Keywords

1 Introduction

As a consequence of the increasing human population density and more intensive animal farming techniques, fresh water systems have become highly susceptible to be at risk of potential contamination by different pharmaceutical products (PhPs) from both human and veterinary use. Awareness of the presence of PhPs in wastewaters and aquatic ecosystems is growing as investigations regarding new pollutants increase and analytical techniques for detecting these chemicals improve. At present, approximately 3,000 different pharmaceutical ingredients are used in the European Union (EU), including antibiotics, β-blockers, lipid regulators, antidepressants, etc. [1]. Estimations of the potential environmental impact of PhPs are usually based on the quantities produced and consumed, their potency and also on their tendency to bioaccumulate in the environment. The risk posed by antibiotics could be explained in terms of any of these premises. In first place, their role is superlative in modern agriculture and livestock, and this fact is reflected in their high consumption rates. Although information on their usage is not available to the general public either in the United States (US) or in the European Union (EU), estimations indicate sales over the 16,000 t in US in 2001, of which 9,300 t are used in animal-feeding operations [2]. According to the European Federation of Animal Health (FEDESA), the annual consumption of antibiotics in the EU in 1999 was in total 13,288 t with 29% for veterinary medicine, 6% as antibiotic feed additives, and 65% in human medicine. In addition, prescription drugs are generally sold in quantities one order of magnitude lower than nonprescription drugs [3]. Regarding their potency, these substances are designed to cause a biological effect in the target organism or patient at relatively low concentrations. Once discharged in the environment, they may have numerous unexpected effects on nontarget, or as yet unknown, receptors. It has been demonstrated in different studies that the environmental presence of antimicrobials leads to the development of antibiotic resistance in bacteria, threat that has been recognized by, among others, the World Health Organization (WHO) and is a well-documented fact nowadays. They can also be toxic to different nontarget organisms, including beneficial bacteria in both natural and urban environments; for instance, wastewater treatment processes may be disrupted [4, 5] or degrading microbiota from different ecosystems can be negatively affected [6]. Finally, antimicrobial resilience and persistence in the environment has been demonstrated [7, 8]. This is a direct consequence of their physicochemical properties such as polarity or liposolubility (they can go through biological membranes), which makes them very persistent compounds in order to stay active and therefore very prone to bioaccumulate

2 Environmental Presence of Sulfonamide Antibiotics: Sources and Occurrence

Sulfonamides (SAs) are one of the most widely used antibiotics in human and especially in animal husbandry and fish farming [9, 10]. They are usually applied in combination with diaminopyrimidines such as trimethoprim due to the enhancement of their activity [11]. In EU, SAs are the second most widely used veterinary antibiotics, representing 21% of the sales in the United Kingdom in 2000, and 11–23% in several other European countries. In US, SAs account for the 2.3% of the total amount of antibiotics used [a2]. SAs are widely used because they are inexpensive, effective against a broad spectrum of common bacterial infections, and have high effectiveness in growth promotion in veterinary applications, although this last use has been banned in the EU since 2006 for all antibiotics [12]. The increase in the number of confined animal-feeding operations (CAFOs), which often lack proper waste management practices, has led to a higher use of these antibiotics and, therefore, to a greater occurrence of these substances in the environment. Following treatment, livestock will excrete 50–90% of the administered dose, the parent drug making up for 9–30%. These amounts of the unchanged substance vary depending on the form of the drug and the animal age and species [13, 14]. Animal excreta are considered one of the major sources of environmental contamination by SAs; residues of these antimicrobials have been detected in manure from medicated animals, which is frequently applied as nutrient amendment in agriculture as it is regarded as a very valuable fertilizer containing essential nutrients for plant growth such as nitrogen, phosphorous, organic carbon or potassium [15–19]. The extensive use of manure in crop fields is among the major routes by which veterinary antibiotics enter the environment [19–21] and, eventually, the different water systems. The consequent diffuse pollution is difficult to prevent and deal with due to the large areas of application. Once on the topsoil and due to their weak sorption to soil tendency and high solubility, the excreted residues of SAs become very mobile and may reach surface waters during runoff episodes and even percolate and contaminate the aquifers [15, 22]. This possibility has already been proved in several publications, showing the presence of SAs at different concentrations in groundwater from various sites close to animal farming facilities [23–30]. On the other hand, although veterinary antibiotics such as SAs only reach wastewater treatment plants (WWTPs) to a limited extent, they have been frequently detected in influent and most importantly, in effluent wastewaters [31–34] due to their generally low biodegradation and elimination efficiency during sewage treatment. As these effluents commonly discharge into natural water courses, in the last decade a growing awareness in the scientific field has been manifested regarding the danger posed by the WWTPs inputs to river ecosystems. River basins and catchment areas can therefore be considered highly vulnerable systems regarding SAs contamination. It should also be considered the frequent application of biosolids from WWTPs as organic amendments in agriculture, opening a different entrance pathway into the environment for these substances [35]. Other secondary input pathways are waste effluents of the manufacturing processes or hospitals, the disposal of unused or expired drug products (solid waste or “flushing”), accidental spills during manufacturing or distribution and leakage from septic systems and agricultural waste-storage facilities [36–38]. Antibiotics that reach landfill sites as solid waste are subjected to biologic degradation processes, but some may persist and leach into surrounding groundwater or reach river courses after flood episodes [39–41]. Another critical scenario is that of aquaculture and antibiotics direct addition to receiving waters, formulated as feed additives, with 70–80% of the administered amount entering the environment [42].

2.1 Presence of Sulfonamides in Wastewater Treatment Plants

Given the relevance of WWTPs discharges as indirect entrance pathway for SAs and many other pollutants onto surface waters, a first step to evaluate the health of a river ecosystem would be to determine the loads of pollutants in these WWTPs effluents. Degradation and vulnerability of river systems are directly dependent on the removal efficiencies (RE%) of the WWTPs regarding these contaminants; however, data on the RE% of these compounds during wastewater treatment is still scarce. In general, Spanish WWTPs apply primary and secondary biologic treatments, the latter usually based on conventional activated sludge (CAS). Tertiary treatments such as ozonation, which have demonstrated to be highly efficient in the removal of different PhPs including SAs, are seldom applied [43–45]. Table 1 summarizes some of the RE% values found recently in the literature. Recently, frequencies of detection and RE%s were reported for the seven main WWTPs located along the Ebro River Basin [31]. SAs of human application such as sulfamethoxazole (SMX), sulfapyridine (SPY) and sulfadiazine (SDZ, also used in veterinary therapies) were the most frequently detected (>85%) and at the highest concentrations (650 ng L−1 for SMX and 227 ng L−1 for SPY) in both influent and effluent samples. RE% values obtained were hard to interpret, as SAs were not regularly present in all the WWTPs, and values ranged from negative removals to 100% elimination. SDZ was in average the SA eliminated most efficiently in these seven WWTPs, whereas SPY showed intermediate to high RE% values. SMX showed both RE% higher than 50% but also negative values in many WWTPs. These higher concentrations detected in the effluents are usually attributed to the presence of SA conjugates and metabolites, which usually are not comprised within the scope of the different studies; these conjugates can be transformed back during treatment into the original compound, as demonstrated recently [46] and could therefore explain higher concentrations of SAs in effluents than in influent waters [47, 48]. Alternative secondary treatments, such as membrane bioreactors (MBRs), have been investigated in recent years to obtain an improvement in the RE% values. However, this treatment technology has proved not to be especially good, in particular for SMX and SPY, the two most relevant SAs in terms of frequencies of detection and concentration. Recent works demonstrated that although elimination rates for SMX were higher in the MBRs than in CAS, removal was only partial as nearly half of the SMX input could still be detected in their respective effluents [49–52]. On the contrary, MBRs worked more efficiently than CAS for other SAs, such as SDZ, which was completely removed after MBR treatment, whereas it was removed only 49% during the CAS treatment. Regarding acetylated metabolites, N4-acetylsulfamethazine (AcSMZ) was 100% removed after MBR treatment, and only in a 54% after the CAS treatment [51]. Tertiary treatments such as ozonation and nanofiltration have demonstrated high efficiencies in SAs removal [44, 53–56], but still its application in WWTPs is scarce and the fate of the transformation products generated unknown [57].

Table 1 Removal efficiencies (RE%) for sulfonamides upon different wastewater treatments
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2.2 Presence of Sulfonamides in Surface Waters

The first reported case of surface water contamination by SAs was in England in 1982, when Watts et al. detected at least one compound from SAs family in river water at concentrations of 1 μg L−1 [58]. Nowadays, in Europe the EU Water Framework Directive (WFD) specifies the need to monitor PPs (SAs among them) in surface waters as an informative step to protect and improve the quality of the European water resources [59]. Given that SAs have been frequently detected in WWTP effluents, several studies have aimed to highlight the state and vulnerability of the receiving freshwaters downstream of urban areas and WWTP facilities, focusing especially in the presence in these water matrices of SAs of human consumption, which are the most commonly detected in the wastewater effluents. The low natural biodegradation of SAs [60], and low tendency to adsorb to solid matrices (from the river bed) [61, 62] together with the SAs inputs, both agricultural and urban, that the river may receive all along the basin would lead to a marked concentration gradient from the source to the mouth of the water course. When interpreting the obtained data, seasonal changes should also be taken into account. Generally, the highest concentrations of human SAs (SMX, SPY) are expected during the dry seasons, as the dilution exerted by the receiving streams is lower. For instance, Kim and Carlson [63] detected SMX at a maximum average concentration of 230 ng L−1 during the winter and of 320 ng L−1 during the summer, in the dry season, in Cache La Poudre River, in northern Colorado. During the rainy season, whereas concentrations of SAs from human use would be more diluted, runoff from irrigated rural areas may increase the concentrations in freshwater of veterinary SAs, denoting its runoff origin from crop lands after heavy rain periods. For instance, in the study by Kim et al., runoff from irrigated rural areas increased the concentrations in freshwater of sulfamerazine (SMR) and sulfadimethoxine (SDM), veterinary SAs (40 and −60 ng L−1), respectively. Cold conditions can also contribute to higher concentrations due to reduced biodegradation of these contaminants in water. Other studies on the distribution of SAs in surface waters yielded similar outcomes, with higher levels of human SAs (SMX,SPY) detected during the dry periods and higher levels of veterinary SAs, such as SDM, sulfamethazine (SMZ), or SDZ during high flow conditions [37, 64, 65]. In some occasions, the release of untreated wastewaters due to strong rainfall events can also contribute to higher concentrations of human SAs than expected [66]. In Europe, the impact of urban inputs was also demonstrated during two sampling campaigns carried out along the Ebro River Basin (Spain) in 2007–2008 [31]. Samples corresponding to tributaries of the main water course presented the highest total concentration of SAs due to their lower flows and dilution exerted on the effluents loads. In 2008, strong rainfall and subsequent runoff events from agricultural land accounted for the highest total SAs concentrations detected in two sampling points in the Ebro River located upstream of two WWTPs (without urban influence). SMX was again the SA most frequently detected in the different surface water samples investigated, being present in the 100% of the samples during the dryer period (2007), with an average concentration of 89.8 ng L−1, and in the 69% of the samples during a higher waterfall period (2008), with an average concentration of 25.5 ng L−1. SPY was also detected in the 100% of the samples during the dry period, at an average concentration of 11 ng L−1, and in the 62% of the samples during the rainy season, with a lower average concentration of 2.7 ng L−1. Another 16 SAs and one acetylated metabolite were detected at concentrations ranging from 0.1 to 127 ng L−1. SMX was also present in freshwater from the Douro River in Portugal, with a maximum concentration of 53.3 ng L−1 and an occurrence of 33% [67]. SMX was detected in the Seine River in all the samples investigated over a period of 6 months in 2006, with average concentrations between 37 and 140 ng L−1 [66]. In this study, the concentration of SMX seemed to increase after heavy rain episodes, which was attributed to the release of untreated wastewaters and not to surface runoff in agriculture areas, as SMX is mainly used in human medicine. Lower concentrations of SMX were detected by the same author in the Oise River, Marne River, and again Seine River (12–26 ng L−1) [68]. SMX was detected also different sampling sites along the Elbe River in Germany and the Czech Republic during 1999 and 2000 at concentrations in the range of 30–70 ng L−1 [69]. The presence of SAs not only in river water samples but also in their sediments [7, 63, 65, 70–72], despite their low distribution coefficients (K d), highlights the river systems vulnerability against these antimicrobials. Furthermore, the presence of SAs metabolites such as their acetylated or glucuronidated moieties has been already demonstrated and the neglection of these compounds would mean to underestimate the real SAs concentration in the water matrix under study, and also the potential adverse effects derivated from the ecosystems exposure to these substances. For instance, the acetylated form of SMX has been detected in natural streams at higher frequencies and concentrations than its parent molecule [73]. A recent study has also demonstrated that N4-acetyl-SPY is more toxic than its parent compound, SPY, to aquatic bacteria [74].

3 Sulfonamide Presence in the Mediterranean Region: The Case of the Llobregat River Basin

The semiarid conditions present in the Mediterranean region aggravate the adverse ecological effects derived from the presence of SAs and other PPs in natural water courses [75]. The hydrology of streams and rivers of these regions are characterized by high seasonal variability with periods of low or intermittent flow disrupted by acute floods [76, 77]. The increasing population density has resulted in not only a higher water demand for irrigation or human consumption, but also in the intensification of wastewater inputs on the receiving streams, which usually present low natural base flows due to the aforementioned long draught periods. These inputs are among the major stressors of receiving streams and rivers, as they contain an excess of nutrients together with a wide range of emerging contaminants. The Llobregat River is an illustrative example of the hydrological pattern of Mediterranean rivers, with low winter and summer discharges and periodic floods in spring and autumn. It is located in the northeast of Catalonia (Spain) and flows into the Mediterranean Sea south of the city of Barcelona. Along its 156 km, it covers a catchment area of about 4,957 km2, which is densely populated (3,089,465 inhabitants, data from 1999), especially in its middle and lower sections. Together with its two main tributaries, the Cardener River and the Anoia River, the Llobregat is subjected to heavy anthropogenic pressure, receiving extensive industrial and urban discharges from more than 50 WWTPs (137 Hm3 year−1; 92% from WWTPs) [78]. These inputs are only partially diluted by its natural flow (0.68–6.5 m3 s−1 basal flow). Furthermore, 30% of the annual discharge of the river (693 Hm3) is used for drinking water supply, including the city of Barcelona. The average monthly flow registered in 2000–2008 period showed peaks of 100 m3 s−1 together with minimum values of 1 m3 s−1 (www.gencat.cat/aca). The Llobregat has therefore been chosen in several studies as the typical case study of the problematic of a Mediterranean overexploited river. Recently, within the framework of the European project MODELKEY, several works have been devoted to study the presence of emerging contaminants [79–81]. During three sampling campaigns carried out along the Llobregat River and one of its main tributaries, the Anoia River, different types of emerging contaminants, including PhPs and SAs, were monitored [82]. Samples were taken in June and November of 2005 and May of 2006, covering spring and autumn periods (maximum flow periods). In the case of SAs, the highest concentrations were detected in the low course of the river and near its mouth (Fig. 1). Due to the cumulative effect along the basin mentioned in Sect. 2.2, SAs followed a pollution gradient and these high concentrations are due to both frequent WWTP discharges and accumulation. SMX was present at a maximum concentration of 4,297 ng L−1, followed by SMZ at 2,482 ng L−1, and its acetylated metabolite that was present at a concentration of 695 ng L−1. Furthermore, estimated values for SPY, sulfamethoxypyridazine (SMP), SDZ, and SPY were out of the analytical calibration range (>5,000 ng L−1) in the sampling location. These values are over two orders of magnitude above the values obtained in continental rivers (Fig. 2, Table 2) and, as can be observed, correspond to SAs of both veterinary and human use. In a recent work, SAs have been detected in effluents of four different WWTPs along this basin, but their concentrations were never higher than 300 ng L−1 [74]. In the Anoia River, despite its lower flow and dilution factor exerted on the incoming pollutants, concentrations were markedly lower, results that can be explained in terms of the lower number of discharging WWTPs to this tributary in comparison with the Llobregat. Urban inputs play a major role in both the hydrology and the presence of pollutants in this basin, as demonstrated with SAs.

Fig. 1
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Sampling sites studied in the Llobregat and Anoia River

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Fig. 2
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Sulfonamide concentrations (ng L−1) detected in the Llobregat and Anoia River

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Table 2 Maximum sulfonamide concentrations detected in different river waters worldwide. Concentrations during high and low flow periods are given when possible (high flow/low flow)
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Table 2 (Continued)
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4 Ecotoxicological Effects of Sulfonamides in the Aquatic Environment

There is a substantial lack of ecotoxicological data regarding adverse effects of SAs and their metabolites, which is probably one of the main reasons for the absence of European regulation on maximum levels of this family of antibiotics in any environmental compartment. Nowadays, none of the PhPs detected in surface water are considered in any of the Drinking Water Directives worldwide [1]. Recently, different PhPs such as carbamazepine or diclofenac were considered to be included in the list of priority substances of the new European Directive 2008/105/EC on environmental quality standards, although they were finally withdrawn. Whereas SAs are probably not pharmacologically active in humans at the concentrations detected so far (usually at the ng L−1 level), they might be potential micropollutants to key living organisms in aquatic ecosystems (e.g., fish, aquatic invertebrates and unicellular algae). These different taxonomic groups, belonging to different trophic levels, may be exposed and negatively affected to different extents. For example, severe toxic effects in primary producers may imply loss of the whole food-chain structure, as they represent a significant portion of the total biomass of the ecosystem and are important as a source of carbon for the rest of the aquatic biosphere. Despite the lack of toxicity data available in the literature, it has been demonstrated that generally microalgae are more sensitive than crustaceans and fish to antibacterial agents (e.g., triclosan and ciprofloxacin). However, SAs have proved to hardly pose any toxicity against green algae [83, 84]; estimated inhibitory concentration (IC) values were much higher than those expected in surface waters and SAs have been considered unlikely to be toxic to algae at environmental concentrations. SMX, as one of the most consumed SAs in human medicine and most frequently detected in natural waters, has been the target of different toxicity evaluations. Median effective concentrations (EC50) range from 80 mg L−1 against green algae [85] to values of 0.52 mg L−1 for algae and 0.21 mg L−1 for crustaceans [85], indicating that the risk posed by this substance should not be excluded in real environmental conditions. It has been demonstrated that aquatics plants [86], crustaceans and fish are also vulnerable to SAs; SMX also showed toxicity against rainbow trout (Oncorhynchus mykiss), but at concentrations so high that were not representative of the real situation in freshwaters [87]. Bioaccumulation of SMZ in sturgeon (Acipenser schrenkii) was also demonstrated, but considered of little environmental concern regarding presence in tissues consumed by humans or to biomagnification in fish consumed by fish predators [88]. On the other hand, toxicity and bioaccumulation in marine environment were observed in brine shrimp exposed to SDM, with the potential implications for the rest of the food chain in the marine community [89].

SAs are usually not detected as isolated drugs in the aquatic environment but together with other SAs, and synergistic effects could be expected when residues of different SAs are detected in the same study site [84, 90]. Belonging to the same family of compounds implies similar molecular structure and modes of action, so “concentration addition” is likely. Furthermore, it is necessary to take into account that the degradation products and metabolites of SAs may also be involved in the final toxic effects on the algae, making the interpretation of the toxic data more complex. Recently, EC50 values for Vibrio fischerii were calculated for SPY and its acetylated metabolite; concentrations of 27.4 mg L−1 and of 8.2 mg L−1 for SPY and the metabolite, respectively, after 15 min exposure were reported [46, 74]. According to the EU legislation (Directive 447 93/67/EEC) that categorizes the toxicity to aquatic organism depending on the EC50, SPY would be classified as harmful, and its metabolite as toxic. To the author’s knowledge, the only reference regarding harmful effects of acetylated SAs is that by Eguchi et al. [90], in which the metabolites of SDM, SMX, and SDZ showed much weaker growth inhibitory effects than the corresponding parent SA against microalgae, usually the more sensitive taxa. The simultaneous presence of the corresponding acetylated metabolites enhanced the inhibitory effect of the three SAs, and also the addition of the diaminopirimidine trimethoprim.

4.1 Ecotoxicity of Sulfonamide Intermediate Products

Whether or not SAs are biodegraded in the aquatic environment would settle the very first step for a complete environmental risk assessment (ERA). At the same time, the toxicity of the intermediate by-products of both biotic and abiotic degradation should be taken into account when evaluating the derived ecological risk. SAs undergo photocatalytic degradation [91, 92] and, if the photodegradation products generated are biodegradable, they can be removed during wastewater treatment using biological methods. If these products are persistent or not readily biodegradable, risks of ecotoxicity should be considered. Both inhibitory and stimulatory effects could be expected, as demonstrated by Baran et al. for sulfathiazole (STZ), SMX, SDZ, and sulfachloropyridazine (SCM) against green algae growth [93]. Photoenhanced toxicity under natural sunlight has already been demonstrated for three SAs (SMX, STZ, and SMZ) against crustacean Daphnia magna, suggesting that the photodegradation of the parent compound leads to the formation of more toxic by-products [94]. Also, the by-products of SMX after ozonation treatment were toxic against D. magna and P. subcapicata [95]. In both cases, the assayed concentrations of SAs that were acutely toxic to D. magna were much higher than levels detected in the environment and the ecological risks associated were considered to be limited.

4.2 Bacterial Resistance

So far, environmental research on antibiotics in general has focused mainly on the bacterial resistance acquired against antimicrobials in the different environmental compartments. Nowadays, the widespread presence of resistant bacterial strains has been demonstrated in several scientific works. In river ecosystems, the frequent presence of SMX has led to the detection of SMX-resistant bacteria belonging to Aeromonas spp., typical waterborne bacteria [96]. The Acinetobacter genera were also affected by the presence of this SA [97], and a correlation was established between SMX environmental concentration and occurrence of SMX-resistant bacteria. SAs-resistant genes have been found not only in surface water but also in river sediments [98]. The concentration of these genes was up to 1,200 times higher in sediments, indicating that they can be considered as important antibiotic resistance genes (ARGs) reservoirs. SAs may have qualitative and quantitative effects upon the resident microbial community found in sediment, which can in turn affect the degradation of organic matter. WWTP effluents have been considered as ARG sources in different works too [4, 99, 100].

4.3 Environmental Risk Assessment for Sulfonamides in Surface Waters

As mentioned above, little information is available regarding the ecological effects of SAs and other PhPs, due mainly to the fact that such investigations are not legally required as part of the licensing procedures for human medicaments. The risk assessment guidelines set up by the European Medicines Agency (EMEA) for the marketing authorization of new medicinal products have been used in a few occasions to prioritize the risk from drugs that are already in use and to assess the potential impact of drugs yet to be released [31, 48, 101–106]. Although they are designed as part of the process for registering new drugs, they are used nowadays as the only restrictive measure established so far to evaluate environmental risk from drugs that are already being consumed and that are being excreted in aquatic or terrestrial environments. The ERA protocol is a two-phase tiered process that begins with an approximate calculation of the predicted environmental concentration (PEC) of the drug in water. These guidelines recommend that any drug exceeding 10 ng L−1 in surface water should progress to Phase II, where standard acute toxicity tests will be carried out in order to estimate predicted no-effect concentration (PNEC) or nonobserved effect concentration (NOEC) [107]. Finally, the ratio of the PEC to PNEC, known as the hazard quotient (HQ), indicates whether a potential environmental impact is implicit and further testing might be needed (HQ > 1). It is also recommended that when the total concentration of metabolites is a 10% greater than the concentration of the corresponding parent drug, the metabolites are also to be further investigated (phase II tier B) in order to determine their ecotoxicological effects. The EMEA Committee for Medicinal Products for Veterinary Use also established similar guidelines to assess the potential for veterinary medicines to affect nontarget species in the environment, including both aquatic and terrestrial species [108]. When PNEC values are not available, an alternative PNEC can be derived by dividing EC50 or median lethal concentration (LC50) values (acute toxicity data) by an uncertainty factor of up to 1,000 [109], and so converting acute to chronic toxicity values, since data on chronic toxicity for SAs is lacking. Likewise, measured environmental concentrations (MECs) are used in the calculation instead of PECs. In order to set up a worst case scenario, maximum MECs and the lowest EC50 or LC50 values are used. In all cases, the MECs should be higher than the boundary value of 10 ng L−1 established by EMEA in Tier 1. Table 3 summarizes the HQ values reported to date in the literature. As can be observed, HQs > 1 were detected only for SMX and only for blue green algae. The highest risk corresponded to the exposure to concentrations detected in the Llobregat River, highlighting once more the vulnerability of water courses located in the Mediterranean climate region.

Table 3 Estimation of hazard quotients (HQ) for the different sulfonamides present in surface waters, following the EMEA guidelines
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