Uptake and elimination of brevetoxin in the invasive green mussel, Perna viridis, during natural Karenia brevis blooms in southwest Florida

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History, origin and transport of green mussels

The green mussel, Perna viridis, is native to the Indo Pacific and is found abundantly in coastal waters of India, Malaysia, Papua New Guinea, Indonesia, China, Japan, and the Philippines (Appukuttan, 1977; Sivalingam, 1977; Siddall, 1980; Haung et al., 1983; Vakily, 1989); reviewed by Baker et al. (2007). High protein content and rapid growth rates, reaching marketable size within 6 – 8 months, make P. viridis an ideal aquaculture species (Sivalingam 1977; Vakily 1989) serving a major food source in these regions, harvested through various aquaculture operations and from wild populations. It is however, also a pesky biofouling organism and if left untreated can cause serious damage to infrastructure including the clogging of intake pipes used in various forms of industry and water cooling systems (Rajagopal et al. 1991a,b). Green mussels frequently occur in densities of 1000 – 4000 individuals m-2 (Fajans and Baker 2005) and as high as 35,000 individuals m2 (Huang et al., 1983). This coupled with high growth rates, can result in serious damage to infrastructure and high costs for removal.
Perna viridis made it first appearance in Caribbean waters in the early 1990’s along the coast of Trinidad and Tobago (Agard et al., 1992) and has since spread throughout the Caribbean. In 1993 they were recorded along the coast of Venezuela (Rylander et al., 1996; Segnini de Bravo, 1998) and in 1998 populations were observed in Kingston Harbour, Jamaica (Buddo et al., 2003). The first occurrence of P. viridis in the United States was observed in 1999 in the water intake pipes of the Tampa Electric Company Power Stations (Gannon and Big Bend) in Tampa Bay, Florida with an estimated initial invasion in early fall of 1998 based on size of individuals collected (Benson et al., 2001; Ingrao et al., 2001). Green mussels have since spread throughout southwest Florida coastal waters via a free swimming larval stage and boat traffic (Baker et al., 2007). Dense populations were recorded in the Estero Bay / Fort Myers, Florida area as early as 2002 and sporadic siting’s as far south as Marco Island, Florida in 2003 (Baker et al., 2007), approximately 240 km south of Tampa Bay, indicating an ability for rapid spread and colonization across long distances. Northern spread on the Gulf coast has not been observed likely due to a combination of colder waters and primarily southern flowing currents favoring a southern distribution. A separate Florida invasion has been documented on the Atlantic coast of Florida in St. Augustine in 2002 in which green mussel populations have spread south to West Palm Beach and as far north as Charleston, South Carolina (Rajagopal et al. 2006; Baker et al. 2007). Northern limits on the Atlantic coast may be more extensive for subtidal offshore reefs due to northern flowing currents and proximity to the Gulf Stream and its prevailing warmer waters (Urian et al. 2010). More recently green mussel populations have been reported throughout Cuba’s coastal waters in 2005 (Fernández-Garcés and Rolán 2005), but suffered a mass mortaltiy event in 2012 resulting in reduced densities possibly due in part to high frequency harvesting and low juvenile recruitment (Lopeztegui-Castillo et al., 2014).
Through genetic testing of green mussels from both native and invaded regions, Gilg et al. (2012) suggests a single invasion to the Caribbean, likely from ships traveling from China to Trinidad and Tobago, and a successive stepping stone spreading effect throughout the Caribbean and southeastern United States. Larval transport via ballast water and / or adults colonizing boat hulls are both potential vectors of transport (Benson et al., 2001; Ingaro et al., 2001; Buddo et al., 2003; Rajagopal et al., 2006; Baker et al., 2007). Boat hulls provide hard substrate for settlement and with rapid growth, green mussels reach reproductive maturity within 3 – 4 months (Rao et al., 1975; Parulekar et al., 1982; Walter, 1982) boats traveling from one port to another are likely to carry reproductively active populations providing a brood stock population when they arrive to a new harbor. With a pelagic larvae larval stage lasting 18 – 28 days (Tan, 1975a; Sreenivasan et al., 1988) long distance dispersion via ballast water and / or ocean currents is possible. This allows for accelerated spread of new populations as boat traffic moves along the coasts, between port cities, and across ocean basins. With a history of rapid spread and colonization, concerns abound regarding invasion potential in other US coastal regions and remains to be elucidated.

Study site: Estero Bay, Florida, US

Classified as an aquatic preserve in 1966, Estero Bay spans 11,000 acres of protected submerged land, creating critical habitat for many commercially and ecologically important species (Florida Department of Environmental Protection, 2015). The bay is protected from the Gulf of Mexico (GOM) by barrier islands and is densely populated by mangrove islands, oyster beds and mud flats (Byrne and Gabaldon, 2008). Six inlets allow for a strong tidal influence and extensive flushing between the estuary and GOM with six major freshwater tributaries from the mainland, resulting in a brackish mixing (Byrne and Gabaldon, 2008). While Estero Bay has a protected status, several of its major tributaries run through urban areas leading to anthropogenic input due to increased development along the rivers amplifying concerns of nutrient loading (Tolley et al., 2006).
Southwest Florida has a unique landscape, dramatically altered by anthropogenic forces and coastal urbanization. Through development, water management, and dredging activities, drainage patterns from the Florida Everglades have been altered from slow meandering streams and creeks to large, heavy flow rivers (Volety et al., 2009). Such hydrological changes have created inconsistent environmental conditions, especially when coupled with the two very dramatic seasons southwest Florida experiences. In the winter-dry season there is little to no rainfall, and the average salinity in the estuary may become hypersaline ranging from 28-38 ppt while in the summer-wet season, high rainfall and large freshwater flushes result in salinities as low as 0-10 ppt (Barnes et al., 2007; Volety et al., 2009). Many of southwest Florida’s estuaries are shallow, allowing for rapid changes in salinity which create an acute and sometimes prolonged exposure to depressed salinities. Besides increased fresh water input, these drainage canals and rivers now flow through areas of increased urban development and agricultural land, accumulating nutrients from fertilizers, pesticides, and increased loads of suspended solids (Volety et al., 2009; Abeels et al., 2012). This high input of nutrients allows for a productive system, however when in excess can initiate and support phytoplankton blooms which may cause sags in dissolved oxygen resulting in benthic mortalities due to anoxic conditions and / or fuel toxic algae blooms, such as Karenia brevis (Heisler et al., 2008), which are typically accompanied by a high incidence of marine life mortality and human health concerns.
Estero Bay is a shallow, mud bottom estuary with an average depth of 1 m (Byrne and Gabaldon, 2008) leaving limited natural substrate (mangrove roots and oyster reefs) for biofouling organisms and thus, primarily only artificial substrate (boat hulls, docks, intake pipes, navigational structures, seawalls) available, primarily located in dredged channels, marinas, and bridge passes. The shallow muddy conditions in Estero Bay result in turbid waters when wind or current flow agitates the water column leaving subtital regions hidden from view. To monitor the establishment and spread of invasive species exploratory divers must inspect these regions, which can be costly and the lack of such monitoring is the primary reason P. viridis has successfully spread to densely populate new regions and go unseen until they have become fully established.

Biology of Perna viridis

Perna viridis is a dioecious species with rare cases of hermaphrodites and while sexual organs are not distinguishable, sex can be determined by tissue color (Lee, 1988). The gonad is dispersed throughout the mantle lobes, mesosoma and intermixed between digestive glands (Rajagopal, 2006) leaving males appearing creamy white in color and females orange or brick red in appearance (Walter, 1982; Sreenivasan et al., 1989; Lee, 1986; Narasimham, 1981). During resting periods of gametogenesis this coloration is not as distinct leading to misidentification without histological confirmation, however in tropical to subtropical regions, reproductive activity frequently persists year round (Parulekar et al., 1982; Walter, 1982). Perna viridis are broadcast spawners, releasing egg and sperm into the water column for external fertilization and may be cued not only by environmental factors (temperature, salinity, food), but also from chemical cues (Widdows, 1991; Barber and Blake, 2006) which can be initiated by either sex (Stephen and Shetty, 1981). Larval densities have been recorded from approximately 20,000 larvae m-3 (Rajagopal et al., 1998a) to as high as 40,000 larvae m-3 (Rajagopal et al., 1998b), similar to that of Mytilus edulis, its cold water relative (Schram, 1970).
The life cycle begins with a pelagic larval phase reported to last from as short as 8 – 15 days (Tan, 1975a; Nair and Appukuttan, 2003) to 15-18 days (Sreenivasan et al., 1988; Nair and Appukutan, 2003) and as long as 24-35 days (Nair and Appukuttan, 2003; Laxmilatha et al., 2011) (Table 1). This wide range is due to variations in environmental conditions and substrate availability, which play a critical role in larval development and settlement (Sreenivasan et al., 1988; Nair and Appukutan, 2003). Marine mussels can delay metamorphosis for several weeks if suitable substrate is not encountered, however this delay results in decreased survival due to the interruption of feeding and growth leading to a prolonged pelagic phase accompanied by increased predation risks (Bayne, 1965; Widdows, 1991).
During the planktonic stage they are at the bottom of the food chain leading to increased predation pressure from plankton grazers and susceptibility to external stressors is at its high (Bayne, 1976). Food and temperature are the two of most important factors affecting growth and development of bivalve larvae (Widdows, 1991), however any environmental altercation outside the threshold will effect growth, development and survival, including temperature, salinity, pH, dissolved oxygen, food availability, HAB’s pollutants, etc. (Nair and Appukuttan, 2003; Bayne, 1965; Gilg et al., 2014). Nair and Appukutan (2003) observed significant difference in larval growth reared under different temperatures with a more narrow tolerance range than that of post metamorphosed mussels. Decreased embryogenesis, larval development and survival have been observed during exposure of local species of oysters, clams and scallops to K. brevis (Leverone et al., 2006; Rolton et al., 2014).
For the first several days post hatching, early larval stages depend on lipids from the egg as the only source of energy for development before feeding is possible (Bayne et al., 1975; Holland, 1978). Stress in adults during egg production can lead to reduced fecundity and success of progeny due to decreased lipid content and viability in eggs (Bayne, 1972; Bayne et al., 1975; 1978). Thus, larval stages are not only affected by environmental conditions in which they are living, but also the physiological state of the broodstock population.
While larvae are free swimming, mobility is minimal and distribution occurs primarily through current and tidal flow. Depending on the local hydrology, mussel larvae may be transported great distances with currents or may be confined to a more concentrated area (Porri et al., 2006; Gilg et al., 2014). Mussel larvae can secrete specialized byssal threads to aid in floatation and transport, different from those secreted for attachment, within 10 – 12 days of hatching (Siddall, 1980). These long byssal or mucous threads can allow for post-metamorphosis migration via floatation when void of suitable substrate (Widdows, 1991; Rajagopal et al., 2006). Primary settlement onto filamentous drift algae followed by secondary settlement to an established mussel bed is common in Mytilid species (Widdows, 1991; Buchanan and Babcock, 1997; Alfaro et al., 2004). Primary settlement substrate, which commonly includes filamentous drift algae, allows for a prolonged searching period when suitable substrate is not available and transport over greater distances (de Vooys, 1999; Cáceres-Martínez et al., 1993; Buchanan and Babcock, 1997). Following primary settlement, mussels can to move from one substrate to another through the use of the foot (Bayne, 1964). This crawling stage can last for several weeks (Dare and Davis, 1975); 5 – 6 weeks for M. edulis (Bayne, 1965) and early settled spat may attach and detach themselves several times before settling in a permanent position (Bayne, 1964, 1965; Seed, 1969; Tan, 1975). A prolonged searching phase allows for more selective behavior for substrate and / or environmental conditions prior to settlement and has been observed previously in several species of bivalve larvae (Muus, 1973).
It is not fully understood what induces the larvae to “choose” a substrate and settle, however, as a gregarious species, chemical cues from adults are believed to play a role and increased recruitment is often observed on or near established adult populations (Bayne, 1964; Widdows, 1991; de Vooys et al., 2003; Baker et al., 2006;). Preferential use of filamentous algae for primary settlement also suggests that byssal threads of adult mussels may attract larvae ready to settle (Eyster and Pechenik, 1987; deVooys et al., 1999; Cáceres-Martínez et al., 1993). By settling within clumps of established adults, new juveniles may find refuge from predation and the presence of adult populations may indicate suitable habitat, increasing survival and proximity giving a reproductive advantage as broadcast spawners (Alfaro, 1994). Once settled, mussels maintain a permanent position through the production of byssal threads allowing them to remain securely fastened to hard substrate and are frequently intertwined with each other forming tight clumps or mats. The byssal gland is in the foot and allows for repositioning of the individual when necessary (Banu et al., 1979). Byssal production is a continuous process allowing for the replacement of broken or damaged threads and have shown increased production and thread thickness as a defense mechanism under predation pressure and / or the presence of damaged conspecifics (Young, 1985; Leonard et al., 1999; Cheung et al., 2004).
Once settled P. viridis has shown rapid growth rates averaging 7 – 10 mm month-1 (Qasim et al., 1977; Lee, 1986; Haung et al., 1983; Sreenvivasan et al., 1989) and as high as 12 – 15 mm month-1 (Rajagopal et al., 1998; Hawkins et al., 1998). Kuriakose and Appukuttan (1980) recorded growth of 15.4 mm in the first month of settlement followed by rates of 11.2 – 13.6 mm in the following four months. These rates are similar to the Mediterranean mussel Mytilus galloprovincialis which exhibits growth rates 6.3 – 11 mm month-1 in NW Spain (Camacho et al., 1995). However, growth of P. viridis far exceeds its cold water relative, M. edulis, which has reported growth of 5 cm in 5 years in the cold waters of Plymouth, England (Bayne and Worrall, 1980) and from 11 – 24 months to reach market size depending on the geographical location (Hawkins et al., 1999).
Rapid growth in sessile organisms serves several purposes particularly within the first year of settlement. As smaller mussels will be most vulnerable to predation pressures, fast growth increases the ability to avoid predation. This also allows individuals to quickly reach maturity and reproduce in a short amount of time allowing for rapid colonization of new habitat. Both high growth and reproductive rates may also allow for reduction in competition for food and substrate from similar species allowing for their success as an invasive species.
Growth rate increases with increasing temperature, within the animals threshold and provided food availability is sufficient, thus tropical species typically have prolonged growth periods ( Rajagopal et al., 1998a; Lodeiros and Himmelman, 2000). Temperature and food availability are often interrelated leaving predictions difficult due to the complexity of environmental conditions. Environmental factors play role in eliciting physiological responses and if one or more is above or below the threshold, growth will be hindered. A decrease in growth rate is often observed in the winter months when temperature and food are reduced (Narashiham 1981; Lee 1986; Richardson et al., 1990; Cheung et al. 1991; 1993). For P. viridis, Cheung (1991) suggested lower threshold for growth of 20 ˚C and a complete cessation in growth has been documented at 17˚C and below (Lee, 1986). Further, Lee (1985) observed that while growth may occur at lower temperatures, highest growth occurs at temperatures of 24 – 29 ˚C and in regions where temperature remains above this threshold decreases in growth can be attributed to a decrease in food availability.
In areas where food and temperature remain high (ie: power plant facilities), P. viridis has been observed to reach 119 mm in the first year, far exceeding the typical values recorded throughout its geographic distribution (Rajagopal et al., 1998b) which typically closer to90 mm in the first year (Roa et al., 1975; Narasimham, 1980; Rivonker et al.,1993). Records of maximal length vary widely with reported values in invaded regions ranging from 161 mm in Trinity Harbour, Austratlia (Stafford et al., 2007) to 172 mm in Tampa Bay, Florida (Baker et al., 2012) and previous work in native regions reporting values from 156 – 230 mm (Roa et al., 1975; Appukuttan, 1977; Narasimham 1981; Cheung, 1990).
High fecundity has been reported in both native and invaded ranges. Many studies within tropical regions report year round gametogenesis with one or two major peaks in spawning that could be linked with environmental variables (Roa et al., 1975; Walter, 1982; Fatima et al.,
1985). Even in more temperate regions where mussels undergo more distinct cycles, highly productive spawning seasons are observed with two major peaks and intermittent spawning throughout the season (Lee, 1986; Rajagopal et al., 1998a, b; Barber et al., 2005). Perna viridis is capable of reaching reproductive maturity within the first 2 – 3 months of settlement at lengths of 15 – 20 mm (Parulekar et al., 1982; Roa et al., 1975; Sreenivasan et al., 1989; Siddall, 1980), potentially allowing for cohorts to spawn within the first year. Populations in Tampa Bay, Florida were found to have a long and productive spawning season, but experienced a resting phase in December and January (Barber et al., 2005) when local water temperatures drop as low as 13 ˚C (Badylak et al., 2007). However, on the Atlantic coast in St. Augustine, Florida where observed winter lows averaged 16 ˚C, P. viridis was observed to maintain high reproductive activity year round with no observed resting stage (Urian, 2009).
High growth and reproductive activity comes at a high energetic cost. The ability to maintain year round gametogenesis and allocation of energy towards growth requires that the individual receives enough energy (food) to support these processes. Green mussels populate eutrophic estuaries and bays with high nutrient availability (Vakily, 1989; Wong and Cheung, 2001) and have exhibited high clearance rates when compared to other bivalves (Hawkins, 1998; McFarland et al., 2013). This high efficiency feeding is needed to supply energy required to fuel the high metabolic demand of rapid growth and year round gametogenesis. Fast growing bivalves have been shown to have high feeding rates and metabolic efficiency (Bayne, 2000). Perna viridis has been reported to have clearance rates ranging from 2.62 – 4.21 L h-1g-1 (Wang et al., 2005) to as high as 8.06 – 9.68 L h-1g-1 (Blackmore and Wang, 2003) with a maximum clearance rate recorded at 15 L g-1h (Hawkins et al., 1998). In the invaded region of southwest Florida P. viridis has been shown to exhibit clearance rates 2 – 3 times greater than the native oyster C. virginica (0.08 – 1.2 and 0.09 – 0.43 L h-1g-1, respectively) at salinities within its optimal range and maintained rates similar to that of the oyster at depressed salinities as low as 10 ppt (McFarland et al., 2013). Hawkins et al. (1998) compared the clearance rate of P. viridis and four other oyster species (Crassostrea belcheri, Crassostrea iradelei, Saccostrea cucculata and Pinctada margarifera) from its native range and found P. viridis to have both higher feeding efficiencies and clearance rates; 7.2 ± 3.1 L h-1 compared to 4.1 – 5.5 L h-1 for the oysters. In comparing feeding rates and efficiencies, Hawkins et al. (1998) also observed higher retention efficiency (63 +/- 3%), ingestion rate (24.8 +/- 3.6 mg g-1h-1) and absorption rate (21.5 +/- 3.4 mg g-1h-1) in P. viridis compared to all species of oysters (4 – 11%; 2.0 – 9.5 mg g-1h-1; 1.3 – 6.0 mg g-1h-1 respectively). Indicating that high clearance rates observed for P. viridis are accompanied by a high retention efficiency and energy assimilation rate fueling the increased metabolic demands of high growth and reproductive activity.  »
When compared to clearance rates of other Mytilids, P. viridis exceeds that of M. edulis which ranges from 1.66 – 4.12 L h-1 under different diets (Bayne et al., 1987) and from different field populations ranging from 1.12 – 2.55 L h-1 (Okumuş and Stirling, 1994), but is similar to the more closely related green-lipped mussel Perna canaliculus which has clearance rates of 6.8 L h-1 (Hawkins et al., 1999) to 8.6 L h-1 (James et al., 2001). While others report a clearance rate higher than that of oysters when compared directly (Hawkins et al., 1998; McFarland et al., 2013), the clearance rate of C. virginica has been reported to range from 1.6 – 8 L h-1 (Riisg!rd, 1988; Grizzel et al., 2008) which is more comparable to that reported for P. viridis. While comparison between studies is difficult due to differences in methodologies and environmental conditions, it is clear that P. viridis can rapidly clear food particles from the water column with high assimilation efficiencies and are likely to compete with other bivalves.
Feeding behavior and pseudofeces production of P. viridis, and many other bivalves, has shown to vary significantly depending on the food quality and quantity (Hawkins et al., 1998; Wong and Cheung, 1999; Wong and Cheung, 2001). Ward et al. (1998) used endoscopy to show selective ability in the labial palps of several species and adaptive capabilities for rapidly and efficiently sorting, allowing for the maintenance of increased filtration rates typically observed (Ward et al., 1998; Ward et al., 2003). Production of pseudofeces aids in the selective ability of bivalves which allows for the filtration of all particles in the water column and selectivity in choosing particles to be ingested and those to be discarded as pseudofeces based on size and organic content (Ward et al., 1998; Baker et al., 1998; Hawkins et al., 1999; Ward et al., 2003). This is especially helpful in shallow estuaries, common in southwest Florida, which may have large amounts of suspended solids (ie: silt) in the water column that must be sorted from food particles prior to ingestion (Morton, 1987; Seed and Richardson, 1999). During periods of low food (either quality or quantity) or conditions of starvation mussels may also increase food gut time to increase nutrient extraction efficiency (Hawkins and Bayne, 1984; Bayne et al., 1987). These behaviors may allow for a greater increase in energy acquisition in the winter when food is low allowing for the maintenance of year round gametogenesis.
While the commonly accepted life span of P. viridis is 3 years (Lee et al., 1985; García et al., 2005) others have documented at least 4 years with a growth rate of 4 mm year-1 in the fourth year (Narasimham, 1981). Other Perna species have been documented to have a higher longevity. McQuaid and Lindsay (2000) determined an average lifespan of 6.7 years with a maximum of 9 years for Perna perna in sheltered areas. Perna perna, however, have a lower estimated ultimate length of 65 – 117 mm (McQuaid and Lindsay, 2000) compared to P. viridis (156 – 230 mm; Roa et al., 1975; Appukuttan, 1977; Narasimham 1981; Cheung, 1990).
Mortality rates vary between geographical locations, Al-Barwani et al. (2007) reported approximately 1.69 / year and Choo and Speiser (1979) report mortality rates of 3.66 (~93%) for mussels ≤15 mm and 0.52 (~41%) for mussels ≥15 mm. In comparing densities between caged and rope cultured P. viridis, Lee (1986) attributed the 20 – 40% loss in rope culture mussels to predation and / or dislodgement rather than natural mortality. Studies which report high mortality rates often occur during periods of environmental stress such as extreme variations in temperature (Power et al., 2004; Urian et al., 2010; Baker et al., 2012), HAB events (Gacutan et al., 1985; Tracey, 1988; Baker et al., 2012), pollution (Lee, 1986), temperature (Power et al., 2004; Baker et al., 2012), salinity (Gilg et al., 2014) and / or a combination of stressors (Cheung, 1993).

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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).

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


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