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The effects of Bisphenol A on the seagrass Cymodocea nodosa: Leaf elongation impairment and cytoskeleton disturbance
Ecotoxicology and Environmental Safety 157 (2018) 431–440
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Ecotoxicology and Environmental Safety journal homepage: www.elsevier.com/locate/ecoenv
The effects of Bisphenol A on the seagrass Cymodocea nodosa: Leaf elongation impairment and cytoskeleton disturbance Ioannis-Dimosthenis S. Adamakisa,1, Paraskevi Maleab, a b
⁎,1
T
, Emmanuel Panterisb
Department of Botany, Faculty of Biology, National and Kapodistrian University of Athens, 15784 Athens, Greece Department of Botany, School of Biology, Aristotle University of Thessaloniki, 54124 Thessaloniki, Macedonia, Greece
A R T I C LE I N FO
A B S T R A C T
Keywords: Bisphenol A Cymodocea nodosa Cytoskeleton disruption Leaf growth Seagrass Stress biomarkers Toxicity
Bisphenol A (BPA) is an emerging pollutant of environmental concern, classified as “moderately toxic” and “toxic”, causing adverse effects on aquatic biota. Although information about BPA toxicity on aquatic fauna is available, the data about BPA effects on aquatic flora remain scarce, missing for marine macrophytes. The effects of environmentally relevant BPA concentrations (ranging from 0.03 to 3 μg L−1) on juvenile leaf elongation and the cytoskeleton (microtubules, MTs and actin filaments, AFs) were studied in the seagrass Cymodocea nodosa for 1–10 days. The suitability of cytoskeleton disturbance and leaf elongation impairment as “biomarkers” for BPA stress were tested. The highest BPA concentrations (0.3, 0.5, 1 and 3 μg L−1) affected significantly leaf elongation from the onset of the experiment, while defects of the cytoskeleton were observed even at lower concentrations. In particular, MTs were initially disrupted (i.e. “lowest observed effect concentrations”, LOECs) at 0.1 μg L−1, while AFs were damaged even at 0.03 μg L−1. AFs appeared thus to be more sensitive to lower BPA concentrations, while there was a correlation between leaf elongation impairment and MT defects. Thus, AF damages, MT disruption and leaf elongation impairment in C. nodosa, in this particular order, appear to be sensitive “biomarkers” of BPA stress, at the above environmentally relevant BPA concentrations.
1. Introduction Worldwide human development and urbanization have led to increased demand for food and beverage packaging, remedial equipment, electronics, flame retardants, adhesives, construction materials, plastic containers and paper coverings (among others see Corrales et al., 2015). The manufacture of the above products requires the substance known as bisphenol A (CAS No. 80-05-7) (BPA; 2,2-bis(4-hydroxyphenyl) propane). As a result, worldwide BPA production (e.g. 3.7 million metric tons per year) and consumption have steadily increased over the years (Corrales et al., 2015). Similarly to other chemicals, marine environment is the major disposal site of BPA pollution. The major inputs of BPA enter the environment from production and processing industries, landfill wastes and effluents of wastewater treatment facilities. BPA discharges are percolated in rivers and streams towards estuarine and coastal systems (e.g. Yamada, 1999; Cousins et al., 2002; Kang et al., 2007; Mihaich et al., 2018). It has been recorded that BPA in municipal and industrial wastes can reach up to 23.02 μg L−1 (Al-Rifai et al., 2007), while in sewage pulps its concentration can be much higher 10-
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1
Corresponding author. E-mail address: [email protected] (P. Malea). These authors contributed equally.
https://doi.org/10.1016/j.ecoenv.2018.04.005 Received 18 February 2018; Received in revised form 2 April 2018; Accepted 4 April 2018 0147-6513/ © 2018 Elsevier Inc. All rights reserved.
> 100,000 μg kg−1 d.wt. (Corrales et al., 2015), and in some cases the reported values extended up to 3.2 × 107 μg kg−1 d.wt. (Harrison et al., 2006). The reported concentrations of BPA in aquatic environments vary, ranging from < 0.0088–1.0 μg L−1 in marine water (Belfroid et al., 2002; Kawahata et al., 2004; Vethaak et al., 2005; Staniszewska et al., 2014), 0.0015–0.145 μg L−1 in estuarine and lagoon water (Fu et al., 2007), > 1 μg L−1 in freshwater systems (Oehlmann et al., 2008; Flint et al., 2012), to elevated concentrations of BPA (100–1000 up to 20,136 and 13,392 μg kg−1 d.wt in Asia and Europe, respectively) in anaerobic sediments (see review by Corrales et al., 2015). Given its presence in the environment, earlier reports by the European Commission and the United States Environmental Protection Agency have classified this compound as “moderately toxic” and “toxic” to aquatic biota (Alexander et al., 1988). Few studies in either laboratory or field settings have examined wildlife organisms’ measurable responses to environmentally relevant BPA concentrations (0.08–12.5 µg L−1; Flint et al., 2012) and their predicted no-effect concentrations have been determined at 0.06 µg L−1 (Wright-Walters et al., 2011). Most of the evidence derived mainly from studies on aquatic invertebrates, such as crustaceans and echinoderms,
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with available data also from insects, fish, amphibians, reptiles and birds (Mihaich et al., 2009; Flint et al., 2012). Plants, especially aquatic ones, have received far less attention (e.g. Mihaich et al., 2009). In particular, the research on BPA toxicity so far has been limited to the freshwater angiosperm Lemna gibba (Mihaich et al., 2009) and to some microalgae species (e.g. Chlorella fusca, Stephanodiscus hantzschii, Chlamydomonas mexicana, Chlorella vulgaris) (Alexander et al., 1988; Hirooka et al., 2005; Li et al., 2009; Ji et al., 2014), in which BPA induced defects in growth, antioxidant enzyme activity and photosynthetic pigments, respectively. In land plants, studies have examined BPA effects on vegetative growth (Ferrara et al., 2006) and some researchers have related the defects of microtubules (MTs) and actin filaments (AFs) induced by BPA treatment with the observed growth inhibition (Adamakis et al., 2013, 2016; Stavropoulou et al., 2018). These potential disturbances, induced by BPA to the cytoskeleton and/or the vegetative growth, could be used as “biomarkers” for the evaluation of BPA contamination in specific marine “bioindicator” organisms. Marine macrophytes, especially the seagrasses, are of particular interest, given their well-known biological and ecological relevance and their sensitivity to anthropogenic disturbances (e.g. Malea, 1994; Malea and Haritonidis, 1995, 1999; Malea et al., 1995; Malea and Zikidou, 2011; Malea et al., 2013a, 2013b, 2013c, 2014). These characteristics mark them as excellent candidate “bioindicator” organisms (see review by Ferrat et al., 2003). Although some “biomarkers” have been determined for BPA toxicity, mainly in aquatic vertebrates and invertebrates, in some microalgae species and in one aquatic angiosperm (Lemna gibba) (Kang et al., 2007; Mihaich et al., 2009 and literature therein), none has been determined in marine macrophyte species and especially in seagrasses. Cymodocea nodosa is a seagrass resilient to anthropogenic disturbances, displaying a fast growth rate (Cabaço et al., 2010). In seagrasses species, among which C. nodosa, the effectiveness of a number of “biomarkers” (e.g. photosynthetic activity, enzymatic activity, heat shock proteins, phenolic compounds, MT arrays, cell mortality, oxidative stress “biomarkers”, biomarkers of detoxification) of various stressors (especially metals, metallic nanoparticles (NPs), pesticides, nutrients, etc.) has been tested and most of them appeared to be valuable and early warning signals for general or particular stress (see Ferrat et al., 2003 for review; Malea et al., 2013a, 2013b, 2014; Papathanasiou et al., 2015; Moustakas et al., 2016). Growth impairment has been examined as an assessor of the aquatic angiosperm health against various chemicals (e.g. Flores et al., 2013; Negri et al., 2015; Papathanasiou et al., 2015; Llagostera et al., 2016). However, the extent by which the effects on the cytoskeleton and ultrastructural and mitotic defects (Adamakis et al., 2013, 2016) may result in more obvious plant physiological and morphological responses (e.g. leaf elongation) remain unexplored. Considering the above, in the present study we sought to investigate the effects of environmentally relevant BPA concentrations on the juvenile leaf growth (expressed as elongation), in relation to the defects in cytoskeleton organization in leaf cells of the seagrass C. nodosa. Moreover, this study contributes to the existing knowledge on the “lowest observed effect concentrations” (LOECs) and hence on “no observed effect concentrations” (NOECs) on this marine angiosperm, which can be used in risk assessment programmes. The suitability of MT and AF disturbance and of leaf elongation impairment as suitable “biomarkers” of BPA-induced stress is discussed.
monospecific meadow. Leaf biomass and leaf blade length display an almost unimodal annual pattern attaining maximum values in JulyAugust (e.g. Malea and Zikidou, 2011). Plant collection was done at 0.7–1.0 m depth in July 2017 with a 20 cm diameter acrylic corer, which penetrated to a depth of 30 cm. All plants were rinsed in seawater at the collection site and transported to the laboratory in plastic containers containing seawater. 2.2. Treatments and experimental conditions Collected plants without epiphytes consisting of roots, plagiotropic (horizontal) and orthotropic (vertical) rhizomes and leaves were kept for 24 h in filtered seawater (0.45 µm Whatman GF/C) under laboratory conditions to equilibrate, under a constant 16 h day/8 h night regime at an ambient temperature of 21 ± 1 °C with light intensity set at 120 μmοl m−2 s−1. The seawater used for all experiments was collected from a nearby coastal area (Epanomi site, Thermaikos Gulf). All the chemicals and reagents used were purchased from Applichem (Darmstadt, Germany), Merck (Darmstadt, Germany) and Sigma (Taufkirchen, Germany), unless stated otherwise. Two different sets of experiments were performed, both run in triplicates. In the first set about 24 orthotropic rhizomes attached to part of plagiotropic rhizomes with roots were incubated in non-BPA-based polypropylene (PP) copolymer (PPCO) aquaria covered with plastic foil to prevent evaporation. Plants were treated with 0.03, 0.1, 0.3, 0.5, 1 and 3 μg L−1 bisphenol (BPA) solutions in filtered seawater. For control, plants incubated in filtered seawater, were used. The solutions in the aquaria were changed every two days and were constantly aerated using aquarium air pumps. From each aquarium, four seagrass shoots (totally 12 shoots per BPA treatment) were randomly removed after 1, 2, 4, 6, 8 and 10 days and the meristematic region of the youngest juvenile leaf blade from each shoot was used for either tubulin immunostaining or AF staining. In the second set of experiments, a similar experimental design, as above was generally applied. In each aquarium, in about four orthotropic rhizomes of C. nodosa attached to part of plagiotropic rhizomes with roots (about 12 shoots in each BPA treatment) the youngest accessible juvenile blades were marked by an inert sealant at their base and their initial lengths were measured (mean ± SE: 33.908 ± 2.034 mm). Leaf elongation (mm) was carefully measured as the newly formed leaf segments, hence the distance between the marker and the juvenile leaf base on the 2nd, 4th, 6th, 8th and 10th day (Mateos Naranjo et al., 2008). Leaf elongation rate (Negri et al., 2015) (expressed as mm day−1) were also calculated as the leaf elongation (mm) among two consecutive days divided by the number of days that they mediate. The final length of each measured juvenile leaf, at the end of the experiment were also measured. 2.3. Imaging of MTs and AFs Whole mount MT immunostaining was conducted in hand cut small juvenile leaf pieces, following the protocol of Katsaros and Galatis (1992) using anti-α-tubulin (YOL1/34, AbD Serotec, Kidlington, UK) and FITC-anti-rat secondary antibody (Invitrogen, Carlsbad, CA), both diluted at 1:80 in phosphate buffered saline (PBS). DNA was counterstained with 10 μgmL-1 DAPI (4′,6-diamidino-2-phenylindole) in PBS and the leaf pieces were finally mounted in an anti-fade solution. AF staining with TRITC-phalloidin was conducted according to Panteris et al. (2009), with some modifications as follows: Leaf pieces were firstly incubated with 300 μM m-maleimidobenzoyl- N-hydroxysuccinimide ester (MBS) in PBS + 0.5% (v/v) Triton X-100 for 30 min in the dark for F-actin stabilization. After AF stabilization, fixation was performed in 4% (w/v) paraformaldehyde (PFA) in the same buffer for 60 min. In the PFA solution 1% (v/v; diluted from a 10 μΜ stock solution in methanol) TRITC-phalloidin was added. The tissue sheets were then rinsed with PBS and extracted in 5% (v/v) DMSO
2. Materials and methods 2.1. Plant collection Cymodocea nodosa (Ucria) Ascherson 1870 was collected from the eastern coast of the Gulf of Thessaloniki, Northern Aegean Sea at Viamyl site (40°33 ´N, 22°58´E), where it forms a continuous 432
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+ 5% (v/v) Triton X-100 in the same buffer for 1 h. Finally, AFs were stained with 10% (v/v; same as before) TRITC-phalloidin in PBS in the dark, for 1 h at 37 °C or overnight at room temperature. As above, following the DNA counterstaining step leaf pieces were mounted with an anti-fade solution. All the fluorescent specimens were examined with a Zeiss Observer. Z1 microscope, equipped with the LSM780 confocal laser scanning system, and images were acquired with ZEN2011 software following the manufacturer's instructions. The laser gain was kept stable in all experiments and all digital images were improved for contrast and color with Adobe Photoshop CS2 software with only linear settings. 2.4. Fluorescence intensity measurements Fluorescence intensity of total tubulin was measured at single cortical or central CLSM sections of variously treated leaf tissues, using the Image J ((http://rsbweb.nih.gov/ij/) software, as already described by (Adamakis et al., 2014). From each treatment, fluorescence intensity from 12 individual cells was measured and the mean values were calculated and expressed as fluorescence intensity percentage of control. 2.5. Data analysis The data on leaf blade elongation in time were fitted to different regression models (linear, logarithmic, inverse, exponential, power), using IBM Statistics SPSS® 24; in each case, the model that provided the best fit (i.e., at the highest significance level) was chosen. Statistical analysis was also carried out using Statistica v.12 (Statsoft Inc.). The significance of the variation in the juvenile leaf elongation (mm), the leaf elongation rate (mm day−1) and the microtubule integrity (%) of juvenile blade cells over the incubation time was tested with the non parametric test, Kruskal-Wallis one-way analysis of variance. MannWhitney U-test was applied to compare the mean values of the studied parameters (juvenile leaf elongation, leaf elongation rate, tolerance index, microtubule integrity) between control and various BPA concentrations or various BPA concentrations by twos. Spearman's Rank correlation coefficient was calculated to identify correlations. Tolerance indices of the juvenile leaves were calculated by the following formula (e.g. Idris et al., 2012):
Tolerance index (%) = (blade length of stressed plant / blade length of control plant )* 100 “No observed effect concentration” (NOEC) is the highest concentration of a toxicant that causes no observable adverse effects on the test organisms and calculated as the highest concentration in the test with a mean response not differing significantly from the mean response of the control. “Low observed effect concentration” (LOEC) is the lowest test concentration having a mean response that differs significantly from that of the control (e.g. Crane and Newman, 2000). For both leaf elongation tolerance and microtubule integrity the 10% and 50% effect concentrations (EC10 and EC50 with 95% confidence limits) at the end of the experiments (day 10) were estimated by probit analysis with XLSTAT 2018 (New York, NY, USA).
Fig. 1. (A) Time course of juvenile leaf elongation (mm), (B) leaf elongation rate (mm day−1) (mean ± SE) in control and in various BPA concentrations and (C) of the Tolerance index of juvenile leaf (mean ± SE).
logarithmic equation provided the best fit (i.e. at the highest significance level) r2: 0.994 for control; 0.987 for 0.03 μg L−1, p < 0.001; r2: 0.987 for 0.3 μg L−1; 0.992 for 0.5 μg L−1; 0.969 for 3 μg L−1, p < 0.01; however, at the 0.1 and 1 μg L−1 treatments, a linear equation best described leaf elongation (r2: 0.988 and 0.965, respectively, p < 0.01. Mann Whitney U test showed that leaf elongation, at each incubation time, did not significantly differ (p > 0.05) between the lowest BPA concentrations (0.03 and 0.1 μg L−1) (with the exception of 0.03 μg L−1 on the 4th day) and those measured in the control, which corresponds to the non-observed inhibition of the leaf elongation at the lowest BPA levels (Table 1, Fig. 1A). In contrast, at the highest BPA concentrations (0.3, 0.5, 1 and 3 μg L−1) the inhibition of leaf elongation started from the beginning and continued up to the end of the experiment (p < 0.05) (Table 1, Fig. 1A).
3. Results 3.1. BPA effects on juvenile leaf elongation Generally, a gradual inhibition of juvenile blade elongation with increasing external BPA concentrations was observed from the beginning of the experiment (Fig. 1A). The Kruskal-Wallis analysis of variance showed that the above time variation was only significant on control (Chi square=16.0, df=4, p < 0.01) and on the lowest BPA concentrations (0.03, 0.1 and 0.3 μg L−1) (Chi square: 10.577, 9.643 and 12.180, respectively, df = 4, p < 0.05). In most of the cases, a
3.2. Leaf elongation rates Time variation in leaf elongation rate (mm day−1) (Fig. 1B) was 433
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of MT immunofluorescence are presented in Fig. 3. In untreated interphase cells cortical MTs were randomly oriented, organized in a dense network (Fig. 3A), as already reported by Malea et al. (2013b). The above network remained unaltered even at the end of the experiment (Fig. 3B). In BPA-affected juvenile leaves of C. nodosa MTs appeared fragmented in all the leaf blade cells studied and at all the concentrations studied, except for 0.03 and 0.1 μg L−1. At 0.03 μg L−1 no effect was observed even at the last day (Figs. 2 and 3C), while at 0.1 μg L−1 interphase cortical MTs appeared bundled at the 8th day of treatment and thereafter (Figs. 2 and 3D). Higher BPA concentrations (1 and 3 μg L−1) affected MTs more quickly and rigorously (already from the 2nd day and thereafter) than the lower ones (0.3, 0.5 μg L−1), which required relatively longer exposure times (6th and 4th day, respectively) to manifest their deleterious effect (Figs. 2 and 3E-H). Fluorescence intensity measurements (Fig. 4A) further consolidated the above results, clearly indicating the depolymerizing effect of BPA. Fluorescence intensity dropped in a time- and concentration-dependent way; the Kruskal-Wallis analysis showed that the time variations in MT integrity (%) up to ten days were only significant at the highest BPA treatments (at 0.3, 0.5, 1 μg L−1, chi square = 19.412, 15.447, 15.772, respectively, p < 0.01 and at 3 μg L−1, chi square=21.473, p < 0.001). Moreover, Mann Whitney U test showed that at low BPA concentrations (0.1, 0.3 and 0.5 μg L−1) significant differences in fluorescence intensity measurements, in comparison to the control, were observed after the sixth day of the experiment, whereas at the highest concentrations (1 and 3 μg L−1) even from the first day of the experiment (Fig. 4B).
Table 1 Comparison (Mann-Whitney U-test) of juvenile leaf elongation (mm) between control and various BPA concentrations each day of the experiment. Exposure concentrations (μg L−1) Control-0.03 Control-0.1 Control-0.3 Control-0.5 Control-1 Control-3
Z Z Z Z Z Z
adjusted adjusted adjusted adjusted adjusted adjusted
2nd day
4th day
6th day
8th day
10th day
1.225 ns 0.707 ns 2.449* 2.449* 2.121* 2.309*
2.205* 1.768 ns 2.205* 2.449* 2.121* 2.309*
1.715 ns 1.414 ns 1.960* 2.449* 2.121* 2.309*
1.715 ns 1.061 ns 1.960* 2.449* 2.327* 2.309*
1.225 ns 1.061 ns 1.960* 2.449* 2.121* 2.309*
* : p < 0.05, ns: non significant. Table 2 Leaf elongation rate (mm day−1) and Tolerance index in each BPA concentration over the incubation period.
Leaf elongation rate
Tolerance index
Exposure concentrations (μg L−1)
Mean ( ± SE)
Range
0.03 0.1 0.3 0.5 1 3 0.03 0.1 0.3 0.5 1 3
75.585 ( ± 1.723) 76.170 ( ± 2.016) 58.906 ( ± 1.372) 50.715 ( ± 0.328) 43.680 ( ± 1.860) 30.277 ( ± 1.142) 4.647 ( ± 0.904) 4.701 ( ± 0.986) 3.663 ( ± 0.778) 3.126 ( ± 0.627) 2.715 ( ± 0.598) 1.896 ( ± 0.421)
70.843–79.627 71.814–82.917 56.231–63.119 49.575–51.284 37.375–46.703 26.891–32.859 1.540–6.506 1.653–7.220 1.222–5.496 1.068–4.454 0.960–4.067 0.630–2.855
3.5. Effects on AFs In general, interphase meristematic cells of untreated C. nodosa leaves exhibited abundant cortical and endoplasmic AFs (Fig. 5A). Cortical AFs were randomly oriented, forming a rather complicated mesh, while dividing cells exhibited intense F-actin signal at both sides of the division plane (Fig. 5A). In leaf cells exposed to BPA, AFs were variously affected, depending on treatment duration and BPA concentration. At 1 and 3 μg L−1 for 2 days, AFs tended to aggregate in thick bundles (Fig. 5E, F). Especially at 3 μg L−1 BPA treatment thick Factin aggregates were formed (Fig. 5F). Lower concentrations (0.03, 0.1 and 0.5 μg L−1) also caused damage on AFs (diminishing or even almost complete depolymerization), but required longer exposure times (6th, 6th and 4th day, respectively) to exert their toxic effect (Fig. 5B-D).
only significant at the control treatment (Kruskal Wallis analysis, Chisquare = 10.000 df = 4, p < 0.05). However, at the control and at the lowest BPA treatments (0.03–0.5 μg L−1) the elongation rate increased from 2nd to 4th day and subsequently gradually decreased up to the end of the experiment (Fig. 1B). At the highest concentrations (0.3–3 μg L−1), significantly (p < 0.01 in 0.3 μg L−1 and p < 0.05 in the rest treatments) decreases were observed between the 8th and 10th day (Fig. 1B). A gradual decrease in the mean ( ± SE) leaf elongation rate (mm day−1) of C. nodosa with increasing BPA concentrations was observed (Table 2). Mann Whitney analysis showed that, at the highest concentrations (0.5 up to 3 μg L−1) the mean elongation rate significantly (p < 0.05) differed from those in control treatment throughout the incubation time.
3.6. Relationship between inhibition of leaf elongation and cytoskeleton disruption At the lowest BPA concentrations (0.03 and 0.1 μg L−1), only a slight (p > 0.05) gradual inhibition of leaf elongation was obvious, even from the first day of the experiment (Fig. 1A, Table 1); in fact, elongation inhibition was significant (p < 0.05) at concentrations equal/higher than 0.3 μg L−1, even from the beginning of the experiment and continued up to the end of the experiment (Fig. 1A, Table 1). Cortical MT disruption at 0.1 μg L−1 BPA was obvious at extended treatment, as confirmed from both fluorescence intensity measurements (6th day, Fig. 4B) and immunofluorescence images (8th day, Figs. 2 and 3D), showing that MT disturbance occurred at lower concentrations than those required to impair leaf growth. A significant positive correlation (Spearman's rank order correlation, r = 0.714, p < 0.001) among the juvenile leaf elongation rate and MT integrity (%; based on fluorescence intensity) was recorded. Similarly, a positive correlation (r = 0.769 p < 0.001) among the leaf tolerance index and MT integrity (%) over the incubation time was observed. Moreover, AF damages were recorded earlier than MT disturbance and leaf growth
3.3. Tolerance index A high significant (p < 0.001) gradual decrease of the mean ( ± SE) “Tolerance Index” with increasing BPA concentrations (chisquare=30.000 df=5, p < 0.001) was observed (Fig. 1C, Table 2). In particular, the “Tolerance Index” differed significantly (p < 0.01) between the two highest concentrations. In contrast, the variation of this index was not time-dependent (p > 0.05), as shows Kruskal-Wallis analysis. 3.4. Effects on MT integrity The concentration- and time-depended MT disturbance are given in a schematic representation (Fig. 2), while representative micrographs
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Fig. 2. Schematic representation of microtubule disruption in regard with time and BPA concentrations. The explanations of the schematics appear at the bottom of the figure.
treatments or after shorter treatments (up 6 days) with concentrations much higher (6–100 mg L−1) than those applied in our experiments (Ferrara et al., 2006; Adamakis et al., 2013, 2016; Frejd et al., 2016; Jayasri et al., 2016; Li et al., 2018; Stavropoulou et al., 2018). In C. nodosa juvenile blade leaves, a gradual inhibition of elongation with increasing external BPA concentrations was observed (Fig. 1A). This inhibition seemed to be significant only at higher BPA concentrations, as juvenile leaf elongation at the lowest BPA concentrations (0.03 and 0.1 μg L−1) was similar to that of the control. On the 4th day at 0.03 μg L−1 a significant elongation inhibition was recorded (Τable 1). Nevertheless, this tendency did not continue for the following days (Table 1). As a result, the overall leaf elongation was not inhibited at the lowest BPA concentrations (Kruskal-Wallis analysis of variance), indicating that leaf growth may not be significantly inhibited at such mild concentrations. This finding is in accordance with previous studies, in which BPA, at such environmentally relevant concentrations, did not affect other aquatic organisms (chironomids, copepods, gobies), according to other measurable “stress biomarkers” (e.g. larval emergence, estrogen synthesis), (e.g. Marcial et al., 2003; Watts et al., 2003; Baek et al., 2007). On the contrary, at relatively higher BPA concentrations (0.3, 0.5, 1 and 3 μg L−1) the inhibition of C. nodosa leaf elongation started from the onset of the experiment and continued up to the end of the experiment (Fig. 1A, Table 1). Similarly, at these BPA levels leaf elongation rates (mm day−1) and their “Tolerance Indices” gradually decreased by increasing BPA exposure (Fig. 1B, C). In addition, the elongation rate and the “Tolerance Index” seemed not to be time-dependent under acute BPA toxicity, which is in accordance with the findings of Mihaich et al. (2009) in Lemna gibba. The inhibition of leaf elongation, as mentioned in previous studies (e.g. under metal or herbicide effects), may be due to adverse effects on the metabolism and physiological reactions (e.g. Chesworth et al., 2004; Idris et al., 2012). On the other hand, disruption of the cytoskeleton may also be a cause for decreased leaf elongation. In particular,
impairment, while their disruption occurred even at lower BPA concentrations (0.03 μg L−1) on the 6th day (Fig. 5B). According to leaf elongation impairment, MT disruption (fluorescence intensity, immunofluorescence images) and AFs damages, at various BPA concentrations, at the end of the experiment (10th day) (Table 1, Figs. 2 and 4 B, 5), it is obvious that the “no observed effect concentration” value (NOEC) was < 0.03 μg L−1, according to AF disruption, 0.03 μg L−1, according to MT disturbance, and 0.1 μg L−1, according to leaf elongation impairment. The respective “low observed effect concentrations” (LOECs) values were 0.03 μg L−1, 0.1 μg L−1 and 0.3 μg L−1 Based on MT fluorescence intensity (%), EC50 (95% confidence limits) on the 10th day was 0.225 (0.211–0.241) μg L−1 and EC10 was 0.570 (0.531–0.616) μg L−1, while based on leaf elongation impairment the respective EC50 (95% confidence limits) was 0.140 (0.129–0.151) μg L−1 and the EC10 was 0.500 (0.464–0.544) μg L−1.
4. Discussion 4.1. BPA effects on leaf elongation and the cytoskeleton MTs and AFs have been shown to be some of the prime sub-cellular targets of BPA toxicity in land plants (Adamakis et al., 2013, 2016; Stavropoulou et al., 2018). Similarly, in the present study BPA caused a concentration-dependent disruption of MTs and AFs (Figs. 2–5), further supporting the conclusion that plant MTs and AFs are a target of BPA toxicity, also in a marine angiosperm. Notably, however, the BPA concentrations applied on C. nodosa were significantly lower than those used in other studies on land plants (corresponding to those used in experiments in which BPA affected MT organization, e.g. George et al., 2008). In land plants, BPA arrested root growth in seedlings of various plant species in a concentration- and time-dependent manner. Root growth inhibition was drastic after prolonged (up to 21 days)
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expanding cell wall (Mathur et al., 1999; Mathur and Hülskamp, 2002). Experimental disruption of AFs and/or arrest of actomyosin-mediated mobility have been shown to result in inhibition of elongation in root cells (Baskin and Bivens, 1995; Baluška et al., 2001; Panteris et al., 2013). Accordingly, inhibition of C. nodosa leaf elongation could be also attributed to AF disruption by BPA. In such a view, the effects on both AFs and MTs could result in a general disruption of leaf elongation due to BPA. 4.2. Establishing reliable “biomarkers” Most of the BPA concentrations selected for the treatments performed in this study were within the range reported in seawater (< 0.0088–1.081 μg L−1) (e.g. Mihaich et al., 2018), in estuarine and lagoon water (0.0015–0.145 μg L−1) (Belfroid et al., 2002; Kawahata et al., 2004; Fu et al., 2007), in freshwater (> 1 μg L−1) (e.g. Flint et al., 2012) and in ground-water (1.9 μg L−1) (e.g. Loos et al., 2010), whereas the highest BPA concentration (3 μg L−1) was selected to be in the range of those recorded in wastewater treatment facilities (WWTP) (0.087 – 5.625 μg L−1) (e.g. Pothitou and Voutsa, 2008). Moreover, BPA is moderately soluble, with a measured solubility 120 mg L−1 at 20 °C in sea water (among other see Staples et al., 1997; Ying and Kookana, 2003). So, all of the BPA concentrations applied here were at environmentally relevant levels, thus our results have environmental significance, contributing to the existing knowledge on BPA toxicity by determining the NOEC, LOEC, EC50 and EC10 values. BPA effects on aquatic animals (invertebrates, fish, amphibians, reptiles, birds, mammals etc), even at environmentally relevant BPA concentrations (0.08–12.5 μg L−1), have been previously reported (Flint et al., 2012; Canesi and Fabbri, 2015). Among the reported effects on animals were: delayed emergence of chironomid larvae (at 0.08 μg L−1), developmental inhibition of marine copepods (at 0.1 μg L−1), mortality of freshwater snails (at > 1 μg L−1), altered blood sex hormone ratios (at > 1 μg L−1), spontaneous metamorphosis inhibition of the clawed frog (at 2.3 μg L−1), premature metamorphosis of marine polychaetes (11.4 μg L−1), estrogen synthesis inhibition in longchin goby (0.1 μg L−1), accelerated larval development in marine copepods (12.5 μg L−1) (Flint et al., 2012 and literature therein). An aquatic hazard assessment determined the predicted no-effect BPA concentration to 0.06 μg L−1, using as “biomarker” the development, reproduction and survival of wildlife (Wright-Walters et al., 2011). The NOECs, in various invertebrate species, ranged from 0.025 to 4.1 mg L−1 (Mihaich et al., 2009, 2018) and in vertebrate species from 0.016 to 3.6 mg L−1 (Staples et al., 2011). Accordingly, the available literature provides evidence that relatively low BPA concentrations (encompassing the environmental range; ≤12 µg L−1) affected the physiological homeostasis and gene expression of aquatic animals (Canesi and Fabbri, 2015). However, in aquatic microalgae and angiosperms, the effective BPA doses were significantly higher from the reported environmentally relevant concentrations, as it will be further on analyzed. In the microalgae Skeletonema costatum and Selenastrum capricornutum, the NOEC values were 0.4 and 1.2 mg L−1, respectively (Alexander et al., 1988), whereas in the aquatic angiosperm Lemna gibba the NOEC value, based on growth rate, was 7.8 mg L−1 (Mihaich et al., 2009). As far as it concerns EC50, the acute invertebrate BPA toxicity presented values ranged from 1.1 to 16 mg L−1 (Mihaich et al., 2009 and literature within). In fresh- or seawater microalgae (Chlorella, Cyclotela, Scenedesmus, Selenastrum, Skeletonema, Stephanodiscus spp) the respective EC50 values ranged from 1 mg L−1 (96 h) to 89 mg L−1 (72 h) (Alexander et al., 1988; Li et al., 2009; Zhang et al., 2012). Concerning the aquatic angiosperm L. gibba the ECs50, based on growth rate and on frond density were 32 and 29 mg L−1, respectively (Mihaich
Fig. 3. Tubulin immunofluorescence images of interphase protodermal cells at cortical CLSM sections. The treatment and relevant duration are stated on each figure. Untreated cells bare a dense network of randomly oriented MTs at both the onset (0 days; A) and the end (10 days; B) of the experiment. Treatment with 0.03 μg L−1 BPA did not have any effect even at the 10th day of exposure (C). Upon treatment with BPA 0.1 μg L−1, MTs appear bundled at longer exposure (8th day; D), while in the rest exposure concentrations (0.3–3 μg L−1; EH) MTs were depolymerized/fragmented, with higher concentrations (1, 3 μg L−1) exerting their effect earlier (4th, 2nd day, respectively; G, H). Scale bar: 5 µm.
cortical MTs control cellulose microfibril orientation in the cell wall, establishing thus elongation axes (Baskin, 2001; Nick, 2008), and are reorganized in response to endogenous signals or various environmental stimuli (e.g. Hashimoto, 2015). Interestingly, in the present study the first obvious signs of MT disturbance were observed before any leaf growth impairment could be recorded. Accordingly, inhibition of elongation could be an outcome of disordered cellulose deposition in the cell wall, due to disrupted cortical microtubules (Fujita et al., 2011). Furthermore, AFs of C. nodosa leaf cells were also affected by BPA, even before MT disruption. In plants, AFs provide the principal tracks on which organelles and vesicles move (Duan and Tominaga, 2018). There is substantial genetic and pharmacological evidence that the dynamic plant actin cytoskeleton has an important role in cell morphogenesis, regulating the delivery of Golgi-derived materials to the
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Fig. 4. (A) Time course of microtubule integrity expressed as percentage (%) of cell fluorescence intensity (mean of 12 individual cells ± SE) in the control and at various BPA concentrations and (B) comparisons (Mann Whitney U-test) of microtubule integrity (%) (based of fluorescence intensity measurements) of juvenile blade cells between control and various BPA concentrations each day of the experiment.
a “biomarker” for a given stressor. Changes in MT pattern were suggested as a suitable specific “biomarker” of cadmium or tungsten stress in land plants (Fusconi et al., 2007; Adamakis et al., 2010). Similarly, in C. nodosa MT defects could be used as an early warning signal of emerging lead, copper, nickel, chromium and cadmium stress (Malea et al., 2013a, 2013b, 2014). Taking into account the determined NOEC values (terminal time: 10th day) of C. nodosa (0.03 μg L−1, based on MT disturbance and < 0.03 μg L−1, based on AF damages) and the EC50 (95% confidence limits), based on MT disruption (fluorescence intensity) (0.225 (0.211–0.241) μg L−1), it is obvious that these values are much lower than those of other “stress biomarkers” observed in other aquatic organisms. According to all the above findings, AF damages, MT disruption and leaf elongation impairment in C. nodosa, in this particular order, appear
et al., 2009). In the present study, the NOEC and LOEC values (0.1 μg L−1 and 0.3 μg L−1, respectively) determined on the basis of C. nodosa leaf elongation, at the 10th day were much lower than those earlier reported for the freshwater angiosperm L. gibba, based on the 7th day frond growth rate, as well as for other aquatic organisms on the basis of various “biomarkers”. In addition, EC50 (95% confidence limits) based on leaf elongation impairment: 0.1340 (0.129–0.151) μg L−1 was much lower than that recorded in other aquatic organisms, as previously mentioned. Thus, leaf elongation of C. nodosa seems to be an appropriate “biomarker” under BPA effect. Furthermore, since reorganization of the cytoskeleton under abiotic stress is crucial for cell survival responses (among others see Soda et al., 2016), it could serve as an early measurable warning sign, constituting 437
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Fig. 5. TRITC-phalloidin staining of F-actin. Maximum intensity projections of serial CLSM sections of untreated (A) and BPA-treated (B-F) protodermal leaf cells. In the control AFs form a mesh of randomly aligned AFs (A). Dividing cells (asterisks in A) exhibit intense F-actin signal at either sides of the division plane (arrows in A). In BPAtreated leaves, AFs appear depolymerized, except from the thickest AF bundles. This effect occurs early (2nd day; E, F) at higher concentrations (1–3 μg L−1), while at lower concentrations (0.03, 0.1, 0.5 μg L−1) it occurs at longer exposure duration (6th, 4th day respectively; B, C,D). Note that at 3 μg L−1 prominent F-actin aggregations are organized in many cells (F). Scale bar: 5 µm.
to be sensitive “biomarkers” of BPA stress, even at environmentally relevant BPA concentrations. Importantly, the data presented highlight for the first time in the relevant literature the suitability of AF damages as a “stress biomarker”. This study, in accordance to previous work on land plants (among others see Adamakis et al., 2013, 2016; Stavropoulou et al., 2018), showed that BPA caused growth inhibition and targeted the cytoskeleton at environmentally relevant concentrations. While no information about BPA toxicity on marine macrophytes existed, this study extended the knowledge about BPA effects on a seagrass. The focus on a wildlife plant species, such as C. nodosa could be useful in understanding BPA effects in nature. Future studies may investigate the specific pathway(s) involved in the pre-mentioned effects as well as the organisms’ ability to recover after BPA removal.
Acknowledgements IDSA received a postdoc scholarship through the State Scholarships Foundation (IKY) with funds of the program: “Development of Human Resources, Education and Lifelong Learning” References Adamakis, I.-D.S., Panteris, E., Eleftheriou, E.P., 2010. The cortical microtubules are a universal target of tungsten toxicity among land plant taxa. J. Biol. Res. 13, 59–66. Adamakis, I.-D.S., Panteris, E., Cherianidou, A., Eleftheriou, E.P., 2013. Effects of bisphenol A on the microtubule arrays in root meristematic cells of Pisum sativum L. Mutat. Res.-Genet. Toxicol. Environ. 750, 111–120. Adamakis, I.D., Panteris, E., Eleftheriou, E.P., 2014. The nitrate reductase inhibitor, tungsten, disrupts actin microfilaments in Zea mays L. Protoplasma 251, 567–574. Adamakis, I.-D.S., Panteris, E., Eleftheriou, E.P., 2016. Bisphenol A disrupts microtubules and induces multipolar spindles in dividing root tip cells of the gymnosperm Abies cephalonica. Chemosphere 149, 202–210. Al-Rifai, J.H., Gabelish, C.L., Schäfer, A.I., 2007. Occurrence of pharmaceutically active and non-steroidal estrogenic compounds in three different wastewater recycling schemes in Australia. Chemosphere 69, 803–815. Alexander, H.C., Dill, D.C., Smith, L.W., Guiney, P.D., Dorn, P., 1988. Bisphenol a: acute aquatic toxicity. Environ. Toxicol. Chem. 7, 19–26. Baek, H.J., Hwang, I.J., Kim, K.S., Lee, Y.D., Kim, H.B., Yoo, M.S., 2007. Effects of BPA and DES on longchin goby (Chasmichthys dolichognathus) in vitro during the oocyte maturation. Mar. Environ. Res. 64, 79–86. Baluška, F., Jasik, J., Edelmann, H.G., Salajová, T., Volkmann, D., 2001. Latrunculin B induced plant dwarfism: plant cell elongation is f-actin-dependent. Dev. Biol. 231, 113–124. Baskin, T.I., 2001. On the alignment of cellulose microfibrils by cortical microtubules: a review and a model. Protoplasma 215, 150–171. Baskin, T.I., Bivens, N.J., 1995. Stimulation of radial expansion in arabidopsis roots by inhibitors of actomyosin and vesicle secretion but not by various inhibitors of metabolism. Planta 197, 514–521. Belfroid, A., van Velzen, M., van der Horst, B., Vethaak, D., 2002. Occurrence of bisphenol A in surface water and uptake in fish: evaluation of field measurements. Chemosphere 49, 97–103. Cabaço, S., Ferreira, Ó., Santos, R., 2010. Population dynamics of the seagrass Cymodocea
5. Conclusions Only the highest BPA concentrations ( > 0.3 μg L−1) significantly affected juvenile leaf elongation of C. nodosa from the onset of the experiment, whereas defects of the cytoskeleton were observed at lower concentrations. The LOEC values, determined from leaf elongation impairment (10th day) was 0.3 μg L−1, based on MT disturbance was 0.1 μg L−1 and based on AF damages was 0.03 μg L−1, so AFs seemed to be more sensitive to BPA exposure. Moreover, a correlation between leaf elongation impairment and MTs disorder was obvious. According to all the above findings, AF damages, MT disruption and juvenile leaf elongation impairment in C. nodosa, in this particular order, appear to be sensitive “biomarkers” of BPA stress, even at environmentally relevant BPA concentrations.
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The effects of Bisphenol A on the seagrass Cymodocea nodosa: Leaf elongation impairment and cytoskeleton disturbance Ioannis-Dimosthenis S. Adamakisa,1, Paraskevi Maleab, a b
⁎,1
T
, Emmanuel Panterisb
Department of Botany, Faculty of Biology, National and Kapodistrian University of Athens, 15784 Athens, Greece Department of Botany, School of Biology, Aristotle University of Thessaloniki, 54124 Thessaloniki, Macedonia, Greece
A R T I C LE I N FO
A B S T R A C T
Keywords: Bisphenol A Cymodocea nodosa Cytoskeleton disruption Leaf growth Seagrass Stress biomarkers Toxicity
Bisphenol A (BPA) is an emerging pollutant of environmental concern, classified as “moderately toxic” and “toxic”, causing adverse effects on aquatic biota. Although information about BPA toxicity on aquatic fauna is available, the data about BPA effects on aquatic flora remain scarce, missing for marine macrophytes. The effects of environmentally relevant BPA concentrations (ranging from 0.03 to 3 μg L−1) on juvenile leaf elongation and the cytoskeleton (microtubules, MTs and actin filaments, AFs) were studied in the seagrass Cymodocea nodosa for 1–10 days. The suitability of cytoskeleton disturbance and leaf elongation impairment as “biomarkers” for BPA stress were tested. The highest BPA concentrations (0.3, 0.5, 1 and 3 μg L−1) affected significantly leaf elongation from the onset of the experiment, while defects of the cytoskeleton were observed even at lower concentrations. In particular, MTs were initially disrupted (i.e. “lowest observed effect concentrations”, LOECs) at 0.1 μg L−1, while AFs were damaged even at 0.03 μg L−1. AFs appeared thus to be more sensitive to lower BPA concentrations, while there was a correlation between leaf elongation impairment and MT defects. Thus, AF damages, MT disruption and leaf elongation impairment in C. nodosa, in this particular order, appear to be sensitive “biomarkers” of BPA stress, at the above environmentally relevant BPA concentrations.
1. Introduction Worldwide human development and urbanization have led to increased demand for food and beverage packaging, remedial equipment, electronics, flame retardants, adhesives, construction materials, plastic containers and paper coverings (among others see Corrales et al., 2015). The manufacture of the above products requires the substance known as bisphenol A (CAS No. 80-05-7) (BPA; 2,2-bis(4-hydroxyphenyl) propane). As a result, worldwide BPA production (e.g. 3.7 million metric tons per year) and consumption have steadily increased over the years (Corrales et al., 2015). Similarly to other chemicals, marine environment is the major disposal site of BPA pollution. The major inputs of BPA enter the environment from production and processing industries, landfill wastes and effluents of wastewater treatment facilities. BPA discharges are percolated in rivers and streams towards estuarine and coastal systems (e.g. Yamada, 1999; Cousins et al., 2002; Kang et al., 2007; Mihaich et al., 2018). It has been recorded that BPA in municipal and industrial wastes can reach up to 23.02 μg L−1 (Al-Rifai et al., 2007), while in sewage pulps its concentration can be much higher 10-
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Corresponding author. E-mail address: [email protected] (P. Malea). These authors contributed equally.
https://doi.org/10.1016/j.ecoenv.2018.04.005 Received 18 February 2018; Received in revised form 2 April 2018; Accepted 4 April 2018 0147-6513/ © 2018 Elsevier Inc. All rights reserved.
> 100,000 μg kg−1 d.wt. (Corrales et al., 2015), and in some cases the reported values extended up to 3.2 × 107 μg kg−1 d.wt. (Harrison et al., 2006). The reported concentrations of BPA in aquatic environments vary, ranging from < 0.0088–1.0 μg L−1 in marine water (Belfroid et al., 2002; Kawahata et al., 2004; Vethaak et al., 2005; Staniszewska et al., 2014), 0.0015–0.145 μg L−1 in estuarine and lagoon water (Fu et al., 2007), > 1 μg L−1 in freshwater systems (Oehlmann et al., 2008; Flint et al., 2012), to elevated concentrations of BPA (100–1000 up to 20,136 and 13,392 μg kg−1 d.wt in Asia and Europe, respectively) in anaerobic sediments (see review by Corrales et al., 2015). Given its presence in the environment, earlier reports by the European Commission and the United States Environmental Protection Agency have classified this compound as “moderately toxic” and “toxic” to aquatic biota (Alexander et al., 1988). Few studies in either laboratory or field settings have examined wildlife organisms’ measurable responses to environmentally relevant BPA concentrations (0.08–12.5 µg L−1; Flint et al., 2012) and their predicted no-effect concentrations have been determined at 0.06 µg L−1 (Wright-Walters et al., 2011). Most of the evidence derived mainly from studies on aquatic invertebrates, such as crustaceans and echinoderms,
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with available data also from insects, fish, amphibians, reptiles and birds (Mihaich et al., 2009; Flint et al., 2012). Plants, especially aquatic ones, have received far less attention (e.g. Mihaich et al., 2009). In particular, the research on BPA toxicity so far has been limited to the freshwater angiosperm Lemna gibba (Mihaich et al., 2009) and to some microalgae species (e.g. Chlorella fusca, Stephanodiscus hantzschii, Chlamydomonas mexicana, Chlorella vulgaris) (Alexander et al., 1988; Hirooka et al., 2005; Li et al., 2009; Ji et al., 2014), in which BPA induced defects in growth, antioxidant enzyme activity and photosynthetic pigments, respectively. In land plants, studies have examined BPA effects on vegetative growth (Ferrara et al., 2006) and some researchers have related the defects of microtubules (MTs) and actin filaments (AFs) induced by BPA treatment with the observed growth inhibition (Adamakis et al., 2013, 2016; Stavropoulou et al., 2018). These potential disturbances, induced by BPA to the cytoskeleton and/or the vegetative growth, could be used as “biomarkers” for the evaluation of BPA contamination in specific marine “bioindicator” organisms. Marine macrophytes, especially the seagrasses, are of particular interest, given their well-known biological and ecological relevance and their sensitivity to anthropogenic disturbances (e.g. Malea, 1994; Malea and Haritonidis, 1995, 1999; Malea et al., 1995; Malea and Zikidou, 2011; Malea et al., 2013a, 2013b, 2013c, 2014). These characteristics mark them as excellent candidate “bioindicator” organisms (see review by Ferrat et al., 2003). Although some “biomarkers” have been determined for BPA toxicity, mainly in aquatic vertebrates and invertebrates, in some microalgae species and in one aquatic angiosperm (Lemna gibba) (Kang et al., 2007; Mihaich et al., 2009 and literature therein), none has been determined in marine macrophyte species and especially in seagrasses. Cymodocea nodosa is a seagrass resilient to anthropogenic disturbances, displaying a fast growth rate (Cabaço et al., 2010). In seagrasses species, among which C. nodosa, the effectiveness of a number of “biomarkers” (e.g. photosynthetic activity, enzymatic activity, heat shock proteins, phenolic compounds, MT arrays, cell mortality, oxidative stress “biomarkers”, biomarkers of detoxification) of various stressors (especially metals, metallic nanoparticles (NPs), pesticides, nutrients, etc.) has been tested and most of them appeared to be valuable and early warning signals for general or particular stress (see Ferrat et al., 2003 for review; Malea et al., 2013a, 2013b, 2014; Papathanasiou et al., 2015; Moustakas et al., 2016). Growth impairment has been examined as an assessor of the aquatic angiosperm health against various chemicals (e.g. Flores et al., 2013; Negri et al., 2015; Papathanasiou et al., 2015; Llagostera et al., 2016). However, the extent by which the effects on the cytoskeleton and ultrastructural and mitotic defects (Adamakis et al., 2013, 2016) may result in more obvious plant physiological and morphological responses (e.g. leaf elongation) remain unexplored. Considering the above, in the present study we sought to investigate the effects of environmentally relevant BPA concentrations on the juvenile leaf growth (expressed as elongation), in relation to the defects in cytoskeleton organization in leaf cells of the seagrass C. nodosa. Moreover, this study contributes to the existing knowledge on the “lowest observed effect concentrations” (LOECs) and hence on “no observed effect concentrations” (NOECs) on this marine angiosperm, which can be used in risk assessment programmes. The suitability of MT and AF disturbance and of leaf elongation impairment as suitable “biomarkers” of BPA-induced stress is discussed.
monospecific meadow. Leaf biomass and leaf blade length display an almost unimodal annual pattern attaining maximum values in JulyAugust (e.g. Malea and Zikidou, 2011). Plant collection was done at 0.7–1.0 m depth in July 2017 with a 20 cm diameter acrylic corer, which penetrated to a depth of 30 cm. All plants were rinsed in seawater at the collection site and transported to the laboratory in plastic containers containing seawater. 2.2. Treatments and experimental conditions Collected plants without epiphytes consisting of roots, plagiotropic (horizontal) and orthotropic (vertical) rhizomes and leaves were kept for 24 h in filtered seawater (0.45 µm Whatman GF/C) under laboratory conditions to equilibrate, under a constant 16 h day/8 h night regime at an ambient temperature of 21 ± 1 °C with light intensity set at 120 μmοl m−2 s−1. The seawater used for all experiments was collected from a nearby coastal area (Epanomi site, Thermaikos Gulf). All the chemicals and reagents used were purchased from Applichem (Darmstadt, Germany), Merck (Darmstadt, Germany) and Sigma (Taufkirchen, Germany), unless stated otherwise. Two different sets of experiments were performed, both run in triplicates. In the first set about 24 orthotropic rhizomes attached to part of plagiotropic rhizomes with roots were incubated in non-BPA-based polypropylene (PP) copolymer (PPCO) aquaria covered with plastic foil to prevent evaporation. Plants were treated with 0.03, 0.1, 0.3, 0.5, 1 and 3 μg L−1 bisphenol (BPA) solutions in filtered seawater. For control, plants incubated in filtered seawater, were used. The solutions in the aquaria were changed every two days and were constantly aerated using aquarium air pumps. From each aquarium, four seagrass shoots (totally 12 shoots per BPA treatment) were randomly removed after 1, 2, 4, 6, 8 and 10 days and the meristematic region of the youngest juvenile leaf blade from each shoot was used for either tubulin immunostaining or AF staining. In the second set of experiments, a similar experimental design, as above was generally applied. In each aquarium, in about four orthotropic rhizomes of C. nodosa attached to part of plagiotropic rhizomes with roots (about 12 shoots in each BPA treatment) the youngest accessible juvenile blades were marked by an inert sealant at their base and their initial lengths were measured (mean ± SE: 33.908 ± 2.034 mm). Leaf elongation (mm) was carefully measured as the newly formed leaf segments, hence the distance between the marker and the juvenile leaf base on the 2nd, 4th, 6th, 8th and 10th day (Mateos Naranjo et al., 2008). Leaf elongation rate (Negri et al., 2015) (expressed as mm day−1) were also calculated as the leaf elongation (mm) among two consecutive days divided by the number of days that they mediate. The final length of each measured juvenile leaf, at the end of the experiment were also measured. 2.3. Imaging of MTs and AFs Whole mount MT immunostaining was conducted in hand cut small juvenile leaf pieces, following the protocol of Katsaros and Galatis (1992) using anti-α-tubulin (YOL1/34, AbD Serotec, Kidlington, UK) and FITC-anti-rat secondary antibody (Invitrogen, Carlsbad, CA), both diluted at 1:80 in phosphate buffered saline (PBS). DNA was counterstained with 10 μgmL-1 DAPI (4′,6-diamidino-2-phenylindole) in PBS and the leaf pieces were finally mounted in an anti-fade solution. AF staining with TRITC-phalloidin was conducted according to Panteris et al. (2009), with some modifications as follows: Leaf pieces were firstly incubated with 300 μM m-maleimidobenzoyl- N-hydroxysuccinimide ester (MBS) in PBS + 0.5% (v/v) Triton X-100 for 30 min in the dark for F-actin stabilization. After AF stabilization, fixation was performed in 4% (w/v) paraformaldehyde (PFA) in the same buffer for 60 min. In the PFA solution 1% (v/v; diluted from a 10 μΜ stock solution in methanol) TRITC-phalloidin was added. The tissue sheets were then rinsed with PBS and extracted in 5% (v/v) DMSO
2. Materials and methods 2.1. Plant collection Cymodocea nodosa (Ucria) Ascherson 1870 was collected from the eastern coast of the Gulf of Thessaloniki, Northern Aegean Sea at Viamyl site (40°33 ´N, 22°58´E), where it forms a continuous 432
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+ 5% (v/v) Triton X-100 in the same buffer for 1 h. Finally, AFs were stained with 10% (v/v; same as before) TRITC-phalloidin in PBS in the dark, for 1 h at 37 °C or overnight at room temperature. As above, following the DNA counterstaining step leaf pieces were mounted with an anti-fade solution. All the fluorescent specimens were examined with a Zeiss Observer. Z1 microscope, equipped with the LSM780 confocal laser scanning system, and images were acquired with ZEN2011 software following the manufacturer's instructions. The laser gain was kept stable in all experiments and all digital images were improved for contrast and color with Adobe Photoshop CS2 software with only linear settings. 2.4. Fluorescence intensity measurements Fluorescence intensity of total tubulin was measured at single cortical or central CLSM sections of variously treated leaf tissues, using the Image J ((http://rsbweb.nih.gov/ij/) software, as already described by (Adamakis et al., 2014). From each treatment, fluorescence intensity from 12 individual cells was measured and the mean values were calculated and expressed as fluorescence intensity percentage of control. 2.5. Data analysis The data on leaf blade elongation in time were fitted to different regression models (linear, logarithmic, inverse, exponential, power), using IBM Statistics SPSS® 24; in each case, the model that provided the best fit (i.e., at the highest significance level) was chosen. Statistical analysis was also carried out using Statistica v.12 (Statsoft Inc.). The significance of the variation in the juvenile leaf elongation (mm), the leaf elongation rate (mm day−1) and the microtubule integrity (%) of juvenile blade cells over the incubation time was tested with the non parametric test, Kruskal-Wallis one-way analysis of variance. MannWhitney U-test was applied to compare the mean values of the studied parameters (juvenile leaf elongation, leaf elongation rate, tolerance index, microtubule integrity) between control and various BPA concentrations or various BPA concentrations by twos. Spearman's Rank correlation coefficient was calculated to identify correlations. Tolerance indices of the juvenile leaves were calculated by the following formula (e.g. Idris et al., 2012):
Tolerance index (%) = (blade length of stressed plant / blade length of control plant )* 100 “No observed effect concentration” (NOEC) is the highest concentration of a toxicant that causes no observable adverse effects on the test organisms and calculated as the highest concentration in the test with a mean response not differing significantly from the mean response of the control. “Low observed effect concentration” (LOEC) is the lowest test concentration having a mean response that differs significantly from that of the control (e.g. Crane and Newman, 2000). For both leaf elongation tolerance and microtubule integrity the 10% and 50% effect concentrations (EC10 and EC50 with 95% confidence limits) at the end of the experiments (day 10) were estimated by probit analysis with XLSTAT 2018 (New York, NY, USA).
Fig. 1. (A) Time course of juvenile leaf elongation (mm), (B) leaf elongation rate (mm day−1) (mean ± SE) in control and in various BPA concentrations and (C) of the Tolerance index of juvenile leaf (mean ± SE).
logarithmic equation provided the best fit (i.e. at the highest significance level) r2: 0.994 for control; 0.987 for 0.03 μg L−1, p < 0.001; r2: 0.987 for 0.3 μg L−1; 0.992 for 0.5 μg L−1; 0.969 for 3 μg L−1, p < 0.01; however, at the 0.1 and 1 μg L−1 treatments, a linear equation best described leaf elongation (r2: 0.988 and 0.965, respectively, p < 0.01. Mann Whitney U test showed that leaf elongation, at each incubation time, did not significantly differ (p > 0.05) between the lowest BPA concentrations (0.03 and 0.1 μg L−1) (with the exception of 0.03 μg L−1 on the 4th day) and those measured in the control, which corresponds to the non-observed inhibition of the leaf elongation at the lowest BPA levels (Table 1, Fig. 1A). In contrast, at the highest BPA concentrations (0.3, 0.5, 1 and 3 μg L−1) the inhibition of leaf elongation started from the beginning and continued up to the end of the experiment (p < 0.05) (Table 1, Fig. 1A).
3. Results 3.1. BPA effects on juvenile leaf elongation Generally, a gradual inhibition of juvenile blade elongation with increasing external BPA concentrations was observed from the beginning of the experiment (Fig. 1A). The Kruskal-Wallis analysis of variance showed that the above time variation was only significant on control (Chi square=16.0, df=4, p < 0.01) and on the lowest BPA concentrations (0.03, 0.1 and 0.3 μg L−1) (Chi square: 10.577, 9.643 and 12.180, respectively, df = 4, p < 0.05). In most of the cases, a
3.2. Leaf elongation rates Time variation in leaf elongation rate (mm day−1) (Fig. 1B) was 433
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of MT immunofluorescence are presented in Fig. 3. In untreated interphase cells cortical MTs were randomly oriented, organized in a dense network (Fig. 3A), as already reported by Malea et al. (2013b). The above network remained unaltered even at the end of the experiment (Fig. 3B). In BPA-affected juvenile leaves of C. nodosa MTs appeared fragmented in all the leaf blade cells studied and at all the concentrations studied, except for 0.03 and 0.1 μg L−1. At 0.03 μg L−1 no effect was observed even at the last day (Figs. 2 and 3C), while at 0.1 μg L−1 interphase cortical MTs appeared bundled at the 8th day of treatment and thereafter (Figs. 2 and 3D). Higher BPA concentrations (1 and 3 μg L−1) affected MTs more quickly and rigorously (already from the 2nd day and thereafter) than the lower ones (0.3, 0.5 μg L−1), which required relatively longer exposure times (6th and 4th day, respectively) to manifest their deleterious effect (Figs. 2 and 3E-H). Fluorescence intensity measurements (Fig. 4A) further consolidated the above results, clearly indicating the depolymerizing effect of BPA. Fluorescence intensity dropped in a time- and concentration-dependent way; the Kruskal-Wallis analysis showed that the time variations in MT integrity (%) up to ten days were only significant at the highest BPA treatments (at 0.3, 0.5, 1 μg L−1, chi square = 19.412, 15.447, 15.772, respectively, p < 0.01 and at 3 μg L−1, chi square=21.473, p < 0.001). Moreover, Mann Whitney U test showed that at low BPA concentrations (0.1, 0.3 and 0.5 μg L−1) significant differences in fluorescence intensity measurements, in comparison to the control, were observed after the sixth day of the experiment, whereas at the highest concentrations (1 and 3 μg L−1) even from the first day of the experiment (Fig. 4B).
Table 1 Comparison (Mann-Whitney U-test) of juvenile leaf elongation (mm) between control and various BPA concentrations each day of the experiment. Exposure concentrations (μg L−1) Control-0.03 Control-0.1 Control-0.3 Control-0.5 Control-1 Control-3
Z Z Z Z Z Z
adjusted adjusted adjusted adjusted adjusted adjusted
2nd day
4th day
6th day
8th day
10th day
1.225 ns 0.707 ns 2.449* 2.449* 2.121* 2.309*
2.205* 1.768 ns 2.205* 2.449* 2.121* 2.309*
1.715 ns 1.414 ns 1.960* 2.449* 2.121* 2.309*
1.715 ns 1.061 ns 1.960* 2.449* 2.327* 2.309*
1.225 ns 1.061 ns 1.960* 2.449* 2.121* 2.309*
* : p < 0.05, ns: non significant. Table 2 Leaf elongation rate (mm day−1) and Tolerance index in each BPA concentration over the incubation period.
Leaf elongation rate
Tolerance index
Exposure concentrations (μg L−1)
Mean ( ± SE)
Range
0.03 0.1 0.3 0.5 1 3 0.03 0.1 0.3 0.5 1 3
75.585 ( ± 1.723) 76.170 ( ± 2.016) 58.906 ( ± 1.372) 50.715 ( ± 0.328) 43.680 ( ± 1.860) 30.277 ( ± 1.142) 4.647 ( ± 0.904) 4.701 ( ± 0.986) 3.663 ( ± 0.778) 3.126 ( ± 0.627) 2.715 ( ± 0.598) 1.896 ( ± 0.421)
70.843–79.627 71.814–82.917 56.231–63.119 49.575–51.284 37.375–46.703 26.891–32.859 1.540–6.506 1.653–7.220 1.222–5.496 1.068–4.454 0.960–4.067 0.630–2.855
3.5. Effects on AFs In general, interphase meristematic cells of untreated C. nodosa leaves exhibited abundant cortical and endoplasmic AFs (Fig. 5A). Cortical AFs were randomly oriented, forming a rather complicated mesh, while dividing cells exhibited intense F-actin signal at both sides of the division plane (Fig. 5A). In leaf cells exposed to BPA, AFs were variously affected, depending on treatment duration and BPA concentration. At 1 and 3 μg L−1 for 2 days, AFs tended to aggregate in thick bundles (Fig. 5E, F). Especially at 3 μg L−1 BPA treatment thick Factin aggregates were formed (Fig. 5F). Lower concentrations (0.03, 0.1 and 0.5 μg L−1) also caused damage on AFs (diminishing or even almost complete depolymerization), but required longer exposure times (6th, 6th and 4th day, respectively) to exert their toxic effect (Fig. 5B-D).
only significant at the control treatment (Kruskal Wallis analysis, Chisquare = 10.000 df = 4, p < 0.05). However, at the control and at the lowest BPA treatments (0.03–0.5 μg L−1) the elongation rate increased from 2nd to 4th day and subsequently gradually decreased up to the end of the experiment (Fig. 1B). At the highest concentrations (0.3–3 μg L−1), significantly (p < 0.01 in 0.3 μg L−1 and p < 0.05 in the rest treatments) decreases were observed between the 8th and 10th day (Fig. 1B). A gradual decrease in the mean ( ± SE) leaf elongation rate (mm day−1) of C. nodosa with increasing BPA concentrations was observed (Table 2). Mann Whitney analysis showed that, at the highest concentrations (0.5 up to 3 μg L−1) the mean elongation rate significantly (p < 0.05) differed from those in control treatment throughout the incubation time.
3.6. Relationship between inhibition of leaf elongation and cytoskeleton disruption At the lowest BPA concentrations (0.03 and 0.1 μg L−1), only a slight (p > 0.05) gradual inhibition of leaf elongation was obvious, even from the first day of the experiment (Fig. 1A, Table 1); in fact, elongation inhibition was significant (p < 0.05) at concentrations equal/higher than 0.3 μg L−1, even from the beginning of the experiment and continued up to the end of the experiment (Fig. 1A, Table 1). Cortical MT disruption at 0.1 μg L−1 BPA was obvious at extended treatment, as confirmed from both fluorescence intensity measurements (6th day, Fig. 4B) and immunofluorescence images (8th day, Figs. 2 and 3D), showing that MT disturbance occurred at lower concentrations than those required to impair leaf growth. A significant positive correlation (Spearman's rank order correlation, r = 0.714, p < 0.001) among the juvenile leaf elongation rate and MT integrity (%; based on fluorescence intensity) was recorded. Similarly, a positive correlation (r = 0.769 p < 0.001) among the leaf tolerance index and MT integrity (%) over the incubation time was observed. Moreover, AF damages were recorded earlier than MT disturbance and leaf growth
3.3. Tolerance index A high significant (p < 0.001) gradual decrease of the mean ( ± SE) “Tolerance Index” with increasing BPA concentrations (chisquare=30.000 df=5, p < 0.001) was observed (Fig. 1C, Table 2). In particular, the “Tolerance Index” differed significantly (p < 0.01) between the two highest concentrations. In contrast, the variation of this index was not time-dependent (p > 0.05), as shows Kruskal-Wallis analysis. 3.4. Effects on MT integrity The concentration- and time-depended MT disturbance are given in a schematic representation (Fig. 2), while representative micrographs
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Fig. 2. Schematic representation of microtubule disruption in regard with time and BPA concentrations. The explanations of the schematics appear at the bottom of the figure.
treatments or after shorter treatments (up 6 days) with concentrations much higher (6–100 mg L−1) than those applied in our experiments (Ferrara et al., 2006; Adamakis et al., 2013, 2016; Frejd et al., 2016; Jayasri et al., 2016; Li et al., 2018; Stavropoulou et al., 2018). In C. nodosa juvenile blade leaves, a gradual inhibition of elongation with increasing external BPA concentrations was observed (Fig. 1A). This inhibition seemed to be significant only at higher BPA concentrations, as juvenile leaf elongation at the lowest BPA concentrations (0.03 and 0.1 μg L−1) was similar to that of the control. On the 4th day at 0.03 μg L−1 a significant elongation inhibition was recorded (Τable 1). Nevertheless, this tendency did not continue for the following days (Table 1). As a result, the overall leaf elongation was not inhibited at the lowest BPA concentrations (Kruskal-Wallis analysis of variance), indicating that leaf growth may not be significantly inhibited at such mild concentrations. This finding is in accordance with previous studies, in which BPA, at such environmentally relevant concentrations, did not affect other aquatic organisms (chironomids, copepods, gobies), according to other measurable “stress biomarkers” (e.g. larval emergence, estrogen synthesis), (e.g. Marcial et al., 2003; Watts et al., 2003; Baek et al., 2007). On the contrary, at relatively higher BPA concentrations (0.3, 0.5, 1 and 3 μg L−1) the inhibition of C. nodosa leaf elongation started from the onset of the experiment and continued up to the end of the experiment (Fig. 1A, Table 1). Similarly, at these BPA levels leaf elongation rates (mm day−1) and their “Tolerance Indices” gradually decreased by increasing BPA exposure (Fig. 1B, C). In addition, the elongation rate and the “Tolerance Index” seemed not to be time-dependent under acute BPA toxicity, which is in accordance with the findings of Mihaich et al. (2009) in Lemna gibba. The inhibition of leaf elongation, as mentioned in previous studies (e.g. under metal or herbicide effects), may be due to adverse effects on the metabolism and physiological reactions (e.g. Chesworth et al., 2004; Idris et al., 2012). On the other hand, disruption of the cytoskeleton may also be a cause for decreased leaf elongation. In particular,
impairment, while their disruption occurred even at lower BPA concentrations (0.03 μg L−1) on the 6th day (Fig. 5B). According to leaf elongation impairment, MT disruption (fluorescence intensity, immunofluorescence images) and AFs damages, at various BPA concentrations, at the end of the experiment (10th day) (Table 1, Figs. 2 and 4 B, 5), it is obvious that the “no observed effect concentration” value (NOEC) was < 0.03 μg L−1, according to AF disruption, 0.03 μg L−1, according to MT disturbance, and 0.1 μg L−1, according to leaf elongation impairment. The respective “low observed effect concentrations” (LOECs) values were 0.03 μg L−1, 0.1 μg L−1 and 0.3 μg L−1 Based on MT fluorescence intensity (%), EC50 (95% confidence limits) on the 10th day was 0.225 (0.211–0.241) μg L−1 and EC10 was 0.570 (0.531–0.616) μg L−1, while based on leaf elongation impairment the respective EC50 (95% confidence limits) was 0.140 (0.129–0.151) μg L−1 and the EC10 was 0.500 (0.464–0.544) μg L−1.
4. Discussion 4.1. BPA effects on leaf elongation and the cytoskeleton MTs and AFs have been shown to be some of the prime sub-cellular targets of BPA toxicity in land plants (Adamakis et al., 2013, 2016; Stavropoulou et al., 2018). Similarly, in the present study BPA caused a concentration-dependent disruption of MTs and AFs (Figs. 2–5), further supporting the conclusion that plant MTs and AFs are a target of BPA toxicity, also in a marine angiosperm. Notably, however, the BPA concentrations applied on C. nodosa were significantly lower than those used in other studies on land plants (corresponding to those used in experiments in which BPA affected MT organization, e.g. George et al., 2008). In land plants, BPA arrested root growth in seedlings of various plant species in a concentration- and time-dependent manner. Root growth inhibition was drastic after prolonged (up to 21 days)
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expanding cell wall (Mathur et al., 1999; Mathur and Hülskamp, 2002). Experimental disruption of AFs and/or arrest of actomyosin-mediated mobility have been shown to result in inhibition of elongation in root cells (Baskin and Bivens, 1995; Baluška et al., 2001; Panteris et al., 2013). Accordingly, inhibition of C. nodosa leaf elongation could be also attributed to AF disruption by BPA. In such a view, the effects on both AFs and MTs could result in a general disruption of leaf elongation due to BPA. 4.2. Establishing reliable “biomarkers” Most of the BPA concentrations selected for the treatments performed in this study were within the range reported in seawater (< 0.0088–1.081 μg L−1) (e.g. Mihaich et al., 2018), in estuarine and lagoon water (0.0015–0.145 μg L−1) (Belfroid et al., 2002; Kawahata et al., 2004; Fu et al., 2007), in freshwater (> 1 μg L−1) (e.g. Flint et al., 2012) and in ground-water (1.9 μg L−1) (e.g. Loos et al., 2010), whereas the highest BPA concentration (3 μg L−1) was selected to be in the range of those recorded in wastewater treatment facilities (WWTP) (0.087 – 5.625 μg L−1) (e.g. Pothitou and Voutsa, 2008). Moreover, BPA is moderately soluble, with a measured solubility 120 mg L−1 at 20 °C in sea water (among other see Staples et al., 1997; Ying and Kookana, 2003). So, all of the BPA concentrations applied here were at environmentally relevant levels, thus our results have environmental significance, contributing to the existing knowledge on BPA toxicity by determining the NOEC, LOEC, EC50 and EC10 values. BPA effects on aquatic animals (invertebrates, fish, amphibians, reptiles, birds, mammals etc), even at environmentally relevant BPA concentrations (0.08–12.5 μg L−1), have been previously reported (Flint et al., 2012; Canesi and Fabbri, 2015). Among the reported effects on animals were: delayed emergence of chironomid larvae (at 0.08 μg L−1), developmental inhibition of marine copepods (at 0.1 μg L−1), mortality of freshwater snails (at > 1 μg L−1), altered blood sex hormone ratios (at > 1 μg L−1), spontaneous metamorphosis inhibition of the clawed frog (at 2.3 μg L−1), premature metamorphosis of marine polychaetes (11.4 μg L−1), estrogen synthesis inhibition in longchin goby (0.1 μg L−1), accelerated larval development in marine copepods (12.5 μg L−1) (Flint et al., 2012 and literature therein). An aquatic hazard assessment determined the predicted no-effect BPA concentration to 0.06 μg L−1, using as “biomarker” the development, reproduction and survival of wildlife (Wright-Walters et al., 2011). The NOECs, in various invertebrate species, ranged from 0.025 to 4.1 mg L−1 (Mihaich et al., 2009, 2018) and in vertebrate species from 0.016 to 3.6 mg L−1 (Staples et al., 2011). Accordingly, the available literature provides evidence that relatively low BPA concentrations (encompassing the environmental range; ≤12 µg L−1) affected the physiological homeostasis and gene expression of aquatic animals (Canesi and Fabbri, 2015). However, in aquatic microalgae and angiosperms, the effective BPA doses were significantly higher from the reported environmentally relevant concentrations, as it will be further on analyzed. In the microalgae Skeletonema costatum and Selenastrum capricornutum, the NOEC values were 0.4 and 1.2 mg L−1, respectively (Alexander et al., 1988), whereas in the aquatic angiosperm Lemna gibba the NOEC value, based on growth rate, was 7.8 mg L−1 (Mihaich et al., 2009). As far as it concerns EC50, the acute invertebrate BPA toxicity presented values ranged from 1.1 to 16 mg L−1 (Mihaich et al., 2009 and literature within). In fresh- or seawater microalgae (Chlorella, Cyclotela, Scenedesmus, Selenastrum, Skeletonema, Stephanodiscus spp) the respective EC50 values ranged from 1 mg L−1 (96 h) to 89 mg L−1 (72 h) (Alexander et al., 1988; Li et al., 2009; Zhang et al., 2012). Concerning the aquatic angiosperm L. gibba the ECs50, based on growth rate and on frond density were 32 and 29 mg L−1, respectively (Mihaich
Fig. 3. Tubulin immunofluorescence images of interphase protodermal cells at cortical CLSM sections. The treatment and relevant duration are stated on each figure. Untreated cells bare a dense network of randomly oriented MTs at both the onset (0 days; A) and the end (10 days; B) of the experiment. Treatment with 0.03 μg L−1 BPA did not have any effect even at the 10th day of exposure (C). Upon treatment with BPA 0.1 μg L−1, MTs appear bundled at longer exposure (8th day; D), while in the rest exposure concentrations (0.3–3 μg L−1; EH) MTs were depolymerized/fragmented, with higher concentrations (1, 3 μg L−1) exerting their effect earlier (4th, 2nd day, respectively; G, H). Scale bar: 5 µm.
cortical MTs control cellulose microfibril orientation in the cell wall, establishing thus elongation axes (Baskin, 2001; Nick, 2008), and are reorganized in response to endogenous signals or various environmental stimuli (e.g. Hashimoto, 2015). Interestingly, in the present study the first obvious signs of MT disturbance were observed before any leaf growth impairment could be recorded. Accordingly, inhibition of elongation could be an outcome of disordered cellulose deposition in the cell wall, due to disrupted cortical microtubules (Fujita et al., 2011). Furthermore, AFs of C. nodosa leaf cells were also affected by BPA, even before MT disruption. In plants, AFs provide the principal tracks on which organelles and vesicles move (Duan and Tominaga, 2018). There is substantial genetic and pharmacological evidence that the dynamic plant actin cytoskeleton has an important role in cell morphogenesis, regulating the delivery of Golgi-derived materials to the
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Fig. 4. (A) Time course of microtubule integrity expressed as percentage (%) of cell fluorescence intensity (mean of 12 individual cells ± SE) in the control and at various BPA concentrations and (B) comparisons (Mann Whitney U-test) of microtubule integrity (%) (based of fluorescence intensity measurements) of juvenile blade cells between control and various BPA concentrations each day of the experiment.
a “biomarker” for a given stressor. Changes in MT pattern were suggested as a suitable specific “biomarker” of cadmium or tungsten stress in land plants (Fusconi et al., 2007; Adamakis et al., 2010). Similarly, in C. nodosa MT defects could be used as an early warning signal of emerging lead, copper, nickel, chromium and cadmium stress (Malea et al., 2013a, 2013b, 2014). Taking into account the determined NOEC values (terminal time: 10th day) of C. nodosa (0.03 μg L−1, based on MT disturbance and < 0.03 μg L−1, based on AF damages) and the EC50 (95% confidence limits), based on MT disruption (fluorescence intensity) (0.225 (0.211–0.241) μg L−1), it is obvious that these values are much lower than those of other “stress biomarkers” observed in other aquatic organisms. According to all the above findings, AF damages, MT disruption and leaf elongation impairment in C. nodosa, in this particular order, appear
et al., 2009). In the present study, the NOEC and LOEC values (0.1 μg L−1 and 0.3 μg L−1, respectively) determined on the basis of C. nodosa leaf elongation, at the 10th day were much lower than those earlier reported for the freshwater angiosperm L. gibba, based on the 7th day frond growth rate, as well as for other aquatic organisms on the basis of various “biomarkers”. In addition, EC50 (95% confidence limits) based on leaf elongation impairment: 0.1340 (0.129–0.151) μg L−1 was much lower than that recorded in other aquatic organisms, as previously mentioned. Thus, leaf elongation of C. nodosa seems to be an appropriate “biomarker” under BPA effect. Furthermore, since reorganization of the cytoskeleton under abiotic stress is crucial for cell survival responses (among others see Soda et al., 2016), it could serve as an early measurable warning sign, constituting 437
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Fig. 5. TRITC-phalloidin staining of F-actin. Maximum intensity projections of serial CLSM sections of untreated (A) and BPA-treated (B-F) protodermal leaf cells. In the control AFs form a mesh of randomly aligned AFs (A). Dividing cells (asterisks in A) exhibit intense F-actin signal at either sides of the division plane (arrows in A). In BPAtreated leaves, AFs appear depolymerized, except from the thickest AF bundles. This effect occurs early (2nd day; E, F) at higher concentrations (1–3 μg L−1), while at lower concentrations (0.03, 0.1, 0.5 μg L−1) it occurs at longer exposure duration (6th, 4th day respectively; B, C,D). Note that at 3 μg L−1 prominent F-actin aggregations are organized in many cells (F). Scale bar: 5 µm.
to be sensitive “biomarkers” of BPA stress, even at environmentally relevant BPA concentrations. Importantly, the data presented highlight for the first time in the relevant literature the suitability of AF damages as a “stress biomarker”. This study, in accordance to previous work on land plants (among others see Adamakis et al., 2013, 2016; Stavropoulou et al., 2018), showed that BPA caused growth inhibition and targeted the cytoskeleton at environmentally relevant concentrations. While no information about BPA toxicity on marine macrophytes existed, this study extended the knowledge about BPA effects on a seagrass. The focus on a wildlife plant species, such as C. nodosa could be useful in understanding BPA effects in nature. Future studies may investigate the specific pathway(s) involved in the pre-mentioned effects as well as the organisms’ ability to recover after BPA removal.
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5. Conclusions Only the highest BPA concentrations ( > 0.3 μg L−1) significantly affected juvenile leaf elongation of C. nodosa from the onset of the experiment, whereas defects of the cytoskeleton were observed at lower concentrations. The LOEC values, determined from leaf elongation impairment (10th day) was 0.3 μg L−1, based on MT disturbance was 0.1 μg L−1 and based on AF damages was 0.03 μg L−1, so AFs seemed to be more sensitive to BPA exposure. Moreover, a correlation between leaf elongation impairment and MTs disorder was obvious. According to all the above findings, AF damages, MT disruption and juvenile leaf elongation impairment in C. nodosa, in this particular order, appear to be sensitive “biomarkers” of BPA stress, even at environmentally relevant BPA concentrations.
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