant families, 12 showed a higher abundance in SS-amended vs. unamended soils, pointing out to an increased risk of antibiotic resistance.

The network analysis is a powerful tool to identify key ARGs (with potential as indicators or proxies) and to delineate potential ARGs hosts (Chen et al., 2016). In our network analysis, the families Chromatiaceae, Alcanivoracaceae, and Opitutaceae were identified as the most common potential hosts of ARGs and MGE-genes in SS-amended soils. Many species belonging to Chromatiaceae are photosynthetic sulfur bacteria which contribute to water purification in sewage lagoons (Holm and Vennes, 1970). The family Alcanivoracaceae contains bacterial species that harbor Type IV and Type II secretion system genes (Yakimov et al., 2019). The Type II pathway is associated with organisms that form biofilms (Sandkvist, 2001). Bacteria within biofilms can be 10–1,000 times more resistant to antibiotics, compared to corresponding planktonic cells (Ceri et al., 1999). In accordance with Chen Q. L. et al. (2019), the family Opitutaceae was strongly correlated with a tetracycline resistance gene (tetG). The higher relative abundance shown by these three families in SS-amended soils, together with their positive correlation with the abundance of ARGs and MGE-genes, suggests their potentially important role in AR in SS-amended agricultural soils.

Structural equation models facilitate the study of complex interactions. In our study, soil physicochemical parameters (OM, CEC, EC, clay content, pH, soil nutrients) were major predictors of the HM pollution index. According to our SEM analysis, the composition and relative abundance of ARGs were mainly driven by MGE-genes, pointing out to the well-known role of HGT for the observed resistome (Smillie et al., 2011). The elapsed time after the last application contributed to the observed variation in MGE-genes diversity and abundance, suggesting that the acquisition of genetic material from non-parental lineages by HGT is dependent on such elapsed time. Our SEM confirmed the recognized crucial role of MGEs in shaping the patterns of ARGs and facilitating their dissemination. Indeed, in this study, MGE-genes (which were themselves affected by parameters such as soil OM and nutrient contents, the elapsed time after the last application of SS, and the distribution of ARGs) have shown their critical role for AR dissemination in agricultural soils.

In our SS-amended soils, the composition of prokaryotic communities was mainly influenced by the soil’s physicochemical properties. The 30 most abundant families within the soil prokaryotic community accounted for a considerable percentage (66%) of the total variation of ARG and MGE-gene relative abundances. Soil prokaryotic α-diversity was negatively correlated with the relative abundance of ARGs and MGE-genes, suggesting that a decrease in soil prokaryotic diversity could facilitate the proliferation and spread of ARGs and MGE-genes in soil. The families Chromatiaceae, Alcanivoracaceae, and Opitutaceae were identified as the most common potential hosts of multi-resistant genes in SS-amended soils. According to the SEM performed here, the diversity and abundance of MGE-genes are critical for AR dissemination in agricultural soils. We concluded that agricultural soils amended with our thermally-dried anaerobically-digested SS showed increased risk of antibiotic resistance dissemination. Much research and awareness is needed for (i) a safe reduction of antibiotic consumption and (ii) the development of innovative treatments for SS (prior to its application) in order to reduce the risk of emergence and spread of ARGs and MGE-genes in SS-amended agricultural soils and then crops. Until then, SS should be applied conservatively in agricultural soils.

The datasets presented in the study are publicly available. The sequencing data can be found here: https://www.ebi.ac.uk/ena, PRJEB40822. The HT-qPCR datased is included in Supplementary Table 2.

CG, IA, and LE designed the study. LJ and LE performed the analytical work and analyzed data. LJ, CG, IA, and LE wrote the manuscript. All authors revised the final version of the manuscript.

The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.