Antibiotics in the Soil Environment—Degradation and Their Impact on Microbial Activity and Diversity

Antibiotics play a key role in the management of infectious diseases in humans, animals, livestock, and aquacultures all over the world. The release of increasing amount of antibiotics into waters and soils creates a potential threat to all microorganisms in these environments. This review addresses issues related to the fate and degradation of antibiotics in soils and the impact of antibiotics on the structural, genetic and functional diversity of microbial communities. Due to the emergence of bacterial resistance to antibiotics, which is considered a worldwide public health problem, the abundance and diversity of antibiotic resistance genes (ARGs) in soils are also discussed. When antibiotic residues enter the soil, the main processes determining their persistence are sorption to organic particles and degradation/transformation. The wide range of DT50 values for antibiotic residues in soils shows that the processes governing persistence depend on a number of different factors, e.g., physico-chemical properties of the residue, characteristics of the soil, and climatic factors (temperature, rainfall, and humidity). The results presented in this review show that antibiotics affect soil microorganisms by changing their enzyme activity and ability to metabolize different carbon sources, as well as by altering the overall microbial biomass and the relative abundance of different groups (i.e., Gram-negative bacteria, Gram-positive bacteria, and fungi) in microbial communities. Studies using methods based on analyses of nucleic acids prove that antibiotics alter the biodiversity of microbial communities and the presence of many types of ARGs in soil are affected by agricultural and human activities. It is worth emphasizing that studies on ARGs in soil have resulted in the discovery of new genes and enzymes responsible for bacterial resistance to antibiotics. However, many ambiguous results indicate that precise estimation of the impact of antibiotics on the activity and diversity of soil microbial communities is a great challenge.

Antibiotics play a key role in the management of infectious diseases in humans, animals, livestock, and aquacultures all over the world. The release of increasing amount of antibiotics into waters and soils creates a potential threat to all microorganisms in these environments. This review addresses issues related to the fate and degradation of antibiotics in soils and the impact of antibiotics on the structural, genetic and functional diversity of microbial communities. Due to the emergence of bacterial resistance to antibiotics, which is considered a worldwide public health problem, the abundance and diversity of antibiotic resistance genes (ARGs) in soils are also discussed. When antibiotic residues enter the soil, the main processes determining their persistence are sorption to organic particles and degradation/transformation. The wide range of DT50 values for antibiotic residues in soils shows that the processes governing persistence depend on a number of different factors, e.g., physico-chemical properties of the residue, characteristics of the soil, and climatic factors (temperature, rainfall, and humidity). The results presented in this review show that antibiotics affect soil microorganisms by changing their enzyme activity and ability to metabolize different carbon sources, as well as by altering the overall microbial biomass and the relative abundance of different groups (i.e., Gram-negative bacteria, Gram-positive bacteria, and fungi) in microbial communities. Studies using methods based on analyses of nucleic acids prove that antibiotics alter the biodiversity of microbial communities and the presence of many types of ARGs in soil are affected by agricultural and human activities. It is worth emphasizing that studies on ARGs in soil have resulted in the discovery of new genes and enzymes responsible for bacterial resistance to antibiotics. However, many ambiguous results indicate that precise estimation of the impact of antibiotics on the activity and diversity of soil microbial communities is a great challenge.

INTRODUCTION
Antibiotics are complex molecules with different functional groups in their chemical structures and are divided into several classes (Figure 1, Table 1) depending on the mechanisms of action, i.e., inhibition of cell wall synthesis, alteration of cell membranes, inhibition of protein synthesis, inhibition of nucleic acids synthesis, competitive antagonism, and antimetabolite activity (Kümmerer, 2009). Antibiotics are widely prescribed for treatment of infectious diseases in humans and animals. Moreover, they are used in livestock to increase meat production by preventing infections or outbreaks of diseases and promoting growth at a global scale. The production of antibiotics is still increasing, and the total annual usage has reached from 100,000 to 200,000 tons worldwide (Gelband et al., 2015). Between 2000 and 2015 antibiotic consumption in 76 countries around the world, expressed in defined daily doses (DDDs), increased 65% and, in 2015, reached 42 billion DDDs. Among high-income countries, the leading consumers of antibiotics in 2015 were the United States, France, and Italy. Leading consumers of antibiotics between low and middle-income countries were India, China, and Pakistan (Klein et al., 2018). It has been predicted that in 2030 global antibiotics consumption will be 200% higher than in 2015, with the greatest increase coming from low and middleincome countries. There are significant differences in trends in the antibiotic consumption in European countries. According to the Antimicrobial Consumption-Annual Epidemiological Report for 2016 published by the European Center for Disease Prevention and Control (ECDC), a statistically significant trend of increasing antibiotics usage was observed for Greece and Spain from 2012 to 2016, while over the same time period a statistically significant decreasing antibiotics usage trends were observed for Finland, Luxembourg, Norway and Sweden (ECDC, 2018). The most prescribed antibiotic classes in the US and EU are penicillins, macrolides, cephalosporins, and fluoroquinolones (CDC Centers for Disease, 2015;ECDC, 2018). More detailed information about consumption of different antibiotics in some countries of the EU and US was presented by Singer et al. (2016).
Despite their benefits, a continuous release of antibiotics into the environment and their potential adverse impact on living organisms is of great concern (Fatta- Kassinos et al., 2011;De la Torre et al., 2012;Larsson, 2014;Barra Caracciolo et al., 2015;Brandt et al., 2015). Because the majority of antibiotics are not completely metabolized in the bodies of humans and animals, a high percentage of administered drugs is discharged into water and soil through municipal wastewater, animal manure, sewage sludge, and biosolids (nutrient-rich organic materials resulting from the treatment of sewage) that are frequently used to irrigate and fertilize agricultural lands (Bouki et al., 2013;Daghrir and Drogui, 2013;Wu et al., 2014) (Figure 2). It has been reported that 75-80, 50-90, and 60% of the doses of tetracyclines, erythromycin, and lincomycin, respectively, are excreted in urine and feces (Kumar et al., 2005a;Sarmah et al., 2006). Reported antibiotic concentrations in wastewater vary significantly and range from nanograms to micrograms per mL (Gulkowska et al., 2008;Michael et al., 2013;Kulkarni et al., 2017). Though some wastewater treatment processes can degrade antibiotics, there is notable variability in antibiotic removal rates. This can be attributed to differences in treatment processes, such as nature of influent, treatment plant capacity, and the type of technology used (Forsberg et al., 2012;Wu et al., 2014).
The concentrations of antibiotic residues in manure, sewage sludge, biosolids, and soil show large variations ( Table 2) and depend on the type of drug, metabolism of the drug in animals, the duration of treatment and the time of sampling relative to the treatment period. Tetracyclines have the highest concentrations and are most frequently reported antibiotic residues in manure (Pan et al., 2011;Chen et al., 2012;Massé et al., 2014). Other groups of antibiotics with considerable concentrations in manure are fluoroquinolones Van Doorslaer et al., 2014) and sulfonamides (Martínez-Carballo et al., 2007). Among the macrolide antibiotics, the highest concentration in manure was measured for tylosin (Dolliver et al., 2008). Compared to manure, biosolids contain much lower amounts of antibiotics (Jones-Lepp and Stevens, 2007) (Table 2).
Scientists have long been aware of potential problems from the presence of antibiotics in soil. Determined antibiotic concentrations in soil matrices have ranged from a few nanograms to milligrams per kg of soil ( Table 2). The highest concentrations are usually found in areas treated with manure or used for livestock (Kay et al., 2004;Zhou et al., 2013a;Hou et al., 2015;DeVries and Zhang, 2016). The concentrations of oxytetracycline and chlortetracycline in some agricultural lands may reach extremely high levels, whereas the concentrations of ciprofloxacin, norfloxacin, and tetracycline are typically significantly lower. Accurate quantification of antibiotics and their transformation products in the soil is of the utmost importance and requires advanced analytical methods, such as high-performance liquid chromatography with tandem mass spectrometry (HPLC/MS) (Aga et al., 2016).
The aim of this review is to evaluate recent literature on (1) the degradation of antibiotics in soil and (2) their impact on the microbial community function, (3) the structural and genetic diversity as well as (4) the abundance and diversity of ARGs in soil.
As many authors have suggested that environmental transformation or degradation depends on the molecular structure (Figure 1) and physicochemical properties ( Table 1) of antibiotics are the most important properties governing the fate of different antibiotics in soils (Thiele-Bruhn, 2003;Sassman and Lee, 2007;Crane et al., 2010;Leal et al., 2013;Awad et al., 2014;Pan and Chu, 2016;Zhang et al., 2017a). Since many abiotic and biotic factors affect the degradation of antibiotics, different groups of these pharmaceuticals differ in their rates of degradation in soils, as is evidenced by the large range of half-lives in soil, between <1 and 3,466 days (Figure 4). For example, Braschi et al. (2013) found that amoxicillin was easily degradable, with a half-life of 0.43-0.57 days. Similarly, a study by Liu et al. (2014) revealed that chlortetracycline (10 mg/kg soil) was quickly degraded in soil with a DT50 value <1 day. In contrast, an outdoor mesocosm study showed long-term persistence of azithromycin, ofloxacin, and tetracycline in soils,  Hamscher et al., 2002;Thiele-Bruhn, 2003;Karci and Balcioglu, 2009;Hu et al., 2010;Liu et al., 2016;Tasho and Cho, 2016;Pan and Chu, 2017b;Łukaszewicz et al., 2018 dw, dry weight.
with half-lives of 408−3466, 866-1733, and 578 days, respectively (Walters et al., 2010). Detailed information about the degradation rates of antibiotics and obtained DT50 values in different soils is presented in Table 3. Notable, even for antibiotics within the same groups or, in some cases for particular antibiotics, DT50 values differ significantly. Observed differences in persistence or similar compounds are probably due to variations in soil composition, doses of antibiotics and conditions used in these studies. However, based on a review of the literature, it can be concluded that fluoroquinolones, macrolides, and tetracyclines are characterized by high DT50 or half-life values (Figure 4). The sorption coefficient (K oc ) is a very important parameter for estimating environmental distribution and environmental exposure level of antibiotics. As was proposed by Crane et al. (2010), antibiotics with values of K oc > 4,000 l/kg are non-mobile and degradable to a very low degree in soils (very persistent) and the length of time needed for the degradation of 50% of an initial dose is >60 days. In contrast, antibiotics characterized by values of K oc <15 l/kg are highly mobile and are more easily degraded; these can be classified as compounds with low persistence in soils (DT50 < 5 days). The relatively low persistence of some antibiotics in soils may be related to their low affinity to various organic and non-organic soil components. In contrast, due to their properties tetracyclines, fluoroquinolones, macrolides, and sulfonamides, bind strongly to soil components and form stable residues. For example, Hammesfahr et al. (2008) showed that the recovery rate of sulfadiazine on day 1 of an experiment was 27 and 45% of the initial antibiotic dosages of 10 and 100 mg/kg soil, respectively. Similarly, sulfamethazine applied at concentrations of 20 and 100 mg/kg soil strongly absorbed to soil components and measured concentrations of antibiotic were 62.1 and 31.5 µg/kg soil at the beginning of the experiment and, 255.5 and 129.8 µg/kg soil on day 56, respectively (Awad et al., 2016). Due to the binding of antibiotics to soil particles or the occurrence in a form of complexes, sorbed antibiotics often cannot be detected and may lose their antibacterial activity (Kümmerer, 2009). The adsorption and desorption of antibiotics are also associated with other soil parameters, such as pH and water content. For example, sulfonamides show a decrease in sorption with an increase of soil pH, whereas the binding of macrolides by soil components increases with a decrease of pH. In the first case, the sorption behavior is consistent with changes in the fraction of ionization of the sulfonamide as it converts from its cationic form to the neutral and anionic forms. The behavior of sulfonamides contrasts with tetracyclines and fluoroquinolones, which interact with soil primarily through cation exchange, surface complexation and cation bridging sorption mechanisms. In general, decreases in pH resulted in increased sorption of the cationic forms of antibiotics, suggesting that electrostatic interactions are the favored sorption mechanism for sulfonamides and macrolides (Schauss et al., 2009;Hammesfahr et al., 2011;Kong et al., 2012;Teixidó et al., 2012;Sittig et al., 2014;Wegst-Uhrich et al., 2014;Fernández-Calviño et al., 2015;Wang and Wang, 2015;Liu et al., 2017).
In addition to abiotic processes, microbial degradation may contribute to disappearance of antibiotics in soil. Some bacteria that degrade antibiotics have been isolated from antibioticscontaminated soils. For example, strains belonging to the genera Microbacterium (Topp et al., 2013), Burkholderia (Zhang and Dick, 2014), Stenotrophomonas (Leng et al., 2016), Labrys (Mulla et al., 2018), Ochrobactrum (Zhang et al., 2017b;Mulla et al., 2018), and Escherichia (Mulla et al., 2018;Wen et al., 2018) were capable of degrading sulfamethazine, penicillin G, tetracycline, erythromycin and doxycycline in liquid cultures, respectively. Other bacteria belonging to the genera Klebsiella (Xin et al., 2012), Acinetobacter, Escherichia , Microbacterium (Kim et al., 2011), Labrys (Amorim et al., 2014), and Bacillus (Rafii et al., 2009;Erickson et al., 2014) that were capable of degrading chloramphenicol, sulfapyridine, sulfamethazine, ciprofloxacin, norfloxacin, and ceftiofur have been isolated from patients, sediments, sludge, animal feces, and seawater. In particular, a Microbacterium sp. exhibited degradation of sulfamethazine in soil and, when introduced into agricultural soil, increased the mineralization of that antibiotic by 44-57% (Hirth et al., 2016). The central role of microorganisms in antibiotic degradation or transformation in soil has been confirmed by results of many studies carried out in sterile and non-sterile soils. As depicted in Table 3, the half-life or DT50 values of antibiotics were much lower in soils with autochthonous microorganisms compared to those obtained from sterilized soils. For example, Pan and Chu (2016) showed that when applied to soil at a dosage of 0.1 mg/kg soil, erythromycin disappeared faster in non-sterile soil compared to sterile soil, with DT50 of 6.4 and 40.8 days, respectively. Sulfachloropyridazine applied at 10 mg/kg soil was degraded almost three times faster in soils with autochthonous microorganisms (half-life 20-26 days) compared to sterile soils (half-life 68-71 days) (Accinelli et al., 2007). Zhang et al. (2017a) also showed that microbial activity plays a major role in the biotransformation of sulfadiazine in soil s, with a DT50 of 8.48, 8.97, and 10.22 days (non-sterile soil) and 30.09, 26.55, and 21.21 days (sterile soil) for concentrations of 4, 10, and 20 mg/kg soil, respectively. A similar phenomenon has also been observed for other antibiotics, such as norfloxacin (Pan and Chu, 2016), sulfamethoxazole (Lin and Gan, 2011;Srinivasan and Sarmah, 2014;Zhang et al., 2017a), sulfamethazine (Accinelli et al., 2007;Pan and Chu, 2016), and tetracycline (Lin and Gan, 2011;Pan and Chu, 2016).
Special attention should be paid to the analytical methods for extraction and determination of antibiotics residue when evaluating and comparing degradation experiments. Some techniques may not always be capable of differentiating between degradation and sorption, and insufficient or improper extraction procedures may result in incorrect interpretation of the fate of antibiotics in the soil as antibiotics that are tightly bound to soil particles could be erroneously considered to have been transformed or degraded.  (Table 3) irrespective of the type of soil and the concentration of antibiotic used. The absence of bars in some cases means no data available.

Factors Affecting the Degradation of Antibiotics in Soil
Degradation of antibiotics depends not only on the catabolic activity of soil microorganisms, but also, to a large extent, on the properties of soil, i.e., organic matter content, pH, moisture, temperature, oxygen status, and soil texture ( Table 3). For example, Li et al. (2010a) observed differences in the degradation rate of oxytetracycline in two agricultural soils with different characteristics. The calculated DT50 for this antibiotic reached values of 30.2 and 39.4 days for soils with a low or high organic carbon content, respectively. Soil type was also found to significantly affect antibiotics degradation, as demonstrated by Koba et al. (2017) who studied the degradation rate of clindamycin, sulfamethoxazole, and trimethoprim in 12 different soils. The authors characteristics, 44-98, 25-99, and 13-84% of clindamycin, sulfamethoxazole and trimethoprim (2 mg/kg soil), respectively, were degraded within 61 days during the experiment. A similar dependence of the rate of antibiotics degradation on soil type of has also been found for chlortetracycline (Halling-Sørensen et al., 2005;Li et al., 2010a), sulfachloropyridazine (Accinelli et al., 2007), sulfadiazine (Förster et al., 2009;Sittig et al., 2014), sulfamethoxazole (Accinelli et al., 2007;Lin and Gan, 2011;Srinivasan and Sarmah, 2014), and tetracycline (Li et al., 2010a). Another study revealed a significant decrease in the rate of chlortetracycline, erythromycin, and tylosin degradation introduced at a dosage of 5.6 mg/kg soil into a sandy loam soil. These degradation experiments showed that 56, 12, and 0%, 100, 75, and 0%, or 100, 100, and 60% of the initial concentration of chlortetracycline, erythromycin, and tylosin were degraded at temperatures of 30, 20, and 4 • C, respectively, within 30 days (Gavalchin and Katz, 1994). In addition, Srinivasan and Sarmah (2014) showed that a lower temperature resulted in a reduced decomposition rate of sulfamethoxazole, irrespective of the soil depth. Some published data have also shown the influence of oxygen in soils on the degradation of antibiotics. For example, Pan and Chu (2016) showed that the half-lives for erythromycin, norfloxacin, sulfamethazine, and tetracycline, all of which had been applied at 0.1 mg/kg soil, increased from 6.4, 2.9, 24.8, and 31.5 days under aerobic conditions to 11.0, 5.6, 34.7, and 43.3 days under anaerobic conditions, respectively. Anaerobic conditions caused an increase in the half-life for sulfamethoxazole by seven days compared to 11.4 days for aerobic degradation of this antibiotic. In turn, no significant effects of incubation conditions on degradation rate were shown for trimethoprim, for which the calculated half-life was about 26 days under both aerobic and anaerobic conditions (Lin and Gan, 2011). The degradation of antibiotics may also be influenced by the pH and moisture content of soils. Weerasinghe and Towner (1997) showed that the half-life of for virginiamycin was negatively correlated with the pH of soils and ranged from 87 to 173 days for different agricultural soils. A study by Wang et al. (2006) revealed that the half-life of sulfadimethoxine in soil decreased from 10.4 days to 6.9, and again to 4.9 days when the soil moisture content was increased from 15, to 20, and 25%, respectively. A similar effect of temperature was shown in the case of the degradation of norfloxacin (Yang et al., 2012).
Degradation of antibiotics strongly depends on their concentration in soil. Increasing dosages of ciprofloxacin (from 1 to 5 and 50 mg/kg soil) led to a reduction of degradation from 75, to 62, and 40% within 40 days (Cui et al., 2014). A similar tendency was also shown by Demoling et al. (2009), who applied sulfamethoxazole at 500, 20, and 1 mg/kg soil and observed a reduction of its removal from 153, to 1.5, and 0.04 mg/kg after 5 weeks. The same trend was reported in degradation of azithromycin, ofloxacin, tetracycline (Walters et al., 2010), sulfadimethoxine (Wang et al., 2006), chlortetracycline (Fang et al., 2016), and SDZ (Zhang et al., 2017a). These results suggest that high concentrations of antibiotics may prolong their persistence in soils, due to the inhibition of the activity of soil microorganisms Pan and Chu, 2016). However, application of unrealistically high concentrations tends to overestimate half-lives, which may not reflect realistic situations (Pan and Chu, 2016). The history of antibiotics application also plays a role in the further disappearance of antibiotics in soils. Repeated application of the macrolide antibiotics clarithromycin and erythromycin into soil resulted in a decrease of the DT50 values from 36.48 and 69.93 days to 15.85 and 4.36 days (soil with a history of clarithromycin and erythromycin application at 0.1 mg/kg soil) and to 9.51 and 0.94 days (soil with a history of clarithromycin and erythromycin application at 10 mg/kg soil), respectively . This phenomenon of pre-adaptation of soil microorganisms was also found for the degradation of chlortetracycline, sulfamethazine and tylosin (Topp et al., 2013).
Most published data have also indicated that the addition of different organic compounds, such as manure, biosolids, slurry, sludge, and compost, into soils may contribute to changes in the rate of antibiotics degradation ( Table 3). Sassman and Lee (2007) showed that the addition of manure (20 mg/kg soil) increased the half-life of lasalocid from 3.6 to 4.3 days. In contrast, Yang et al. (2012) showed that norfloxacin (10 mg/kg soil) was degraded faster (half-life 24-39 days) in soil with manure (3-9%) compared to the control soil (half-life 48 days). In turn, Li et al. (2010a) found no significant effects from the addition of manure on degradation of chlortetracycline in soil.

IMPACT OF ANTIBIOTICS ON SOIL MICROORGANISMS
Soil microorganisms perform many vital processes and participate in the maintenance of soil health and quality. They play a crucial role in organic matter turnover, nutrients release, and stabilization of the soil structure and ensure soil fertility. Moreover, many microorganisms act as biological control agents by inhibiting the growth of pathogens (Varma and Buscot, 2005). The homeostasis of soil may be disturbed by biotic and abiotic factors, including bacteriophages, predation, competition, pesticide, heavy metals, toxic hydrocarbons, and antibiotics (Cycoń et al., 2011;Piotrowska-Seget, 2015, 2016;Sułowicz and Piotrowska-Seget, 2016;Xu et al., 2016;Chen et al., 2017;Wepking et al., 2017;Orlewska et al., 2018a). The high antimicrobial activity of antibiotics in soil should differentially inhibit the growth of soil microorganisms and thus influence the soil microbial community composition, which may result in alterations of the ecological functionality of the soil (Kotzerke et al., 2008;Keen and Patrick, 2013;Molaei et al., 2017) (Figure 3).
There is abundant data on the impact of antibiotics on microorganisms and on soil processes mediated by bacteria and fungi. Both the effects of antibiotics on individual microbial populations as well as on the composition of entire microbial communities have been documented. A wide range of methods based on measurements of parameters that reflect the activity and abundance of total microbial communities, such as soil organic matter turnover, respiration, and microbial biomass have been used to characterize such effects. Other methods have focused on selected microbial processes such as nitrification, denitrification, sulfate, and iron reduction, methanogenesis and the activity of enzymes responsible for C, N, and P turnover (Brandt et al., 2009;Chen et al., 2013;Cui et al., 2014;Liu et al., 2015;Ma et al., 2016). A large number of studies have instead focused on assessing changes in microbial diversity, using metagenomics or 16S rRNA Frontiers in Microbiology | www.frontiersin.org gene amplicon sequencing, as well as analysis of phospholipid fatty acids (PLFAs) isolated from soil (Zielezny et al., 2006;Hammesfahr et al., 2008;Reichel et al., 2015;Xu et al., 2016). In the following sections, the methods and parameters utilized to assess effects of antibiotics on the function and structure of soil microbial communities are presented and discussed (Figure 3).

Soil Processes
Many studies have found that even low concentrations (below the MIC) of antibiotics influence various soil processes mediated by microorganisms ( Table 4). A significant decrease in soil respiration (SR) was reported for soils containing sulfamethoxazole, sulfamethazine, sulfadiazine, and trimethoprim (Kotzerke et al., 2008;Liu et al., 2009), however, this effect was transient and depended on the disappearance rate of these antibiotics (DT50 2-5 days). The authors concluded that the effects of antibiotics on SR diminished due to decreased bioavailability of the antibiotic. In a study by Wepking et al. (2017), the response of microbial respiration to cephapirin, tetracycline, or erythromycin was dependent on both the type of antibiotic and the exposure of soil to dairy cattle manure. In other studies, no obvious effects of tetracycline, chlortetracycline, oxytetracycline, sulfadiazine, sulfapyridine, or tylosin on SR were observed (Thiele-Bruhn and Beck, 2005;Liu et al., 2009;Toth et al., 2011). Nitrification and/or denitrification rates were also influenced by antibiotic exposure, and the effects were strongly dependent on the type of antibiotic and the length of exposure. For example, sulfadimethoxine inhibited soil nitrification; however, this effect was only observed on some sampling days and only for a high sulfadimethoxine treatment (Toth et al., 2011). A decrease in the nitrification rate caused by a high oxytetracycline concentration (30 mg/kg) and sulfadiazine (100 mg/kg) in a single application was also observed by Ma et al. (2016) and Kotzerke et al. (2008), whereas ciprofloxacin and norfloxacin were reported to stimulate the rate of nitrification in a soil microcosm, but only at the lowest concentration of antibiotic (1 mg/kg soil) (Yang et al., 2012;Cui et al., 2014). A study by DeVries et al. (2015) revealed that low doses of sulfamethoxazole, sulfadiazine, narasin, or gentamicin (500 µg/kg soil) inhibited denitrification, but dosages <1 µg/kg soil actually stimulated the process transiently. Toth et al. (2011) did not observe any effects of monensin and chlortetracycline, on soil nitrification at concentrations of 0.01-0.1 and 0.0003-0.3 mg/kg soil, respectively.
Antibiotics may also change the turnover rate of iron in soil. Sulfadiazine and monensin blocked Fe(III) reduction in soil over periods ranging from a few days to the end of a 50-day experiment (Toth et al., 2011). Strong inhibition of Fe(III) reduction was also observed in soil contaminated with sulfamethoxazole and oxytetracycline at concentrations >10 mg/kg soil (Molaei et al., 2017). Thiele-Bruhn and Beck (2005) calculated the EC50 value of sulfapyridine for the microbial reduction of Fe(III) in two different soils at 12.4 and 0.310 mg/kg soil. It should be noted that the lack of standardized tests hinders comparisons that would lead to general conclusions about the effects of antibiotics on biogeochemical cycles and the turnover of iron.

Enzyme Activities
Specific enzyme activity is considered to be a useful indicator of the response of microorganisms to stress caused in the soil by antibiotics (Unger et al., 2013;Liu et al., 2014Liu et al., , 2015Ma et al., 2016) (Table 5). Enzyme activity indicates the potential of microbial communities to carry out biochemical processes that are essential to maintain soil quality. Any application of a toxicant that might affect the growth of soil microorganisms can induce alterations in the general activity of enzymes, such as dehydrogenases (DHAs), phosphatases (PHOSs), and urease (URE) (Gil-Sotres et al., 2005;Hammesfahr et al., 2011;Cycoń et al., 2016a). Inhibition of DHAs and URE was observed in soil amended with tetracycline at a concentration of 1 µg/kg soil, however this dosage did not affect acid phosphatase (PHOS-H) activity. Sulfamethazine applied at 53.6 µg/g soil had a significant short-term negative impact on the activities of DHAs and URE (Pinna et al., 2012). Inhibition of DHAs and arylsulphatase activities with increasing concentration of oxytetracycline at 1 to 200 mg/kg over 7 weeks was also reported by Chen et al. (2013). Benzylpenicillin, tylosin and sulfadiazine inhibited soil DHAs and PHOSs from 35 to 70% compared to the non-antibiotic amended control (Reichel et al., 2014;Akimenko et al., 2015). The results obtained by Unger et al. (2013) showed that the application of oxytetracycline or lincomycin (50 and 200 mg/kg soil) resulted in a temporary decrease in DHA activity in soil. A temporary increase, followed by a decrease, in DHA activity was found in soils containing chlortetracycline (1, 10, and 100 mg/kg soil) by Liu et al. (2015). In another study, DHA activity was unaffected by sulfapyridine or oxytetracycline, even at a dosage of 1 mg/kg soil (Thiele-Bruhn and Beck, 2005).
Various effects on the activity of PHOSs from different antibiotics applied to soil have been shown. For example, six antibiotics, i.e., chlortetracycline, tetracycline, tylosin, sulfamethoxazole, sulfamethazine, and trimethoprim at dosages of 1-300 mg/kg soil inhibited the activity of PHOS-H . In contrast, results from Yang et al. (2009) indicated that only alkaline phosphatase (PHOS-OH) was sensitive to the application of oxytetracycline, with a 41.3% decline in enzyme activity at a concentration of 10 mg/kg soil and a further decrease of 64.3-80.8% when the dose of the antibiotic exceeded 30 mg/kg. Ma et al. (2016) found that OTC at dosages up to 30 mg/kg soil had no effect on the activity of neutral PHOS over a 120-day experimental period.
The inhibition of enzyme activity in antibiotic-treated soils may be related to inhibition of growth or death of sensitive microorganisms (Boyd and Mortland, 1990;Gianfreda et al., 1994;Alef, 1995;Marx et al., 2005). In turn, the increased activity of enzymes under antibiotics pressure may result from the ability of many bacteria to subsist on antibiotics as a carbon source (Dantas et al., 2008). We can speculate that enzymes produced by such bacteria compensate for the negative effects of antibiotics on enzyme activity by increasing the activity of the antibioticresistant microbial community. Similarly, an increase in the activity of some enzymes in soils treated with different antibiotics might be related to the capability of some microorganisms to make use of the antibiotics in their metabolisms, thus resulting in an increase in the abundance of some soil microorganisms and enzyme production. In addition, the presence of some antibiotics in soil may cause an overgrowth of fungi, which are generally less sensitive to antibiotics than bacteria. Fungi are a major producer of enzymes in soils and so may be responsible for observed increases in enzyme activity (Tabatabai and Bremner, 1969;Westergaard et al., 2001;Hammesfahr et al., 2011;Ding et al., 2014).

Functional Capacity Evaluation
Functional microbial diversity reflects the ability of an entire microbial community to utilize a suite of substrates (Zak et al., 1994). The response of microorganisms to the presence of antibiotics in soils, expressed as the community level physiological profile (CLPP), can be evaluated using the Biolog (EcoPlates TM ) or MicroResp TM methods, which are based on the utilization of 31 or 15 carbon sources, respectively, by organisms present in soil extracts (Toth et al., 2011;Ma et al., 2014 (Table 6). Although the exact numbers and taxonomic identities of the bacterial species responsible for the Biolog reactions remain uncertain, patterns of functional diversity indicate how biodiversity affects specific ecosystem functions (Laureto et al., 2015). For determination of CLPP, the biodiversity (Shannon H') and average well-color development (AWCD) indices have been used as indicators of changes in the catabolic potential of soil microorganisms exposed to antibiotics. For example, chlortetracycline (1 and 10 mg/kg soil) caused a decrease in AWCD values, showing low catabolic potential in the microbial community, however, this effect was seen only at the beginning of the experiment. After 35 days, irrespective of the frequency of chlortetracycline application, the AWCD values gradually recovered to the control level (Fang et al., 2014(Fang et al., , 2016. A slight inhibitory effect on microbial activity (expressed as the H ′ index) in soil was observed over a range from 1 to 300 mg oxytetracycline /kg soil. Even lower concentrations of this antibiotic (0.5 to 90 mg/kg of soil) significantly decreased the functional activity (expressed as AWCD) of the soil microbial communities (Kong et al., 2006). Sulfonamide antibiotics, such as sulfamethoxazole and sulfamethazine, can also alter the activity of microbial populations, however, they were observed to only have short-term detrimental effects (Demoling et al., 2009;Pinna et al., 2012;Pino-Otín et al., 2017). In turn, the application of chlortetracycline or sulfadimethoxine to soil did not significantly change AWCD, whereas monensin slightly increased the value of the H' index (Toth et al., 2011). In many studies, changes in the preferential degradation of groups of substrates in Eco-plates, e.g., amino acids, carbohydrates, carboxylic acids, aromatic acids, and miscellaneous, by microorganisms were observed over the course of an experimental period. Xu et al. (2016) found that a high sulfadiazine concentration decreased utilization rates of amino acids, carbohydrates, carboxylic acids, and aromatic acids by bacteria present in a soil suspension. A significant decrease in the utilization of carbohydrates and miscellaneous substrates was also observed in soil that had been treated with sulfamethoxazole (Liu et al., 2012a). However, this effect was only observed 7 days after the antibiotic application; by day 21 substrate utilization had increased compared to the first sampling day. The negative impact of antibiotics on CLPP may be mitigated by the enrichment of soil with organic matter from animal wastes, antibiotic-free manure or plant wastes (Demoling et al., 2009;Pinna et al., 2012;Liu et al., 2014). Contrarily, Wang et al. (2016) found that doxycycline application at 10 mg/kg soil increased utilization of the substrates. CLPP patterns do not always show changes in the functional diversity of microorganisms in response to antibiotic. In such situations, the lack of observable effects in CLPP might be because indices expressing total microbial activity do not detect possible changes caused by low antibiotic concentrations. Moreover, repeated application of antibiotics leads to the adaptation of bacteria to these compounds and proliferation of antibiotic-resistant microorganisms. Another limitation is that the Biolog technique measures only the activity of catabolically active cells and omits non-culturable populations as well as microorganisms in a dormant state. Moreover, fast-growing microorganisms are mainly responsible for effects observed by this test (Floch et al., 2011).

PLFA Fingerprinting
The PLFA method is a rapid tool for assessing the biomass and composition of microbial communities in soil, because various phospholipid-derived fatty acids (PLFAs) are indicative of species or microbial groups in soil. In addition, the ratios of biomass of bacteria/fungi and Gram-positive/Gram-negative bacteria are often used to evaluate the response of microorganisms to organic pollutants (Frostegård et al., 1993(Frostegård et al., , 2011. In general, application of antibiotics into soil show strong and dose-dependent effects on the microbial structure expressed as the total PLFA biomass ( Table 7). For example, Chen et al. (2013) found a significant increase in bacterial and fungal PLFA biomass 7 weeks after oxytetracycline contamination of 1 and 15 mg/kg soil, but a strong decrease in the case of 200 mg antibiotic/kg treatment. A decrease in microbial PLFA biomass content was also found in soils that had been treated with penicillin G (10 and 100 mg/kg soil) and tetracycline (100 mg/kg soil) (Zhang et al., 2016a). Simultaneously, the ratios of Gram-negative/Grampositive bacteria increased, indicating that the bacteria that were resistant to antibiotics in tested soils were more likely to be Gramnegative. A study by Unger et al. (2013) found only short-term effects (35 days) from lincomycin or oxytetracycline applications (5-200 mg/kg soil), suggesting that the resilience of soil microbial communities toward effects of these antibiotics.
An increase in total PLFA biomass was also observed in soils treated with sulfadiazine and difloxacin in slurry from pigs medicated with these antibiotics (Reichel et al., 2013). Influences from the difloxacin-slurry on PLFA biomass were observed at the beginning of the experiment (7 and 14 days), but only on the last sampling day (63 day) in the sulfadiazine-slurry. The authors hypothesized that the delayed effect of sulfadiazine was a result of continuous remobilization of antibiotic residues bound to soil particles by microorganisms. Hammesfahr et al. (2008) and Kotzerke et al. (2011), in contrast, found no changes in soil microbial communities upon addition of the above-mentioned antibiotics into soil spiked with slurry.
Several studies showed changes in the soil fungal PLFA signatures upon amendment with antibiotics. For example, the content of fungal marker fatty acids (18:1ω9 and 18:2ω6,9) significantly increased in alfalfa-amended soil and both alfalfaand manure-amended soils with 20 and 500 mg sulfamethoxazole /kg of soil (Demoling et al., 2009). Gutiérrez et al. (2010) mixed three sulfonamides, sulfadimethoxine, sulfamethoxazole, and sulfamethazine, and observed a general increase in the proportion of fungal PLFAs among the total soil biomass PLFAs. A shift in the soil microbial community toward more fungi because of sulfapyridine application was also noted by Thiele-Bruhn and Beck (2005). The increase in the abundance of fungi may ultimately result in a decline in the productivity and quality of soils and agricultural products (Ding and He, 2010).
The PLFA method, previously used extensively for determination of microbial community composition, has been largely replaced by techniques based on the analysis of nucleic acids extracted from soil. However, PLFA-based analyses remain an adequately sensitive, efficient way to rapidly screen whether a microbial community has been affected by antibiotics.

Phylogenetic Analyses
Many authors have used PCR-dependent DNA fingerprinting techniques such as a terminal restriction fragment length polymorphism (T-RFLP) or denaturing gradient gel electrophoresis (DGGE) to evaluate the impact of antibiotics on diversity of soil microorganisms (Müller et al., 2002;Reichel et al., 2014;Chessa et al., 2016a;Orlewska et al., 2018a,b). Most of these studies target 16S rRNA genes and show altered diversity upon antibiotics applications. These changes may be explained by the disappearance of sensitive bacteria and outgrowth of resistant bacteria present in the antibiotic-contaminated soils ( Table 8). For example, the application of tylosin (2 mg/kg soil) strongly modified the DGGE pattern, reducing the number of bands observed on days 15 and 22 of the experiment compared to non-exposed soil. However, on day 33 the difference was smaller, yet detectable (Westergaard et al., 2001). In a study by Müller et al. (2002), the DGGE pattern of 16S rDNA in tylosin-treated soil differed slightly to the diversity to the control soil, although colony morphology typing and potential of microbial communities to utilize selected substrates did not reveal any differences. Cycoń et al. (2016a) found that vancomycin dosed to soil at 10 mg/kg resulted in quantitative changes in microbial community patterns, suggesting selection exerted by the antibiotic. The changes in DGGE patterns suggest a decrease in populations of a species or group of species; revealing that some bacterial species were more sensitive to vancomycin than others were. This was corroborated by a decline in the values of H' and S (richness) indices. Another study showed significant effects of chlortetracycline on the microbial community that lasted for no longer than 1 week after the application of the antibiotic as the communities recovered over time until the end of the experiment (42 day) (Nelson et al., 2011). Contrarily, using PCR- DGGE Zielezny et al. (2006) showed that the application of sulfadiazine and chlortetracycline at concentrations of 1, 10, and 50 mg/kg soil had no significant effect on the structure of the soil microbial community. In another study, changes in DGGE profiles showed that sulfadiazine (10 and 100 µg/g soil) applied to soil in manure altered bacterial diversity (Hammesfahr et al., 2008). Although the DGGE profiles proved the impact of manure and sulfadiazine on the total soil bacterial communities, these effects were less distinct for pseudomonads and β-Proteobacteria. This observation may be explained by the resistance of many strains to sulfonamides, which may be only indirectly affected by sulfadiazine. Genetic changes within the β-Proteobacteria and the pseudomonads, expressed by the appearance or loss of a band in the DGGE profiles from manure, and sulfadiazineamended soils were also reported by Reichel et al. (2014). However, statistical analysis showed that the β-Proteobacteria significantly responded to the time and moisture regimes, whereas the pseudomonads responded to the factors of time and treatment. The effects of sulfadiazine-containing manure on microbial diversity resulted in a low stability of soil bacterial communities and dominance of taxa that are known to contain human pathogens, such as Gemmatimonas, Leifsonia, Devosia, Clostridium, Shinella, and Peptostreptococcus .
No apparent effects on microbial diversity in response to streptomycin and difloxacin were observed in similar studies (Shade et al., 2013;Jechalke et al., 2014a).

DIVERSITY OF ANTIMICROBIAL RESISTANCE GENES
Studies are increasingly focusing on genes for antimicrobial resistance itself, which can be followed by metagenomic sequencing approaches. Resistance determinants present in the soil, defined as the soil resistome, include genes that confer antibiotic resistance to pathogenic and non-pathogenic species present in the environment, as well as proto-resistance genes that serve as a source of resistance elements (D'Costa et al., 2006(D'Costa et al., , 2007Perry et al., 2014). The potential transfer of antibiotic resistance among microbial populations in soil and the risk of animal and human infection are major concerns (Figure 5). Special attention is given to the detection and quantitation of ARGs, identifying resistance mechanisms, discovery of novel enzymes with unexpected activities, and the impact of pollutants and agricultural practices on the abundance of ARGs in soils. Resistance to antibiotics has been documented even in bacteria inhabiting pristine soils never exposed to these compounds, for example genes conferring resistance to common antibiotics (βlactams, fluoroquinolones, and tetracyclines) were found in the Tibet . Interestingly, a tiny fraction of ARGs located on MGEs has been discovered, indicating a low, but real, potential for these ARGs to be transferred among bacteria  . Additionally, some ARGs that have been discovered encode efflux pumps that are completely different from previously known efflux pumps. A large number (177) of ARGs encoding mostly single or multi-drug efflux pumps conferring resistance to aminoglycosides, chloramphenicol and β-lactams were reported in pristine Antarctic soils (Van Goethem et al., 2018). A low number and diversity of ARGs was detected in glacial soil and permafrost as compared to environments highly impacted by human activities in a study by Zhang et al. (2018). It has been proposed that ARGs in pristine environments most likely represent functional historical genes conferring resistance to natural antibiotics. Antibiotic resistance appeared to be transferred vertically over generations with limited to no horizontal transfer of ARGs between species (Van Goethem et al., 2018). Diversity and abundance of ARGs is strongly influenced by agricultural practices (Wepking et al., 2017;Xiong et al., 2018). It has been documented that the spreading of manure from animals treated with different antibiotics on fields affects the abundance and diversity of ARGs (Su et al., 2014;Kyselková et al., 2015). However, it has also been documented that organic matter content, pH, and the history of soil management are important factors influencing the fate and abundance of ARGs in the soil (Cermák et al., 2008;Popowska et al., 2012).
The genes conferring resistance to minocycline, tetracycline, streptomycin, gentamycin, kanamycin, amikacin, chloramphenicol, and rifampicin in soils subjected to manure application accounted for ∼70% of the total ARGs. More than half of the ARGs shared low similarity, <60%, with the sequences of their closest proteins in GeneBank (Su et al., 2014). Among the ARGs studied (ampC, tetO, tetW, and ermB), the average abundance of ampC and tetO in the manure-treated soil was 421 and 3.3% greater in comparison with a non-treated control soil, respectively (Wepking et al., 2017). A broad-spectrum profile of ARGs has been found in 5 paddy soils of South China (Xiao et al., 2016). Based on Antibiotic Resistance Genes Database and next generation sequencing, 16 types of ARGs corresponding to 110 ARG subtypes were identified. Multidrug resistance genes dominated in all soils (up to 47.5%), followed mainly by specific resistance genes for acriflavine, macrolide-lincosamide-streptogramin and bacitracin. Efflux pumps, antibiotic deactivation, and cellular protection were three major resistance mechanisms inferred from the resistome characterized Xiao et al. (2016).
An increase in the number and proportions of ARGs in the bacterial community was also shown by Nõlvak et al. (2016), who studied the effect of animal manure on the rate of the dissemination of ARGs (sul1, tetA, blaCTX-M, blaOXA2, and qnrS) and integron-integrase genes in soil. The abundance of genes for tetracycline resistance (tetM, tetO, and tetW) in soil that had had manure applied was significantly higher than in untreated soil (Xiong et al., 2018). It seems that an increasing abundance of ARGs resulted from inputs of ARGs already present in manure. This is because antibiotics may provide a significant selective advantage for some species or groups of antibiotic-resistant bacteria that then come to dominate the microbial community.
By contrast, many studies have shown that the dissemination of ARGs in soils is more correlated with the application of manure than with the presence of antibiotics. Though no effect on the abundance of tetG, tetO, and tetW was observed in one study, despite the presence of these genes in manure applied to the soil (Fahrenfeld et al., 2013), other studies have found an impact of manure from never antibiotic-treated animals on the number, diversity, and spread potential of ARGs (Udikovic-Kolic et al., 2014;Hu et al., 2016). This phenomenon can be explained by the stimulation of growth autochthonous bacteria and accelerated spread of ARGs by fertilizers free from antibiotics. Due to the ambiguous results of studies concerning the spread of ARGs in manure-treated soil, there is a need for future research to clearly assess the relative contribution of manure and antibiotics on microbial community structure and processes mediated by soil microorganisms.
The studies on the fate of ARGs in soil allowed not only to assess the diversity and abundance of genes of interest, but also to discover new ARGs, and described new resistance mechanisms and novel enzymes responsible for resistance of bacteria to antibiotics (Lau et al., 2017;Wang et al., 2017). Functional metagenomics for the investigation of antibiotic resistance led to the discovery of several tetracycline resistance genes from different classes in soil. Efflux pump genes, including 21 major facilitator superfamily efflux pump genes, were found to dominate the sequence libraries derived from studied soils. In addition, two new genes involved in the enzymatic inactivation of tetracycline were identified (Wang et al., 2017). Similarly, Torres-Cortés et al. (2011) described a new type of dihydrofolate reductase (protein Tm8-3; 26.8 kDa) conferring resistance to trimethoprim in soil microbiome. Using functional metagenomics they also identified three new antibiotic resistance genes conferring resistance to ampicillin, two to gentamicin, two to chloramphenicol, and four to trimethoprim in libraries generated from three different soil samples. Nine carbapenem-hydrolizing metallo-beta-lactamases (MBLs) including seven novel enzymes showing 33-59% similarity to previously described MBLs were discovered. Six originated from Proteobacteria, two from Bacteroidetes, and one from Gemmatimonadetes. MBLs were detected more frequently and exhibited higher diversity in grassland than in agricultural soil (Gudeta et al., 2016). In a study of resistome of bacterial communities in Canadian soils by functional metagenomics, 34 new ARGs with high homology to determinants conferring resistance aminoglycosides, sulfonamides and broad range of β-lactams were identified (Lau et al., 2017). High-resolution proteomics in combination with functional metagenomics resulted in the discovery of a new proline-rich peptide PPP AZI promoting resistance to various macrolides, but not to other ribosome-targeting antibiotics (Lau et al., 2017). Similarly, Donato et al. (2010) identified two novel bifunctional proteins responsible for bacteria resistance to ceftazidime and kanamycin, respectively.
Other synthetic pollutants can also influence the acquisition and maintenance of ARGs that can confer resistance not only to antibiotics but also to a number of structurally unrelated contaminants present in soil. As an example for this, ARGs were ∼15 times more abundant in polycyclic aromatic hydrocarbon (PAH) contaminated soils than in similar non-contaminated soils . It resulted from the selection by PAHs of Proteobacteria containing ARGs encoded multi-drug efflux pumps leading to simultaneously enriching of ARGs carried by them in the soils. Moreover, most of ARGs (70%) found in the PAHs-contaminated soils were not carried by plasmids, indicating a low possibility of horizontal gene transfer (HGT) between bacteria .
It has recently been reported that heavy metals are also as a factor contributing to maintenance of ARGs in the environment. A study of dairy farms showed significant correlations between the abundance of ARGs, metal resistance genes and the content of Cu and Zn in cattle feces . The role of heavy metals can be explained by cross-resistance, tolerance of bacteria to more than one antimicrobial agent (Chapman, 2003;Baker-Austin et al., 2006), and co-resistance, the localization of ARGs and genes encoding resistance to heavy metals and other antibacterial agents on mobile genetic elements (Chapman, 2003), mechanisms. Multi-drug efflux pumps, which mediate rapid extrusion of antibiotics and heavy metals from the cell decrease susceptibility toward these compounds, are an example of cross-resistance (Martinez et al., 2009). Co-resistance can occur due to the close arrangement of genes on a chromosome or extrachromosomal element, which increases the likelihood that genes are subject to combined transmission via HGT. Moreover, heavy metals not only trigger co-selection processes, but also increase the level of tolerance to antibiotics due to co-regulation of resistance genes (Baker-Austin et al., 2006). Such mechanisms may affect the spread of ARGs from environmental populations to bacteria of clinical importance even in the absence of direct antibiotics selection (Herrick et al., 2014).
Antibiotics concentrations in the environment are much lower than that used in therapeutic doses. However, even at sub-inhibitory concentrations (below the MIC), antibiotics may select for resistant phenotypes (Gullberg et al., 2011(Gullberg et al., , 2014. It has been reported that resistant bacteria can be selected at antibiotic concentrations even 100-fold below lethal concentrations. Using data on MICs from the EUCAST database for 111 antibiotics, Bengtsson-Palme and Larsson (2016) estimated that the predicted no-effect concentrations (PNECs) range from 8 ng/L to 64 µg/L. Mutants selected by antibiotics at minimum selective concentration appeared to be more fit than those selected at high concentration and still highly resistant (Gullberg et al., 2011;Sandegren, 2014). Antibiotics at such concentrations play multifaceted roles in the soil, including influencing inter-species competition, signal communication, host-parasite interactions, virulence modulation, and biofilm formation (Sengupta et al., 2013). Moreover, they induce mutations, genetic recombination, and HGT processes Hughes, 2012, 2014). Sublethal concentrations of drugs can also increase the mutation rate and potential for enrichment of stable genetic mutants by the induction of SOS responses, increasing translation misreading, and generation of oxygen radicals. It has been reported that sub-lethal concentrations of tetracycline stimulated up to 1,000-fold higher horizontal transfer rates of mobile genetic elements in Listeria monocytogenes (Bahlm et al., 2004) and integron recombination rates in Vibrio cholerae (Guerin et al., 2009). However, the rate of spread of ARGs depends on the antibiotic concentration in the soil. Huang et al. (2016) showed that the dissipation rate of plasmid-located genes [i.e., qnrS, oqxA, aac(6 ′ )-Ib-cr] was significantly lower in soil treated with ciprofloxacin at concentrations of 0.04 and 0.4 mg/kg soil compared to soil treated with antibiotic at 4 mg/kg and controls. Forsberg et al. (2012) also found little evidence for HGT in soil.
Because of the risks posed by the dissemination of ARGs among bacteria and plants, studies investigating new practices designed to decrease the accumulation and transport of ARGs within soil are being conducted. For example, the use of biochar as a soil amendment significantly reduced the abundance of tetracyclines (tetC, tetG, tetW, and tetX) and sulfonamides (sulI and sulII) resistance genes in soil and lettuce tissues (Ye et al., 2016;Duan et al., 2017).

CONCLUSION
Environmental antibiotic pollution is a problem that is expected to gain more attention in the near future since antibiotic consumption is still increasing around the world.
A review of available literature shows that transformation and/or degradation are the most important processes that determine the fate of antibiotics in soils and that soil microorganisms play an important role in these processes. However, the rate of these transformation and degradation depends largely on the antibiotic structure and is affected by many abiotic and biotic factors, as is most evident in the large range of DT50 values of antibiotics in different soils. Studies have shown that within groups of similar antibiotics or even particular antibiotics, DT50 values differ significantly because of various soil properties, antibiotic dosage, and environmental conditions. Current literature reviewed here indicates that the input of antibiotics into soil alters the structure and activity of microbial communities and the abundance of ARGs. However, results from studies in this field are often ambiguous, making environmental risk assessments related to the presence of antibiotics depend upon too many different factors to be reliable. Interactions between soil, antibiotics, and microorganisms are multifarious and many environmental factors may influence the value of tested parameters. The effect of antibiotics on the activity and diversity of microbial communities depends on the physicochemical parameters of the soil, the antimicrobial activity, and dosage of the antibiotic, as well as the time of exposure. It has become clear that microorganisms that are sensitive to different antibiotics are killed or inhibited in the presence of antibiotics, which may result in outgrowth of resistant bacteria resistant. In turn, this may alter the diversity of the soil microbial communities. Contrarily, there is some evidence that certain microorganisms can adapt to and possibly transform antibiotics. The less toxic transformation products would favor recovery of the original microbial communities from the initial disturbances caused by antibiotics exposure. Several studies found transient negative effects of antibiotics on the functional, structural and genetic diversity of soil microbial communities; a temporary loss of soil functionality with subsequent recovery.
Existing OECD and ISO ecotoxicological tests used to evaluate the toxicity of chemicals on the activity of a single species (e.g., Microtox, MARA, measurements of selected enzyme activities) and some of the processes mediated by soil microorganisms are not sufficient to gauge the effects of antibiotics on soil microbial communities. Therefore, ecotoxicological tests that combine various microbial community parameters should be developed. Moreover, future studies should be based on "omics" approaches such as genomics, transcriptomics, and proteomics, that allow for deeper examination of microbial communities than CLPP, PLFA, and PCR-DGGE methods. "Omics" methods are better tools for tracking the fate and determining dissipation rate of ARGs, as well as for recognizing new mechanisms of antibiotic resistance. Future frameworks for antibiotics should focus on the effect of soil properties on the maintenance and development of antibiotic resistance, the fate of antibiotics and ARGs in manure-fertilized soils, the rate of HGT between antibiotic-resistant and autochthonous bacteria, the link between antibiotic concentrations and genetic changes in soil resistomes, and the role of various contaminants and co-selecting agents on maintenance of ARGs in soil. Moreover, there is a need to complement biological assays with advanced analytical methods (such as LC-MS/MS, isotope dilution mass-spectrometry) and proper sample preparation, allowing assessment of antibiotic concentrations and, especially, of bioavailability in complex environmental matrices. Other problems that should be taken into consideration are variations on the procedures for estimating the limits of antibiotic detection and the lack of standard analytical methods for monitoring antibiotic in the environment. Future studies should also be focused on design and management practices that minimize the spread of ARGs from harvested crops in antibiotic-treated soils into the food chain. Finally, though it is rarely studied, the genetic potential of microorganisms involved in the degradation of antibiotics in soil warrant wider investigation.

AUTHOR CONTRIBUTIONS
MC, AM, and ZP-S conceived and designed content of the paper and wrote the paper. MC prepared the figures and tables.