Thanks to their low cost and their broad spectrum of activity in preventing or treating bacterial infections, sulfonamides (SAs) are one of the oldest groups of veterinary chemotherapeutics, having been used for more than fifty years. To a lesser extent they are also applied in human medicine. After tetracyclines, they are the most commonly consumed veterinary antibiotics in the European Union. As these compounds are not completely metabolized, a high proportion of them are excreted unchanged in feces and urine. Therefore, both the unmetabolized antibiotics as well as their metabolites are released either directly to the environment in aquacultures and by grazing animals or indirectly during the application of manure or slurry [1-3].
Physico-chemical properties and chemical structures of selected SAs are presented in Table 1. They are fairly water-soluble polar compounds, the ionization of which depends on the matrix pH. All the sulfonamides, apart from sulfaguanidine, are compounds with two basic and one acidic functional group. The basic functional groups are the amine group of aniline (all the SAs) and the respective heterocyclic base, specific to each SA. The acidic functional group in the SAs is the sulfonamide group. With such an SA structure, these compounds may be described by the
|M = 214.2 g mol-1|
pKa2 = 2.8
pKa3 = 12.0
logP = -1.22
|M = 249.2 g mol-1|
pKa2 = 2.37
pKa3 = 7.48
logP = 0.03
|M = 250.3 g mol-1|
pKa2 = 1.98
pKa3 = 6.01
logP = -0.09
|M = 253.3 g mol-1|
pKa2 = 1.81
pKa3 = 5.46
logP = 0.89
|M = 255.3 g mol-1|
pKa2 = 2.06
logP = -0.04
|M = 264.3 g mol-1|
pKa2 = 2.16
pKa3 = 6.80
logP = 0.11
|M = 267.3 g mol-1|
pKa2 = 2.15
logP = 1.01
|M = 270.3 g mol-1|
pKa2 = 2.24
pKa3 = 5.30
logP = 0.47
|M = 278.3 g mol-1|
pKa2 = 2.46
pKa3 = 7.45
logP = 0.27
|M = 280.3 g mol-1|
pKa2 = 2.20
pKa3 = 7.20
logP = 0.32
|M = 284.7 g mol-1|
pKa2 = 1.72
pKa3 = 6.39
logP = 0.71
|M = 310.3 g mol-1|
pKa2 = 2.5
pKa3 = 6.0
logP = 1.63
Due to their properties, after disposal in soils, these compounds may enter surface run-off or be leached into the groundwater. Moreover, they are also quite persistent, non-biodegradable and hydrolytically stable, which explains why in the last ten years they have been regularly detected not only in aquatic but also in terrestrial environments [1-3,7,14]. Although SAs concentrations in environmental samples are quite low (at the µg L-1 or ng L-1 level), they are continuously being released [3,15]. Therefore, the kind of exposure organisms may be subjected to will resemble that of traditional pollutants (e.g. pesticides, detergents), even those of limited persistence. Consequently, SAs as well as other pharmaceuticals may be considered pseudo-persistent.
SAs are designed to target specific metabolic pathways (they competitively inhibit the conversion of
However, knowledge of the potential effects of SAs on the environment is very limited. Recently, a few review papers have been published that summarize the available ecotoxicity data of pharmaceuticals, including some sulfonamides [16-17,19-21]. Such data as are available on the potential effects of pharmaceuticals in the environment appear to indicate a possible negative impact on different ecosystems and imply a threat to public health. However, if we look just at the sulfonamides, most current studies have investigated acute effects mainly of single compounds and mostly with reference to sulfamethoxazole (SMX), one of the most common SAs, used in both veterinary and human medicine [16-17,20]. Available information on the ecotoxicity of selected sulfonamides has been review and is presented in Table 2.
|>50(30 min)||43.56(96h, P. subcapitata)||30.30(7d, L.gibba)||0.87(48h, D. magna)|
|3.40(96h, S. dimorphus)||0.22(7d, L. minor)|
|16.59(96h, S. leopoliensis)|
|>50(30 min)||5.28(24h, S.vacuolatus)||0.46(7d, L. minor)|
|>25(30 min)||7.80(72h, P. subcapitata)||0.07(7d, L. minor)||221(48h, D. magna)|
|2.19(72h, P. subcapitata)||13.7(21d, D. magna)|
|0.135(72h, M. aeruginosa)||212(48h, D. magna)|
|23.3(30 min)||1.53(72h, P. subcapitata)||0.081(7d, L.gibba)||189.2(48h, D.magna)||123.1(48h, D.magna)||>750(48h, O. latipes) a|
|>84(30 min)||0.15(96h, P. subcapitata)||0.132(7d, L.gibba)||177.3(96h, D.magna)||205.1(48h, D.magna)||562.5(96h, O. latipes) a|
|78.1(15 min)||0.52(72h, P. subcapitata)||0.0627(14d, D.carota)||25.2(24h, D. magna)||70.4(48h, M.macrocopa)||27.36(24h, O. myskiss)|
|74.2(5 min)||2.4(96h, C. meneghiniana)||0.0612(21d, D.carota)||15.51(48h, C. dubia)||84.9(24h, M.macrocopa)|
|>100(30 min)||0.0268(96h, S. leopoliensis)||0.0454(28d, D.carota)||0.21(7d, C. dubia)||9.63(48h, B.calyciflorus)|
|1.54(24h, S.vacuolatus)||0.21(7d, L. minor)||100(48h, C. dubia)||35.36(24h, T.platyurus) a|
|>1000(15min)||13.10(24h, S.vacuolatus)||3.552(7d, L.gibba)||149.3/85.4(48h/96h, D. magna)||>500(48h, O. latipes) a|
|>50(30 min)||4.89(7d, L. minor)||616.7(24h, D. magna)||>500(96h, O. latipes) a|
|391/430(48h/24h, M. macrocopa)||>100(48h, O. myskiss) a|
|135.7/78.9(48h/96h, D. magna)|
|>50(30 min)||11.90(24h, S.vacuolatus)||0.68(7d, L. minor)|
|>50(30 min)||18.98(24h, S.vacuolatus)||0.62(7d, L. minor)|
|>100(30 min)||24.94(24h, S.vacuolatus)||2.54(7d, L. minor)|
|344.7(15 min)||19.52(24h, S.vacuolatus)||1.277(7d, L. gibba)||174.4/158.8(48h/96h, D. magna)||>500(48h, O. latipes) a|
|>100(30 min)||1.74(7d, L. minor)||215.9/506.3(48h/24h, D. magna)||>500(96h, O. latipes) a|
|111/311(48h/24h, M. macrocopa)|
|185.3/147.5(48h/96h, D. magna)|
|>100(30 min)||3.82(24h, S.vacuolatus)||1.51(7d, L. minor)|
|26.4(15 min)||32.25(24h, S.vacuolatus)||2.33(7d, L. minor)||375.3/233.5(48h/96h, D. magna)||589.3(48h, O. latipes) a|
|>50(30 min)||2.48(7d, L. minor)||535.7(96h, O. latipes) a|
|>500(15 min)||2.30 (72h, P. subcapitata)||0.445(7d, L.gibba)||248.0/204.5(48h/96h, D. magna)||>100(48h, O. latipes) a|
|>500(5 min)||11.2 (72h, C. vulgaris)||0.248(7d, L.gibba)||270/639.8(48h/24h, D. magna)||>100 (96h O. latipes) a|
|>50(30 min)||9.85(24h, S.vacuolatus)||0.02(7d, L. minor)||184/297(48h/24h, M. macrocopa)|
|0.25(96h, P. subcapitata)||13.55(7d, L.gibba)||3.47(48h, D. magna)|
|0.45(96h, S. dimorphus)||2.33(7d, L. minor)|
|2.83(96h, S. leopoliensis)|
This demonstrates the lack of data relating to the long-term exposure of non-target organisms, and especially how continuous exposure for several generations may affect a whole population. Moreover, as these compounds occur in natural media not as a single, isolated drug but usually together with other compounds of the same family or the same type, accumulated concentrations or synergistic-antagonistic effects can be also observed. The simultaneous presence of several pharmaceuticals in the environment may result in a higher level of toxicity towards non-target organisms than that predicted for individual active substances.
Therefore, the main aim of this chapter was to review the existing knowledge on the chronic and mixture toxicity of the residues of sulfonamides in the environment, since it has not been done yet. This will be achieved by: (1) presenting current approaches for Environmental Risk Assessment (ERA) for pharmaceuticals with respect to the evaluation of chronic and mixture toxicity of these compounds; (2) introducing the reader to basic concepts of chemical mixture toxicology; and finally (3) by discussing detailed available information on chronic and mixture toxicity of the residues of sulfonamides in the environment.
2. Environmental risk assessment of pharmaceuticals vs. chronic and mixture toxicity of pharmaceuticals
The approaches currently being used to assess the potential environmental effects of human and veterinary drugs in the U.S. and in the European Union are in some respects dissimilar [34-39]. The Environmental Risk Assessment (ERA) process usually starts with an initial exposure assessment (Phase I). But with some exceptions, a fate and effects analysis (Phase II) is only required when exposure-based threshold values, the so-called action limits, are exceeded in different environmental compartments. Thus risk assessment, described by Risk Quotient (RQ), is performed by the calculation the ratio of the predicted (or measured) environmental concentration (PEC or MEC respectively) and predicted biological non-effective concentrations (PNEC) on non-target organisms. If RQ is less than one it indicates that no further testing is recommended. Calculations of environmental concentrations rely e.g. on information on treatment dosage and intensity along with default values for standard husbandry practices, and are based on a total residue approach reflecting worst-case assumptions. For example, the recently introduced European guidance on assessing the risks of human drugs excludes the testing of pharmaceuticals whose PECsurface water is below an action limit of 0.01 µg L-1; in the U.S. this threshold value is 0.1 µg L-1. Moreover, there are two different action limits for veterinary pharmaceuticals, one each for the terrestrial and the aquatic compartments. No fate and effect analysis is required for veterinary pharmaceuticals used to treat animals if the PECsoil is < 100 µg kg-1 dry weight of soil. However, a Phase II assessment is not required for veterinary medicines used in an aquaculture facility if the estimated concentration of the compound is < 1 µg L-1 [40-41]. If the PECsurface water of a pharmaceutical is above the action limit, effects on algae, crustaceans and fish are investigated. However, if PECsoil is higher than the action limit, then Phase II, divided into two parts: Tier A, in which the possible fate of the pharmaceutical or its metabolites and its effects on earthworms (mortality) and plants (germination and growth) as well as the effects of the test compound on the rate of nitrate mineralization in soil are determined; and Tier B in which only effect studies are recommended for affected taxonomic levels [34-39].
The main problem associated with this approach is the fact that the no actual sales figures or measured environmental concentrations are at hand when a risk assessment is conducted. Therefore, only crude PEC calculations are performed . Moreover, the (eco)toxicity tests included in Phase II focus on acute toxicity of only single compounds. Chronic and mixture toxicity is not obligatory. As the risk of an acute toxic effect from pharmaceuticals in the environment is unlikely and organisms in the environment are exposed to mixtures of pharmaceuticals, such limited focus results in important uncertainties. Additionally, same drugs (like sulfonamides) are used to treat both humans and animals. Although the exposures may differ, their potential effects on non-target organisms will be the same, and so the effect-testing approaches should be similar. For these reasons, many scientists have already pointed out the need for more reliable PEC and PNEC calculations for more realistic ERA of pharmaceutical [40-42].
3. Basic concepts of chemical mixture toxicology
To predict the toxicity of mixtures, ecotoxicologists use concepts originally developed by pharmacologists in the first half of the 20th century [43-48]. Since more than 20 years, they have been trying to elucidate the problem of risk assessment for complex mixtures of various substances. As a result a lot of excellent studies have been performed in this topic [49-51]. One of the main interests of scientists in the field of combination toxicology is to find out whether the toxicity of a mixture is different from the sum of the toxicities of the single compounds; in other words, will the toxic effect of a mixture be determined by additivity of dose or effect or by supra-additivity (synergism - an effect stronger than expected on the basis of additivity) or by infra-additivity (antagonism - an effect lower than the sum of the toxicities of the single compounds) The toxic effect of a mixture appears to be highly dependent on the dose (exposure level), the mechanism of action, and the target (receptor) of each of the mixture constituents. Thus, information on these aspects is a prerequisite for predicting the toxic effect of a mixture [46-47, 52].
Generally, three basic concepts for the description of the toxicological action of constituents of a mixture have been defined by Bliss and are still valid half a century later: (1) simple similar action (concentration addition, CA), (2) simple dissimilar action (independent action, IA), (3) interactions (synergism, potentiation, antagonism) .
Concentration addition (CA), also known as ‘simple joint action’, is based on the idea of a similar action of single compounds, whereas interpretations of this term can differ considerably. From mechanistic point of view, similar action means in a strict sense that single substance should show the same specific interaction with a molecular target site in the observed organisms. This is a nonintereactive process, which means that the chemicals in the mixture do not affect the toxicity of one another. Each of the chemicals in the mixture contributes to the toxicity of the mixture in proportion to its dose, expressed as the percentage of the dose of that chemical alone that would be required to obtain the given effect of the mixture. All chemicals of concern in a mixture act in the same way, by the same mechanisms, and differ only in their potencies [46-47, 52].
It has been shown that the concept of concentration addition is also applicable to nonreactive, nonionized organic chemicals, which show no specific mode of action but whose toxicity toward aquatic species is governed be hydrophobicity. The mode of action of such compounds is called narcosis or baseline toxicity [53-54]. The potency of a chemical to induce narcosis is entirely dependent on its hydrophobicity, generally expressed by its octanol-water partition coefficient logKow. As a result, in the absence of any specific mechanism of toxicity, a chemical will, within certain boundaries, always be as toxic as its logKow indicates. Mathematically, the concept of concentration addition for a mixture of
The alternative concept of independent action (IA), also known as ‘independent joint action’ was already formulated by Bliss . IA is when toxicants act independently and have different modes of toxic action [43, 46-47]. In this case the agents of a mixture do not affect each other’s toxic effect. As a result of such a dissimiliar action, the relative effect of one of the toxicants in a mixture should remain unchanged in the presence of another one. For binary mixture the combination effect can be calculated by the equitation :
Additionally, compounds may interact with one another, modifying the magnitude and sometimes the nature of the toxic effect. This modification may make the composite effect stronger or weaker. An interaction might occur in the toxicokinetic phase (processes of uptake, distribution, metabolism, and excretion) or in the toxicodynamic phase (effects of chemicals on the receptor, cellular target, or organ).These include terms such as synergism and potentiation (i.e., resulting in a more than additive effect), or antagonism (i.e., resulting in a less than additive effect) . It must be highlighted that at given concentrations of the single compounds in a mixture the combination effect will in general be higher if the substances follow the concept of concentration addition. Thus, misleadingly the different concepts were sometimes brought in correlation to the term synergism and antagonism. But synergism or antagonisms between the used substances and their effects can occur independently of a similar or dissimilar mode of action .
For these reasons, prediction of the effect of a mixture based on the knowledge of each of the constituents requires detailed information on the composition of the mixture, exposure level, mechanism of action, and receptor of the individual compounds. However, often such information is not or is only partially available and additional studies are needed. In addition to considering which of these concepts should be used to evaluate combined toxic effects, the design of the study is important in quantifying the combined effects. Most of such studies are based on a comparison of observed values with those predicted by a reference mode (IA or CA). An important aspect of toxicity studies of mixtures is the impracticability of ‘complete’ testing. If all combinations are to be studied at different dose levels, an increasing number of chemicals in a mixture results in an exponential increase in number of test groups: to test all possible combinations (in a complete experimental design) at only one dose level of each chemical in a mixture consisting of 4 or 6 chemicals, 16 (24–1) or 64 (26–1) test groups, respectively, would be required. Such
One approach is to test the toxicity of the mixture without assessing the type of interactions. This is the simplest way to study effects of mixtures by comparing the effect of a mixture with the effects of all its constituents at comparable concentrations and duration of exposure at one dose level without testing all possible combinations of two or more chemicals. This approach requires a minimum number of experimental groups (n + 1, the number of compounds in a mixture plus the mixture itself). If there are no dose-effect curves of each of the single compounds it is impossible to describe the effect of the mixture in terms of synergism, potentiation, antagonism, etc. This strategy would be of interest for a first screening of adverse effects of a mixture.
The second approach is based on assessment of interactive effects between two or three compounds which can be identified by physiologically based toxicokinetic modeling, isobolographic or dose-effect surface analysis, or comparison of dose-effect curves. However, interactive effects of compounds in mixtures with more than three compounds can be best ascertained with the help of statistical designs such as (fractionated) factorial designs, ray designs or dose-effect surface analysis. Here we would like to described shortly only the isobole methods as so far they are mainly used in the studies concerning the determination of pharmaceutical mixture toxicity [25-26,46-47].
An isobole, originally developed by Loewe and Muischnek , is a contour line that represents equi-effective quantities of two agents or their mixtures . The theoretical line of additivity is the straight line connecting the individual doses of each of the single agents that produce the fixed effect alone. The method requires a number of mixtures to be tested and is used for a graphical representation to find out if mixtures of two compounds behave in a dose-additive manner and subsequently can be regarded as chemicals with a similar mode of action. When all equi-effect concentrations are connected by a downward concave line, the effect of the combinations is antagonistic, and a concave upward curve indicates synergism. The use of the isobole procedure to evaluate the effects of binary mixtures is widely used, but is very laborious and requires large data sets in order to produce sufficiently reliable results .
4. State of the knowledge concerning mixture and chronic toxicity of the residues of sulfonamides in the environment
4.1. What do we know about the long-term effects of the presence of the residues of sulfonamides in the environment?
Chronic toxicity tests are studies in which organisms are exposed to different concentrations of a chemical and observed over a long period, or a substantial part of their lifespan. In contrast to acute toxicity tests, which often use mortality as the only measured effect, chronic tests usually include additional measures of effect such as growth rates, reproduction or changes in organism behavior [55-56]. Therefore, the standard acute toxicity tests do not seem appropriate for risk assessment of pharmaceuticals, because of the nature of these compounds. The use of chronic tests over the life-cycle of organisms for different trophic levels could be more appropriate . However, there is still an ongoing debate between ecotoxicologists over the determination which tests should be considered to be chronic or acute (based on their duration). This applies not only to aquatic animal testing with invertebrates and fish, but also to standard 96-h algal and 7-d higher plant test methods.
Molander et al.  reviewed the data published in the
Looking at the available acute toxicity data, it can be concluded that SAs are practically non-toxic to most microorganisms tested including selected strains of bacteria, such as
However, data relating to the long-term exposure of non-target organisms, and especially how continuous exposure for several generations may affect a whole population is very limited. Most chronic toxicity data for sulfonamides, is available for invertebrates, probably because these are the briefest and therefore least expensive chronic toxicity tests to run. Available chronic toxicity data for sulfonamides is summarized in Table 3 and discussed below.
The major concern over the effects of all antimicrobials (including sulfonamides) on microbial assemblages is the development of antimicrobial resistance and the effect of this on public health. Recently, Baran et al.  has reviewed the papers concerning the influence of presence of SAs in the environment to antimicrobial resistance. They concluded that SAs in the environment increase the antimicrobial resistance of microorganisms and the number of bacterial strains resistant to SAs increases systematically in recent years. Resistant bacterial species commonly carried single genes, but in recent years, an increased number of pathogens that possess three SAs-resistant genes have been observed. Moreover, they have also highlighted that these drugs have shown the highest drug resistance, almost twice as high as tetracyclines and many times higher than other antibiotics. Most often, bacterial resistance to SAs has been described in
Additionaly, Heuer and Smalla  investigated the effects of pig manure and sulfadiazine on bacterial communities in soil microcosms using two soil types. In both soils, manure and sulfadiazine positively affected the quotients of total and sulfadiazine-resistant culturable bacteria after two months. The results suggest that manure from treated pigs enhances spread of antibiotic resistances in soil bacterial communities. Monteiro and Boxall  have recently examined the indirect effects of sulfamethoxazole on the degradation of a range of human medicines in soils. It was observed that the addition of SMX significantly reduce the rate of degradation of human non-steroidal anti-inflammatory drugs, naproxen. This observation may have serious implications for the risks of other compounds that are applied to the soil environment such as pesticides.
|EC50, 48h = 221 mg L-1||EC50, 21d = 13.7 mg L-1|||
|(166 – 568 mg L-1)||(12.2 – 15.3 mg L-1)|
|EC50, 24h = 26.27 mg L-1||EC50, 48h = 9.63 mg L-1|||
|(16.32 – 42.28 mg L-1)||(7.00 – 13.25 mg L-1)|
|EC50, 48h = 15.51 mg L-1||EC50, 7d = 0.21 mg L-1|
|(12.97 – 18.55 mg L-1||(0.14 – 0.39 mg L-1)|
|-logEC50, 15 min||-logEC50, 24h|
|3.12 (± 0.04) M|
2.92 (± 0.05) M
3.32 (± 0.04) M
3.32 (± 0.02) M
3.81 (± 0.02) M
3.67 (± 0.03) M
4.30 (± 0.04) M
|4.08 (± 0.06) M|
3.84 (± 0.04) M
4.45 (± 0.05) M
4.50 (± 0.06) M
4.43 (± 0.03) M
5.05 (± 0.05) M
4.78 (± 0.04) M
|EC50, 48h = 131 mg L-1||EC50, 21d = 3.466 mg L-1||[25-26]|
|(119 – 143 mg L-1)||(2.642 – 4.469 mg L-1)|
|EC50, 48h = 3.86 mg L-1||EC50, 21d = 0.869 mg L-1|
|(3.19 – 5.08 mg L-1)||(0.630 – 1.097 mg L-1)|
|EC50, 48h = 202 mg L-1||EC50, 21d = 4.25 mg L-1|
|(179 – 223 mg L-1)||(3.84 – 4.62 mg L-1)|
|EC50, 48h = 215.9 mg L-1||EC50, 21d|||
|(169.6 – 274.9 mg L-1)||no effect up to 30 mg L-1|
|EC50, 48h = 616.9 mg L-1||LOEC = 35 mg L-1|
|(291.7 – 1303.6 mg L-1)|
|EC50, 48h = 110.7 mg L-1||EC50, 8d|
|(89.5 – 136.9 mg L-1)||no effect up to 35 mg L-1|
|EC50, 48h = 391.1 mg L-1||no effect up to 30 mg L-1|
|(341.9 – 440.3 mg L-1)|
Only few studies have also explored effects of SAs on aquatic microbes. It must be highlighted that it was already proved that the effects of antibiotics like SAs on bacteria should not be determined using acute tests. These compounds possess specific mode of action and impacts frequently became evident upon extending the incubation period. Most of the toxicity data available for
Studies conducted on the toxicity mechanism of single SAa indicated that the pKa played a vital role in the toxic effect of SAs or their antibacterial activity . Because LUC (Lucyferase) is an endoenzyme, and SAs have to be transported into the cell before bind with LUC, it was clear that the antibiotic toxicity included both LUC-binding and a toxic transportation effect (which can be described using pKa). pKa is a decisive factor in transporting SAs into the cell. Three species (neutral, cationic and anionic) of SAs depend on the pKa and surrounding pH values. The neutral species have higher cell membrane permeability than anionic species. Therefore, pKa was the key parameter of sulfonamides toxic effects. Some similarity in acute and chronic toxicity mechanisms was observed. However, in conclusion the distinct receptor proteins of SAs in acute toxicity and chronic toxicity led to the different toxicity mechanisms of single antibiotics . A comparison of the results of short and long term bioassays with
Similar conclusions can be obtained if only acute toxicity of SAs to invertebrates is taken into consideration. Detailed information is presented in Table 3. No acute effects on
Unfortunately, there is no information about long-term effects of the residues of these compounds to higher plants and other aquatic as well as terrestrial organisms. Therefore, it seems to be necessary for researchers to study the chronic toxicity of antibiotic [46-47, 55] because of their widespread use and continuous emissions into the environment .
4.2. What do we know about mixture toxicity of the residues of sulfonamides in the environment?
As SAs occur in natural environment not as a single, isolated drug but usually together with other compounds of the same family or the same type, accumulated concentrations or synergistic-antagonistic effects need to be considered. Sulfonamides are widely used in combination therapy together with their potentiator (mostly trimethoprim, TMP) in human and veterinary medicine [1-3]; thus, the occurrence of TMP together with other antibiotics has been commonly detected .
Santos et al.  pointed out that, ecotoxicological data show that the effects of mixtures may differ from those of single compounds. For example, Cleuvers  showed that a mixture of diclofenac and ibuprofen exhibited a greater than predicted toxicity to
|SDM||Evaluation of the toxicity of the mixture of selected compounds based on concentration addition concept.||2.30 mg L-1||Synergistic growth inhibition between SAs and TMP or PMT and for SMX:AcSMX:TMP mixture.|||
|SMX||1.50 mg L-1|
|SDZ||2.19 mg L-1|
|80.8 mg L-1|
|Pyrimethanine (PMT)||5.06 mg L-1|
|SMZ + TMP||according to|
|0.275 mg L-1|
|SDA + TMP||0.465 mg L-1|
|SDM + PMT||2.36 mg L-1|
|SMX:AcSMX:TMP (20:105:3)||0.784 mg L-1|
|SDM:AcSDM:TMP (176:8:1)||2.17 mg L-1|
|SDZ:AcSDZ:TMP (42:24:1)||2.08 mg L-1|
|SDZ||Assessment of interactive effects between two compounds identified by isobologram method.|
|212 mg L-1||Antagonistic interaction for mixtures:|
SMZ + SDZ
SMZ + SGD
SMZ + SMR
SMT + SDM
Complex interaction (synergism additivity and antagonism) for mixture of SMT + SQO
|SGD||3.86 mg L-1|
|SMR||277 mg L-1|
|SDM||202 mg L-1|
|SDMD (SMZ)||270 mg L-1|
|SQO||131 mg L-1|
|TMP||149 mg L-1|
|binary mixtures of SMZ+6 compounds|
|Simple additivity for SMZ + TMP|
|SGD||Assessment of interactive effects between two compounds identified by isobologram method.|
|0.896 mg L-1||Additive (antagonistic) interaction between SQO and SGD.|||
|SQO||3.466 mg L-1|
|43.559 mg L-1|
|SQO||0.246 mg L-1|
|Toxicity of single compound|
(-logEC50, 15 min)
|Toxicity of binary mixture SAs and TMP(-logEC50, mixture)|||
|Antagonistic interaction between SAs and TMP in acute toxicity test.|
15 min and 24 h
|3.12 (± 0.04) M||2.78 (± 0.02) M|
|SPY||Evaluation of the toxicity of the mixture of selected compounds based on concentration addition concept||2.92 (± 0.05) M||2.89 (± 0.02) M|
|SMX||3.32 (± 0.04) M||2.79 (± 0.01) M|
|SDZ||3.32 (± 0.02) M||2.76 (± 0.03) M|
|SSX||3.81 (± 0.02) M||2.79 (± 0.07) M|
|SMM||3.67 (± 0.03) M||2.73 (± 0.02) M|
|SCP||4.30 (± 0.04) M||3.00 (± 0.03) M|
|TMP||3.22 (± 0.07) M|
|Toxicity of single compound|
|Toxicity of binary mixture SAs and TMP|
|SDMD||4.08 (± 0.06) M||5.08 (± 0.05) M||Synergistic interaction between SAs and TMP in chronic toxicity test.|
|SPY||3.84 (± 0.04) M||4.85 (± 0.07) M|
|SMX||4.45 (± 0.05) M||5.50 (± 0.07) M|
|SDZ||4.50 (± 0.06) M||5.42 (± 0.03) M|
|SSX||4.43 (± 0.03) M||5.45 (± 0.03) M|
|SMM||5.05 (± 0.05) M||6.01 (± 0.05) M|
|SCP||4.78 (± 0.04) M||5.73 (± 0.05) M|
|TMP||5.37 (± 0.02) M|
The toxicity of mixture of sulfonamides to non-target organisms was firstly reported by Brain et al.  and Eguchi et al. . Brain et al. investigated the toxicity of the mixture of eight most commonly used pharmaceuticals belonging to different groups (atorvastatin, acetaminophen, caffeine, sulfamethoxazole, carbamazepine, levofloxacin, sertraline and trimethoprim) to the aquatic macrophytes
On the other hand, Eguchi et al.  found that a mixture of trimethoprim or pyrimethamine (pyrimethamine is often used as a substitute for trimethoprim), sulfamethoxazole and sulfadiazine significantly increased growth inhibition (synergistic effect of the mixture was observed) in the algae
Both SAs and TMPs inhibit the folate synthesis pathway in bacteria, but their inhibition sites are different. SAs inhibit dihydropterinic acid synthetase (DHPS), thereby inhibiting the synthesis of folic acid. On the other hand, TMPs inhibits dihydrofolic acid reductase (DHFR), which converts folic acid to 7,8-dihydrofolic acid (7,8- DHF) and 5,6,7,8-tetrahydrofolic acid (5,6,7,8-THF), both active forms of folic acid suitable for utilization. Therefore, the synergistic effect of the combination of SAs and TMPs is likely to be due to the cumulative effect of their actions on two different sites in the folate biosynthesis pathway. Since SAs block the synthesis of folate, the growth inhibitory effect of this compound can be reversed by the addition of folate. In contrast, TMP blocks enzymes downstream of folate in the synthesis pathway, thus addition of folate will not reverse the growth-inhibiting effect of this compound. Since algea also have a similar folate synthesis pathway, the growth inhibitory effect of SAs on these organisms is likely to be the result of the same inhibitory mechanism. Therefore, algal cells could survive in the presence of SAs, but not TMP, when folic acid was added to the medium.
De Liguoro et al. [25-26] evaluated the acute mixture toxicity of combining sulfamethazine with TMP towards
Zou et al.  have recently highlighted that these results cannot represent the mixture toxicity between the SAs and TMP in an actual environment because non-target organisms (microlage and
These examples highlight the fact that the simultaneous presence of several pharmaceuticals in the environment may result in a higher level of toxicity towards non-target organisms than that predicted for individual active substances. More ecotoxicological studies should therefore be done to evaluate the impact of different mixtures of pharmaceuticals in non-target organisms.
The reason for concern regarding risks of mixtures is obvious. Man is always exposed to more than one chemical at a time. This dictates the necessity of exposure assessment, hazard identification, and risk assessment of chemical mixtures. However, for most chemical mixtures data on exposure and toxicity are fragmentary, and roughly over 95% of the resources in toxicology is still devoted to studies of single chemicals. Moreover, organisms are typically exposed to mixtures of chemicals over long periods of time; thus, chronic mixture toxicity analysis is the best way to perform risk assessment in regards to organisms.
However, testing of all kinds of (complex) mixtures of chemicals existing in the real world or of all possible combinations of chemicals of a simple (defined) mixture at different dose levels is virtually impossible. Moreover, even if toxicity data on individual compounds are available, we are still facing the immense problem of extrapolation of findings obtained at relatively high exposure concentration in laboratory animals to man being exposed to (much) lower concentrations.
As stated by several authors, it is essential to investigate if mixtures of pharmaceuticals interact, leading to a larger effect in the environment than would be predicted when each compound is considered individually. Mixtures with antibiotics in the environment may be very complex (e.g. wastewater effluent) but they also may be simple. Although the latter may be more easily studied experimentally, in both cases the identification and quantitative description of synergism caused by specific substances is crucial.
Over past 10 years there has been increasing interest in the impacts of SAs and other veterinary medicines in the environment and there is now a much better understanding about their environmental fate and their impacts on aquatic and terrestrial organisms. However, there are still a number of uncertainties that require addressing before there can be a full understanding of the environmental risks of these compounds. Areas requiring further research are presented below.
The assessment of the potential impacts of those SAs for which ecotoxicity data is lacking but are seen to regularly occur in the environment.
More information about the ecotoxicity of these compounds to soil organisms should be provided. This regards to acute, chronic and single/mixture toxicity of most of the veterinary pharmaceuticals.
Information on the potential environmental effects of parent compounds (drugs) as well as metabolites and transformation products. This includes the single and joint effects evaluation.
Further research is required on the mixture toxicity of SAs in combination with other medicines and non-medicinal substances.
The possible indirect effects of SAs should be identified.
Data from acute and chronic ecotoxicity tests on species belonging to different trophic levels such as bacteria, algea, crustaceans and fish among others, is relevant to illustrate the several adverse effects that environmental exposure to measured concentrations of these contaminants can have. The principal toxicological endpoints/studies that are described are growth, survival, reproduction and immobilization of species, comparatively to trangenerational and population level studies that are still sparse. In the near future, the evaluation of chronic toxicity effects should be set out as a priority for the scientific community since simultaneous exposure to pharmaceuticals, metabolites and transformation products of several therapeutic classes are unknown and whose probable effects on subsequent generations should be assumed.
|DHFR||Dihydrofolic Acid Reductase|
|DHPS||Dihydropterinic Acid Synthetase|
|EPA||Environmental Protection Agency|
|ERA||Environmental Risk Assessment|
|MEC||Measured Environmental Concentration|
|NOEC||No Observable Effect Concentrations|
|PEC||Predicted Environmental Concentration|
|PNEC||Predicted Non-Effective Concentrations|
|SDMD (SMZ)||Sulfadimidine (Sulfamethazine)|
AcknowledgmentsFinancial support was provided by the Polish National Science Centre under grant DEC-2011/03/B/NZ8/03009 “Determining the potential effects of pharmaceuticals in the environment: an ecotoxicity evaluation of selected veterinary drugs and their mixtures” (2012-2015).
Sukul P., Spiteller M. Sulfonamides in the environment as veterinary drugs. Reviews of Environmental Contamination and Toxicology 2006; 187 67-101.
García-Galán MJ., Díaz-Cruz MS., Barceló D. Identification and determination of metabolites and degradation products of sulfonamide antibiotics. Trends in Analytical Chemistry 2008; 27 1008-1022.
Baran W., Adamek E., Ziemiańska J., Sobczak A. Effects of the presence of sulfonamides in the environment and their influence on human health. Journal of Hazardous Materials 2011; 196 1-15.
Briuce PY. Organic Chemistry. Prentice-Hall, Inc., Upper Saddle River; 1995.
Bell PH., Roblin RO. Studies in chemotherapy. Journal of the American Chemical Society1942;64 2905-2917.
Şanli S., Altun Y., Şanli N., Alsancak G., Baltran JL. Solvent Effects on pKa values of Some Substituted Sulfonamides in Acetonitrile-Water Binary Mixtures by the UV-Spectroscopy Method. Journal of Chemical and Engineering Data 2009;54 3014-3021.
Białk-Bielińska A., Stolte S., Matzke M., Fabiańska A., Maszkowska J., Kołodziejska M., Liberek B., Stepnowski P., Kumirska J. Hydrolysis of sulphonamides in aqueous solutions. Journal of Hazardous Materials 2012.; doi:10.1016/j.jhazmat.2012.04.044.
Stoob K. Veterinary sulfonamide antibiotics in the environemnt: fate in grassland soils and transorpt to surface waters. PhD thesis. Swiss Federal Institute of Technology Zurich, Zurich; 2005.
http://www.vcclab.org/lab/alogps/start.html (ALOGPS 2.1 program)
http://www.syrres.com/what-we-do/databaseforms (SRC PhysProp Database)
Babić S., Horvat A. J. M., Mutavdžić Pavlović D., Kaštelan-Macan M. Determination of pKa values of active pharmaceutical ingredients. Trends in Analytical Chemistry 2007;26 1043-1061.
Carda-Broch S., Berthod A. Countercurrent chromatography for the measurement of the hydrophobicity of sulfonamide amphoteric compounds. Chromatographia 2004;59 79-87.
Ruiz-Angel MJ., Carda-Broch S., García-Alvarez-Coque MC., Berthod A. Effect of ionization and the nature of the mobile phase in quantitive structure-retention relationship studies. Journal of Chromatogrphy A 2005;1063 25-34.
Kümmerer, K. Antibiotics in the aquatic environment – A review – Part I, II. Chemosphere 2009;75 417-434.
Petrović M., Barceló D., editors. Analysis, fate and removal of pharmaceutical in the water cycle. Comprehensive Analytical Chemistry. Amsterdam: Elsevier; 2007.
Park S., Choi K. Hazard assessment of commonly used agricultural antibiotics on aquatic ecosystems. Ecotoxicology 2008;17 526-538.
García-Galán MJ., Díaz-Cruz MS., Barceló D. Combining chemical analysis and ecotoxicity to determine environmental exposure and assess risk from sulfonamides. Trends in Analytical Chemistry 2009;28 804-819.
Schauss K., Focks A., Heuer H., Kotzerke A., Schmitt H., Thiele-Bruhn S., Smalla K., Wilke B.M., Matthies M., Amelung W., Klasmeier J., Schloter M. Analysis, fate and effects of the antibiotic sulfadiazine in soil ecosystems. Trends in Analytical Chemistry 2009;28 612-618.
Molander L., Ågerstrand M., Rudén C. WikiPharma – A freely available, easily accessible, interactive and comprehensive database for environmental effect data for pharmaceuticals. Regulatory Toxicology and Pharmacology 2009;55 367-371.
Santos LHLM., Araújo AN., Fachini A., Pena A., Delerue-Matos C., Montenegro, MCBSM. Ecotoxicological aspects related to the presence of pharmaceuticals in the aquatic environment. Journal of Hazardous Materials 2010;175 45-95.
Carlsson C., Johansson AK., Alvan G., Bergman K., Kühler T. Are pharmaceuticals potent environmental pollutants? Part II: Environmental risk assessments of selected pharmaceutical excipients. Science of the Total Environment 2006;364 67-87.
Brain RA., Johnson DJ., Richards SM., Hanson ML., Sanderson H., Lam MW., Young C., Mabury SA., Sibley PK., Solomon KR. Microcosm evaluation of the effects of an eight pharmaceutical mixture to the aquatic macrophytes Lemna gibba and Myriophyllum sibiricum. Aquatic Toxicology 2004;70 23-40.
Brain RA., Ramirez AJ., Fulton BA., Chambliss CK., Brooks BW. Herbicidal effects of sulfamethoxazole in Lemna gibba: Using p-aminobenzoic acid as a biomarker effect. Environmental Science and Technology 2008;42 8965-8970.
Grote M., Chwanke-Anduschus C., Michel R., Stevens H., Heyser W., Langenkämper G., Betsche T., Freitag M. Incorporation of veterinary antibiotics into crops from manured soil. Lanbauforschung Völkenrode 2007;1 25-32.
De Liguoro M., Di Leva V., Gallina G., Faccio E., Pinto G., Pollio A. Evaluation of the aquatic toxicity of two veterinary sulfonamides using five test organisms. Chemosphere 2010;81 788-793.
De Liguoro M., Fioretto B., Poltronieri C., Gallina G. The toxicity of sulfamethazine to Daphnia magna and its additivity to other veterinary sulfonamides and trimethoprim. Chemosphere 2009;75 1519-1524.
Baran W., Sochacka J., Wardas W. Toxicity and biodegradability of sulfonamides and products of their photocatalytic degradation in aqueous solutions. Chemosphere 2006;65 1295-1299.
Eguchi K., Nagase H., Ozawa M., Endoh YS., Goto K., Hirata K., Miyamoto K., Yoshimura H. Evaluation of antimicrobial agents for veterinary use in the ecotoxicity test using microalgae. Chemosphere 2004;57 1733-1738.
Isidori M., Lavorgna M., Nardelli A., Pascarella L., Parrella A. Toxic and genotoxic evaluation of six antibiotics on non-target organisms. Science of the Total Environment 2005;346 87-98.
Kim Y., Choi K., Jung J., Park S., Kim PG., Park J. Aquatic toxicity of acetaminophen, carbamazepine, cimetidine, diltiazem and six major sulfonamides, and their potential ecological risks in Korea. Environment International 2007;33 370-375.
Pro J., Ortiz JA., Boleas S., Fernández C., Carbonell G., Tarazona JV. Effect assessment of antimicrobial pharmaceuticals on the aquatic plant Lemna minor. Bulletin of Environmental Contamination and Toxicology 2003;70 290-295.
Tappe W., Zarfl C., Kummer S., Burauel P., Vereecken H., Groeneweg J. Growth-inhibitory effects of sulfonamides at different pH: Dissimilar susceptibility patters of a soil bacterium and a test bacterium used for antibiotic assays. Chemosphere 2008;72 836-843.
Białk-Bielińska A., Stolte S., Arning J., Uebers U., Böschen A., Stepnowski P., Matzke M. Ecotoxicity evaluation of selected sulfonamides. Chemosphere 2011;85 928-933.
U.S. Deparment of Health and Human Services, Food and Drug Administration. Guidance for Indrustry: Environmental Assessment of Human Drug and Biologics Application; July 1998; www.fda.gov/cder/guidance/1730fnl.pdf.
U.S. Deparment of Health and Human Services, Food and Drug Administration. Guidance for Indrustry: Environmental Assessment for Veterniary Medical Products – Phase I; March 2001; www.fda.gov/cvm/Guidance/guide89.pdf.
CHMP (Committee for Medical Products for Human Use), Guideline on the Environmental Risk Assessment of Medical Products for Human Use, EMEA/CHMP/SWP/4447/00, London 2006.
CVMP (Committee for Medical Products for Veterinary Use), Revised Guideline on the Environmental Risk Assessment for Veterinary Medicinal Products in Support of the VICH GL6 and GL 38, EMEA/CVMP/ERA/418282/2005-Rev.1, London 2008.
VICH (International Cooperation on Harmonization of Technical Requirements for Registration of Veterinary Medical Products), Guideline GL 6 on Environmental Impact Assessment (EIAs) for Veterinary Medicinal Products - Phase I, CVMP/VICH/592/98-FINAL, London 2000.
VICH (International Cooperation on Harmonization of Technical Requirements for Registration of Veterinary Medical Products), Guideline GL 38 on Environmental Impact Assessment for Veterinary Medicinal Products - Phase II, CVMP/VICH/790/03-FINAL, London 2005.
Schmitt H., Boucard T., Garric J., Jensen J., Parrott J., Péry A., Römbke J., Straub JO., Hutchinson TH., Sánchez-Argüello P., Wennmalm Å., Duis K. Recommendations on the Environmental Risk Assesment of Pharmaceuticals – effect characterization. Integrated Environmental Assessment and Management 2010;6 588-602.
Tarazona JV., Escher BI., Giltrow E., Sumpter J., Knacker T. Targeting the environmental risk assessment of pharmaceuticals: fact and fantasies. Integrated Environmental Assessment and Management 2010;6 603-613.
Ankley GT., Brooks BW., Huggett DB., Sumpter JP. Repeating history: Pharmaceuticals in the environment. Environmental Science and Technology 2007;15 8211-8217.
Jonker MJ., Svendsen C., Bedaux JJM., Bongers M., Kammenga JE. Siginificance testing of synergisitic/antagonistic, dose level-dependent, or dose ratio-dependent effects in mixture dose-response analysis. Environmental Toxicology and Chemistry 2005;24 2701-2713.
Loewe S., Muischnek H. über Kombinationswirkungen. 1. Mitteilung: hilfsmittel der fragestellung. Naunyn-Schmiedebergs Archiv für Experimentelle Pathologie und Pharmakologie 1926;114 313-326.
Bliss CI. The toxicity of poisons applied jointly. Annual Review of Applied Biology 1939;26 585-615.
Cleuvers M. Aquatic ecotoxicity of pharmaceuticals including the assessment of combination effects. Toxicology Letters 2003;142 185-194.
Cleuvers M. Mixture toxicity of the anti-inflamatory drugs diclofenac, ibuprofen, naproxen and acetylsalicylic acid, Ecotoxicology and Environmental Safety 2004;59 309-315.
Berenbaum MC. The expected effect of a combination of agents: the general solution. Journal of Theoretical Biology1985;114 413-431.
Altenburger R., Backhaus T., Boedeker W., Faust M., Scholze M., Grimme LH. Predictability of the toxicity of multiple chemical mixtures to Vibrio fischeri: mixtures composed of similarly acting chemicals. Environmental Toxicology and Chemistry 2000;19 2341-2347.
Backhaus T., Altenburger R., Boedeker W., Faust M., Scholze M., Grimme LH. Predictability of the toxicity of a multiple mixture of dissimilarly acting chemicals to Vibrio fischeri. Environmental Toxicology and Chemistry 2000;19 2348-2356.
Faust M., Altenburger R., Bachaus T., Blanck H., Boedeker W., Gramatica P., Hamer V., Scholze M., Vighi M., Grimme LH. Predicitng the joint algal toxicity of multicomponenet s-triazine mixtures at low-effect concentrations of individual toxicants. Aquatic Toxicology 2001;56 13-32.
Cassee FR., Groten JP., van Bladeren PJ., Feron VJ. Toxicological Evaluation and Risk Assessment of Chemical Mixtures, Critical Reviews in Toxicology 1998;28(1) 73–101.
Van Loon WMGM., Verwoerd ME., Eijnker FG., van Leeuwen CJ., van Duyn P., van deGuchte, Hermens JLM. Estimating total body residues and baseline toxicity of complex organic mixtures in effluents and surface waters. Environmental Toxicology and Chemistry 1997;16 1358-1365.
Verhaar HJM., van Leeuwen CJ., Hermens JLM. Classifying environmental pollutants. Chemosphere 1992;25 471-491.
Crane M., Watts C., Boucard T. Chronic aquatic environmental risks from exposure to human pharmaceuticals. Science of the Total Environment 2006;367 23–41.
Waller WT., Allen HJ. Acute and Chronic Toxicity. Ecotoxicology 2008; 32-43.
Blasco J., DelValls A. Impact of emergent contaminant in the environment: Environmental Risk Assessment. Handbook of Environmental Chemistry 2008;5 169-188.
Heuer H., Smalla K. Manure and sulfadiazine synergistically increased bacterial antibiotic resistance in soil over at least two months. Environmental Microbiology 2007;9(3) 657-666.
Monteiro SC., Boxall ABA. Factors affecting the degradation of pharmaceuticals in agricultural soils. Environmental Toxicology and Chemistry 2009;28(12) 2546-2554.
Wollenberger L., Halling-Sørensen B., Kusk KO. Acute and chronic toxicity of veterinary antibiotics to Daphnia magna. Chemosphere 2000;40 723-730.
Zou X., Lin Z., Deng Z., Yin D., Zhang Y. The joint effects of sulfonamides and their potentiator on Photobacterium phosphoreum: Differences between the acute and chronic mixture toxicity mechanisms. Chemosphere 2012;86 30-35.
Kümmerer K., Alexy R., Hüttig J., Schöll A. Standardized tests fail to assess the effects of antibiotics on environmental bacteria, Water Research 2004;38 2111-2116.
Backhaus T., Grimme LH. The toxicity of antibiotic agents to the bioluminescent bacterium Vibrio fischeri. Chemosphere 1999;38 3291-3301.
Bolelli L., Bobrovová Z., Ferri E., Fini F., Menotta S., Scandurra S., Fedrizzi G., Girotti S. Bioluminescent bacteria assay of veterinary drugs in excreta of food-producing animals. Journal of Pharmaceutical and Biomedical Analysis 2006;42 88-93.