Size-selective toxicity effects of the antimicrobial tylosin on estuarine phytoplankton communities

Environmental Pollution 216 (2016) 806e810

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Size-selective toxicity effects of the antimicrobial tylosin on estuarine phytoplankton communities* Allison Kline, James L. Pinckney* Marine Science Program and Department of Biological Sciences, University of South Carolina, Columbia, SC, 29208, USA

a r t i c l e i n f o

a b s t r a c t

Article history: Received 28 April 2016 Received in revised form 16 June 2016 Accepted 21 June 2016 Available online 1 July 2016

The purpose of this study was to determine the lethal and sublethal effects of the antimicrobial tylosin on natural estuarine phytoplankton communities. Bioassays were used in experimental treatments with final concentrations of 5 to 1000 mg tylosin l
1. Maximum percent inhibition ranged from 57 to 85% at concentrations of 200e400 mg tylosin l
1. Half maximum inhibition concentrations of tylosin were ca. 5x lower for small phytoplankton (<20 mm) relative to larger phytoplankton (>20 mm) and suggests that small phytoplankton are more sensitive to tylosin exposure. Sublethal effects occurred at concentrations as low as 5 mg tylosin l
1. Environmental concentrations of tylosin (e.g., 0.2e3 mg l
1) may have a significant sublethal effect that alters the size structure and composition of phytoplankton communities. The results of this study highlight the potential importance of cell size on toxicity responses of estuarine phytoplankton. © 2016 Elsevier Ltd. All rights reserved.

Keywords: Pharmaceutical Tylosin Diatom Cryptophyte Sublethal Inhibition

1. Introduction Salt marsh estuaries are a common feature along the southeastern US and the northern Gulf of Mexico coastlines. These systems provide critical habitat for many larval and juvenile species of fish and invertebrates that use phytoplankton as their primary food source. Changes in the structure and function of phytoplankton communities in these habitats may have major implications for ecosystem trophodynamics and water quality. Furthermore, phytoplankton community changes could have cascading impacts on higher trophic levels and affect both recreational and commercial fisheries. Estuaries are receiving increasing amounts of pollutants including antibiotics and pharmaceuticals that enter the system via sewage treatment facilities and wastewater effluent (HallingSørensen et al., 1998; Kolpin et al., 2002; Daughton, 2004; Benotti and Brownawell, 2007, 2009; Kemper, 2008; Nakada et al., 2008). Many of these compounds and their derivatives are persistent in the environment due to the long degradation half-lives and retention in estuarine waters. As antimicrobial agents, these compounds are effective at eliminating natural microbiota (including

* This paper has been recommended for acceptance by Harmon Sarah Michele. * Corresponding author. E-mail address: [email protected] (J.L. Pinckney).

http://dx.doi.org/10.1016/j.envpol.2016.06.050 0269-7491/© 2016 Elsevier Ltd. All rights reserved.

phytoplankton) under environmental conditions (Halling-Sørensen et al., 1998; Kümmerer, 2003; Kostich and Lazorchak, 2008). Antimicrobial compounds are commonly found in urban estuaries at concentrations high enough to affect the indigenous microbiota (Hirsch et al., 1999; Kolpin et al., 2002; Ashton et al., 2004; Benotti and Brownawell, 2007, 2009; Kemper, 2008). Furthermore, the rapidly growing human population in the coastal zone will likely result in increasing inputs of antibiotics and pharmaceuticals via treated wastewater inputs. As such, these contaminants are an emerging area of concern for the effective management of water quality in estuarine ecosystems. Potential impacts include negative effects on the structure and function of estuarine microbiota and possibly fostering growth of antibiotic-resistant strains or species in the microbiota community of these ecosystems (HallingSørensen et al., 1998; Ellis, 2006). One antimicrobial of concern is the macrolide antibiotic tylosin, which is used worldwide as a prophylactic veterinary treatment as well as a growth additive to animal feed. This compound affects binding to the 50S ribosomal subunit by interfering with prokaryotic protein synthesis. Concentrations of 0.31e3.02 nmol l
1 (0.28e2.77 mg l
1) and 2.84 nmol kg
1 (2.6 mg kg
1) have been reported in water and sediments, respectively (Halling-Sørensen et al., 2000; Kolpin et al., 2002; Calamari et al., 2003; Kim and Carlson, 2007b). Tylosin exhibits minimal photolysis in natural waters (Werner et al., 2007) and thereby is resistant to natural

A. Kline, J.L. Pinckney / Environmental Pollution 216 (2016) 806e810

degradation by sunlight. Additionally, tylosin readily binds to sediment particles. The high binding affinity in combination with a high sediment/water partitioning coefficient results in active residence times exceeding 100 þ days (Halling-Sørensen et al., 2003; Kim and Carlson, 2007a,b). Several studies have demonstrated that tylosin is toxic to many phytoplankton species at nearenvironmental concentrations (Ebringer, 1972; Halling-Sørensen et al., 2000; Eguchi et al., 2004; Yang et al., 2008; Swenson et al., 2012). The purpose of this study was to determine the lethal and sublethal effects of tylosin on natural estuarine phytoplankton communities over a range of tylosin concentrations. We further examined the possibility of size-selective toxicity effects for large vs. small phytoplankton. The goal of this research, in contrast to the more common unialgal culture methods, was to provide insights into potential responses of mixed, natural phytoplankton communities exposed to environmentally realistic concentrations of tylosin.

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0.46  25 cm, 5 mm packing) reverse-phase C18 column in series as the solid phase. A non-linear binary gradient consisting of solvent A (80% methanol: 20% 0.5 M ammonium acetate) and solvent B (80% methanol: 20% acetone) was used for the mobile phase (Pinckney et al., 2001). Absorption spectra and chromatograms (440 ± 4 nm) were obtained using a Shimadzu SPD-M10av photodiode array detector and pigment peaks were identified by comparing retention times and absorption spectra with pure standards (DHI, Denmark). The synthetic carotenoid b-apo-8’-carotenal (Sigma) was used as an internal standard. QA/QC protocols follow the guidelines of Van Heukelem and Hooker (2011). The effective limits of detection for pigments ranged from 0.003 to 0.011 mg l
1 and effective limits of quantification ranged from 0.010 to 0.036 mg l
1. The percent inhibition (%inhibition) for each sample was calculated using the equation:

  rtreatment %inhibition ¼ 100 1
rcontrol

2. Materials and methods The North Inlet e Winyah Bay National Estuarine Research Reserve (NERR), near Georgetown, South Carolina, USA is a euhaline Spartina marsh system with minimal anthropogenic impacts (Allen et al., 2014) (http://www.northinlet.sc.edu). The nearly pristine conditions of this estuary minimize potential experimental artifacts due to acclimation of the local phytoplankton communities to chronic antibiotic exposure (Wirth et al., 1998; Sanger et al., 1999). Water samples were collected from the upper 1 m of the water column using an integrated water sampler (PVC baler) on 4 separate dates in July 2015 at low tide from Oyster Landing (33.349N, 79.189W) in the North Inlet Estuary. Samples were transported (in an ice-cooled container) to the lab and dispensed into 250 ml clear polystyrene culture flasks (VWR, cat. # 10062862) within 3 h of collection. Tylosin (tylosin tartrate salt; MP Biomedicals, cat. # 193454) was dissolved in Milli-Q de-ionized water to make a stock solution (500 mg tylosin ml
1) then added to the experimental treatments to achieve final concentrations of 5, 10, 50, 75, 100, 125, 150, 200, 300, 500, 750, and 1000 mg tylosin l
1. There were a minimum of 3 (usually 5) replicates for each concentration. Controls consisted of sample water without any addition of tylosin. Nitrate and phosphate (20 mM NaNO3, 10 mM KH2PO4 final concentrations) were added to all treatments including the control to minimize nutrient limitation during the bioassays. Light for incubations was supplied using a 91 cm, 4  39 W Ocean Light T5 hood (10,000 K 39 W eTRU fluorescent bulbs) to achieve an irradiance of ca.130 mmol quanta m2 s
1. Light was cycled according to times of sunrise and sunset on the dates the water samples were collected. Temperature was a constant 22  C and approximated the water temperature at the time of collection (20  C). Bioassays were incubated for a period of 48 h. At the end of the incubations, samples were vacuum (
50 kPa) filtered onto Whatman GF/F glass microfiber filters. One half of the sample was pre-filtered through 20 mm nitex mesh to size fractionate samples. Thus each sample was divided into two size fractions; phytoplankton <20 mm, and phytoplankton >20 mm (i.e., whole water <20 mm fraction). Phytoplankton photopigment concentrations were measured using HPLC (Roy et al., 2011). Filters were first lyophilized for 18e24 h at
50  C. Photopigments were then extracted by adding 750 mL of 90% aqueous acetone solvent followed by storage for 12e20 h at
20  C. Filtered extracts (250 mL) were injected into a Shimadzu HPLC with a single monomeric column (Rainin Microsorb, 0.46  1.5 cm, 3 mm packing) and a polymeric (Vydac 201TP54,

Where rtreatment and rcontrol are the pigment concentrations in the treatment and the corresponding control, respectively. The data were fit to a saturating hyperbolic function in the form of:

 y ¼ Imax

x ðKI þ xÞ



Where y is the %inhibition, x is the concentration (mg l
1) of tylosin, Imax is the maximum %inhibition, and KI is the tylosin concentration at which the percent inhibition is one-half of the maximum %inhibition. Non-linear curve-fitting was accomplished using an iterative Levenberg-Marquardt procedure (IBM SPSS Statistics, v. 22). 3. Results Qualitative microscopy revealed that the natural phytoplankton community used for the bioassays was composed primarily of diatoms and cryptophytes with minor concentrations (<5%) of chlorophytes and cyanobacteria. Analysis was limited to diatoms and cryptophytes due to the low numbers of other algal groups. The relative abundances of diatoms and cryptophytes was determined using the biomarker pigments fucoxanthin and alloxanthin, respectively. Chlorophyll a (chl a) was used as an indicator for total phytoplankton biomass. The percent inhibition (%inhibition) relative to the respective controls was calculated for total chl a and each algal group. Control treatments in the bioassays exhibited increases of 430e624% in chl a relative to the initial (time 0) values, showing a strong growth response to the nutrient (N and P) additions. Percent inhibition of chl a for both the >20 and < 20 mm size fractions was plotted to show the overall response of the phytoplankton community to 13 levels of tylosin exposure ranging from 0 to 1000 mg l
1 (Fig. 1). Data for each size fraction were fitted to a saturating hyperbolic curve yielding adjusted coefficients of determination (adj R2) of 0.88 and 0.78 for the >20 and <20 size fractions, respectively (Fig. 1, Table 1). Maximum %inhibition (Imax) was similar for both size fractions (85 vs. 78%), but the half maximum %inhibition (KI) was much higher for the >20 fraction (104 ± 17.0 SE mg tylosin l
1) relative to the <20 fraction (28 ± 5.9) (Table 1). The two curve fits for chl a were significantly different (F2,139 ¼ 19.66, p < 0.01). These results suggest that the smaller phytoplankton (<20 mm fraction) were more sensitive than the larger phytoplankton (>20 mm fraction) to tylosin exposure. Furthermore, the maximum %inhibition was ca. 80%, even at concentrations as high as 1000 mg tylosin l
1, and shows that a small

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A. Kline, J.L. Pinckney / Environmental Pollution 216 (2016) 806e810

the fitted curves for the two size fractions showed an Imax of ca. 80% but the KI values were 123 ± 17.7 and 26 ± 5.6 mg tylosin l
1 for the >20 and < 20 mm fractions, respectively (Table 1). The two curve fits for fucoxanthin were significantly different (F2,139 ¼ 34.91, p < 0.01). The close agreement between the chl a and diatom data is consistent with the observation that most of the phytoplankton community consisted of diatoms. Cryptophytes (alloxanthin) were retained on the 20 mm screen only. Thus the analysis of tylosin exposure responses was limited to the >20 mm size fraction. Response data were fitted to a saturating hyperbolic curve with a resulting adj R2 of 0.67 (Fig. 3, Table 1). The Imax for the fitted data was 57 ± 3.2%inhibition and the KI was 31 ± 9.6 mg tylosin l
1. The cryptophyte response was similar to the smaller (<20 mm) fraction responses of the diatoms and total phytoplankton community. 4. Discussion Fig. 1. Saturating hyperbolic curve fits for the two size fractions of the total phytoplankton community as indicated by concentrations of chlorophyll a. The horizontal dashed lines indicate the half-maximum percent inhibition concentrations (KI) for tylosin.

portion of the phytoplankton community (ca. 20%) was apparently resistant to high levels of tylosin exposure during the 48 h incubation. Diatom (fucoxanthin) %inhibition was calculated for the two size fractions (Fig. 2). Data were fit to saturating hyperbolic curves resulting in adj R2 values of 0.91 and 0.75 for the >20 and <20 mm size fractions, respectively (Table 1). Similar to the results for chl a,

The percent maximum inhibition (Imax) for all algal groups ranged from 57 to 85% at tylosin concentrations as high as 1000 mg l
1. Higher tylosin concentrations were not tested because natural exposure levels are usually much lower than 1000 mg tylosin l
1 and would be unrealistic for estuarine ecosystems (Kolpin et al., 2002; Calamari et al., 2003). A saturating hyperbolic equation, rather than the customary Hill-Slope model (Endrenyi et al., 1975), was used to fit the %inhibition data. We chose this approach because the %inhibition never achieved 100% and the hyperbolic curve appeared to more closely fit the phytoplankton responses. In these experiments, the measured KI is analogous to the more commonly used half maximal effective concentration

Table 1 Summary statistics for saturating hyperbolic curve fits. Size classes were determined by size-fractionated filtration of the incubation water. The sample size for each was 75 samples, Imax is the maximum percent inhibition (±1 standard error), KI is the tylosin concentration (in mg l
1) at which the percent inhibition was one-half of the maximum inhibition (±1 standard error), and Adj R2 is the adjusted coefficient of determination for the fitted curve. Parameter

Size class

Imax

±1 S E

KI

±1 S E

Adj R2

Chlorophyll a (all phytoplankton)

>20 mm <20 mm >20 mm <20 mm Whole water

85 78 81 75 57

4.2 3.1 3.7 2.9 3.2

104 28 123 26 31

17.0 5.9 17.7 5.6 9.6

0.88 0.78 0.91 0.75 0.67

Fucoxanthin (diatoms) Alloxanthin (cryptophytes)

Fig. 2. Saturating hyperbolic curve fits for the two size fractions of diatoms as indicated by concentrations of the accessory photopigment fucoxanthin. The horizontal dashed lines indicate the half-maximum percent inhibition concentrations (KI) for tylosin.

Fig. 3. Saturating hyperbolic curve fits for cryptophytes as indicated by concentrations of the accessory photopigment alloxanthin. The horizontal dashed line indicates the half-maximum percent inhibition concentration (KI) for tylosin.

A. Kline, J.L. Pinckney / Environmental Pollution 216 (2016) 806e810

(EC50). The tylosin concentrations at one-half of the maximum %inhibition (KI) for the two size classes of the algal groups was significantly lower for the smaller size fraction and suggests that smaller phytoplankton are more sensitive to tylosin exposure than larger cells. To our knowledge, this is the first time that this size-specific response to tylosin has been reported for estuarine phytoplankton. We speculate that the larger surface area to volume ratio for smaller phytoplankton may increase their susceptibility to tylosin exposure. The smaller cell volume relative to surface area may result in higher contaminant levels within the cytoplasm and thus the “effective” exposure level would be higher for smaller cells (Mullin et al., 1966; Khoshmanesh et al., 1997). In a similar study, Weiner et al. (2004) report that smaller size phytoplankton with greater surface area to volume ratios incorporate more of the herbicide atrazine and are more sensitive to exposure. Other studies have demonstrated the direct effects of tylosin on microalgae using dose-response experiments but found IC50’s approximating 1 mg l
1 using unialgal cultures (Hagenbuch and Pinckney, 2012). Tylosin is toxic to the freshwater chlorophyte Selenastrum capricornutum (¼ Pseudokirchnerella subcapitata) (Halling-Sørensen, 2000; Eguchi et al., 2004) and the freshwater cyanobacterium Microcystis aeruginosa (Halling-Sørensen, 2000) at EC50 concentrations ranging from 34 to 1380 mg l
1. Division rates of the benthic marine diatom Cylindrotheca closterium were depressed by the common antibiotics ciprofloxacin and lincomycin (Swenson et al., 2012). Similar trends were seen using unialgal cultures of C. closterium and the diatom Navicula ramosissima (Hagenbuch and Pinckney, 2012). Tylosin concentrations ranging from 0.28 to 2.77 mg l
1 in water and 2.6 mg kg
1 in sediments have been reported for other aquatic ecosystems (Halling-Sørensen et al., 2000; Kolpin et al., 2002; Calamari et al., 2003; Kim and Carlson, 2007b). These concentrations are well below the KI values for our experiments (26e123 mg tylosin l
1). However, we show that the <20 size fraction of total phytoplankton (chl a), diatoms (fucoxanthin), and cryptophytes (alloxanthin) were inhibited by 26.8, 27.3, and 14.8% at a tylosin concentration of 5 mg l
1, respectively. This result suggests a significant sublethal impact of tylosin on small phytoplankton at concentrations approximating the upper end of environmental levels. The natural phytoplankton community used in our experiments also contained heterotrophic bacteria in the incubation water. One alternative explanation for our observed responses could be an indirect effect of alterations in the bacterial community. However, the size-specific response of the phytoplankton community suggests that the primary response we observed was likely a direct effect of the tylosin additions. Anthropogenic impacts on estuarine ecosystems are a growing concern due to the importance of these systems as nursery areas for many fish and shellfish species as well as being sites of important biogeochemical processes. Phytoplankton provide a good sentinel for assessing the anthropogenic impacts on these ecosystems due to the large collective surface area of individual cells and their sensitivity to both waterborne and deposited pollutants. Reductions in phytoplankton biomass or changes in community composition resulting from these pollutants may have major implications for the trophodynamics and biogeochemistry of salt marsh estuaries. This study demonstrates not only the sensitivity of estuarine phytoplankton communities to tylosin, but suggests that small cells may be more sensitive to exposure than larger cells. Thus environmental impact studies should include not only a range of species, but also a range of sizes for the indicator species.

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5. Conclusions In the bioassays performed in this study, natural estuarine phytoplankton communities showed both lethal and sublethal responses to exposures to the antimicrobial tylosin. Maximum percent inhibition ranged from 57 to 85% at tylosin concentrations of 200e400 mg tylosin l
1. Half maximum inhibition concentrations of tylosin were ca. 5 lower for small phytoplankton (<20 mm) relative to larger phytoplankton (>20 mm) and suggests that small phytoplankton are more sensitive to tylosin exposure. Sublethal effects occurred at concentrations as low as 5 mg tylosin l
1. Environmental concentrations of tylosin (e.g. 0.2e3 mg l
1) may have a significant sublethal effect that alters the size structure and composition of phytoplankton communities. The results of this study highlight the potential importance of cell size on toxicity responses of estuarine phytoplankton. Acknowledgements Funding for this project was provided by OCE 1156831 (National Science Foundation, Research Experiences for Undergraduates). This is publication 1811 from the Belle W. Baruch Institute for Marine and Coastal Sciences. References Allen, D.M., Allen, W.B., Feller, R.F., Plunket, J.S. (Eds.). 2014. Site profile of the North Inlet e Winyah Bay National Estuarine Research Reserve, Georgetown, SC. 432 pgs. Ashton, D., Hilton, M., Thomas, K.V., 2004. Investigating the environmental transport of human pharmaceuticals to streams in the United Kingdom. Sci. Total Environ. 333, 167e184. Benotti, M.J., Brownawell, B.J., 2007. Distributions of pharmaceuticals in an urban estuary during both dry- and wet-weather conditions. Environ. Sci. Technol. 41, 5795e5802. Benotti, M.J., Brownawell, B.J., 2009. Microbial degradation of pharmaceuticals in estuarine and coastal water. Environ. Pollut. 157, 994e1002. Calamari, D., Zuccato, E., Castiglioni, S., Bagnati, R., Fanelli, R., 2003. Strategic survey of therapeutic drugs in the rivers Po and Lambro in northern Italy. Environ. Sci. Technol. 37, 1241e1248. Daughton, C.G., 2004. PPCPs in the environment: future research - beginning with the end always in mind. In: Kümmerer, K. (Ed.), Pharmaceuticals in the Environment. Springer, Berlin, pp. 463e495. Ebringer, L., 1972. Are plastids derived from prokaryotic micro-organisms? Action of antibiotics on chloroplasts of Euglena gracilis. J. General Microbiol. 71, 31e52. Eguchi, K., Nagase, H., Ozawa, M., Endoh, Y.S., Goto, K., Hirata, K., Miyamoto, K., Yoshimura, H., 2004. Evaluation of antimicrobial agents for veterinary use in the ecotoxicity test using microalgae. Chemosphere 57, 1733e1738. Ellis, J.B., 2006. Pharmaceutical and personal care products (PPCPs) in urban receiving waters. Environ. Pollut. 144, 184e189. Endrenyi, L., Fajszi, C., Kwong, F.H.F., 1975. Evaluation of hill slopes and Hill coefficients when the saturation binding or velocity is not known. Eur. J. Biochem. 51, 317e328. Hagenbuch, I.M., Pinckney, J., 2012. Toxic effect of the combined antibiotics ciprofloxacin, lincomycin, and tylosin on two species of marine diatoms. Water Res. 46, 5028e5036. Halling-Sørensen, B., 2000. Algal toxicity of antibacterial agents used in intensive farming. Chemosphere 40, 731e739. Halling-Sørensen, B., Nielsen, S.N., Lanzky, P.F., Ingerslev, F., Lützhøft, H.C.H., Jørgensen, S.E., 1998. Occurrence, fate and effects of pharmaceutical substances in the environment e a review. Chemosphere 36, 357e393. Halling-Sørensen, B., Lützhøft, H., Andersen, H., Ingerslev, F., 2000. Environmental risk assessment of antibiotics: comparison of mecillinam, trimethoprim and ciprofloxacin. J. Antimicrob. Chemother. 46, 53e58. Halling-Sørensen, B., Sengeløv, G., Jensen, L., 2003. Reduced antimicrobial potencies of oxytetracycline, tylosin, sulfadiazine, streptomycin, ciprofloxacin, and olaquindox due to environmental processes. Arch. Environ. Contam. Toxicol. 44, 7e16. Hirsch, R., Ternes, T., Haberer, K., Kratz, K.L., 1999. Occurrence of antibiotics in the aquatic environment. Sci. Total Environ. 225, 109e118. Kemper, N., 2008. Veterinary antibiotics in the aquatic and terrestrial environment. Ecol. Indic. 8, 1e13. Khoshmanesh, A., Lawson, F., Prince, I.G., 1997. Cell surface area as a major parameter in the uptake of cadmium by unicellular green microalgae. Chem.

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