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Environmental boundaries
The niche of a new species is defined by several factors. Not only must the habitat fit the species regarding to physical attributes (substrate type and location, salinity, temperature, food availability), but the population must be able to successfully compete with local species for these resources and avoid predation. Green mussel populations are primarily found in habitats ranging from oceanic and high salinity estuarine waters favoring areas with high phytoplankton and / organic matter (Morton, 1987; Vakily, 1989; Wong and Cheung, 2001) and high current flow / flushing aiding in the removal of waste and continuous food supply (Rajagopal et al., 1998b; Buddo et al., 2003; Rajagopal et al., 2006). Green mussels commonly dominate biofouling communities on and near power plants where water temperatures remain high and flushing is extensive (Rajagopal et al, 1991a,b). High density settlement has caused damage to power plants in their native range due to extensive fouling of water intake pipes (Rajagopal et al, 1991a) and the water cooling systems of power plants in Tampa Bay are believed to be the first point of invasion creating a broodstock population allowing for the further spread throughout the bay (Benson et al., 2001).
As an important aquaculture species, much attention has been given to production and environmental tolerances within their native range. In both native an invaded regions they are aggressive biofoulers and have been found coating hard substrate in densities as high as 4,000 – 35,000 individuals m-2 (Haung et al., 1983; Baker and Benson, 2002; Fajans and Baker, 2005) and clogging water intake pipes at biomasses exceeding 200 g m-2 (Rajagopal et al., 1991a). As a gregarious species, high growth rates and fecundity have allowed for the rapid colonization of new habitat and decreased native fauna through resource competition. Thriving in a range of tropical to subtropical waters, conditions frequently allow for year round gametogenesis and high food availability in productive bays and estuaries allows for continuous input to energy reserves and ample food supply for developing larvae.
Perna viridis populations are most commonly found on hard substrate including mangrove prop roots and submerged rock and shell, and other bivalves including oyster reefs, but also frequent artificial substrate such as pilings, piers, and floating objects such as buoys (Vakily, 1989; Buddo et al., 2003; Baker et al., 2007; 2012). Locally, Estero Bay comprises of soft bottom sediments with little hard substrate available leaving intertidal oyster reefs and artificial structure as the only available substrate. However, green mussel populations have been observed on the soft bottom sediments within seagrass beds in Tampa Bay and Hillsborough Bay Florida covering areas as large as 2300 m2 (Johansson and Avery, 2004) and in Kingston Harbor Jamaica (Buddo et al., 2003) often attaching to shell fragments, filamentous root structure, and / or other bivalves (Ingrao et al., 2001).
Although often referred to as an intertidal species (Shafee, 1978; Vakily, 1989) P. viridis is primarily found on subtidal substrate or areas with minimal emersion time (Tan, 1975b; Rajagopal et al., 1998b). Juvenile recruitment is often observed in the intertidal regions, however these individuals rarely survive to maturity (Baker et al., 2012; personal observation) and dense populations are primarily found subtidally (Nair and Appukuttan 2003) with the highest growth observed at depths below the mean low tide mark (Sivalingam, 1977). Tan (1975b) found the highest population densities at depths of 1.5 – 11.7 m below the high water spring tide mark and no living mussels were observed above the high tide mark. This distribution is likely due to desiccation stress from aerial exposure in which normal physiological requirements, (feeding, exchange of gases, and removal of waste) are inhibited. Green mussels found in the intertidal zone have been shown to be very sensitive to even short term extremes in air temperature (Urian et al., 2010; Baker et al., 2012; McFarland et al., 2014). Both field and lab observations suggest that P. viridis cannot tolerate aerial exposure at temperatures below 14˚C (Baker et al. 2012; Firth et al., 2011; Power et al., 2004) and Urian et al. (2010) observed 100% mortality at 3 ˚C within the first 24 hours of exposure. Likewise green mussels have shown an inability to cope with aerial exposure under high temperature stress with mortality rates (≥97%) at air temperatures as low as 25 ˚C in which surface temperature often exceed 30 ˚C when under direct sunlight (McFarland et al., 2014).
As a tropical species, P. viridis thrives in warm waters with an optimal range of 26 – 32 ˚C and lower and upper limits showing 50% survival at 10˚C and 35˚C, respectfully (Silvalingam, 1977). Populations in invaded regions, however, may adapt to changing environments and develop localized tolerances. Segnini de Bravo et al. (1998) reported lethal lower and upper temperatures of 6 ˚C and 37.5 ˚C in the invaded region of Venezuela while Ueda et al. (2013) reported a lower tolerance of 12 ˚C in Japan. Locally, southwest Florida populations have an observed tolerance of 13- 30 ˚C determined through both laboratory exposures and field observations (Baker et al., 2012; McFarland et al., 2014). Likewise, populations on the Atlantic coast of Florida do not survive extreme winter water temperatures (Power et al., 2004; Spinuzzi et al., 2013) experiencing 93.8% mortality after 30 days at 14 ˚C and 100% mortality at 10 ˚C within 13 days (Urian et al., 2010). Likewise an intolerance to high temperature extremes is reported. Nicholson (2002) observed 30% and 70% mortality at 31 ˚C and 34 ˚C, respectively. Similarly, McFarland et al. (2014) observed 100% mortality within 12 days at 35 ˚C, indicating reduced tolerance with rising temperatures. Reduced filtration rates, byssal production and gonad development have been observed at temperatures ≥ 35 ˚C (Rajagopal et al., 1995; Sreedevi et al., 2014).
However, wide ranges reported from short term laboratory studies may overstate the boundaries for long term growth and survival. Urian et al. (2010) found a significant increase in heat shock proteins (Hsp70) after only 2 hours of exposure to 10 ˚C suggesting that stress from short term exposures may have prolonged metabolic effects especially if other stressors are involved or repeat exposures occur. The various ranges reported in the literature are likely due to methodologies including the rate of change and exposure duration. Temperature range is also dependent upon salinity in which low salinity conditions reduced the temperature range at which P. viridis can survive (Rajagopal et al., 1995; unpublished data cited by Spinuzzi et al., 2013).
A wide range of salinity tolerances have been reported for P. viridis throughout both native and invaded regions. The optimal salinity in its native range has been reported as 27 – 33 ppt with a tolerance of 19 – 44 ppt and is reflected accordingly in population distribution within bays and estuaries (Silvalingam, 1977; Sundaram and Shafee, 1973; Huang et al., 1983; Vakily, 1989), however, wider tolerances have been reported through laboratory studies. Silvalingam (1977) observed 50% survival at salinities of 24 and 80 ppt and Sengini de Bravo et al. (1998) determined a range from 0 – 64 ppt by decreasing the salinity by 1 ppt / day in the laboratory. These were however, short term exposures and physiological changes such as increased valve closure, inability to osmoregulate, and reduced clearance rates suggest long term survival at these extremes is unlikely (McFarland et al., 2013). In Tampa Bay, established populations are found in areas which remain above 14 ppt and the highest densities are found in regions where the salinity ranges from 20 – 28 ppt (Baker et al., 2012). Through both acute and gradual salinity decreases, southwest Florida populations have an observed lower threshold of 12 ppt (McFarland et al., 2014), however at salinities of 15 ppt and below osmotic stress may reduce long term survival (McFarland et al., 2013). Likewise, upper salinity tolerance may be more limited in natural populations versus laboratory experimentation. Buddo et al. (2003) found dense populations in regions with average salinities of 27.7 ppt, but no mussels were found near salt ponds where salinities averaged 33.8 ppt.
Harmful algal blooms may impose additional environmental constraints. Blooms of the toxic dinoflagellate, K. brevis, are a common occurrence in southwest Florida and the response of P. viridis to these events is unknown. Karenia brevis produces a suite of potent neurotoxins known as brevetoxins (PbTx) and blooms are commonly accompanied by massive fish kills (Ray & Wilson 1957; McFarren et al. 1965; Naar et al. 2007) and in some cases marine mammal, sea turtle and seabird strandings and mortalities (Adams et al. 1968; Forrester et al. 1977; Flewelling et al., 2005). While local bivalves are tolerant to these bloom events, they accumulate PbTx in their soft tissue body parts posing a threat to predators through trophic transfer and human consumption (Plakas et al., 2002; 2004; Pierce and Henry, 2008; Rolton et al., 2014). Locally, clams Mercenaria mercenaria and oysters C. virginica are monitored routinely during these events as a human health precaution. PbTx levels ≥800 ng g-1 PbTx-3 equivalent in shellfish soft tissue and / or K. brevis cell counts ≥5,000 cells L-1 result in mandatory closure of shellfish harvesting (Tester & Steidinger, 1997; Steidinger, 2009; Plakas et al., 2008). Clams and oysters have been shown to depurate the toxin within 2 – 8 weeks post bloom when tissue concentrations ranged from 1,500 – 28,000 ng g-1 PbTx-3 equivalent (Morton and Burklew, 1969; Steidinger and Ingle, 1972; Plakas et al., 2002; Plakas et al., 2004; Plakas et al., 2008; Bricelj et al., 2012; Griffith et al., 2013), however PbTx accumulation and depuration rates are unknown for P. viridis. Previous work addressing the effects of toxic dinoflagellates on bivalves has shown mussel species to accumulate high toxin levels and increased sensitivity to HAB exposure when compared with species of oysters and clams (Ingham et al., 1986; Shumway and Cucci, 1987; Shumway et al., 1988, 1990, 1995; Bricelj and Shumway, 1998; Lesser and Shumway, 1993).
Through observations of established populations of P. viridis in Tampa Bay, Baker et al.
(2012) reported ˃90% mortality in regions of the bay exposed to a natural K. brevis bloom, while populations in portions further removed from the bloom appeared unaffected. However, this mortality event was followed by a rapid repopulation of juvenile green mussels in the following year post bloom dissipation (Leverone, 2007) Although a population rebound was observed, green mussels have yet to rebound and consistently maintain the high densities observed in Tampa Bay upon their arrival (Dr. S. Baker, personal communication).
In their native range, P. viridis has shown tolerance to several HAB species including Karenia mikimotoi (Robin et al., 2013) and Gymnodinium nagasakiense (Karunasagar and Karunasaga, 1992), but has resulted in high mortality rates during exposure to Alexandrium monilatum (Hégaret et al., 2008) and both increased mortality and prolonged harvesting bans due to accumulated tissue toxins were observed during and following a natural Pyrodinium bahamense var. compressa bloom in the Philippines (Gacutan et al., 1984). Cheung et al. (1993) documented high spring mortalities affecting 30 and 70% of adult P. viridis in Hong Kong during May – September of 1987 and 1988, respectively and, though not directly attributed, these mortality events occurred during periods of high red tide frequency. Juveniles from cohorts spawned during this period (June – July) had low survival. Similar mortality events have been documented in other bivalve species sensitive to toxic algae blooms. Tracey (1988) monitored M. edulis populations during and after a brown tide bloom caused by Aureococcus anophagefferens in the summer of 1985 which caused nearly 100% mortality and inhibited juvenile recruitment. Following this event, recruitment did not resume until the fall of 1986 at which time only sporadic spat were observed. This reduction in larval recruitment was attributed to the severe decline in density of spawning adults and poor quality of eggs and larvae produced due to stress and reduced feeding in the adults during the bloom (Tracey, 1988). While an intolerance to K. brevis in south Florida populations is a positive for limiting population spread, it may pose trophic transfer and human health concerns. Perna viridis is not harvested for human consumption, however illegal harvesting occurs and several species of fish, crab and seabirds may be at risk for toxin accumulation if mussels harbor the toxins for periods of time following exposure without succumbing to death. Tolerance of P. viridis to K. brevis blooms is currently unclear and the extent to which they accumulate associated toxins in their tissue is unknown.
Problems with invasive species: Biofouling and potential harm to oysters and ecosystem disruption
Introducing non-native species to new environments is a rising problem world-wide. Increased boat traffic and international shipping have increased vectors for spread over wide geographic distances and increased urbanization and tourism in coastal areas has created new substrate for species including navigational structures, piers, and break walls (Bax et al., 2003). While Florida coastal waters already have a significant biofouling problem, consisting mainly of barnacles, tunicates and other native bivalves, none reach the same sizes and densities observed in P. viridis populations (Benson et al., 2001). Increased biofouling communities lead to increased costs for removal and can halt industrial production in water cooling systems. Pipes and pumps may become clogged leaving a reduced flow causing pumps to overheat and burn out, thereby increasing cost of both removal of the species and replacement of damaged infrastructure. Rajagopal et al. (1997) found dense populations of P. viridis coating water cooling systems of power plants in densities reaching 211 kg m-2 causing damage to equipment and clogging of pipes reducing water flow for the cooling system. Clogging is caused not only by dense settlement of mussels but also from byssal mats which can get sucked into the pumps and block protecting screens (Ingrao et al., 2001). While green mussels are only contributing to an existing biofouling problem not creating one, their fast growth and reproduction may increase this problem and frequency of treatment.
From an ecosystem perspective, invasive species may cause alterations in ecosystem services including nutrient cycling; habitat modification (or loss); competition and displacement of native species, which may include aquaculture facilities and fisheries; act as vectors for disease, viruses, parasites, bacteria, and harmful algae (Bax et al., 2003; Hégaret et al., 2008).
Bivalves may transport and introduce parasites or diseases from their native range not normally found in the new ecosystem or may be immune to parasites in the new ecosystem leaving themselves free of parasites giving them a leg up on the competition (Branch and Steffani, 2004). Competition for resources, such as food and settling space, can lead to displacement and reduction in numbers of native species (Dulvy et al., 2003; Karatayev et al., 2007). As ecosystem engineers, bivalve invasions have been shown to drastically alter ecosystems. For example the zebra mussel Dreissena polymorphia in the Great Lakes resulted in alterations of the phytoplankton community, reduction of native species and an estimated cost of $1 billion per year in biofouling damage and control (Pimentel et al., 2005; Karatayev et al., 2007). Similarly, the Mediterranean mussel, Mytilys galloprovincialis reduced native bivalves and limpets by out competing for space on the coast of South Africa (Branch and Steffani, 2004) and P. viridis invasion in Venezuela has resulted in heavy competition and a population decrease of the brown mussel Perna Perna due to substrate competition (Segnini de Bravo et al., 1998; Rylander et al., 1996). In Tampa Bay Florida, dead oyster shell was found under green mussel populations on bridge pilings and the subtidal, outer crest of a reef system (Baker et al., 2007).
Through rapid growth and vigorous reproductive activity, P. viridis is an aggressive marine invader rapidly reaching high densities allowing for heavy competition with other bivalves for hard substrate. The primary concern surrounding green mussel population spread in the southeastern United States is competition with the native oyster Crassostrea virginica (Fig. 3). Oysters are a keystone species in the soft bottom bays of south Florida where they form permanent three-dimensional habitat which creates a refuge and / or foraging grounds for over 300 species of fish and crab (Wells, 1961) including several recreational and commercially important species (Henderson and O’Neil, 2003; Tolley et al., 2005). Many estuarine species spend their entire lives on the oyster reefs while others utilize the intricate reef system as a safe hiding place to lay their eggs and nursery for juveniles (Tolley and Volety, 2005). Many offshore fish and crab species enter the estuary for breeding where their young use oyster reefs as a refuge (Beck et al., 2011). This refuge for early life stages and small organisms creates an extensive feeding grounds for larger species, both estuarine and oceanic (Peterson et al., 2003; Grabowski et al., 2005). Peterson et al. (2003) found that oyster reefs enhanced growth and increased the abundance of 19 species of fish and large mobile crustaceans, with an estimated contribution to fish production of 2.57 kg 10m-2 yr-1.
Table of contents :
Chapter1: General Introduction
History, origin and transport of green mussels
Study site: Estero Bay, Florida, USA
Biology of P. viridis
Environmental boundaries
Problems with invasive species: Biofouling and potential harm to oysters and ecosystem disruption
Understanding population dynamics through applying the Dynamic Energy Budget Theory to predict and prevent spread
Chapter 2: Uptake and elimination of brevetoxin in the invasive green mussel, Perna viridis, during natural Karenia brevis blooms in southwest Florida
Chapter 3: Potential impacts on growth, survival and juvenile recruitment of the green mussel Perna viridis during blooms of the toxic dinoflagellate Karenia brevis in southwest Florida
Chapter 4: Seasonal variation in gametogenesis and energy storage of the invasive green mussel, Perna viridis, in southwest Florida
Chapter 5: Application of the Dynamic Energy Budget theory to model growth and reproduction of established population of the non-native green mussel Perna viridis in southwest Florida coastal waters
Chapter 6: General Discussion
