The environmental release and fate of antibiotics

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of foreign molecules (Kafafi et al., 1993; Hankinson, 1995; Deni- son and Nagy, 2003; Desta et ... its concentrations in meat samples remain stable after cooking.
Marine Pollution Bulletin 79 (2014) 7–15

Contents lists available at ScienceDirect

Marine Pollution Bulletin journal homepage: www.elsevier.com/locate/marpolbul

Review

The environmental release and fate of antibiotics Sergio Manzetti a,b,⇑, Rossella Ghisi c a

Fjordforsk A.S. Midtun, 6894 Vangsnes, Norway1 Science for Life Laboratory, Department for Cell and Molecular Biology, University of Uppsala, Box 596, 751 24 Uppsala, Sweden c Department of Agronomy, Food, Natural Resources, Animals and the Environment, University of Padova, Agripolis, viale dell’Università 16, 35020 Legnaro, PD, Italy b

a r t i c l e

i n f o

Keywords: Antibiotics Environment Release Fate Transformation Accumulation

a b s t r a c t Antibiotics have been used as medical remedies for over 50 years and have recently emerged as new pollutants in the environment. This review encompasses the fate of several antibiotics in the environment, including sulfonamides, nitrofurans, terfenadines, cephalosporins and cyclosporins. It investigates the cycle of transfer from humans and animals including their metabolic transformation. The results show that antibiotic metabolites are of considerable persistence and are localized to ground-water and drinking water supplies. Furthermore, the results also show that several phases of the cycle of antibiotics in the environment are not well understood, such as how low concentrations of antibiotic metabolites in the diet affect humans and animals. This review also shows that improved wastewater decontamination processes are remediating factors for these emerging pollutants. The results obtained here may help legislators and authorities in understanding the fate and transformation of antibiotics in the environment. Ó 2014 Elsevier Ltd. All rights reserved.

Contents 1. 2. 3. 4. 5. 6. 7.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Antibiotic usage statistics . . . . . . . . . . . . . . . . . . . . . . . . . . . . Catabolic fate of antibiotics in animals and humans: origins Metabolism of antibiotic compounds . . . . . . . . . . . . . . . . . . . Antibiotics in wastewaters . . . . . . . . . . . . . . . . . . . . . . . . . . . Environmental concentrations of antibiotics . . . . . . . . . . . . . Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

..................... ..................... of antibiotic metabolites ..................... ..................... ..................... ..................... .....................

1. Introduction Antibiotics are chemical compounds that have been used since being discovered by Sir Alexander Fleming in 1928, when the structure of penicillin was discovered and later published as a remedy for infections, inflammations and diseases (Fleming, 1944, 1946). Although much attention has been accredited to Fleming, Louis Pasteur suggested the approach of studying the antagonism between bacteria as rationale to develop new medicines at the end of the 19th century (Kingston, 2008). Dr. Pasteur and his colleague Robert Koch observed the first antibiotic effects in vitro in 1877 in the strain B. Antracis (Landsberg, 1949). Dr. John Tyndall was the first to describe the antibiotic reactions by fungi on ⇑ Corresponding author at: Fjordforsk A.S. Midtun, 6894 Vangsnes, Norway. Tel.: +47 484200967. 1

www.fjordforsk.no.

0025-326X/$ - see front matter Ó 2014 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.marpolbul.2014.01.005

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bacteria in England (Kingston, 2008). Since then, antibiotic use in human and veterinary medicine has become a common therapeutic practice that has led to both high consumption and the gradual accumulation of antibiotics in the environment, deriving from sources of contamination such as wastewater, landfills, urban sites as well as industrial and hospital effluents (Renew and Huang, 2004; Brown et al., 2006; Karthikeyan and Meyer, 2006; Batt et al., 2007; Watkinson et al., 2009). Antibiotic compounds are largely composed of structures encompassed by cyclic components, represented by benzene rings, piperazine units, hexahydropyrimidines, as well as sulfonamides, quinolone and morpholine groups (Renew and Huang, 2004). These compounds have meta-stable properties and yield both activated metabolites, conjugates and hydroxylated forms after processing in humans and animals (Kennedy et al., 1980; García-Galán et al., 2008), which leads to a range of diverse active chemical compounds being continuously released in the environment with possible reactive properties

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and largely unknown consequences. Several studies have reported on the fate of antibiotics; however, with emerging chemical modifications and new products in the medical industry, a proper assessment of the environmental fate of antibiotics requires updates and new investigations. This review encompasses new data for the delineation of the cycle of antibiotics in the environment, from industrial synthesis to release and transformation in the environment. 2. Antibiotic usage statistics Consumption of antibiotics has increased steadily (Vander Stichele et al., 2006), however the European Medicines Agency (EMA) (EMA, 2011) reported a negative trend (approximately 11%) in the sales of veterinary antimicrobial agents in nine European countries from 2005 to 2009. The observed decrease was mainly ascribable to a general reduction in tetracycline utilization, of which dosages are substantially higher than that of other antimicrobials for which increased utilization was observed. In this context, the European Centre for Disease Prevention and Control (ECDPC, 2010) cites no increase in the median consumption of antibacterial agents for human medicine in 2010 compared to the 2001–2009 period. Nevertheless, EMA (2012) reports that 4802 tons of active ingredients were sold in 19 EU countries in 2010 for animal medicine, without considering antimicrobial agents for farmed fish (EMA, 2012). Considering the 2:1 ratio proposed by FEDESA between antibiotics used for human and veterinary medicine in Europe still valid (FEDESA, 2001), we can argue that the amount of antibiotics released in the EU environment might be approximately 15,000 tons/year in the EU alone. Consumption in the USA exceeds that of the EU with considerably higher levels, particularly for new antibiotics (Patrick et al., 2004; Llor and Bjerrum, 2005). Approximately 16,200 tons were produced in 2000, 70% of which was used for live-stock farming (Kümmerer, 2003). Worldwide consumption of antibiotics is not known; however, estimates indicate that the total quantity lies between 100,000 and 200,000 tons (Kümmerer, 2003). 3. Catabolic fate of antibiotics in animals and humans: origins of antibiotic metabolites The metabolic fate of xenobiotics and exogenous molecules encompasses a series of pathways and systems that are identical among most species, particularly vertebrates (Bertault-Peres et al., 1987; Thomas, 2007). The most well-known detoxification systems are represented by the cytochrome P-450, which encompasses some 60 members of genes encoded as membrane-bound oxidase enzymes in a superfamily (Thomas, 2007). Several of these 60 members act on xenobiotic substances introduced in the body through food and air, and perform catalytic conversion of these toxic compounds into solubilized forms, which are then excreted via the urine, sweat and other bodily fluids (Chahine and O’Donnell, 2011; Stiborova et al., 2012; Palma et al., 2013). The xenobiotic-selective group of the P-450 system is expressed in the 7th chromosome at the q22 cluster and encompasses 4 known members, with several other potential members with direct specificity to exogenous compounds (Thomas, 2007). Interestingly, the genetic cluster of this superfamily, which is one of the studied gene clusters, has the highest rate of natural mutation, indicating a genomic adaption to different chemical and biochemical environments presented to the species with time, exposure and evolution. Thus, environmental stress and pollution challenges the ability of the species to adapt in terms of detoxification mechanisms, particularly for fish, animals and humans, which have slower adaption rates than microbial and yeast species.

Detoxification is further governed by several other systems for the removal of xenobiotics. One of these systems is represented by the Glutathione S-Transferase enzymes, which belong to a superfamily that encodes for a group of dimeric enzymes predominantly expressed in the liver (Roldan et al., 2013; Vyskocˇilová et al., 2013). As with the P-450 system, Glutathione S-Transferase members are both expressed for endogenous substrates and for exogenous and xenobiotic substrates including heavy metals (Eum et al., 2013) and persistent organic compounds, which they conjugate to thiol groups for solubilized excretion through the bile duct (Alnouti, 2009) and urinary tract (Kropp et al., 1982; Terrier et al., 1990; Vree et al., 1995). In other organs, such as the liver, microsomes are particularly active in the hydroxylation of specific types of antibiotics, such as cyclosporin A (Bertault-Peres et al., 1987). Here, the monooxygenases CYP1A, CYP2A, CYP3A and CYP2D6 enzymes, as parts of the P450 family, catalyze the non-specific oxidation, hydroxylation and conversion of the xenobiotic to solubilized forms. These are catalyzed after the introduction of the exogenous compound by the Aryl hydrocarbon receptor, a key receptor for the processing and conversion of foreign molecules (Kafafi et al., 1993; Hankinson, 1995; Denison and Nagy, 2003; Desta et al., 2004; de Montellano, 2005). The biological conversion of antibiotics then yields solubilized forms, including glucuronated, glutathione-conjugated and possibly arachidonated conjugates, as well as hydroxylated and nitroreduced forms (Sasabe et al., 1998; Giamarellos-Bourboulis et al., 1999; Yamamura et al., 2000; Farkas et al., 2007). Some of these forms are bioassimilable (Giamarellos-Bourboulis et al., 1999; Yamamura et al., 2000; Ferguson et al., 2005; Farkas et al., 2007) and are therefore potentially absorbed in more advanced species, such as fish, animals and humans through nutrition and diet, a feature that has been recently observed, as chloramphenicol glucuronic-conjugates have been found in the tissues of poultry, in bee-honey, and shrimp from Asian countries (Ferguson et al., 2005).

4. Metabolism of antibiotic compounds The metabolic fate of antibiotics depends on their chemical properties, functional groups and the reactive atoms in the structures. Various types of antibiotics can therefore be treated differently by the body and thus precede different types of metabolites in the excretion (urine/feces). A recent study reported the specific identification of metabolites of sulfonamides, where ten different compounds were identified in urine samples, including 5-hydroxysulfadiazine, 4-hydroxysulfadiazine, 5-hydroxysulfadiazine glucuronide and 5-hydroxysulfadiazinesulfate among other forms, deriving from a single antibiotic (Fig. 1) (García-Galán et al., 2008). This study indicated that hydroxylation of the central amine group, conjugation to glucuronic acid to the amide group, and also acetylation of the amine group were metabolic processes acting on sulfonamides in humans (García-Galán et al., 2008). Sulfonamides are metabolized similarly in animals as in humans; however, animals such as livestock include also other forms, such as the N4-acetyl-conjugates, de-aminated metabolites and N4-glucoseconjugates (Fig. 2) (García-Galán et al. (2008)). Other antibiotics, such as cyclosporin, are metabolized in a similar manner, resulting in conjugation to water-soluble molecules (Bertault-Peres et al., 1987). The detailed structural data of this conjugation are not known; however, indications suggest that cyclosporin is initially attacked by cytochrome enzymes at the amide and propyl sites (Fig. 3) (Bertault-Peres et al., 1987). Furthermore, in Fig. 3, a comprehensive review of the most likely sites for hydroxylation is shown after a quantum chemical analysis, which shows a specific region of susceptibility from hydroxylating enzymes.

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Fig. 1. Selected metabolites of sulfonamide antibiotics found in urine samples (García-Galán et al., 2008).

Fig. 2. Metabolites of sulfonamide antibiotics found in feces and urine samples from livestock (García-Galán et al., 2008).

Nitrofurans are another group of antibiotics commonly used in agricultural and veterinary medicine and represent structures that are composed of oxazolidine groups coupled with a furan group through a methylenamine bridge. Studies from Leitner et al. (2001) show that the nitrofuran structure is attacked by the cytochrome P450 system and hydrolyzed into single-ringed moieties, where the oxazolidin rings remain largely conserved after detoxification (Leitner et al., 2001; Cooper et al., 2005). Such a metabolic conversion indicates that the stable metabolite oxazolidin is not conjugated (Fig. 4) and is not further processed upon excretion in the urine. This process implies that the nitrofuran compounds generate meta-stable metabolites in the urine of animals, fish

and humans, which are less water-soluble than conjugated variants. Reduced water-solubility also indicates a higher toxicity. Other types of metabolites are also generated from nitrofurans, such as the common 3-amino-2-oxazolidinone (Fig. 4), which accumulates in a second species (rats) feeding off the tissues of pigs from farms treated with the original nitrofuran antibiotic (McCracken and Kennedy, 1997). This metabolite has been found to accumulate in the liver, kidneys and muscle tissue of rats, and its concentrations in meat samples remain stable after cooking or microwaving (McCracken and Kennedy, 1997). Interestingly, another nitrofuran-based metabolite, semicarbazide (Hydrazine carboxamide), which resides in livestock (Leitner et al., 2001), has been found to trigger oxidative deamination in endothelial cells and diabetes (Yu and Zuo, 1993; Yu et al., 2003). This metabolite has also been related to vascular disorders and potentially Alzheimer’s disease (Unzeta et al., 2007) and requires several days for complete excretion (Leitner et al., 2001). N4-acetyl-sulfamethazine is another antibiotic-metabolite deriving from sulfamethazine that has received attention in recent studies. This metabolite was recently found in samples from cattle and pig farms (Haller et al., 2002), where the extracts from the samples could reach 2.6 mg pr. kg grab samples (Haller et al., 2002). The concentrations of the metabolite were 2–50 times lower than the parent antibiotic, showing different rates of biotransformation. Interestingly, N4-acetyl-sulfamethazine transforms back to the form of its parent antibiotic sulfonamide (Berger et al., 1986), which calls for additional attention to extend the research on the metabolites of compounds in the environment, as this indicates generation of bioactive compounds in the environment from a single type of antibiotic. An additional antibiotic studied for metabolic fate is thienamycin (Kropp et al., 1982). Thienamycin (Fig. 5) is known to be degraded by a renal dehydropeptidase-I, yielding a formimidoyl derivative of thienamycin. The further metabolic fate of this

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Fig. 3. Cyclosporin structure (Weber et al., 1991) and primary site of oxidation. Left: 2D-structure; right: 3D-structure. Circled in red: primary site of susceptibility to oxidation by OH ions (derived from quantum chemical calculations susceptibility to hydroxylation). (For interpretation of the references to color in this figure legend, the reader is referred to the web version of this article.)

Fig. 4. Catalytic oxidation of a nitrofuran antibiotic type. The resulting complex of 3-amino-2-oxazolidinone is documented (Leitner et al., 2001), however the conjugate of 2-Methyl-5-nitro-furan is hypothetically deduced.

structure is not elucidated (Kropp et al., 1982), however its structure indicates a potential persistent ring-moiety, similar to 3-amino-2-oxazolidinone, which is preserved through the metabolic process in livestock and accumulated in second animals. Further types of antibiotics are terfenadine antibiotics, which encompass bi-cyclic structures with interconnecting bridges (Jurima-Romet et al., 1994). Their metabolism has been found to produce alcohol and ketone metabolites, as well as a carboxylic acid metabolites, which are predominantly found in human excrement (Garteiz et al., 1982). Interestingly, the metabolism of terfenadine has been found to be inhibited or altered by co-administration of other antibiotics (Jurima-Romet et al., 1994), implying that antibiotic cocktails may inhibit the conversion of antibiotics to inactive (or active) products and may increase the concentration of nonmetabolized antibiotics in the environment following the metabolic process deriving from the animals/humans. Even more strikingly, cephalosporin antibiotics do not readily metabolize, thus demonstrating that rats, mice and dogs excrete the antibiotic wholly unmodified through the renal-urinary tract (Sullivan et al., 1976). In this study, minor differences among the tested species were found; however, the main trend was that this antibiotic retained its structure throughout the metabolic process in the tested animals (Sullivan et al., 1976). This study shows that some structures can be particularly resistant to detoxification attempts from organisms, which can thus be excreted and released into the environment as active antibiotic compounds. Additionally, nutrition chemists have expressed concern in regards to cephalosporins (Donoghue, 2003). Interestingly, compounds such as chlorsulfuron and sulfosulfuron, which are good examples of sulfonylurea herbicides utilized in

agriculture for weed control, are subjected to both biotic and abiotic conversion to sulfonamides in the environment (Sukul and Spiteller, 2006). Plants have also been found to release metabolites from antibiotics and can therefore represent a significant source of release to the environment in larger settings such as crop fields and agricultural sites, particularly considering that some antibiotic metabolites have also been found to have antibiotic effects (Wallace, 2004). Agricultural sites can therefore be areas of particular importance for studies examining the release in nutritive sources, including both food and nutrients for commercial production, as well as the release and diffusion of antibiotic metabolites to the soil and drinking water supplies – as several sulfonamide metabolites have been found recently in commercial bottled water, ground-water and surface waters (Díaz-Cruz et al., 2008). The studies on surface-water samples showed several compounds present, encompassing sulfadimethoxine, sulfathiazole, sulfamethazine, sulfamethoxazole, sulfamethoxypyridazine which were all found at ng/mL concentrations, and additional 316 ng/L concentrations of the metabolite N4-acetylsulfamethazine (Díaz-Cruz et al., 2008). Similar findings have also been reported by Chang and coworkers, where 16 sulfonamide types were found in lakes and ponds in Japan at concentrations of up to 8.9 ng/L (Chang et al., 2008). Parallel to these studies, a recent study of 164 sampling sites of ground-water reserves in Europe showed the presence of many pharmaceuticals and pesticides. Among these compounds sulfamethoxazole (deriving from animal and human antibiotic treatment), was found at ground-water locations with a 24% frequency and at concentrations up to 38 ng/L (Loos et al., 2010). Other studies also show metabolites of antibiotics found in drinking water (Schaider et al., 2014).

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Fig. 5. The potential catalytic degradation step of thienamycin. The first step involving renal dehydropeptidase, where the antibiotic is subjected to the first step of ring-opening is confirmed (Kropp et al., 1982). The last structure is derived from the findings reported in (Kropp et al., 1982), with no structural data to confirm this structure.

Metabolites of antibiotics are therefore apparent sources of pollution in the environment, with a focus on the water phase and ground-water reserves. With increasing and poorly regulated use of antibiotics in the Western world, the accumulation of antibiotics and their metabolites in animals and humans can thus represent an environmental and toxicological threat, with wastewaters as central transfer mediums. Thus, renewing and developing better decontamination approaches for wastewaters is critical, as wastewaters are frequently released to the environment without purification, even in industrialized nations (Manzetti and Stenersen, 2010). 5. Antibiotics in wastewaters Wastewaters are the major carriers of antibiotics to the environment in urban regions and metropolises, and the decontamination of wastewaters is critical for purification and avoidance of their release to the environment (Hu, 2013). The accumulation of antibiotics in the environment is also overrepresented by discharges from industrial sites of production of pharmaceuticals, hospitals and medical centers, which all generate high concentrations of antibiotics and metabolites in their effluents (Kümmerer et al., 2000; Karthikeyan and Meyer, 2006). The presence of antibiotics in wastewaters is therefore a critical theme in environmental sciences and was defined as a poorly researched area as recently as 1998 (Halling-Sørensen et al., 1998). Poor decontamination of wastewaters (Kümmerer et al., 2000; Batt et al., 2006), and the resistance of antibiotic compounds to purification methods (Kimura et al., 2004; Yoon et al., 2006) present important issues to be considered when reducing antibiotics in the environment.

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This problem is particularly important, and expected to grow, as several sites in modernized nations release wastewaters without decontamination steps (Manzetti and Stenersen, 2010). Therefore, the research on the release of antibiotics through wastewaters calls for a more frequent implementation of wastewater decontamination and more stringent regulations, as wastewaters can contain high concentrations of these chemical compounds, as concentrations of sulfamethoxazole, trimethoprim, ciprofloxacin, tetracycline, and clindamycin have recently been documented in wastewaters in the US (Batt et al., 2006). The measured concentrations of the abovementioned antibiotics ranged from 0.090 to 6.0 lg/L. The concentrations of these compounds were interestingly measured far from the site of discharge of the wastewater plant, indicating a persistence of the metabolites of antibiotics in waters (Batt et al., 2006). Another group reported the presence of ciprofloxacin, ofloxacin, sulfamethoxazole and trimethoprim in wastewaters at respective concentrations of 100–160, 205–305, 395–575, and 40–705 ng/L in samples from wastewaters taken in three US states (Renew and Huang, 2004). Ofloxacin, ciprofloxacin and trimethoprim were generally found in excessive concentrations for both the secondary and final effluents from wastewater plants. A study by Renew and Huang (Renew and Huang, 2004) indicated that chlorination significantly reduces the levels of antibiotics compared to UV; however, different functional groups in the mixtures of antibiotics may require other approaches, which therefore introduces a series of challenges for providing complete systems of decontamination of low and medium–high concentrations of antibiotics in wastewaters. This finding is of crucial importance, as risk-assessments have been requested for any antibiotic exceeding the concentration of 1 lg/L in the diet by FDA (1998). Levels of antibiotics have also been found at considerable concentrations in the Great Lakes in the US, where carbamazepine, naproxen, diclofenac, bezafibrate, gemfibrozil, and ketoprofen were identified, with carbamazepine found at the highest concentration of 0.65 lg/L at 400 m from the wastewater discharge site (Metcalfe et al., 2003). Carbamazepine was detected in all surface water samples taken by the group, at various sites across the shores of the Great Lakes (Metcalfe et al., 2003). According to Kolpin et al. (2002), more than 20 different antibiotics including metabolites were found in American wastewaters with the metabolite erythromycin-H2O, and the antibiotics trimethoprim, sulfamethoxazole, sulfamethazine and lincomycin at the highest concentrations, reaching the notable level of 27 lg/L. Compared to other studies, the study by Kolpin and colleagues shows considerably higher concentrations, reaching up to 25 times more than what other studies have documented in wastewater analyses. The study by Kolpin also reports on the concentration of a metabolite (erythromycin-H2O), an aspect that is missing in antibiotic detection studies in wastewater. As metabolites encompass a large fraction of the excreted compounds from humans and animals (see above), new studies of antibiotic concentrations in wastewaters should focus also on their metabolites, as several of these can be of higher concentrations than their parent antibiotics. This feature of research is also important as the rate of uptake of the antibiotic in the animal/human compared to its release in an non-metabolized form is generally higher, sometimes reaching an antibiotic/inherent metabolite ratio as high as 1:50 (Haller et al., 2002). This finding implies that the concentration of metabolites can be far higher than the concentration of the parent antibiotics. Another recent study by Batt and colleagues showed concentrations of 0.2–1.4 lg/L of ciprofloxacin, 0.2–2.8 lg/L of sulfamethoxazole, 0.06–1.1 lg/L of tetracycline, and 0.2–7.9 lg/L of trimethoprim in samples taken from various wastewater sites in New York state (Batt et al., 2007). In Switzerland, levels of antibiotics have been confirmed in wastewaters at ng/L concentrations (Göbel et al., 2004).

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Fig. 6. Chart of purification patterns of four wastewater purification plants in the US with relevance to antibiotics decontamination. The second step (sludge activation) is the most crucial step for antibiotics decontamination, and can be expanded to a third step to reach zero-values of antibiotics, by the use of additional antibiotic-degrading bacterial species.

The applied decontamination methods are therefore critical for removal of antibiotics in wastewater treatment plant effluents. In this context Batt et al. (2007) reports on an important comparison of the concentrations of antibiotics depending on several decontamination methods of wastewater. These compared methods of purification include sludge activation, nitrification, sand-filtering, aeration of sludge with ferrous chloride, oxygen activation and biological contractors and chlorination (Fig. 6). A comparison of

purification methods shows that the majority of the antibiotics, which were found at original concentrations of lg/L in the influents, were removed after the second step of purification, and the most efficient method of removal was a two-step sludge activation, which required 49 days of processing (Batt et al., 2007). The degradation process, a two-step approach catalyzed by nitrifying bacteria, induced the degradation of trimethoprim from levels of 70% to 25%, however lower degradation rates for the other antibiotics such as ciprofloxacin and tetracyclin were observed. The purification methods are therefore more efficient on some antibiotics compared to others, and require blends of nitrifying and oxidizing bacterial cultures or alternatively, several repetitions of the nitrification step for a full removal of the antibiotics. Novel methods for purifying wastewaters from antibiotics can be therefore be considered for additional steps of bacterial degradation including deployment of the species Verrucomicrobia, Actinobacteria, a-Proteobacteria, b-Proteobacteria, and Nitrospirae as candidates for three-step nitrification, as these species have been found to degrade antibiotics in Chinese lakes in a recent study (Jiang et al., 2010). These issues are critical, as the low concentrations of antibiotics that remain in circulation reach ground waters and drinking waters, and although the application of decontamination approaches remove large parts of the contaminants, the methods

Fig. 7. The estimated life-cycle of antibiotics in the environment. The life-cycle is fed by the release of antibiotics from wastewater and non-purified sources. The life-cycle is also weakened by microbial degradation and general animal/human biotransformation. However, the balance between feeding and degradation is eventually mapped by concentrations in the environment. As the data demonstrate for antibiotic metabolites in ground and drinking water (Stackelberg et al., 2004; Loos et al., 2010), the feeding factors are stronger than the degrading influences, thus indicating a significant deposition in the environment.

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do not remove very low concentrations of pharmaceuticals and antibiotics (Batt et al., 2006). These low concentrations of antibiotics are known to affect various ecological parameters such as bacterial strain development (Glassmeyer et al., 2005; Batt et al., 2006) and also induce potential long-term effects on humans as a continuous part of their diet through water and food, and as a general contaminant in the environment. The degree of transfer from wastewater to drinking water also requires further investigation, as Stackelberg et al. recently showed it to be a viable transfer route for these contaminants because concentrations reaching up to 2 lg/L of twenty-five antibiotics in drinking-water supplies were found (Stackelberg et al., 2004), with roxarsone, oxytetracycline and virginiamycin overrepresented in these concentrations.

6. Environmental concentrations of antibiotics The diffusion of antibiotic compounds from sources to humans and animals and ultimately to the environment goes through a cycle of partial biotransformation and bioaccumulation and gradual deposition in soil and ground-waters (Fig. 7). The complete consumption of antibiotics at the global scale is estimated to 100,000–200,000 tons, and specific regions such as China and the US, which are particularly large consumers of antibiotics, release high quantities into the environment, feeding the cycle of biotransformation and bioaccumulation of antibiotics in the environment. Several studies have been organized in Asia, including the Hong Kong area (Gulkowska et al., 2007), the rivers of Sindian and Dahan in neighboring Taiwan (Lin and Tsai, 2009), the Yellow river, the Haihe river and Pearl River in China (Xu et al., 2007, 2009; Wang et al., 2010; Luo et al., 2011). Additionally, other continents such as Europe, in Italian rivers (Calamari et al., 2003; Lalumera et al., 2004; Zuccato et al., 2005), in Germany (Hirsch et al., 1999; Sacher et al., 2001; Kümmerer, 2003; Kemper, 2008), in the United Kingdom (Ashton et al., 2004; Bound and Voulvoulis, 2005; Roberts and Thomas, 2006; Zhang et al., 2008) have already organized studies. Finally, and the United States (Kolpin et al., 2002; Cabello, 2006; Focazio et al., 2008) and Australia (Costanzo et al., 2005) have also conducted research on this topic. These studies show that antibiotics deriving from the veterinary and agricultural industries, from wastewater plants, from landfills as well as effluents from hospitals and industrial sites of production of antibiotics with poor decontamination methods are the main sources. Most samples were taken from surface waters, lakes, rivers and deltas near wastewater plants (Kolpin et al., 2002), and they have also been taken near industrial production sites (Larsson et al., 2007) and residential and urban sites in other studies (Brown et al., 2006). The ecosystems most vulnerable to antibiotic contamination from these exposures are confined aquatic ecosystems such as ponds, lakes, soils close to urban and agricultural sites, where antibiotic compounds as tetracyclines, tylosin, sulfamethazine, amprolium, and nicarbazine have been found (De Liguoro et al., 2003; Kumar et al., 2004, 2005). In these ecosystems, the sources of contamination can reach as high as 216 mg pr. liter (manure slurry), resulting in contaminated soils with known concentration reaching 400 g antibiotic pr. hectare (Kumar et al., 2005). The accumulation of antibiotics in the sources of contamination has also been found to yield as much as 19 mg antibiotics pr. kg manure, as measured after 30 days of administration of antibiotics to live stock. Furthermore, the biodegradation and biotransformation of antibiotics at exposed agricultural sites can take as long as 150 days (De Liguoro et al., 2003). The type of antibiotic (its chemical structure and persistency), the pH levels in the ground, the temperature, humidity and microbial populations in the soils determine the rates of biotransformation (De Liguoro et al., 2003) and also interestingly, the physical properties of the soil play a

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major role in the biotransformation of antibiotics, as they regulate the uptake and retention periods of the antibiotics (Rabølle and Spliid, 2000; Thiele-Bruhn, 2003). These parameters of environmental transformation utterly determine the half-life and diffusion rates of the antibiotic compounds to secondary phases, such as the aquatic phase and sediments and result in a continuous cycle where antibiotics are released from animals and humans through wastewaters and are absorbed through drinking water, through fish and meat sources, and through agricultural products (Fig. 7).

7. Conclusions Antibiotics are a group of emerging pollutants that require scrutiny from the environmental science field with particular focus on urban and heavily populated demographic sites. The current studies indicate that metabolites of antibiotics can be persistent, and accumulate in foods and drinking supplies, including groundwaters which were originally expected to be impervious to considerable contamination. The major source of these contaminations is wastewater, and modern decontamination approaches do not remove the antibiotic compounds fully, resulting in a low but constant concentration remaining present in the environment. The improvement of wastewater decontamination methods, and prolonged incubation in biological reactors can significantly reduce the diffusion of antibiotics in the environment. A second and major source of antibiotic-pollution in the environment is the agricultural sector, where antibiotics are transferred to the surrounding environment via manure and sludge, resulting in local cycles of antibiotic transfer from livestock to soil, and from soil back to live stock and ultimately humans. This last event is fed by accumulation in food products, which for some antibiotics, can in turn result in a repetition of transfer to wastewaters and furthermore the environment. The research on antibiotics in the environment therefore requires further studies, as new and different compounds are continuously emerging, with different half-lives, different environmental fates, which may thus require new and different decontamination approaches to prevent further accumulation in the environment.

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