Effects of Lyngbya majuscula blooms on the seagrass Halodule wrightii and resident invertebrates
Introduction
In coastal ecosystems, seagrass foundation species form habitat and stabilize local conditions (Dayton, 1972, Dawes et al., 1995, Bruno and Bertness, 2001, Duffy, 2006). They provide valuable ecosystem services, including acting as nursery habitat and refugia, providing a food source, recycling nutrients, and are highly productive (Dawes et al., 1995, Williams and Heck, 2001, Duarte, 2002, Heck et al., 2003, Duffy, 2006, van Tussenbroek et al., 2006). Worldwide seagrasses are declining due to a series of natural and anthropogenic stressors that have created poor conditions for seagrass success (Short and Wyllie-Echeverria, 1996, Orth et al., 2006, Waycott et al., 2009). It is estimated that 29% of the known areas of seagrasses have disappeared, and continue to decline at a rate of 7% per year (Waycott et al., 2009).
Anthropogenic pressures on coastal ecosystems are increasing worldwide (Halpern et al., 2008), which is one of the primary causes of seagrass losses (Orth et al., 2006, Waycott et al., 2009). Global increases in human populations have increased the frequency and magnitudes algal, phytoplankton, and cyanobacterial blooms, including harmful algae blooms (HABs) as a result of increased eutrophication (Hallegraeff, 1993, Anderson et al., 2002, Heisler et al., 2008). The occurrence and density of blooms can depend on the type of nutrients, availability of nutrients, density of grazers or suspension feeders, and local hydrodynamics (Cloern, 2001, Pittman and Pittman, 2005, Ahern et al., 2006, Ahern et al., 2007, O’Neil et al., 2012, Lapointe et al., 2015). HABs have produced deleterious effects on seagrasses by overgrowing and smothering of seagrasses that lead to reduced light availability and inhibited gas exchange, ultimately decreasing photosynthetic rates (Watkinson et al., 2005, O’Neil et al., 2012, Lapointe et al., 2015, Morris et al., 2015). Therefore, the severity of the HABs is strongly related to seagrass losses.
Lyngbya majuscula is a potential HAB-forming species in marine ecosystems that has been increasing in frequency since the early 2000s (Paul et al., 2005, Paul, 2008, O’Neil et al., 2012). It is a filamentous benthic cyanobacteria found growing attached to sediment, seagrasses, and corals (Albert et al., 2005, Paul et al., 2005, Arthur et al., 2006, Ahern et al., 2007, Martín-García et al., 2014, O’Neil et al., 2012). The extensive benthic mats can become detached and blanket large areas due to oxygen bubbles trapped in their filaments following active photosynthesis (Albert et al., 2005, O’Neil et al., 2012). Toxicity of secondary compounds produced by L. majuscula have been shown to have adverse effects on humans, with severe skin and eye irritations and hampered respiratory functions (Osborne et al., 2001, Watkinson et al., 2005, O’Neil et al., 2012). Effects of Lyngbya spp. on coral reefs are also primarily a function of toxicity rather than shading by the cyanobacterial mats (Titlaynov et al., 2007). In animal communities, L. majuscula produces compounds that act as strong feeding deterrents (Osborne et al., 2001, Paul et al., 2005, Paul, 2008). When L. majuscula grows epiphytically on seagrasses, animals, such as green sea turtles (Chelonia mydas), are negatively impacted by exposure to potentially harmful toxins when ingesting L. majuscula with seagrass leaves (Arthur et al., 2006, Arthur et al., 2008). Fish and meiofaunal communities (i.e. nematodes, copepods, polychaetes) abundances decreased in the presence of L. majuscula (Pittman and Pittman, 2005, García and Johnstone, 2006). Infaunal species also altered their sediment depth distribution, although this effect was less prominent in polychaete species (García and Johnstone, 2006).
Anthropogenic nutrient loading can cause algal blooms which in turn lead to dramatic decreases in seagrass abundance and distribution (Dennison et al., 1989, Hauxwell et al., 2001, Hauxwell et al., 2003, McGlathery, 2001, Lamote and Dunton, 2006, Lapointe et al., 2015, Morris et al., 2015). When algae grows in epiphytic associations with seagrass species, their presence can result in reduced light absorption, nutrient absorption, and gas exchange by seagrasses (Brush and Nixon, 2002, Drake et al., 2003, Brodersen et al., 2015, Costa et al., 2015). Bloom scale events have also lead to decreases in maximum depth limits, coverage, and stability (McGlathery, 2001, Lapointe et al., 2015, Breininger et al., 2016) and can exacerbate the negative effects of other stressors such as grazing (Maciá, 2000). Lyngbya majuscula blooms have caused declines in seagrass biomass, attributable to light reduction, although the extent of decline is highly variable between seagrass species (Watkinson et al., 2005). These reductions in light can be a key factor affecting seagrass distribution (Duarte, 1991, Ralph et al., 2007, Steward et al., 2005, Choice et al., 2014). Due to their high light requirement, seagrasses can be sensitive to shading from algal and cyanobacterial blooms (Duarte, 1991, Orth et al., 2006, Ralph et al., 2007). Duarte (1991) estimated that seagrasses had an average minimum light requirement of 11% of surface light. However, light requirements are highly variable among species due to morphological and physiological differences, ranging from 4.4 to 29.4% of surface light (Dennison et al., 1993). The physical presence of L. majuscula may create barriers that alter light availability and photosynthetic potential for seagrasses, particularly when attached to seagrass leaves (Arthur et al., 2006, Ahern et al., 2007, O’Neil et al., 2012, Fabbri et al., 2015). This light reduction impedes photosynthetic activity of seagrasses (Duarte, 1991, Ralph et al., 2007) resulting in reduced growth and biomass (Tomasko, 1992, Czerny and Dunton, 1995, Shafer, 1999, Biber et al., 2009).
The Indian River Lagoon (IRL) is diverse estuary in Florida that extends 251 km from Ponce Inlet to Jupiter Inlet (Gilbert and Clark, 1981, Zimmermann and Montgomery, 1984, Dawes et al., 1995, Steward et al., 2005). The IRL watershed has drastically changed due to rapid urbanization and population growth and creation of canals that increased the amount of freshwater input into the estuary (Kim et al., 2002, Duncan et al., 2004). As a result, eutrophication of this shallow estuary increased due to the presence of land-based sources of nutrients (Sigua et al., 2000, Sigua and Tweedale, 2003, Lapointe et al., 2015). Point-source sewage pollution was eliminated by 1996 as per the Indian River Lagoon Act of 1990 (sensu Lapointe et al., 2015) to protect the estuary. However, nutrient inputs remain, primarily sourced by storm water and non-point sewage pollution (Belanger et al., 2007, Lapointe et al., 2012, Lapointe et al., 2015). Residency time of water in the lagoon can be extensive, greater than one year, particularly in the northern IRL which has restricted water exchange (Smith, 1993). As a result recent evidence of nutrient accumulation in the IRL has been linked non-point sewage pollution, leading to increases in algal blooms (Phlips et al., 2002, Phlips et al., 2010, Phlips et al., 2011, Phlips et al., 2015, Lapointe et al., 2015). Blooms in the IRL are indicative of these eutrophication events (Benz et al., 1979, Virnstein and Carbonara, 1985) and have resulted in decreases in seagrass abundance and distribution as a result of reduced light availability in the IRL (Lapointe et al., 2015, Morris et al., 2015). Recently, approximately 45% of seagrasses in the northern IRL have been loss since a series of algal and phytoplankton blooms in 2011 and subsequent years (Morris et al., 2015).
One of the dominant seagrass species in the IRL is Halodule wrightii Ascherson, a pioneer seagrass species, found in shallow intertidal and sub-tidal waters (Dawes et al., 1995, Steward et al., 2005). As a foundation species, seagrasses like H. wrightii strongly influences community composition, including benthic infauna abundance (Orth et al., 1984, Coen and Heck, 1991, Omena and Creed, 2004, Johnson and Heck, 2007). Halodule wrightii is also considered to have relatively broad environmental tolerances to factors such as salinity (up to 65 PSU) and temperature (2 °C–39 °C), compared to other seagrass species (McMillan, 1979, McMillan, 1984, Björk et al., 1999, Koch et al., 2007, Mazzotti et al., 2007). As a result H. wrightii is known as a swift colonizer, particularly after stressors, compared to other species (De Oliveria et al., 1983). However, light requirements of H. wrightii are relatively high, at 20 ± 14% surface photosynthetic active radiation (PAR) (Steward et al., 2005). As a result, H. wrightii is known to have rapid responses to light attenuation as observed by changes in morphological traits, reductions in biomass and increased mortality (Tomasko, 1992, Czerny and Dunton, 1995, Shafer, 1999, Biber et al., 2009). In the IRL, H. wrightii abundance has been negatively impacted by series of algal, phytoplankton, and cyanobacterial blooms since 2011 and recovery has been slow (Morris et al., 2015). However, impacts of algal and cyanobacterial blooms on H. wrightii and its associate communities are still largely unknown.
The objective of this study was to test if L. majuscula affects H. wrightii communities and if those effects are a result of shading by L. majuscula mats on seagrass beds. This will provide insight into seagrass persistence in response to blooms. We hypothesize that (i) the presence of L. majuscula will reduce seagrass size and abundance, (ii) artificial shading will cause similar seagrass responses as the effect of L. majuscula presence and (iii) the abundance of macroinfaunal species (the bivalve, Macoma constricta, and two polychaetes Clymenella mucosa and Nereis succinea) will respond negatively L. majuscula presence and artificial shading. These infauna species are abundant bioturbators in H. wrightii meadows and are considered to have a random distribution (Coen and Heck, 1991, Omena and Creed, 2004).
Section snippets
Study area
In the summer of 2006 (April to August), widespread blooms of L. majuscula occurred in the central IRL and were observed from Fort Pierce Inlet (27°28′22.002N, 80°17′16.691W) to Vero Beach (27°35′13.176N, 80°21′50.068W), Florida (Capper and Paul, 2008; personal observations). Our field experiment was established at Harbor Branch Oceanographic Institute (HBOI), located in Fort Pierce, FL (27°31′57.87N, 80°20′53.673W). This area is part of the central IRL, approximately 10 km from the Fort Pierce
Lyngbya majuscula biomass
In June 2006, Lyngbya majuscula average dry biomass in non-shaded plots was 105.53 ± 11.62 g m−2 (Mean ± SE) and in shaded plots was 78.81 ± 14.09 g m−2 (Mean ± SE) before initiation of shading or removal treatments. By the end of the experiment (August 2006), L. majuscula average biomass in non-shaded plots was 17.65 ± 26.17 g m−2 (Mean ± SE) and in shaded plots was 15.77 ± 14.50 g m−2 (Mean ± SE), indicating a decline in the bloom over time. By September of 2006 the bloom had run its course (Capper and Paul, 2008,
Discussion
Harmful cyanobacterial and algal blooms can result in phase shifts in foundation species when seagrasses are replaced by algal-dominated communities (Anderson et al., 2002, Duarte, 2002, Heisler et al., 2008). Our field experiment suggests that HABs can produce complex responses in seagrass communities when coupled with secondary stressors like that produced by artificial shading. We observed changes in the relative amounts of below ground biomass and changes in leaf length over very small
Conclusion
Seagrass communities throughout Florida undergo a barrage of stresses from natural causes such as species interactions or human induced environmental changes such as large-scale releases of freshwater into the IRL. Understanding seagrasses responses to changes in nutrients and light penetration, especially during recovery after a disturbance period will provide valuable information in the recovery and restoration of seagrasses in Florida estuaries. The compounding effects of shading by harmful
Acknowledgements
We thank D. Devlin for helping with experimental design, field work, identification of polychaetes, and for substantive comments on earlier versions of the manuscript. J. Moore provided useful comments on early stages of this paper. We appreciate field assistance by G. Coldren. D. Hanisak provided the LI-COR used for PAR measurements.[CG]
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Present Address: Department of Life Sciences, Texas A&M – Corpus Christi, 6300 Ocean Dr., Corpus Christi, Texas, 78412, USA.