- Email: [email protected]
Does nanosized plastic affect aquatic fungal litter decomposition?
Fungal Ecology xxx (xxxx) xxx
Contents lists available at ScienceDirect
Fungal Ecology journal homepage: www.elsevier.com/locate/funeco
Does nanosized plastic affect aquatic fungal litter decomposition? Sahadevan Seena a, *, Diana Graça a, Anne Bartels a, Julien Cornut a, b a b
MARE-Marine and Environmental Sciences Centre, University of Coimbra, PT-3004-517, Coimbra, Portugal Laboratory of Applied Microbiology, University of Applied Sciences and Arts of Southern Switzerland, via Mirasole 22A, 6501, Bellinzona, Switzerland
a r t i c l e i n f o
a b s t r a c t
Article history: Received 26 October 2018 Received in revised form 18 February 2019 Accepted 22 February 2019 Available online xxx
Nanosized plastics are an emerging concern in freshwater ecosystems, raising the question whether they put freshwater ecological processes at risk. Litter decomposition is a major ecological function in forested streams which is mainly driven by aquatic hyphomycetes. Here we investigated whether increasing concentrations (up to 102.4 mg/L) of nanosized polystyrene plastics (NPPs; 100nm) affect litter decomposition by five widely distributed species of aquatic hyphomycetes. Results showed that average litter decomposition decreased by 8% relative to the control when exposed to 102.4 mg/L NPPs. Aquatic hyphomycete species differed in their sensitivity to NPPs. The greatest inhibition of litter decomposition was found with Tetracladium marchalianum, where it dropped from 37 (control) to 16% (102.4 mg/L of NPP). Overall our study highlights the emerging risks and potential dangers of NPPs to freshwater ecosystem functioning. It also indicates that the impact of NPPs may be species specific. © 2019 Elsevier Ltd and British Mycological Society. All rights reserved.
Corresponding Editor: Prof. L. Boddy Keywords: Freshwater ecosystems Streams Aquatic hyphomycetes Litter decomposition Nanoparticles Polystyrene
1. Introduction Plastic pollution is now acknowledged as a significant threat to aquatic environments (for review see Koelmans et al., 2015). Virgin plastic production has been estimated to be 8300 million metric tons (Mt) and by 2050 around 12,000 Mt of plastic waste is envisioned to be in landfills or in the natural environment (Geyer et al., 2017). It is predicted that every year between 1.15 and 2.41 million metric tonnes of plastic enter the ocean via global riverine systems (Lebreton et al., 2017). Although most studies dealing with the consequences of plastics in ecosystems have been performed in marine systems, rivers are the main source of marine plastic pollution (Lebreton et al., 2017). Plastics can be physically fragmented into nanosized particles whose environmental effects are virtually unknown and of increasing concern. Nanosized plastics are also used in a vast range of products such as toothpaste, waterborne paints and biomedical products (Koelmans et al., 2015). Unlike macroplastics, micro and nanoscale plastics are difficult to detect and quantify (Blair et al., 2017; Alimi et al., 2018). A recent survey conducted after winter
* Corresponding author. E-mail addresses: [email protected], (S. Seena).
[email protected]
floods in the River Tame, UK, reported over 500,000 microplastic particles per square meter (Hurley et al., 2018). At the nanoscale, plastics are even harder to detect; hence their exact environmental concentration and distribution are unknown. This poses severe challenges when studying the impact of nanosized plastics on organisms or ecosystems for using realistic environmental concentrations (Wagner et al., 2014; Koelmans et al., 2015; Blair et al., 2017). Despite this limitation, recent studies have demonstrated nanosized polystyrene plastics (NPPs) to be hazardous, for instance, by interfering with algal photosynthesis (Bhattacharya et al., 2010), reducing filter-feeding activity of mussels (Mytilus adulis) (Wegner et al., 2012) and causing mortality of copepods upon ingestion (Tigriopus japonicus) (Lee et al., 2013). In woodland streams, leaf litter is the primary source of carbon and energy (Graça et al., 2015; Chauvet et al., 2016; Seena et al., 2017). Leaf litter decomposition is a crucial ecosystem process in forested streams; aquatic fungi, especially aquatic hyphomycetes are key players responsible for leaf breakdown (B€ arlocher, 1992; Gessner and Chauvet, 1994). Leaf litter decomposition is sensitive to variations in water quality and has been recommended as a surrogate for assessing the function and health of freshwater ecosystems (B€ arlocher, 1992; Martínez et al., 2014; Chauvet et al., 2016). Previous studies have demonstrated that metals (Ferreira et al., 2016), nanomaterials (Pradhan et al., 2011) and other pollutants ndez et al., 2015) impair freshwater like fungicides (Ferna
https://doi.org/10.1016/j.funeco.2019.02.011 1754-5048/© 2019 Elsevier Ltd and British Mycological Society. All rights reserved.
Please cite this article as: Seena, S et al., Does nanosized plastic affect aquatic fungal litter decomposition?, Fungal Ecology, https://doi.org/ 10.1016/j.funeco.2019.02.011
2
S. Seena et al. / Fungal Ecology xxx (xxxx) xxx
ecosystem functioning. Still, none of the earlier studies have attempted to address the impact of nanosized plastics on freshwater ecosystem functioning. Polystyrene is a versatile plastic used in a wide range of products. Its applications include food-packing, fast-food and cosmetic containers, medical pipettes and Petri dishes (Rosemond et al., 2010). Polystyrene monomers have been reported to show adverse effects on animal genes, lungs and livers (Roe, 1994; Koelmans et al., 2015; Lu et al., 2016). Since they are used mainly in the manufacture of single-use products and not biodegradable, polystyrene monomers are a significant contributor to the debris in aquatic ecosystems (Syranidou et al., 2017). Since litter decomposition is an indicator of freshwater ecosystem functioning and health (Chauvet et al., 2016; Ferreira rold, 2017); we sought to elucidate the effects of NPPs and Gue (100 nm) on freshwaters via a litter decomposition (leaf mass loss) experiment. We hypothesised that NPPs will negatively impact leaf litter decomposition and that the effects of NPPs are more pronounced with increasing concentrations. We also predicted that species identity will affect the impact on litter decomposition (Duarte et al., 2006) by various concentrations of NPPs. We used five species of worldwide-distributed aquatic hyphomycetes in microcosms to simulate a natural stream environment under controlled conditions. We measured leaf mass loss and characterized the NPPs, focusing on their size and morphology. 2. Materials and methods 2.1. Fungal species Five species of aquatic hyphomycetes, namely Anguillospora crassa (ANCR), Tetracladium marchalianum (TEMA) Tetrachaetum elegans (TEEL), Articulospora tetracladia (ARTE) and Tricladium splendens (TRSP) were isolated from pristine streams of Switzerland from submerged leaf litter according to Descals (2005) (Table 1). The isolates were sub-cultured for one month (15 C, 12h photoperiod) in Petri dishes (100 20 mm) with 20 mL of 2% Malt Extract Agar (Difco™ BD™).
VEGA3 SBH from TESCAN, 15 kV. The samples were sputter coated with platinum and all the images were taken at magnifications of 5000. Size of NPPs in stock suspension was measured by dynamic light scattering (DLS) via a Zeta PALS Zeta Potential Analyzer (Brookhaven Instruments Corporation). 2.3. Leaf litter For decomposition experiments, we used leaves of Alnus glutinosa (Black alder: Betulaceae), a common riparian species with high-quality leaves (Boyero et al., 2011). Newly fallen leaves were collected at a single site at Parque Verde do Mondego Coimbra, Portugal (40º1102100 N, 8º250 30W00 ), dried and stored at room temperature. Leaves were leached for 24 h in distilled water and 12 mm discs were punched out, using a cork borer. Sets of 20 leaf discs were autoclaved (120 C for 20 min), dried at 105 C for 48 h (Thermo Scientific Heratherm) and weighed (AS 220/C/2, RADWAG) to obtain their initial mass. 2.4. Experimental set-up To assess the impact of NPPs on litter decomposition, sets of 20 autoclaved and pre-weighed leaf discs were placed in Erlenmeyer flasks (100 mL, 3 replicates) and exposed to 20 mL of increasing nominal concentrations of NPPs. Each flask was inoculated randomly with 15 fungal plugs (5 mm) punched out from onemonth-old fungal cultures. Each microcosm consisted of fungal plugs of only one aquatic hyphomycete species. All microcosms were kept for 25 d in an orbital shaker (114 rpm) at 15 C and under a 12 h photoperiod. The solutions were renewed every 5 d. At each renewal, fungal plugs were retained in the microcosms. Leaf discs in the microcosms release nitrogen and phosphorous to water (Howarth and Fisher, 1976); the agar plugs provide additional
2.2. Suspension of nanosized polystyrene plastics We used 100 nm polystyrene plastics (Sigma-Aldrich; 100,000 mg/L aqueous suspension). Nominal concentrations of NPPs used in microcosms were 0 (control), 1.6, 6.4, 25.6 and 102.4 mg/L (4-fold increase at each step). The selected range of NPPs encompasses the published concentrations of nanosized plastics with impacts on aquatic organisms (Table S1). For higher concentrations (25.6 and 102.4 mg/L) the NPPs used were taken directly from the aqueous suspension (100,000 mg/L) and diluted in sterile distilled water. For lower concentrations, 100 mg/L of NPPs stock suspension was prepared from 100,000 mg/L aqueous suspension. The lower concentrations (1.6 and 6.4 mg/L) were prepared by diluting the NPPs from the stock suspension in distilled water. We characterized the nanosized polystyrene surface topography by scanning electron microscopy (SEM). SEM was performed on a Hitachi TM-1000 table top microscope, operating at 5 kV and on a Table 1 The species used, acronym and location of isolation. Species
Acronym
GPS Coordinates
Tricladium splendens Articulospora tetracladia Anguillospora crassa Tetrachaetum elegans Tetracladium marchalianum
TRSP ARTE ANCR TEEL TEMA
9 010 18.300 E46 120 59.200 N 9 060 43.200 E46 140 30.500 N 8 530 53.400 E46 100 44.500 N 9 010 18.300 E46 120 59.200 N 8 590 40.500 E46 140 06.200 N
Fig. 1. Workflow of litter decomposition in microcosms.
Please cite this article as: Seena, S et al., Does nanosized plastic affect aquatic fungal litter decomposition?, Fungal Ecology, https://doi.org/ 10.1016/j.funeco.2019.02.011
S. Seena et al. / Fungal Ecology xxx (xxxx) xxx
3
Fig. 2. SEM images of NPPs at (A) 1.6 mg/L, (B) 6.4 mg/L, (C) 25.6 mg/L and (D) 102.4 mg/L concentrations.
nutrients necessary for the fungi. The general workflow and experimental set-up of the study are shown in Fig. 1 and S1, respectively. After incubation, leaf discs from each microcosm were dried at 105 C for 48 h and weighed to assess the mass loss.
2.5. Statistical analyses Data were checked for normal distribution (D'Agostino-Pearson test) using GraphPad Prism version 6.00 for Mac (GraphPad Software, La Jolla, CA; www.graphpad.com). Data followed normal distribution but failed to meet the assumptions of homogeneity after transformation (Levene's or Bartlett's test; Statistica 7 software; StatSoft Inc., Tulsa,OK, U.S.A.). Univariate permutational analysis of variance (PERMANOVA) was used to determine the effects of NPPs concentrations and aquatic hyphomycete species on leaf mass loss. PERMANOVA was performed by using the PRIMER version 6.00 (Clarke and Gorley, 2006) with PERMANOVA þ add-on package (Anderson et al., 2008). PERMANOVA is particularly suited when data do not meet the standard assumptions and is formally equivalent to standard ANOVA (Anderson et al., 2008). Univariate PERMANOVA were run on Euclidian distance (Anderson et al., 2008), residuals were permutated under a reduced model, with 9999 permutations. PERMANOVA was followed by post hoc pairwise tests (Anderson et al., 2008), achieved by an additional separate run of the PERMANOVA routine. The p-values for all pairwise tests in PERMANOVA were acquired using permutations; p < 0.05 were considered significant.
3. Results The average diameter of NPPs was 150 nm and 147.98 nm ± 7.28 (Fig. S2) by DLS and SEM respectively. NPPs were spherical and the number of spheres increased with concentration NPPs (Fig. 2). The NPPs suspension was well dispersed with little or no agglomeration. Litter mass loss significantly differed among concentrations (F4,50 ¼ 4.06, p ¼ 0.005) and aquatic hyphomycete species (twoway PERMANOVA, F4,50 ¼ 4.83, p ¼ 0.0016) with no interactions (F16,50 ¼ 0.849, p ¼ 0.643). The average litter mass loss ranged from 21.9% (0 mg/L) to 14.4% (102.4 mg/L) (Fig. 3A). Among the concentrations, significant differences were observed between 0 and 6.4 and 102.4 mg/L (Pair-wise test, p ¼ 0.0001 and 0.0238, respectively) and also between 1.6 and 102.4 mg/L (p ¼ 0.0043) (Fig. 3A). Among the species, average litter mass loss ranged from 13.2% (T. splendens; 102.4 mg/L) to 36.6% (T. marchalianum; 0 mg/L) (Fig. 3B). Litter decomposition by A. crassa differed significantly from those by T. marchalianum (Pair-wise test, p ¼ 0.0075) and T. elegans (p ¼ 0.0074). Significant differences were also observed between T. splendens and T. marchalianum and T. elegans (p ¼ 0.0026 and 0.0004, respectively) (Fig. 3B).
4. Discussion We demonstrated that litter decomposition was impacted by NPP concentrations and that aquatic fungal species differ in their tolerance to NPPs. Litter decomposition by TEMA was markedly
Please cite this article as: Seena, S et al., Does nanosized plastic affect aquatic fungal litter decomposition?, Fungal Ecology, https://doi.org/ 10.1016/j.funeco.2019.02.011
4
S. Seena et al. / Fungal Ecology xxx (xxxx) xxx
However, because of interspecific differences, it remains unclear what will happen in a multi-species natural system. If the remaining tolerant species are capable of functionally replacing the least tolerant species, then global ecological function (litter decomposition) may remain unchanged at relatively high concentrations of NPPs. This points to the potential importance of high fungal diversity in freshwater streams and suggests that such systems will be more resilient when confronted with plastic pollution. Acknowledgments Access to TAIL-UC facility funded under QREN-Mais Centro Project ICT/2009/02/012/1890 is gratefully acknowledged. S.S. was supported by FCT (SFRH/BPD/103865/2014 and UID/MAR/04292/ 2019). Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.funeco.2019.02.011. References
Fig. 3. Litter mass loss when exposed to different concentrations of (A) nanosized polystyrene plastics and (B) species (mean ± standard error; n ¼ 3). Different letters on top indicate significant differences between concentrations or species (p < 0.05).
affected when compared to other species, even at the lowest concentrations used (1.6 mg/L). Generally, TEMA exhibits faster growth and superior decomposition capability, making it often dominant in €rlocher and Kendrick, 1973; aquatic hyphomycete communities (Ba Arsuffi and Suberkropp, 1985). Duarte et al. (2006) previously demonstrated that specific traits of aquatic hyphomycetes affect their ecological functions and that and their effect on the ecosystem may be selectively heightened or weakened at higher trophic levels €rlocher and Kendrick, 1973; Arsuffi and Suberkropp, 1985). For (Ba instance, the identity of fungi that colonized the leaves strongly €rlocher influences selection of food by leaf-eating invertebrates (Ba and Kendrick, 1973; Arsuffi and Suberkropp, 1985; Lecerf et al., 2005). Very few studies have provided any evidence of the effects of NPPs on aquatic organisms at concentrations higher than ~0.5 mg/L NPPs (Koelmans et al., 2015). For example, developmental effects on sea urchins (Paracentrotus lividus) were observed with 2.61 and 3.85 mg/L of NPPs (Della Torre et al., 2014), ingestion and mortality of copepods (Tigriopus japonicus) were seen at concentrations of 12.5 mg/L and 1.25 mg/L (Lee et al., 2013) and absorption of 1.8e6.5 mg/L of NPPs was demonstrated to affect algal photosynthesis (Bhattacharya et al., 2010). Currently, there are no reference data available regarding environmental concentrations of NPPs; nevertheless, the lowest effect in our study was observed at 1.6 mg/L NPPs, which is roughly four to six orders of magnitude higher than the 0.4e34 ng/L microplastic concentrations reported in freshwaters in the USA (Eriksen et al., 2013) and Europe (Besseling et al., 2014). Our findings document the concentrations of NPPs where ecological functions of aquatic hyphomycetes are impaired.
Alimi, O.S., Farner Budarz, J., Hernandez, L.M., Tufenkji, N., 2018. Microplastics and nanoplastics in aquatic environments: aggregation, deposition, and enhanced contaminant transport. Environ. Sci. Technol. 52 (4), 1704e1724. Anderson, M., Gorley, R.N., Clarke, K.R., 2008. PERMANOVAþ for PRIMER: Guide to Software and Statistical Methods. PRIMER-E. Arsuffi, T.L., Suberkropp, K., 1985. Selective feeding by stream caddisfly (Trichoptera) detritivores on leaves with fungal-colonized patches. Oikos 45, 50e58. €rlocher, F., Kendrick, B., 1973. Fungi and food preferences of Gammarus pseudoBa limnaeus. Arch. Hydrobiol. 72, 501e516. €rlocher, F., 1992. Research on aquatic hyphomycetes: historical background and Ba overview. In: B€ arlocher, F. (Ed.), The Ecology of Aquatic Hyphomycetes. Ecological Studies (Analysis and Synthesis). Springer, Berlin, Heidelberg, pp. 1e15. Besseling, E., Foekema, E.M., Koelmans, A.A., 2014. Preliminary Investigation of Micro- Plastic in the Management Area of Water Board Rivierenland (In Dutch). Wageningen University, Wageningen, The Netherlands, pp. 1e18. http://edepot. wur.nl/299787. Bhattacharya, P., Lin, S., Turner, J.P., Ke, P.C., 2010. Physical adsorption of charged plastic nanoparticles affects algal photosynthesis. J. Phys. Chem. C 114 (39), 16556e16561. Blair, R.M., Waldron, S., Phoenix, V., Gauchotte-Lindsay, C., 2017. Micro- and Nanoplastic Pollution of Freshwater and Wastewater Treatment Systems. Springer Science Reviews, pp. 19e30, 5. Boyero, L., Pearson, R.G., Gessner, M.O., Barmuta, L.A., Ferreira, V., Graça, M.A.S., Dudgeon, D., Boulton, A.J., Callisto, M., Chauvet, E., Helson, J.E., Bruder, A., ~ o, R.J., Yule, C.M., Arunachalam, M., Davies, J.N., Figueroa, R., Flecker, A.S., Albarin Ramírez, A., Death, R.G., Iwata, T., Mathooko, J.M., Mathuriau, C., Gonçalves, J.F., Moretti, M.S., Jinggut, T., Lamothe, S., M'Erimba, C., Ratnarajah, L., Schindler, M.H., Castela, J., Buria, L.M., Cornejo, A., Villanueva, V.D., West, D.C., 2011. A global experiment suggests climate warming will not accelerate litter decomposition in streams but might reduce carbon sequestration. Ecol. Lett. 14 (3), 289e294. Chauvet, E., Ferreira, V., Giller, P.S., McKie, B.G., Tiegs, S.D., Woodward, G., Elosegi, A., re, J.O., Dobson, M., Fleituch, T., Graça, M.A.S., Gulis, V., Hladyz, S., Lacoursie Lecerf, A., Pozo, J., Preda, E., Riipinen, M., Rîs¸noveanu, G., Vadineanu, A., Vought, L.B.M., Gessner, M.O., 2016. Chapter three - litter decomposition as an indicator of stream ecosystem functioning at local-to-continental scales: insights from the European RivFunction project. In: Dumbrell, A.J., Kordas, R.L., Woodward, G. (Eds.), Advances in Ecological Research, vol. 55. Academic Press, pp. 99e182. Clarke, K.R., Gorley, R.N., 2006. Primer V6: User Manual - Tutorial. Plymouth Marine Laboratory, Plymouth, UK. Della Torre, C., Bergami, E., Salvati, A., Faleri, C., Cirino, P., Dawson, K.A., Corsi, I., 2014. Accumulation and embryotoxicity of polystyrene nanoparticles at early stage of development of sea urchin embryos Paracentrotus lividus. Environ. Sci. Technol. 48 (20), 12302e12311. Descals, E., 2005. Techniques for handling Ingoldian fungi. In: Graça, M.A.S., B€ arlocher, F., Gessner, M.O. (Eds.), Methods to Study Litter Decomposition: A Practical Guide. Springer, Dordrecht, pp. 129e141. €rlocher, F., 2006. Aquatic hyphomycete diversity Duarte, S., Pascoal, C., C assio, F., Ba and identity affect leaf litter decomposition in microcosms. Oecologia 147 (4), 658e666. Eriksen, M., Mason, S., Wilson, S., Box, C., Zellers, A., Edwards, W., Farley, H., Amato, S., 2013. Microplastic pollution in the surface waters of the Laurentian great lakes. Mar. Pollut. Bull. 77 (1), 177e182.
Please cite this article as: Seena, S et al., Does nanosized plastic affect aquatic fungal litter decomposition?, Fungal Ecology, https://doi.org/ 10.1016/j.funeco.2019.02.011
S. Seena et al. / Fungal Ecology xxx (xxxx) xxx ndez, D., Voss, K., Bundschuh, M., Zubrod, J.P., Sch€ Ferna afer, R.B., 2015. Effects of fungicides on decomposer communities and litter decomposition in vineyard streams. Sci. Total Environ. 533, 40e48. rold, F., 2016. Effects of Ferreira, V., Koricheva, J., Duarte, S., Niyogi, D.K., Gue anthropogenic heavy metal contamination on litter decomposition in streams e a meta-analysis. Environ. Pollut. 210, 261e270. rold, F., 2017. Leaf litter decomposition as a bioassessment tool of Ferreira, V., Gue acidification effects in streams: evidence from a field study and meta-analysis. Ecol. Indicat. 79, 382e390. Gessner, M.O., Chauvet, E., 1994. Importance of stream microfungi in controlling breakdown rates of leaf litter. Ecology 75 (6), 1807e1817. Geyer, R., Jambeck, J.R., Law, K.L., 2017. Production, use, and fate of all plastics ever made. Science Advances 3 (7). ~ o, F., Graça, M.A.S., Ferreira, V., Canhoto, C., Encalada, A.C., Guerrero-Bolan Wantzen, K.M., Boyero, L., 2015. A conceptual model of litter breakdown in low order streams. Int. Rev. Hydrobiol. 100 (1), 1e12. Howarth, R.W., Fisher, S.G., 1976. Carbon, nitrogen, and phosphorus dynamics during leaf decay in nutrient-enriched stream microecosystems. Freshw. Biol. 6 (3), 221e228. Hurley, R., Woodward, J., Rothwell, J.J., 2018. Microplastic contamination of river beds significantly reduced by catchment-wide flooding. Nat. Geosci. 11 (4), 251e257. Koelmans, A., Besseling, E., Shim, W., 2015. Nanoplastics in the aquatic environment. Critical review. In: Bergmann, M., Gutow, L., Klages, M. (Eds.), Marine Anthropogenic Litter. Springer, Cham, pp. 325e340. Lebreton, L.C.M., van der Zwet, J., Damsteeg, J.-W., Slat, B., Andrady, A., Reisser, J., 2017. River plastic emissions to the world's oceans. Nat. Commun. 8, 15611. Lecerf, A., Dobson, M., Dang, C.K., Chauvet, E., 2005. Riparian plant species loss alters trophic dynamics in detritus-based stream ecosystems. Oecologia 146, 432e442. Lee, K.-W., Shim, W.J., Kwon, O.Y., Kang, J.-H., 2013. Size-dependent effects of micro polystyrene particles in the marine copepod Tigriopus japonicus. Environ. Sci.
5
Technol. 47 (19), 11278e11283. Lu, Y., Zhang, Y., Deng, Y., Jiang, W., Zhao, Y., Geng, J., Ding, L., Ren, H., 2016. Uptake and Accumulation of Polystyrene Microplastics in Zebrafish (Danio rerio) and Toxic Effects in Liver. Environmental Science & Technology 50, 4054e4060. ~ aga, A., Pe rez, J., Descals, E., Pozo, J., 2014. Temperature affects Martínez, A., Larran leaf litter decomposition in low-order forest streams: field and microcosm approaches. FEMS (Fed. Eur. Microbiol. Soc.) Microbiol. Ecol. 87 (1), 257e267. ssio, F., 2011. Can metal nanoparticles be a Pradhan, A., Seena, S., Pascoal, C., Ca threat to microbial decomposers of plant litter in streams? Microb. Ecol. 62 (1), 58e68. Roe, F.J.C., 1994. Styrene: toxicity studies d what do they show? Crit. Rev. Toxicol. 24 (Suppl. 1), s117es125. Rosemond, Z.B.A., Chou, S., Wilson, J., Schwartz, M., Tomei-Torres, F., Ingerman, L., Knaebel, D.B., Citra, M., 2010. Toxicological Profile for Styrene. Public Health Service, Agency for Toxic Substances and Disease Registry, U.S. Department of Health and Human Services. https://www.atsdr.cdc.gov/toxprofiles/tp53.pdf. ssio, F., Pascoal, C., 2017. Does the developmental stage and Seena, S., Carvalho, F., Ca composition of riparian forest stand affect ecosystem functioning in streams? Sci. Total Environ. 609, 1500e1511. Syranidou, E., Karkanorachaki, K., Amorotti, F., Franchini, M., Repouskou, E., Kaliva, M., Vamvakaki, M., Kolvenbach, B., Fava, F., Corvini, P.F.X., Kalogerakis, N., 2017. Biodegradation of weathered polystyrene films in seawater microcosms. Sci. Rep. 7, 17991. ~ oz, D., Brennholt, N., Bourrain, X., Buchinger, S., Wagner, M., Scherer, C., Alvarez-Mun Fries, E., Grosbois, C., Klasmeier, J., Marti, T., Rodriguez-Mozaz, S., Urbatzka, R., Vethaak, A.D., Winther-Nielsen, M., Reifferscheid, G., 2014. Microplastics in freshwater ecosystems: what we know and what we need to know. Environ. Sci. Eur. 26 (1), 12. Wegner, A., Besseling, E., Foekema, E.M., Kamermans, P., Koelmans, A.A., 2012. Effects of nanopolystyrene on the feeding behavior of the blue mussel (Mytilus edulis L.). Environ. Toxicol. Chem. 31 (11), 2490e2497.
Please cite this article as: Seena, S et al., Does nanosized plastic affect aquatic fungal litter decomposition?, Fungal Ecology, https://doi.org/ 10.1016/j.funeco.2019.02.011
Contents lists available at ScienceDirect
Fungal Ecology journal homepage: www.elsevier.com/locate/funeco
Does nanosized plastic affect aquatic fungal litter decomposition? Sahadevan Seena a, *, Diana Graça a, Anne Bartels a, Julien Cornut a, b a b
MARE-Marine and Environmental Sciences Centre, University of Coimbra, PT-3004-517, Coimbra, Portugal Laboratory of Applied Microbiology, University of Applied Sciences and Arts of Southern Switzerland, via Mirasole 22A, 6501, Bellinzona, Switzerland
a r t i c l e i n f o
a b s t r a c t
Article history: Received 26 October 2018 Received in revised form 18 February 2019 Accepted 22 February 2019 Available online xxx
Nanosized plastics are an emerging concern in freshwater ecosystems, raising the question whether they put freshwater ecological processes at risk. Litter decomposition is a major ecological function in forested streams which is mainly driven by aquatic hyphomycetes. Here we investigated whether increasing concentrations (up to 102.4 mg/L) of nanosized polystyrene plastics (NPPs; 100nm) affect litter decomposition by five widely distributed species of aquatic hyphomycetes. Results showed that average litter decomposition decreased by 8% relative to the control when exposed to 102.4 mg/L NPPs. Aquatic hyphomycete species differed in their sensitivity to NPPs. The greatest inhibition of litter decomposition was found with Tetracladium marchalianum, where it dropped from 37 (control) to 16% (102.4 mg/L of NPP). Overall our study highlights the emerging risks and potential dangers of NPPs to freshwater ecosystem functioning. It also indicates that the impact of NPPs may be species specific. © 2019 Elsevier Ltd and British Mycological Society. All rights reserved.
Corresponding Editor: Prof. L. Boddy Keywords: Freshwater ecosystems Streams Aquatic hyphomycetes Litter decomposition Nanoparticles Polystyrene
1. Introduction Plastic pollution is now acknowledged as a significant threat to aquatic environments (for review see Koelmans et al., 2015). Virgin plastic production has been estimated to be 8300 million metric tons (Mt) and by 2050 around 12,000 Mt of plastic waste is envisioned to be in landfills or in the natural environment (Geyer et al., 2017). It is predicted that every year between 1.15 and 2.41 million metric tonnes of plastic enter the ocean via global riverine systems (Lebreton et al., 2017). Although most studies dealing with the consequences of plastics in ecosystems have been performed in marine systems, rivers are the main source of marine plastic pollution (Lebreton et al., 2017). Plastics can be physically fragmented into nanosized particles whose environmental effects are virtually unknown and of increasing concern. Nanosized plastics are also used in a vast range of products such as toothpaste, waterborne paints and biomedical products (Koelmans et al., 2015). Unlike macroplastics, micro and nanoscale plastics are difficult to detect and quantify (Blair et al., 2017; Alimi et al., 2018). A recent survey conducted after winter
* Corresponding author. E-mail addresses: [email protected], (S. Seena).
[email protected]
floods in the River Tame, UK, reported over 500,000 microplastic particles per square meter (Hurley et al., 2018). At the nanoscale, plastics are even harder to detect; hence their exact environmental concentration and distribution are unknown. This poses severe challenges when studying the impact of nanosized plastics on organisms or ecosystems for using realistic environmental concentrations (Wagner et al., 2014; Koelmans et al., 2015; Blair et al., 2017). Despite this limitation, recent studies have demonstrated nanosized polystyrene plastics (NPPs) to be hazardous, for instance, by interfering with algal photosynthesis (Bhattacharya et al., 2010), reducing filter-feeding activity of mussels (Mytilus adulis) (Wegner et al., 2012) and causing mortality of copepods upon ingestion (Tigriopus japonicus) (Lee et al., 2013). In woodland streams, leaf litter is the primary source of carbon and energy (Graça et al., 2015; Chauvet et al., 2016; Seena et al., 2017). Leaf litter decomposition is a crucial ecosystem process in forested streams; aquatic fungi, especially aquatic hyphomycetes are key players responsible for leaf breakdown (B€ arlocher, 1992; Gessner and Chauvet, 1994). Leaf litter decomposition is sensitive to variations in water quality and has been recommended as a surrogate for assessing the function and health of freshwater ecosystems (B€ arlocher, 1992; Martínez et al., 2014; Chauvet et al., 2016). Previous studies have demonstrated that metals (Ferreira et al., 2016), nanomaterials (Pradhan et al., 2011) and other pollutants ndez et al., 2015) impair freshwater like fungicides (Ferna
https://doi.org/10.1016/j.funeco.2019.02.011 1754-5048/© 2019 Elsevier Ltd and British Mycological Society. All rights reserved.
Please cite this article as: Seena, S et al., Does nanosized plastic affect aquatic fungal litter decomposition?, Fungal Ecology, https://doi.org/ 10.1016/j.funeco.2019.02.011
2
S. Seena et al. / Fungal Ecology xxx (xxxx) xxx
ecosystem functioning. Still, none of the earlier studies have attempted to address the impact of nanosized plastics on freshwater ecosystem functioning. Polystyrene is a versatile plastic used in a wide range of products. Its applications include food-packing, fast-food and cosmetic containers, medical pipettes and Petri dishes (Rosemond et al., 2010). Polystyrene monomers have been reported to show adverse effects on animal genes, lungs and livers (Roe, 1994; Koelmans et al., 2015; Lu et al., 2016). Since they are used mainly in the manufacture of single-use products and not biodegradable, polystyrene monomers are a significant contributor to the debris in aquatic ecosystems (Syranidou et al., 2017). Since litter decomposition is an indicator of freshwater ecosystem functioning and health (Chauvet et al., 2016; Ferreira rold, 2017); we sought to elucidate the effects of NPPs and Gue (100 nm) on freshwaters via a litter decomposition (leaf mass loss) experiment. We hypothesised that NPPs will negatively impact leaf litter decomposition and that the effects of NPPs are more pronounced with increasing concentrations. We also predicted that species identity will affect the impact on litter decomposition (Duarte et al., 2006) by various concentrations of NPPs. We used five species of worldwide-distributed aquatic hyphomycetes in microcosms to simulate a natural stream environment under controlled conditions. We measured leaf mass loss and characterized the NPPs, focusing on their size and morphology. 2. Materials and methods 2.1. Fungal species Five species of aquatic hyphomycetes, namely Anguillospora crassa (ANCR), Tetracladium marchalianum (TEMA) Tetrachaetum elegans (TEEL), Articulospora tetracladia (ARTE) and Tricladium splendens (TRSP) were isolated from pristine streams of Switzerland from submerged leaf litter according to Descals (2005) (Table 1). The isolates were sub-cultured for one month (15 C, 12h photoperiod) in Petri dishes (100 20 mm) with 20 mL of 2% Malt Extract Agar (Difco™ BD™).
VEGA3 SBH from TESCAN, 15 kV. The samples were sputter coated with platinum and all the images were taken at magnifications of 5000. Size of NPPs in stock suspension was measured by dynamic light scattering (DLS) via a Zeta PALS Zeta Potential Analyzer (Brookhaven Instruments Corporation). 2.3. Leaf litter For decomposition experiments, we used leaves of Alnus glutinosa (Black alder: Betulaceae), a common riparian species with high-quality leaves (Boyero et al., 2011). Newly fallen leaves were collected at a single site at Parque Verde do Mondego Coimbra, Portugal (40º1102100 N, 8º250 30W00 ), dried and stored at room temperature. Leaves were leached for 24 h in distilled water and 12 mm discs were punched out, using a cork borer. Sets of 20 leaf discs were autoclaved (120 C for 20 min), dried at 105 C for 48 h (Thermo Scientific Heratherm) and weighed (AS 220/C/2, RADWAG) to obtain their initial mass. 2.4. Experimental set-up To assess the impact of NPPs on litter decomposition, sets of 20 autoclaved and pre-weighed leaf discs were placed in Erlenmeyer flasks (100 mL, 3 replicates) and exposed to 20 mL of increasing nominal concentrations of NPPs. Each flask was inoculated randomly with 15 fungal plugs (5 mm) punched out from onemonth-old fungal cultures. Each microcosm consisted of fungal plugs of only one aquatic hyphomycete species. All microcosms were kept for 25 d in an orbital shaker (114 rpm) at 15 C and under a 12 h photoperiod. The solutions were renewed every 5 d. At each renewal, fungal plugs were retained in the microcosms. Leaf discs in the microcosms release nitrogen and phosphorous to water (Howarth and Fisher, 1976); the agar plugs provide additional
2.2. Suspension of nanosized polystyrene plastics We used 100 nm polystyrene plastics (Sigma-Aldrich; 100,000 mg/L aqueous suspension). Nominal concentrations of NPPs used in microcosms were 0 (control), 1.6, 6.4, 25.6 and 102.4 mg/L (4-fold increase at each step). The selected range of NPPs encompasses the published concentrations of nanosized plastics with impacts on aquatic organisms (Table S1). For higher concentrations (25.6 and 102.4 mg/L) the NPPs used were taken directly from the aqueous suspension (100,000 mg/L) and diluted in sterile distilled water. For lower concentrations, 100 mg/L of NPPs stock suspension was prepared from 100,000 mg/L aqueous suspension. The lower concentrations (1.6 and 6.4 mg/L) were prepared by diluting the NPPs from the stock suspension in distilled water. We characterized the nanosized polystyrene surface topography by scanning electron microscopy (SEM). SEM was performed on a Hitachi TM-1000 table top microscope, operating at 5 kV and on a Table 1 The species used, acronym and location of isolation. Species
Acronym
GPS Coordinates
Tricladium splendens Articulospora tetracladia Anguillospora crassa Tetrachaetum elegans Tetracladium marchalianum
TRSP ARTE ANCR TEEL TEMA
9 010 18.300 E46 120 59.200 N 9 060 43.200 E46 140 30.500 N 8 530 53.400 E46 100 44.500 N 9 010 18.300 E46 120 59.200 N 8 590 40.500 E46 140 06.200 N
Fig. 1. Workflow of litter decomposition in microcosms.
Please cite this article as: Seena, S et al., Does nanosized plastic affect aquatic fungal litter decomposition?, Fungal Ecology, https://doi.org/ 10.1016/j.funeco.2019.02.011
S. Seena et al. / Fungal Ecology xxx (xxxx) xxx
3
Fig. 2. SEM images of NPPs at (A) 1.6 mg/L, (B) 6.4 mg/L, (C) 25.6 mg/L and (D) 102.4 mg/L concentrations.
nutrients necessary for the fungi. The general workflow and experimental set-up of the study are shown in Fig. 1 and S1, respectively. After incubation, leaf discs from each microcosm were dried at 105 C for 48 h and weighed to assess the mass loss.
2.5. Statistical analyses Data were checked for normal distribution (D'Agostino-Pearson test) using GraphPad Prism version 6.00 for Mac (GraphPad Software, La Jolla, CA; www.graphpad.com). Data followed normal distribution but failed to meet the assumptions of homogeneity after transformation (Levene's or Bartlett's test; Statistica 7 software; StatSoft Inc., Tulsa,OK, U.S.A.). Univariate permutational analysis of variance (PERMANOVA) was used to determine the effects of NPPs concentrations and aquatic hyphomycete species on leaf mass loss. PERMANOVA was performed by using the PRIMER version 6.00 (Clarke and Gorley, 2006) with PERMANOVA þ add-on package (Anderson et al., 2008). PERMANOVA is particularly suited when data do not meet the standard assumptions and is formally equivalent to standard ANOVA (Anderson et al., 2008). Univariate PERMANOVA were run on Euclidian distance (Anderson et al., 2008), residuals were permutated under a reduced model, with 9999 permutations. PERMANOVA was followed by post hoc pairwise tests (Anderson et al., 2008), achieved by an additional separate run of the PERMANOVA routine. The p-values for all pairwise tests in PERMANOVA were acquired using permutations; p < 0.05 were considered significant.
3. Results The average diameter of NPPs was 150 nm and 147.98 nm ± 7.28 (Fig. S2) by DLS and SEM respectively. NPPs were spherical and the number of spheres increased with concentration NPPs (Fig. 2). The NPPs suspension was well dispersed with little or no agglomeration. Litter mass loss significantly differed among concentrations (F4,50 ¼ 4.06, p ¼ 0.005) and aquatic hyphomycete species (twoway PERMANOVA, F4,50 ¼ 4.83, p ¼ 0.0016) with no interactions (F16,50 ¼ 0.849, p ¼ 0.643). The average litter mass loss ranged from 21.9% (0 mg/L) to 14.4% (102.4 mg/L) (Fig. 3A). Among the concentrations, significant differences were observed between 0 and 6.4 and 102.4 mg/L (Pair-wise test, p ¼ 0.0001 and 0.0238, respectively) and also between 1.6 and 102.4 mg/L (p ¼ 0.0043) (Fig. 3A). Among the species, average litter mass loss ranged from 13.2% (T. splendens; 102.4 mg/L) to 36.6% (T. marchalianum; 0 mg/L) (Fig. 3B). Litter decomposition by A. crassa differed significantly from those by T. marchalianum (Pair-wise test, p ¼ 0.0075) and T. elegans (p ¼ 0.0074). Significant differences were also observed between T. splendens and T. marchalianum and T. elegans (p ¼ 0.0026 and 0.0004, respectively) (Fig. 3B).
4. Discussion We demonstrated that litter decomposition was impacted by NPP concentrations and that aquatic fungal species differ in their tolerance to NPPs. Litter decomposition by TEMA was markedly
Please cite this article as: Seena, S et al., Does nanosized plastic affect aquatic fungal litter decomposition?, Fungal Ecology, https://doi.org/ 10.1016/j.funeco.2019.02.011
4
S. Seena et al. / Fungal Ecology xxx (xxxx) xxx
However, because of interspecific differences, it remains unclear what will happen in a multi-species natural system. If the remaining tolerant species are capable of functionally replacing the least tolerant species, then global ecological function (litter decomposition) may remain unchanged at relatively high concentrations of NPPs. This points to the potential importance of high fungal diversity in freshwater streams and suggests that such systems will be more resilient when confronted with plastic pollution. Acknowledgments Access to TAIL-UC facility funded under QREN-Mais Centro Project ICT/2009/02/012/1890 is gratefully acknowledged. S.S. was supported by FCT (SFRH/BPD/103865/2014 and UID/MAR/04292/ 2019). Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.funeco.2019.02.011. References
Fig. 3. Litter mass loss when exposed to different concentrations of (A) nanosized polystyrene plastics and (B) species (mean ± standard error; n ¼ 3). Different letters on top indicate significant differences between concentrations or species (p < 0.05).
affected when compared to other species, even at the lowest concentrations used (1.6 mg/L). Generally, TEMA exhibits faster growth and superior decomposition capability, making it often dominant in €rlocher and Kendrick, 1973; aquatic hyphomycete communities (Ba Arsuffi and Suberkropp, 1985). Duarte et al. (2006) previously demonstrated that specific traits of aquatic hyphomycetes affect their ecological functions and that and their effect on the ecosystem may be selectively heightened or weakened at higher trophic levels €rlocher and Kendrick, 1973; Arsuffi and Suberkropp, 1985). For (Ba instance, the identity of fungi that colonized the leaves strongly €rlocher influences selection of food by leaf-eating invertebrates (Ba and Kendrick, 1973; Arsuffi and Suberkropp, 1985; Lecerf et al., 2005). Very few studies have provided any evidence of the effects of NPPs on aquatic organisms at concentrations higher than ~0.5 mg/L NPPs (Koelmans et al., 2015). For example, developmental effects on sea urchins (Paracentrotus lividus) were observed with 2.61 and 3.85 mg/L of NPPs (Della Torre et al., 2014), ingestion and mortality of copepods (Tigriopus japonicus) were seen at concentrations of 12.5 mg/L and 1.25 mg/L (Lee et al., 2013) and absorption of 1.8e6.5 mg/L of NPPs was demonstrated to affect algal photosynthesis (Bhattacharya et al., 2010). Currently, there are no reference data available regarding environmental concentrations of NPPs; nevertheless, the lowest effect in our study was observed at 1.6 mg/L NPPs, which is roughly four to six orders of magnitude higher than the 0.4e34 ng/L microplastic concentrations reported in freshwaters in the USA (Eriksen et al., 2013) and Europe (Besseling et al., 2014). Our findings document the concentrations of NPPs where ecological functions of aquatic hyphomycetes are impaired.
Alimi, O.S., Farner Budarz, J., Hernandez, L.M., Tufenkji, N., 2018. Microplastics and nanoplastics in aquatic environments: aggregation, deposition, and enhanced contaminant transport. Environ. Sci. Technol. 52 (4), 1704e1724. Anderson, M., Gorley, R.N., Clarke, K.R., 2008. PERMANOVAþ for PRIMER: Guide to Software and Statistical Methods. PRIMER-E. Arsuffi, T.L., Suberkropp, K., 1985. Selective feeding by stream caddisfly (Trichoptera) detritivores on leaves with fungal-colonized patches. Oikos 45, 50e58. €rlocher, F., Kendrick, B., 1973. Fungi and food preferences of Gammarus pseudoBa limnaeus. Arch. Hydrobiol. 72, 501e516. €rlocher, F., 1992. Research on aquatic hyphomycetes: historical background and Ba overview. In: B€ arlocher, F. (Ed.), The Ecology of Aquatic Hyphomycetes. Ecological Studies (Analysis and Synthesis). Springer, Berlin, Heidelberg, pp. 1e15. Besseling, E., Foekema, E.M., Koelmans, A.A., 2014. Preliminary Investigation of Micro- Plastic in the Management Area of Water Board Rivierenland (In Dutch). Wageningen University, Wageningen, The Netherlands, pp. 1e18. http://edepot. wur.nl/299787. Bhattacharya, P., Lin, S., Turner, J.P., Ke, P.C., 2010. Physical adsorption of charged plastic nanoparticles affects algal photosynthesis. J. Phys. Chem. C 114 (39), 16556e16561. Blair, R.M., Waldron, S., Phoenix, V., Gauchotte-Lindsay, C., 2017. Micro- and Nanoplastic Pollution of Freshwater and Wastewater Treatment Systems. Springer Science Reviews, pp. 19e30, 5. Boyero, L., Pearson, R.G., Gessner, M.O., Barmuta, L.A., Ferreira, V., Graça, M.A.S., Dudgeon, D., Boulton, A.J., Callisto, M., Chauvet, E., Helson, J.E., Bruder, A., ~ o, R.J., Yule, C.M., Arunachalam, M., Davies, J.N., Figueroa, R., Flecker, A.S., Albarin Ramírez, A., Death, R.G., Iwata, T., Mathooko, J.M., Mathuriau, C., Gonçalves, J.F., Moretti, M.S., Jinggut, T., Lamothe, S., M'Erimba, C., Ratnarajah, L., Schindler, M.H., Castela, J., Buria, L.M., Cornejo, A., Villanueva, V.D., West, D.C., 2011. A global experiment suggests climate warming will not accelerate litter decomposition in streams but might reduce carbon sequestration. Ecol. Lett. 14 (3), 289e294. Chauvet, E., Ferreira, V., Giller, P.S., McKie, B.G., Tiegs, S.D., Woodward, G., Elosegi, A., re, J.O., Dobson, M., Fleituch, T., Graça, M.A.S., Gulis, V., Hladyz, S., Lacoursie Lecerf, A., Pozo, J., Preda, E., Riipinen, M., Rîs¸noveanu, G., Vadineanu, A., Vought, L.B.M., Gessner, M.O., 2016. Chapter three - litter decomposition as an indicator of stream ecosystem functioning at local-to-continental scales: insights from the European RivFunction project. In: Dumbrell, A.J., Kordas, R.L., Woodward, G. (Eds.), Advances in Ecological Research, vol. 55. Academic Press, pp. 99e182. Clarke, K.R., Gorley, R.N., 2006. Primer V6: User Manual - Tutorial. Plymouth Marine Laboratory, Plymouth, UK. Della Torre, C., Bergami, E., Salvati, A., Faleri, C., Cirino, P., Dawson, K.A., Corsi, I., 2014. Accumulation and embryotoxicity of polystyrene nanoparticles at early stage of development of sea urchin embryos Paracentrotus lividus. Environ. Sci. Technol. 48 (20), 12302e12311. Descals, E., 2005. Techniques for handling Ingoldian fungi. In: Graça, M.A.S., B€ arlocher, F., Gessner, M.O. (Eds.), Methods to Study Litter Decomposition: A Practical Guide. Springer, Dordrecht, pp. 129e141. €rlocher, F., 2006. Aquatic hyphomycete diversity Duarte, S., Pascoal, C., C assio, F., Ba and identity affect leaf litter decomposition in microcosms. Oecologia 147 (4), 658e666. Eriksen, M., Mason, S., Wilson, S., Box, C., Zellers, A., Edwards, W., Farley, H., Amato, S., 2013. Microplastic pollution in the surface waters of the Laurentian great lakes. Mar. Pollut. Bull. 77 (1), 177e182.
Please cite this article as: Seena, S et al., Does nanosized plastic affect aquatic fungal litter decomposition?, Fungal Ecology, https://doi.org/ 10.1016/j.funeco.2019.02.011
S. Seena et al. / Fungal Ecology xxx (xxxx) xxx ndez, D., Voss, K., Bundschuh, M., Zubrod, J.P., Sch€ Ferna afer, R.B., 2015. Effects of fungicides on decomposer communities and litter decomposition in vineyard streams. Sci. Total Environ. 533, 40e48. rold, F., 2016. Effects of Ferreira, V., Koricheva, J., Duarte, S., Niyogi, D.K., Gue anthropogenic heavy metal contamination on litter decomposition in streams e a meta-analysis. Environ. Pollut. 210, 261e270. rold, F., 2017. Leaf litter decomposition as a bioassessment tool of Ferreira, V., Gue acidification effects in streams: evidence from a field study and meta-analysis. Ecol. Indicat. 79, 382e390. Gessner, M.O., Chauvet, E., 1994. Importance of stream microfungi in controlling breakdown rates of leaf litter. Ecology 75 (6), 1807e1817. Geyer, R., Jambeck, J.R., Law, K.L., 2017. Production, use, and fate of all plastics ever made. Science Advances 3 (7). ~ o, F., Graça, M.A.S., Ferreira, V., Canhoto, C., Encalada, A.C., Guerrero-Bolan Wantzen, K.M., Boyero, L., 2015. A conceptual model of litter breakdown in low order streams. Int. Rev. Hydrobiol. 100 (1), 1e12. Howarth, R.W., Fisher, S.G., 1976. Carbon, nitrogen, and phosphorus dynamics during leaf decay in nutrient-enriched stream microecosystems. Freshw. Biol. 6 (3), 221e228. Hurley, R., Woodward, J., Rothwell, J.J., 2018. Microplastic contamination of river beds significantly reduced by catchment-wide flooding. Nat. Geosci. 11 (4), 251e257. Koelmans, A., Besseling, E., Shim, W., 2015. Nanoplastics in the aquatic environment. Critical review. In: Bergmann, M., Gutow, L., Klages, M. (Eds.), Marine Anthropogenic Litter. Springer, Cham, pp. 325e340. Lebreton, L.C.M., van der Zwet, J., Damsteeg, J.-W., Slat, B., Andrady, A., Reisser, J., 2017. River plastic emissions to the world's oceans. Nat. Commun. 8, 15611. Lecerf, A., Dobson, M., Dang, C.K., Chauvet, E., 2005. Riparian plant species loss alters trophic dynamics in detritus-based stream ecosystems. Oecologia 146, 432e442. Lee, K.-W., Shim, W.J., Kwon, O.Y., Kang, J.-H., 2013. Size-dependent effects of micro polystyrene particles in the marine copepod Tigriopus japonicus. Environ. Sci.
5
Technol. 47 (19), 11278e11283. Lu, Y., Zhang, Y., Deng, Y., Jiang, W., Zhao, Y., Geng, J., Ding, L., Ren, H., 2016. Uptake and Accumulation of Polystyrene Microplastics in Zebrafish (Danio rerio) and Toxic Effects in Liver. Environmental Science & Technology 50, 4054e4060. ~ aga, A., Pe rez, J., Descals, E., Pozo, J., 2014. Temperature affects Martínez, A., Larran leaf litter decomposition in low-order forest streams: field and microcosm approaches. FEMS (Fed. Eur. Microbiol. Soc.) Microbiol. Ecol. 87 (1), 257e267. ssio, F., 2011. Can metal nanoparticles be a Pradhan, A., Seena, S., Pascoal, C., Ca threat to microbial decomposers of plant litter in streams? Microb. Ecol. 62 (1), 58e68. Roe, F.J.C., 1994. Styrene: toxicity studies d what do they show? Crit. Rev. Toxicol. 24 (Suppl. 1), s117es125. Rosemond, Z.B.A., Chou, S., Wilson, J., Schwartz, M., Tomei-Torres, F., Ingerman, L., Knaebel, D.B., Citra, M., 2010. Toxicological Profile for Styrene. Public Health Service, Agency for Toxic Substances and Disease Registry, U.S. Department of Health and Human Services. https://www.atsdr.cdc.gov/toxprofiles/tp53.pdf. ssio, F., Pascoal, C., 2017. Does the developmental stage and Seena, S., Carvalho, F., Ca composition of riparian forest stand affect ecosystem functioning in streams? Sci. Total Environ. 609, 1500e1511. Syranidou, E., Karkanorachaki, K., Amorotti, F., Franchini, M., Repouskou, E., Kaliva, M., Vamvakaki, M., Kolvenbach, B., Fava, F., Corvini, P.F.X., Kalogerakis, N., 2017. Biodegradation of weathered polystyrene films in seawater microcosms. Sci. Rep. 7, 17991. ~ oz, D., Brennholt, N., Bourrain, X., Buchinger, S., Wagner, M., Scherer, C., Alvarez-Mun Fries, E., Grosbois, C., Klasmeier, J., Marti, T., Rodriguez-Mozaz, S., Urbatzka, R., Vethaak, A.D., Winther-Nielsen, M., Reifferscheid, G., 2014. Microplastics in freshwater ecosystems: what we know and what we need to know. Environ. Sci. Eur. 26 (1), 12. Wegner, A., Besseling, E., Foekema, E.M., Kamermans, P., Koelmans, A.A., 2012. Effects of nanopolystyrene on the feeding behavior of the blue mussel (Mytilus edulis L.). Environ. Toxicol. Chem. 31 (11), 2490e2497.
Please cite this article as: Seena, S et al., Does nanosized plastic affect aquatic fungal litter decomposition?, Fungal Ecology, https://doi.org/ 10.1016/j.funeco.2019.02.011