The urgent need for risk assessment on the antibiotic resistance spread via sewage sludge land application

Environment International 87 (2016) 49–55

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Environment International journal homepage: www.elsevier.com/locate/envint

Review article

The urgent need for risk assessment on the antibiotic resistance spread via sewage sludge land application Kinga Bondarczuk ⁎, Anna Markowicz, Zofia Piotrowska-Seget Department of Microbiology, Faculty of Biology and Environmental Protection, University of Silesia in Katowice, Poland

a r t i c l e

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Article history: Received 2 September 2015 Received in revised form 10 November 2015 Accepted 11 November 2015 Available online xxxx Keywords: Soil fertilisation Antibiotic-resistant bacteria Antibiotic resistance genes Co-resistance Heavy metals Environmental resistome

a b s t r a c t Sewage sludge is an ever-increasing by-product of the wastewater treatment process frequently used as a soil fertiliser. To control its quality and prevent any possible hazardous impact of fertilisation, some mandatory limits of heavy metal content have been established by the European Commission (Sewage Sludge Directive). However, since the implementation of the limits, new emerging contaminants have been reported worldwide. Regardless of the wastewater treatment process, sewage sludge contains antibiotics, antibiotic-resistant bacteria and antibiotic resistance genes, which can be released into the environment through its land application. Such a practice may even boost the dissemination and further development of antibiotic resistance phenomenon – already a global problem challenging modern medicine. Due to the growing pharmaceutical pollution in the environment, the time is ripe to assess the risk for the human and environmental health of sewage sludge land application in the context of antibiotic resistance spread. In this review we present the current knowledge in the field and we emphasise the necessity for more studies. © 2015 Elsevier Ltd. All rights reserved.

Contents 1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2. Antibiotic resistance in sewage sludge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3. The persistence of antibiotic-resistant bacteria in fertilised soils assessed by culture-based methods . . . . . . 4. The persistence of antibiotic resistance determinants in fertilised soils assessed by culture-independent methods 5. The link between heavy metals and the maintenance of antibiotic resistance . . . . . . . . . . . . . . . . . 6. The environmental resistome . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7. Concluding remarks and perspectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1. Introduction The successive implementation of the Urban Waste Water Treatment Directive 91/271/EC (European Commission, 1991) by all the European Union Members has been causing an increase in the quantity of sewage sludge - a by-product derived from the wastewater treatment process. It has been predicted that the amount of sludge generated in these countries will exceed 13 million

Abbreviations: ARB, antibiotic-resistant bacteria; ARGs, antibiotic resistance genes; CSA, co-selecting agents; MGEs, mobile genetic elements. ⁎ Corresponding author at: Department of Microbiology, Faculty of Biology and Environmental Protection, University of Silesia, 40-032 Katowice, Jagiellońska 28, Poland. E-mail address: [email protected] (K. Bondarczuk).

http://dx.doi.org/10.1016/j.envint.2015.11.011 0160-4120/© 2015 Elsevier Ltd. All rights reserved.

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tonnes by 2020 (Milieu, Ltd., et al., 2010). Since sewage sludge is rich in valuable nutrients and organic matter, it can be used as a fertiliser and for soil remediation purposes (Cortet et al., 2011; Weber et al., 2007). These applications appear to be an excellent solution for waste disposal. However, despite its indisputable advantages, sewage sludge contains pollutants such as heavy metals and diverse organic contaminants that may have toxic effects on all living organisms (Alvarenga et al., 2015; Passuello et al., 2010; Petrie et al., 2014). In 1986 the European Commission implemented Sewage Sludge Directive 86/278/EEC, which encouraged the application of sewage sludge in agriculture provided that the sludge is used correctly and does not impair soil quality and agricultural products. The regulated use of sludge should ‘prevent harmful effects on soil, vegetation, animals and man’. To reach this goal, mandatory limit values for heavy metal

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concentrations in sludge and fertilised soils were established (Annexes IA-IC of the Sewage Sludge Directive 86/278/EEC). Moreover, since the implementation of the Directive, untreated sludge must not be applied, unless it is injected or worked into the soil (European Commission, 1986). Treated sewage sludge, often referred to as biosolids, is subjected to additional processes (such as aerobic or anaerobic digestion) prior to land application in order to minimise the content of pathogenic bacteria (Youngquist et al., 2014). Because of new emerging contaminants, there is a growing debate about the revision of the Sewage Sludge Directive. A study launched by the European Commission revealed emerging risks connected to sludge use on land. The FATE SEES (FATE — fate and impacts of pollutants in terrestrial and aquatic ecosystems, SEES — sewage sludges and effluents for emerging substances) monitoring project is dedicated to assessing the presence and concentration levels of pollutants in sewage sludge. Aside from the well-established contaminants such as heavy metals, polychlorinated biphenyls, polychlorinated dibenzo-p-dioxins and polycyclic aromatic hydrocarbons, the study is focused on the less-investigated emerging pollutants such as brominated flame retardants, the ingredients of personal care products, and last but not least, pharmaceuticals (Gorga et al., 2013; Gurke et al., 2015; Roig et al., 2012). The occurrence of pharmaceuticals in the environment and their potential ecotoxicological effects have drawn special attention in recent years (Malmborg and Magner, 2015; Martín et al., 2015). Drugs are not completely metabolised in human and animal bodies and they are excreted either as parent compounds or as metabolites via urine and faeces. They may enter the environment directly along with the faeces, (e.g., through land application of manure) or via wastewater treatment plants effluents and the application of sewage sludge (Gao et al., 2012; Loganathan et al., 2009; McEneff et al., 2014). Due to the global spread of antibiotic resistance among bacteria, the release of antibiotics, antibiotic-resistant bacteria (ARB) and antibiotic resistance genes (ARGs) is a matter of special concern (Roca et al., 2015). Furthermore, the co-occurrence of antibiotic and metal resistance in bacteria has been observed (Máthé et al., 2012). This effect is caused by the cross- and co-resistance phenomena. Cross-resistance occurs when the same mechanism reduces the susceptibility to metals and antibiotics simultaneously and co-resistance occurs when separate resistance genes are situated on the same genetic element (Baker-Austin et al., 2006; Knapp et al., 2011). This fact may be of great importance in the case of the agricultural use of sewage sludge, since as was already mentioned, it contains significant amounts of heavy metals (Alvarenga et al., 2015; Heck et al., 2015; Milinovic et al., 2014; Seiler and Berendonk, 2012). Therefore, the presence of heavy metals in both sludge and sludge-amended soil may select for antibiotic-resistant bacteria in fertilised soil (Baker-Austin et al., 2006; Gullberg et al., 2014; Seiler and Berendonk, 2012). The aim of this review is to summarise recent advances in the field of land application of sewage sludge in the context of antibiotic resistance phenomenon. 2. Antibiotic resistance in sewage sludge It is well established that sewage sludge contains significant amounts of diverse antibiotics that represent nearly all major classes (excluding labile β-lactams). The concentrations assessed in sewage sludge vary between ng to mg per kg of dried weight (Jelić et al., 2012; Le-Minh et al., 2010; Li et al., 2013; McClellan and Halden, 2010). It has been documented that pharmaceuticals are adsorbed on sewage sludge particles. The rate of this process depends on drug chemical structure, mobility, hydrophobicity, biodegradation and the nature of the sludge itself. Moreover, it has been reported that pharmaceuticals adsorbed to sludge exhibit a higher degree of stability than those found in wastewater (Cheng et al., 2014; Li et al., 2013). Recent findings suggest that even ppb (part per billion) concentrations of

antimicrobials maintain the ARGs in bacterial populations and may favour plasmid transfer (Gullberg et al., 2014; Kim et al., 2014). Due to antibiotic content, it is reasonable to assume that some microorganisms residing in sewage sludge may be intrinsically tolerant to these compounds (e.g. due to cell envelope impermeability) and/or exhibit antibiotic resistance conferred by clinically relevant mechanisms. Indeed, regardless of applied method treated sewage sludge is rich in ARB. However, the most advanced technologies such as anaerobic digestion and lime stabilisation significantly reduce ARB number when compared to simple dewatering and gravity thickening (Munir et al., 2011). Intriguingly, even if ARB lose viability during the treatment process, the frequency of ARGs may simultaneously increase. Studies conducted by Su et al. (2015) revealed 156 unique ARGs and mobile genetic elements encoding resistance to virtually all of the known antibiotic groups in composted sewage sludge, thus suggesting this byproduct is a significant reservoir of antibiotic resistance determinants (Calero-Cáceres et al., 2014; Su et al., 2015). 3. The persistence of antibiotic-resistant bacteria in fertilised soils assessed by culture-based methods Antibiotics are abundant in wastewater and are not completely eliminated during the treatment process. They persist in activated sludge basins at sublethal concentrations, thus leading to the selection of resistant bacteria. The presence of antibiotics together with a high microbial cell density and diversity favour the horizontal gene transfer (HGT) of resistance determinants among the bacteria residing in activated sludge (Michael et al., 2013; Zhang et al., 2011). Antibioticresistant bacteria and antibiotic resistance genes may subsequently be transferred to the environment through the land application of sewage sludge (Hölzel et al., 2010; Rahube et al., 2014; Riber et al., 2014). Culture-based approaches were used to determine the persistence of antibiotic-resistant bacteria in sewage sludge-fertilised soils (Table 1). Even though the culture-based methods allow only a small fraction of the bacteria present in soil to be grown (ca. 1%), they still provide valuable information about the resistance phenomenon. Culture-based methods give the possibility to establish the link between antibiotic resistance genes and their bacterial hosts. Well-known pathogenic and indicator bacteria are easily grown on selective media using standard procedures. That is why assessing potential human health risks is feasible using culture-based approaches. Furthermore, phenotyping provides information about complete and active genes, which is not always the case when using molecular methods such as qPCR. On the other hand, culture-dependent methods are laborious and time consuming, nevertheless their major limitations stem from the narrow fraction of bacteria that can be cultivated (Hill et al., 2000; Torsvik et al., 1998). In order to get information about the fate of culturable ARB in fertilised soils, two approaches may be applied. One considers the persistence of ARB added to soil together with sewage sludge, while the second screens for antibiotic resistance among autochthonous microorganisms. It is expected that although the number of bacteria carried with fertiliser would decrease with time, some changes in ARGs pool among soil-dwelling bacteria might be observed. A three-year study conducted by Rahube et al. (2014) added significant information to the current knowledge in this field. The authors investigated the impact of fertilising with raw sewage sludge and dewatered municipal biosolids (anaerobically digested sewage sludge) on the abundance of ARB, pathogens and ARGs in soils and on vegetables at harvest. In this comprehensive field study, the presence of pathogenic bacteria and antibiotic-resistant coliforms was verified using culture-dependent methods. As expected, the number of viable bacteria evaluated in the sewage sludge was much higher when compared with the biosolids. None of the pathogens detected and quantified in the sewage sludge were found at quantifiable levels in the biosolids. The percentage of antibiotic-resistant coliform bacteria was also higher in

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Table 1 Experimental conditions and methods used in research on antibiotic resistance in sewage sludge-fertilised soils. Fertiliser type

Dose

Type of study

Culture-based methods

Culture-independent methods

Tested genes

Period of study

References

Dewatered untreated and treated (anaerobically digested) municipal sewage sludge

10.8 wet tonnes ha−1 28.6 wet tonnes ha−1

field study – without previous history of sewage sludge application

plate count of: total and faecal coliforms, E. coli, Enterococcus spp., Clostridium perfringens, Aeromonas spp., Yersinia spp., Campylobacter spp., Salmonella spp., Listeria spp.

PCR quantitative PCR

various antibiotic resistance genes, integrases, plasmid incompatibility groups

12 months

Rahube et al. (2014)

Dewatered mesophilic anaerobic digested residual municipal wastewater solids

0, 20, 40, 100 g of residual solids kg−1 soil

microcosms

Real-time PCR

6 months

Burch et al. (2014)

Municipal solid waste compost and sewage sludge (treatment unspecified)

11 tonnes DM ha−1 year−1 accelerated amounts for long-term effects (113 years)

field trial – soil with long-term urban organic waste fertilization

16S rRNA for all Bacteroides spp., human-specific Bacteroides spp., integrase gene of class 1 integrons, antibiotic resistance genes 16S rRNA – total bacterial diversity 16S rRNA – Pseudomonas gene copy number

29 weeks (7 months)

Riber et al. (2014)

plate counts of: total number of Pseudomonas strains, number of tetracycline or gentamicin resistant Pseudomonas

the raw amendment. The study revealed no significant differences in the number of enteric bacteria in the soil and on vegetables at harvest between the treated and untreated plots. However, the only observed impact of the fertilisation with sewage sludge at harvest was the increased number of Clostridium perfringens and an increase in cefoxitin-resistant coliform bacteria on carrots (Rahube et al., 2014). A coherent impact of the biosolids and sewage sludge applications could not be established. Coliforms resistant to some antibiotics were grown frequently but the results varied between the replicates. The plate counts revealed that the enteric bacteria introduced into the fertilised soil lost their viability before the crop harvest and that a year after the harvest no regrowth of these microorganisms was observed (Rahube et al., 2014). In another field approach Riber et al. (2014) used the viable plate counts to investigate the influence of animal and urban organic waste fertilisers on the Pseudomonas population in amended soil. Since the species representing this genus are a part of the natural soil microbial communities and are known to be opportunistic pathogens, the assessment of pseudomonads in the context of antibiotic resistance is of great importance. Moreover, monitoring this well-studied genus using culture-dependent methods is a convenient tool because the standard procedures for the isolation of its representatives are well-established. In their experiment, Riber et al. (2014) treated soil with manure, sewage sludge and composted source-separated organic household waste (compost). They determined the prevalence of cultivable tetracycline- and gentamicin-resistant pseudomonads and tested the resistant isolates for multi-resistance. The number of antibiotic-resistant isolates increased significantly in the sewageand compost-treated soils directly after the fertilisation; however, it decreased rapidly over time. These results may reflect the survival of bacteria introduced into the soil with the amendments or the transient enrichment of indigenous resistant strains caused by the selective pressure of the contaminants derived from the wastes (Riber et al., 2014). It is noteworthy that the majority of tetracycline-resistant pseudomonads (including the isolates from the control plots) exhibited multi-resistance. This fact indicates that even though the vast majority of ARB die out over time, some remaining multi-resistant strains or ARGs that were once introduced into the microbial community may remain in the soil and contribute to the common gene pool. Moreover, it has been documented that even pristine environments may be rich in ARGs and the concept of natural resistome has been proposed (Wright, 2007). For details please see the section ‘The environmental resistome’.

pyrosequencing quantitative PCR exogenous plasmid isolation, biparental matings using P. putida KT2442

Culture-based methods may also be used to study the co-selection of antibiotic and metal resistance in bacteria. Using media amended with antibiotics and heavy metals provides initial information about the co-occurrence of antibiotic and metal resistance in isolated strains. Establishing a link between observed phenotypes may enable us to determine which antibiotic resistance determinants may be maintained in a given environment polluted with heavy metals. Interestingly, Riber et al. (2014) reported the co-selection of mercury and tetracycline resistance among Pseudomonas isolated from sewage sludge-treated soil. Moreover, the authors also stressed the fact that all mercury-resistant isolates were simultaneously resistant to at least two antimicrobials (Riber et al., 2014). Already presented data suggest that the knowledge about the mobile genetic elements conferring the co-resistance might facilitate the prediction of the spread and persistence of antibiotic resistance in polluted environments. 4. The persistence of antibiotic resistance determinants in fertilised soils assessed by culture-independent methods Mobile genetic elements (MGEs), i.e. plasmids, integrons and transposons, are widely recognised as antibiotic and metal resistance carriers. Some of them are omnipresent and are isolated from various samples and species (Allen et al., 2010). MGEs have a formidable nature, which pertains to their ability to shuffle genes, to provide the machinery necessary for expression and to transfer genes between the DNA molecules inside a cell and between different microorganisms (Allen et al., 2010; Hall, 2012). These remarkable genetic elements involved in HGT play a crucial role in the evolution of composite resistance determinants such as the plasmid R100 (Womble and Rownd, 1988). Culture-independent methods have been used to study the persistence of MGEs in fertilised soils. Culture-independent techniques overlap the gene and species diversity that cannot be described using culture-based approaches. However, they are not devoid of limitations. High-quality DNA extraction from matrices such as soil and sewage sludge still remains difficult. Additionally, PCR and qPCR, which were implemented in the studies, cannot distinguish between genes derived from living and dead bacteria or whether a PCR product was amplified on a complete and active gene template. Among the methods recently implemented to study the spread of resistance in fertilised soils, exogenous plasmid isolation deserves special attention. Mating approaches using donor bacteria collected directly from a sample allow to isolate active resistance determinants

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from living cells. Capturing these elements enables their further structural and functional characteristics including the dynamics of their spread to be determined. Nonetheless, the use of a given recipient during mating always results in narrowing the recovery of the plasmid pool from environmental strains. This is due to the specified host range that plasmids can be propagated in. Exogenous plasmid isolation has been used by Riber et al. (2014) to analyse the horizontal transfer frequency of antibiotic (tetracycline, gentamicin) and metal (mercury) resistance genes. The authors demonstrated a higher frequency of HGT in amended soil compared to an untreated control only in the first week after fertilisation. Obtained transconjugants were often multi-resistant. The co-selection of mercury and drug resistance was also observed and, noteworthy, all of the mercury-resistant transconjugants were resistant to all five tested antibiotics. Although the impact of the fertilisation regarding the number of cultivable antibiotic-resistant pseudomonads was negligible, the percentage of multi-resistant transconjugants increased in week 29 compared to week 0. This indicates that autochthonous bacteria might have acquired some MGEs conferring antibiotic resistance, thus suggesting a change in bacterial gene pool (Riber et al., 2014). Resistance plasmids often carry integrons – natural expression systems that capture resistance genes under a common promoter. The integration of gene cassettes is catalysed by a site-specific recombinase (integrase), which is encoded by the intI gene (Hall, 2012). Several studies have revealed the high prevalence of integrons carrying antibiotic resistance gene cassettes in wastewater treatment plants, final effluents and treated sewage sludge (Ma et al., 2013; Miller et al., 2014; Mokracka et al., 2012; Moura et al., 2012). The results obtained by Miller et al. (2014) showed an increase in intI genes in sludge that had been stored for land application within the first two months followed by decrease in the next two months, thereby suggesting at least a two-month-long storage period before fertilisation. Noteworthy, storage at cold temperatures even favoured the observed increase in intI genes (Miller et al., 2014). In a six-month microcosm experiment Burch et al. (2014) investigated the fate of class 1 integrons and five antibiotic resistance genes in biosolid-amended soil. Using qPCR, they revealed a slower decay rate of the chosen genes (half lives longer than two weeks) in soil compared to data reported during the wastewater treatment process (half lives below one week). That implies a potential risk of the dissemination of resistance determinants in treated soils and a need to develop procedures within wastewater treatment plants to minimise the spread of ARGs rather than relying on decay in fertilised soils (Burch et al., 2014). The most persistent gene in the studied microcosms was intI. This suggests that some antibiotic resistance determinants may be maintained and may simultaneously extend the intrinsic microbial capability of HGT. What is interesting is that the different mass loading of fertiliser did not significantly change the observed decay rate, thus suggesting that limiting the mass loading of sewage sludge would not affect the persistence of genes in amended soil. However, gene accumulation cannot be excluded when sewage sludge is applied repeatedly (Burch et al., 2014). To assess the attenuation of the ARGs in fertilised soils, Rahube et al. (2014) prolonged their monitoring until a year after the treatment. Using PCR and qPCR, they showed that even though an increase in the variety and abundance of target genes was observed in the season of fertiliser application it diminished to no observable effect after 15 months. Some of the ARGs were detected in the control soils thereby demonstrating a natural resistome of the studied environment (Rahube et al., 2014) (Please see the section ‘The environmental resistome’). 5. The link between heavy metals and the maintenance of antibiotic resistance The concentrations of heavy metals in sewage sludge are highly variable and depend mainly on the source of the wastewater and the

treatment processes that are used. The wide range of the variability was notably elucidated by Pan-European FATE SEES Project (61 samples of sewage sludge from 15 European countries, Table 2) (Tavazzi et al., 2012). However, the highest heavy metal levels were found in industrial wastewaters and sludge (Chen, 2014; Przewrocki et al., 2004). Prior to its application to soil, sewage sludge is often composted in order to reduce metal availability (Vaca et al., 2011). However, the leaching of heavy metals into sludge-amended soils occurs and has been studied extensively (Milinovic et al., 2014; Toribio and Romanyà, 2006). It is known that leaching depends on several physicochemical characteristics of soil and applied sludge as well as of the metals themselves or climate conditions. The release of some metals has a cyclic behaviour (Ahlberg et al., 2006; Milinovic et al., 2014). However, unlike antibiotics, metals are not subject to degradation and can subsequently cause a long-term selective pressure. Some authors have observed significant correlations between the presence of heavy metals and antibiotics in sludge, especially Zn, Pb and tetracyclines (Cheng et al., 2014). The relationships between Zn and tetracyclines may be due to the use of ZnSO4 as a purification agent during the production of these antibiotics (Foley and Blackwell, 2003). Some metals may cause an increase in the adsorption of antibiotics into high organic matter sludge. For example, Cu (II) may increase the formation of ternary complexes between organic matter and tetracycline ligands (MacKay and Canterbury, 2005; Zhao et al., 2013) and copper ions can enhance the adsorption of tetracyclines into clays (Wang et al., 2008). It is well established that the presence of metals and other stressors in the environment is responsible for the elevated number of antibiotic resistance genes (Gullberg et al., 2014). Recent findings suggest that the number of intrinsic antibiotic resistance genes positively correlates with the content of copper, chromium, nickel, lead and iron even in natural uncontaminated soil environments (Knapp et al., 2011). This effect is caused by the cross and co-resistance phenomena that has already been mentioned (Baker-Austin et al., 2006). This has been well documented in the instance of transposon-carried mer operon, which confers mercury resistance. Operon mer together with gene encoding for streptomycin/ spectinomycin resistance was localised on transposon Tn21 incorporated into plasmid R100. Aside from the Tn21, the plasmid R100 additionally contains Tn9 and Tn10 conferring chloramphenicol and tetracycline resistance, respectively. The resistance conjugative plasmid R100 was originally extracted from a pathogenic Shigella flexneri strain isolated from human faeces (Womble and Rownd, 1988). 6. The environmental resistome Although antibiotic resistance had been primarily and extensively studied in pathogens, it has also been recognised as a common and Table 2 Metal concentrations in 61 studied sewage sludge samples from 15 European countries (A, B, CZ, FIN, D, GR, H, IRL, LT, P, RO, SLO, S, CH, NL) provided in the report by FATE SEES Project. BDL — below detection limit, STD — standard deviation (Tavazzi et al., 2012).

Ag As Ba Cd Co Cr Cu Hg Mn Mo Ni Pb Sb Ti V

Min. conc. [mg kg−1]

Max. conc. [mg kg−1]

Average conc. [mg kg−1]

STD [mg kg−1]

BDL BDL 41.5 BDL 1.5 10.8 27.3 0.1 75.2 1.7 8.6 4 BDL 65.2 2.3

14.7 56.1 580 5.1 16.7 1542 578 1.1 960 12.5 310 430 53.6 1071 135

3.3 – 225 0.9 6.3 79.8 257 0.4 329 4.9 29 47.6 6 440 25

3 – 102 0.7 3.3 215 118 0.2 193 1.9 40.2 59.3 8.2 255 20.3

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ancient feature among environmental strains (D'Costa et al., 2011). Furthermore, it seems that antibiotics have a hormetic effect on bacterial cells. This means that exposure to sublethal doses of antimicrobials may even be beneficial for bacteria (Martínez, 2008). These facts have changed our understanding of antibiotic resistance phenomenon. If antibiotics at low concentrations are involved in cell metabolisms, the genes conferring antibiotic resistance may also have some other metabolic functions as was described for the gene encoding an aminoglycoside acetyltransferase in Providencia stuartii (Debabov, 2013; Franklin and Clarke, 2001; Martinez, 2014). The enzyme catalyses peptidoglycan modifications but (due to the resemblance in chemical structure), it also recognises gentamicin, leading to the antibiotic inactivation (Franklin and Clarke, 2001). Thus ARGs found in environmental samples may not necessarily be involved in resistance in situ. However, when the bacteria harbouring the genes are challenged by antibiotic pressure, the genes may fulfil their resistance role (Martínez, 2008). Moreover, it has been suggested that the ARG pool in the environment is probably much larger than the gene pool observed among pathogens. If that is the case, the question of why the release of ARB into natural environments should be a cause for concern. There are two reasons – firstly, pathogenic bacteria may acquire new ARGs from environmental strains, and secondly, environmental microorganisms may capture clinically important genes evolved under a strong selective pressure from pathogens. Both scenarios threaten the future of antibiotic therapy. Such a transfer has been reported for some clinically relevant genes such as ctx-m encoding for extended spectrum β-lactamase, which originates from the environmental Kluyvera strain (Poirel et al., 2002).

7. Concluding remarks and perspectives In the face of the increasing production of sewage sludge and its possible linkage to the global threat of antibiotic resistance, human and ecological risk assessment and the fate of the ARB and ARGs introduced during the fertilisation of soil should be evaluated. To date, only a few studies have been carried out on the issue and these have used both culture-based and culture-independent methods, which complement each other in order to provide more comprehensive information. The already published results imply a rather negligible effect of sewage sludge amendment on antibiotic resistance in fertilised soils, but after a certain period of time. The established periods varied depending on the type of experiment with the longest period reaching fifteen months (Rahube et al., 2014). The European Commission has set the waiting period at ten months before harvest, thus a prolongation of the time may be needed (European Commission, 1986). Since sewage sludge properties depend on the nature of the raw sewage as well as on the treatment process within a plant, a precise comparison between the different experiments conducted in different parts of the world remains problematic. Furthermore, different experimental conditions (such as a microcosm or a field study, soil with or without a previous history of sewage sludge fertilisation, etc.) highly influence the final results (Table 1). Therefore, some standard analyses of used sewage sludge and the type of soil should be implemented as a rule of thumb together with the fact that the experimental conditions should resemble agricultural practice. Exogenous plasmid isolation was performed to study the co-selection of antibiotic and metal resistance. However, a more detailed characterisation of the isolated replicons would be valuable. The observed co-selection of metal and antibiotic resistance suggests that determining the heavy metal concentrations in soil and fertiliser would allow a potential link between ARB/ARGs and soil metal content to be established. This would be of great importance when sewage sludge is used for soil recultivation purposes in heavily contaminated areas. Due to the fact that the already provided information is insufficient, there is an urgent need to broaden our knowledge in the field.

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As a matter of fact research gaps and uncertainties are among the major hurdles in assessing the risk of environmental antibiotic resistance in general. That issue was risen during the Antimicrobial Resistance in the Environment: Assessing and Managing Effects of Anthropogenic Activities workshop in Canada (2012). The participants proposed key aspects which should be included to assess the risk to human health posed by environmental antibiotic resistance. They highlighted that the following issues should be considered: (i.) elucidating the link between selective pressure and the development of ARB in environmental compartments; (ii.) estimating the rates of horizontal ARGs transfer and ARGs mutations in the compartments; (iii.) quantifying dose–response relationships for environmental ARB (Ashbolt et al., 2013). Regarding the issues raised in their postulate, risk assessment for land application of sewage sludge should include: • Defining agents co-selecting for antibiotic resistance (co-selecting agents, CSA) which are present in sewage sludge and/or soils; finding the dose–response relationships (time, concentrations); • Reporting the concentrations of antibiotics, heavy metals and other possible CSA in soils and sewage sludge which are destined for fertilisation; • Investigating the fate of antibiotics, CSA, ARB and ARGs in fertilised soils; • Assessing the rate of HGT between introduced and autochthonous bacteria (with special attention drawn on pathogens and opportunistic pathogens); • Finding the impact of soil properties/type on the development of antibiotic resistance; • Characterising natural resistome in soils; • Defining waiting periods for particular soil and crop types (Ashbolt et al., 2013). Acknowledgments The research on antibiotic resistance phenomenon carried out in the authors' lab is supported by National Science Centre (NCN) grants: 2014/13/N/NZ9/03915 and 2013/11/D/NZ9/0251. KB has received a scholarship provided by the “DoktoRIS — Scholarship programme for innovative Silesia”, which is co-financed by the European Union under the European Social Fund covered by Human Capital Programme. The authors declare no conflicts of interest. References Ahlberg, G., Gustafsson, O., Wedel, P., 2006. Leaching of metals from sewage sludge during one year and their relationship to particle size. Environ. Pollut. 144, 545–553. http://dx.doi.org/10.1016/j.envpol.2006.01.022. Allen, H.K., Donato, J., Wang, H.H., Cloud-Hansen, K.A., Davies, J., Handelsman, J., 2010. Call of the wild: antibiotic resistance genes in natural environments. Nat. Rev. Microbiol. 8, 251–259. http://dx.doi.org/10.1038/nrmicro2312. Alvarenga, P., Mourinha, C., Farto, M., Santos, T., Palma, P., Sengo, J., Morais, M.-C., Cunha-Queda, C., 2015. Sewage sludge, compost and other representative organic wastes as agricultural soil amendments: benefits versus limiting factors. Waste Manag. 40, 44–52. http://dx.doi.org/10.1016/j.wasman.2015.01.027. Ashbolt, N.J., Amézquita, A., Backhaus, T., Borriello, P., Brandt, K.K., Collignon, P., Coors, A., Finley, R., Gaze, W.H., Heberer, T., Lawrence, J.R., Larsson, D.G.J., McEwen, S.A., Ryan, J.J., Schönfeld, J., Silley, P., Snape, J.R., Van den Eede, C., Topp, E., 2013. Human health risk assessment (HHRA) for environmental development and transfer of antibiotic resistance. Environ. Health Perspect. 121, 993–1001. http://dx.doi. org/10.1289/ehp.1206316. Baker-Austin, C., Wright, M.S., Stepanauskas, R., McArthur, J.V., 2006. Co-selection of antibiotic and metal resistance. Trends Microbiol. 14, 176–182. http://dx.doi.org/10. 1016/j.tim.2006.02.006. Burch, T.R., Sadowsky, M.J., LaPara, T.M., 2014. Fate of antibiotic resistance genes and class 1 integrons in soil microcosms following the application of treated residual municipal wastewater solids. Environ. Sci. Technol. 48, 5620–5627. http://dx.doi. org/10.1021/es501098g. Calero-Cáceres, W., Melgarejo, A., Colomer-Lluch, M., Stoll, C., Lucena, F., Jofre, J., Muniesa, M., 2014. Sludge as a potential important source of antibiotic resistance genes in both the bacterial and bacteriophage fractions. Environ. Sci. Technol. 48, 7602–7611. http://dx.doi.org/10.1021/es501851s.

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