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Suitability of Great South Bay, New York to Blooms of piscicida and P. shumwayae Prior to Superstorm Sandy, October 29, 2012

Pawel Tomasz Zablocki CUNY Hunter College

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This work is made publicly available by the City University of New York (CUNY). Contact: [email protected] Suitability of Great South Bay, New York, to Blooms of and P. shumwayae Prior to Superstorm Sandy, October 29, 2012.

By

Pawel Zablocki

Submitted in partial fulfillment of the requirements for the degree of Master of Arts Hunter College of the City of New York

2015

Thesis sponsor:

__25 July 2015 Peter X. Marcotullio Date First Reader

_2 August 2015 Karl H. Szekielda Date Second Reader

i

Acknowledgements

I would like to thank my advisor, Professor H. Gong and two of my excellent readers—Professor

Peter Marcotullio and Professor Karl Szekielda who provided their invaluable advice, alleviated my concerns, and weathered the avalanche of my questions.

ii

Abstract of the Thesis

Pfiesteria piscicida and P. shumwayae are toxic implicated in massive fish kills

in North Carolina and Maryland during 1990s. A set of physical, chemical, and biological factors

influence population dynamics of these organisms. This study employs information gathered

from relevant literature on temperature, salinity, dissolved oxygen, pH, turbulent mixing, and

dissolved nutrients, bacteria, algae, microzooplankton, mesozooplankton, bivalve mollusks,

finfish, and other toxic dinoflagellates, which influence Pfiesteria population dynamics. The research focused on whether conditions in the Great South Bay, Long Island, New York were suitable to blooms of Pfiesteria species prior to the passage of superstorm Sandy in October

2012. The Great South Bay is a coastal lagoon under intensive physicochemical and biological

study. A comparison of conditions between areas where Pfiesteria blooms emerged and that of

the Great South Bay proved inconclusive. The simply inventory-like analysis was, therefore,

insufficient to be employed in answering the research question. Seemingly, residence time plays

an important role in coastal marine environments characterized by high degree of separation

from coastal ocean. Suitability of Great South Bay to Pfiesteria blooms could not be established based on simple comparative analysis which did not stress the importance of residence time; it proved inconclusive with respect to two determinants of Pfiesteria population dynamics— turbulent mixing and potential interactions of Pfiesteria with zooplankton. However, based on analysis that recognized lagoonal character of Great South Bay and the importance of residence time, prior to the passage of Sandy, the confined areas of Great South Bay’s easternmost reaches constitute environments where Pfiesteria blooms appear to have been feasible. Blooms have never occurred, though. Some of the reasons as to why this might be the case include general

iii complexity of HAB development and persistence as well as interactions with viruses and abundant populations of brown tide species Aureococcus anophagefferens .

iv

Table of Contents

Acknowledgements………………………………………………………………...…………...... ii

List of Figures…………………………………………………………………………………...viii

List of Tables……………………………………………………………………….……….…... ix

List of Terms……………………………………………………………………………………....x

List of Acronyms………………………………………………………………………………...xii

1. Introduction……………………………………………………………………………………..1

2. Background

2.1. Study Site Description……………………………………………………………………..2

2.2. Analytical Procedures…………………………………………………………...... …4

2.3. Dinoflagellates, HABs, and Pfiesteria Species: an Overview………………………..…....4

2.4. Physicochemical and Biological Determinants of Pfiesteria Population Dynamics

2.4.1. Effective Cell Concentration ……..…………………………………………………...…8

2.4.2. Temperature…………………………………………….…………….…………...….8

2.4.3. Salinity…………………………………………………………………….………….9

2.4.4. Dissolved Oxygen Concentration …….…………….…………………...... ….10

2.4.5. pH …….…………………………………………………………………………..…11

2.4.6. Turbulent Mixing……...……………………………………………...……………..13

2.4.7. Availability of Dissolved Nutrients………………………...……………………….14

2.4.8. Availability of Finfish Prey……….…………………………………...………….….14

2.4.9. Availability of Algal Prey…………………………………………..……………….17

2.4.10. Interactions with Bacteria…………………...……………………………………..18

v

2.4.11. Interactions with Microzooplankton ……………………………..……………...... 20

2.4.12. Interactions with Mesozooplankton ……………………………………………….21

2.4.13. Interactions with Bivalve Mollusks

2.4.13a. Eastern Oyster ( Crassostrea virginica)…………………………….……….22

2.4.13b. Hard Clam ( Mercenaria mercenaria )…………………………………..….24

2.4.13c. Bay Scallop ( Aequipecten irradians )………………………………………25

2.4.14. Interactions with Other Toxic Dinoflagellates…………………………………..…27

3. Status of Physicochemical and Biological Determinants of Pfiesteria Population Dynamics in

Great South Bay Prior to Superstorm Sandy

3.1. Temperature……………………………………………………………………...... 28

3.2. Salinity…………………………………………………………………………………...30

3.3. Dissolved Oxygen Concentration …….…………………………………………………31

3.4. pH………………………………………………………………….……………………..33

3.5. Turbulent Mixing ……………….……………………………………………………….34

3.6. Availability of Dissolved Nutrients……………………………………………………...36

3.7. Availability of Finfish Prey……………………………………………………………...39

3.8. Availability of Algal Prey…………………...…………………………………………...42

3.9. Interactions with Bacteria………………………………………………………………..43

3.10. Interactions with Microzooplankton ………………………………………………...…44

3.11. Interactions with Mesozooplankton.……………………………………………………45

3.12. Interactions with Eastern Oysters ( Crassostrea virginica ) …………………………….45

3.13. Interactions with Hard Clams ( Mercenaria mercenaria ) ……………….……………..47

3.14. Interactions with Bay Scallops ( Aequipecten irradians ) ………………………………49

vi

4. Discussion……………………………………………………………………………………..51

5. Conclusion…………………………………………………………………………………….94

References……………………………………………………………………………….…...... 99

vii

List of Figures

Figure 2.1. Constituent Embayments of GSB …………………………………………………….4

Figure 3.1. Average July Sea Surface Temperature Readings along the Eastern Coast of the

United States……………………………………………………………………….30

Figure 3.2. Temperature Measurements at Bellport and Blue Point, Great South Bay……….....31

Figure 3.3. Great South Bay Salinity…………………………………..….……………..…...... 32

Figure 3.4. Great South Bay Dissolved Oxygen Concentrations……...……………………..…..34

Figure 3.5. Great South Bay pH Profile………………………………….…….………………...35

Figure 3.6. Great South Bay Tidal Range…………………….…………………….……………36

Figure 3.7. Great South Bay Tidal Current Velocities………………………..……...... …36

Figure 3.8. Recreational Catches of Black Sea Bass in Great South Bay ………………………42

Figure 3.9. Commercial and Recreational Northern Kingfish Catches in Great South Bay…….42

Figure 3.10. Commercial and Recreational Great South Bay Bluefish Catches.…...…………...42

Figure 3.11. General Decline of Bivalve Mollusk Fisheries in Great South Bay……………...... 45

Figure 3.12. Decline of Commercial Hard Clam Yield in Great South Bay …………………....49

Figure 3.13. Decline of Commercial Bay Scallop Catch in Great South Bay………………..….50

Figure 4.1. Spatial Variability of Dissolved Nitrogen Concentrations in a Typical Coastal Lagoon

(Great South Bay)……...... 68

Figure 4.2. Potential Ingestion of Pfiesteria by Microzooplankton according to Salinity………84

viii

List of Tables

Table 2.1. Occurrence of Pfiesteria in Water and Sediment Samples Collected in Coastal Marine

Systems along the Eastern and Southeastern United States (1998-2000)……………...6

Table 3.1. Eutrophication Status of Great South Bay and Other Coastal Marine Systems in the

United States………………………………………………………….. ………….…37

Table 3.2. Respective Contribution of Main Nitrogen Sources to of Great South Bay...... 38

Table 4.1. Some Differences between Estuaries and Coastal Lagoons…………...... 60

Table 4.2. Status of Some Physicochemical Factors in Celestún Lagoon, Mexico….…………..67

ix

List of Terms:

Advection – horizontal movement of particles in the atmosphere or the sea

Anoxia – a state in which an aquatic/marine waterbody is completely devoid of dissolved oxygen

Axenic – containing no bacteria and other normally co-occurring microorganisms

Autotrophy – an ability of organisms to synthetize their nutritional requirements

Bacterivore – an organism that feeds on bacteria

Benthic – relating to the lowest level of a waterbody

Bloom – large concentration of a single phytoplankton species

Chlorophyll – pigment used by plants to conduct photosynthesis

Chloroplasts – chlorophyll-containing organella algae and other plants use to photosynthesize

Cilia (singular: ) – thin elongated outgrowths ciliated protozoa use to move through water column

Culture – cultivation of organisms in laboratory setting

Cyst – thick-walled spherical formations which a number of species turn into to survive unhospitable environmental conditions.

Diatoms – the most abundant marine algae group

Denitrification – biologically mediated transformation of nitrate to gaseous nitrogen, nitric oxide or nitrous oxide

Encystment – a process by which an organism (such as a dinoflagellate) surrounds itself with a cyst when environmental conditions become inhospitable

Eurythermal – an organism capable of functioning over a broad temperature spectrum

Excystment – a process by which an organism reemerges from a resting cyst

Heterotrophy – fulfillment of nutritional needs through consumption of other organisms

Hypoxia – a state in which an aquatic or marine waterbody experiences a noticeable decline in dissolved oxygen concentration

x

Lysis – breakdown and disintegration of a living cell

Mesohaline – moderately brackish (salinity 5-18 ppt)

Mixotrophy – an ability of organisms to feed autotrophically and heterotrophically pH – a measure of acidity/alkalinity of aqueous solutions

Photosynthesis – a biochemical process conducted by plants to generate nutrients they require

Picoplankton – planktonic organisms within the 02-2 µm size range

Primary Productivity – organic products of photosynthesis

Residence Time – the time a particle (like water) in a given system

Suboxic – characterized by below normal dissolved oxygen concentration

Toxicogenic – capable of producing toxins

Viral – relating to viruses

Zoospore – a stage in dinoflagellate life cycle

xi

List of Acronyms

APES Albemarle-Pamlico Estuarine System

DIN dissolved inorganic nitrogen

DO dissolved oxygen

DOM dissolved organic matter

GSB Great South Bay

HAB harmful algae bloom

LI Long Island

NRE Neuse River Estuary

PS Pamlico Sound

RT residence time

SAV submerged aquatic vegetation

SCDHS Suffolk County Department of Health Services

TPC toxic Pfiesteria complex

xii

1. Introduction

Harmful algae blooms (HABs) have been occurring with increased frequency around the world

(Glibert et al. 2005a; Glibert and Burkholder 2006a; Hallegraeff 2010). Such blooms are

undesirable as some of the microorganisms that form them produce toxic compounds which can

kill marine life and sicken and even kill human beings (Fleming et al. 1999; Landsberg 2002).

During the 1990s, two toxic dinoflagellates—Pfiesteria piscicida and P. shumwayae —were discovered in coastal waters of a number of Atlantic and Gulf Coast states; however, outbreaks of these organisms occurred only in North Carolina and Maryland. In the early 2000s, Pfiesteria were detected in Great South Bay and other waters of Long Island, NY. According to Rublee et al. (2006), prior to superstorm Sandy, GSB did not constitute an environment where Pfiesteria could establish viable populations due to the lack of alignment of physicochemical and biological parameters that have been suggested to determine population dynamics of these organisms. One published study (Rublee et al. 2006) and a webpage based on its findings

(Suffolk County 2014a) seem insufficient to affirm that bloom formation and toxic activity of

Pfiesteria species were implausible in GSB prior to the passage of superstorm Sandy in October

2012. Therefore, based on what is currently known about Pfiesteria and the complex mechanism governing their toxin production, the goal of the current study is to establish whether Pfiesteria constituted a threat in GSB prior to the extreme weather event of October 2012. Analysis of the pre-Sandy status of the physicochemical and biological determinants of Pfiesteria population

dynamics in GSB did not facilitate procurement of a straightforward answer to study’s research

question because while GSB appears to have been hospitable to Pfiesteria blooms with respect to some of these determinants, it does not appear to have been such in regards to others. Additional research, undertaken to qualify information compiled in the preceding section of the document,

1 shows that prior to October 29, 2012, some sites of GSB appear to have constituted environments where Pfiesteria bloom formation was conceivable.

The reminder of the document consists of “Background” (Chapter 2), “Status of

Physicochemical and Biological Determinants of Pfiesteria Population Dynamics in Great South

Bay Prior to Superstorm Sandy” (Chapter 3), “Discussion” (Chapter 4), “Conclusion” (Chapter

5), and “Tables, “Figures,” and “References sections. The first section of the Background ,

“Dinoflagellates, HABs, and Pfiesteria Species: An Overview” (Chapter 2.3) provides some general information on dinoflagellates and—more specifically—Pfiesteria . In the second section of the Background (Chapter 2.2), “Physicochemical and Biological Determinants of Pfiesteria

Population Dynamics,” factors which influence Pfiesteria population dynamics, have been laid out. “Status of Physicochemical and Biological Determinants of Pfiesteria Population Dynamics in GSB Prior to Superstorm Sandy,” which constitutes Chapter 3 of this study, deals with the status of the physicochemical and biological framework as it pertained to Pfiesteria population dynamics in GSB prior to the passage of superstorm Sandy in October 2012. The “Discussion”

(Chapter 4) first point outs insufficiency of the information compiled in Chapter 3 and then goes on to argue that, owing to geomorphologically-driven spatial variability of physicochemical and biological determinants of Pfiesteria population dynamics, as well as general hardiness of these organisms in the face of physical, chemical, and biological challenges commonly encountered in marginal marine environments, Pfiesteria bloom formation may not have been as unlikely in

GSB as it had been speculated in Rublee et al. (2006) and Suffolk County (2014a).

2. Background

2.1 . Study Site Description

2

Great South Bay (GSB) is a shallow waterbody stretching along the coast of southern Long

Island, New York (Beck et al. 2008) east of South Oyster Bay, west of Moriches Bay and about

60 km from New York City (Dowhan et al. 1997) (Figure 2.1).

Figure 2.1. Constituent Embayments of GSB (NYSDOS 2001).

It is approximately 35-40 km long and 9 km wide at its widest point stretching over an area of over 235 sq. km (Beck et al. 2008). It is considered a subregion of the South Shore Estuary

Reserve (NYSDOS 2001) and commonly it is subdivided into eastern and western sectors in relation to the location of the Fire Island Inlet (USACE 2004a). GSB is considered a productive biological resource as its waters are an important habitat for a numerous species of finfish, shellfish, and migratory birds (Dowhan et al. 1997). It is economically important as it constitutes an important transportation route for commercial cargo, a significant commercial and recreational fishing site, and an attractive tourist destination (Greene et al. 1978; Dowhan et al.

1997; Monti and Scorca 2003).

3

2.2. Analytical Procedures

The research conducted in the current document will utilize a broad variety of secondary data

gathered from peer-reviewed scientific literature, databases of non-profit organizations, federal,

state and county governments, as well as books and newspaper articles to establish whether

Pfiesteria outbreaks were plausible in LI’s GSB prior to the passage of superstorm Sandy over these waters in October 2012. Physicochemical and biological determinants of Pfiesteria

population dynamics, delineated in “Physicochemical and Biological Determinants of Pfiesteria

Population Dynamics” (Chapter 2.2), will be examined in the context of a specific geographic

location—Great South Bay, Long Island, NY.

2.3. Dinoflagellates, HABs, and Pfiesteria Species: An Overview

Dinoflagellates are ecologically important and common inhabitants of the marine realm (Jeong

1999; Margalef 1978; Holmes-Farley 2006). They can be autotrophic (by obtaining nutrients through the process of photosynthesis), heterotrophic (by sustaining themselves through ingestion of other organisms), or mixotrophic (when—depending on the availability of suitable prey—they can attain nourishment both autotrophically and heterotrophically (Burkholder et al.

1993; Mender-Deuer and Lessard 2000; Stoecker, Tillman, and Granéli 2006). A number of these organisms produce potent toxins which can harm and kill finfish, shellfish, seabirds, and marine mammals (Hallegraeff 1993; Landsberg 2002; FFWCC 2014), and—as a result—afflict coastal economies (Anderson et al. 2000; Hoagland et al. 2002; Hoagland and Scatasta 2006).

Several dinoflagellate species can also negatively affect human health (Tibbetts 1998; Fleming et. al 1999; Anderson, Glibert, and Burkholder 2002). As coastal populations swell (Bowen and

4

Valiela 2001; Paerl et al. 2006; Wang et al. 2008), international commerce and associated marine

traffic increases (Hallegraeff 1998; Doblin et al. 2004; Drake et al. 2005), and as climate change

advances (Paerl et al. 2006; Hallegraeff 2010; Anderson 2012), many coastal marine ecosystems,

have been becoming more hospitable to HABs. Indeed, HABs have been spreading geographically (McMinn et al. 1997; Mallin 2000; Vila et al. 2001) and becoming increasingly frequent and intense (Smayda 1990, 2002; Hallegraeff 1993). There were 22 identified toxic dinoflagellates species in 1984 (Steidinger and Baden 1984); a little more than a decade later the number of these organisms has increased to 59 (Burkholder 1998; Glibert et al. 2005). In the

United States, HABs occur in every coastal state (Tibbetts 1998; Fleming et al. 1999; Turgeon et al. 2009).

Pfiesteria piscicida and P. shumwayae , member species of the so called Toxic Pfiesteria

Complex (TPC) (Burkholder and Glasgow 2001; Glasgow et al. 2001) have been considered the

culprit of numerous finfish kill events and human health problems in tributaries of Albemarle

Pamlico Estuarine System (APES) and Chesapeake Bay. Pfiesteria piscicida was first discovered

in 1988 in experimental fish tanks at North Carolina State University’s College of Veterinary

Medicine (Burkholder et al. 1992, 1995). In May 1991, P. piscicida was detected in its natural

habitat for the first time during an active finfish kill event (Burkholder et al. 1992, 1995;

Steidinger et al. 1996). , the second TPC species, was discovered by

Howard B. Glasgow and JoAnn M. Burkholder of the Center for Applied Aquatic Ecology

(CAAE) of North Carolina State University (NCSU 2000). Both species have been subdivided

into three functional types based on the status of their toxicity: TOX-A (actively toxic), TOX-B

(temporarily non-toxic), and NON-IND (non-inducible, i.e. unable to produce toxins)

(Burkholder et al. 2001a; Cancellieri et al. 2001; Glibert and Burkholder 2006).

5

Pfiesteria have become a cause for concern because toxins which these organisms release have been found to adversely affect human health. Although there have been no recorded mortalities associated with Pfiesteria toxin exposure (Griffith 1999), fishermen, crabbers, and recreational swimmers who came into contact with waters where a Pfiesteria outbreak was occurring, suffered from headaches, abdominal pains, shortness of breath, blurred vision, and an array of neurologic and cognitive problems (Glasgow et al. 1995; Morris 1999; Grattan, Oldach, and Morris 2001), collectively termed the possible-estuary associated syndrome (Backer et al.

2001; Moe et al. 2001; Shoemaker 2001).

Since their discovery and toxic activity in APES and Chesapeake Bay, potentially toxic

Pfiesteria have been detected in a number of other coastal states (FDEP 1999; Marshall,

Seaborn, and Wolny 1999; SC Task Group on Toxic Algae 1999; NJDEP 2000; Rublee et al.

2001; Lewitus et al. 2002). Both TPC species are now residents of coastal waters as far north as

New York’s Long Island and as far south as Texas (Rublee et al. 2001) (Table 2.1).

Table 2.1. Occurrence of Pfiesteria in Water and Sediment Samples Collected in Coastal Marine Systems along the Eastern and Southeastern United States between 1998 and 2000 (Rublee et al. 2001).

6

Toxicogenic Pfiesteria have also been detected in Argentina (Rublee et al. 2004), Chile (Lin,

Zhang, and Dubois 2006), Indonesia (Park, Bolch, and Hallegraeff 2007), Mexico (Rublee et al.

2004), New Zealand (Rhodes et al. 2002), Norway (Jakobsen et al . 2001), and South Korea

(Jeong et al. 2006). Although Pfiesteria are not sufficiently abundant in these locations to form blooms (Marshall 1999; Coyne et al. 2001; Lewitus et al. 2002; Coyne et al. 2006; Lin, Zhang, and Dubois 2006), and there have been no confirmed Pfiesteria outbreaks in neither the APES nor Chesapeake Bay since 1999 (Moe et al. 2001; Rublee et al. 2001), populations of these organisms have been growing steadily (Burkholder and Glasgow [unpublished data] in Jochems and Martens 2001; Rublee et al. 2005). Rublee et al. (2001) found Pfiesteria cells in only 3% of water and sediment samples collected in coastal waters of Eastern United States. Conducting a study in the same region, Lin, Zhang, and Dubois (2006) found Pfiesteria cells in 33.7% of samples collected.

During the early 2000s, Pfiesteria have been detected in GSB and other embayments of

Long Island’s South Shore Estuarine System. Pfiesteria cells collected by Rublee et al. (2006) in the water column and sediments of GSB were tested for toxicity and some of these intoxicated and killed finfish. Despite these findings, it was concluded that bloom formation and toxic activity of TPC species were unlikely in GSB (Rublee et al. 2006).

HABs are highly unpredictable (Smayda 2002; Hood et al. 2006). There are, however, suites of physical, chemical, and biological parameters, a “bioclimate envelope,” (Cheung 2009,

236), which have been shown to influence occurrence and ecological success of HAB species

(Paerl 1988; Hood et al. 2006; Place et al. 2012). Laboratory and field research have indicated that interplay of specific set of physical, chemical, and biological parameters (“the basic conditions,” Burkholder and Glasgow 2001, 834) determines rate of population growth,

7

geographic distribution, and toxic activity of Pfiesteria species (Glasgow et al. 2001; Hood et al.

2006). The importance of investigating these environmental variables has been stressed in a number of publications. Coyne et al. (2001), for instance, stressed that “an analysis of these factors and their relationships to Pfiesteria community composition is essential to the prediction

and prevention of toxic blooms” (Coyne et al. 2001, 276). The “factors” in question are water

temperature, its salinity, magnitude of its pH, availability of prey and symbiotic bacteria,

availability and nature of dissolved nutrients, magnitude of turbulent mixing, and predation

pressure exacted by bivalve mollusks, micro-, and mesozooplankton.

2.4. Physicochemical and Biological Determinants of Pfiesteria Population Dynamics

2.4.1. Effective Cell Concentration

Similar to other species of toxic dinoflagellates (Steidinger 1975; Paerl 1988; Zigone et al.

2006), to bloom, Pfiesteria population must increase in size (Drgon et al. 2005). Pfiesteria

populations can achieve very high densities. During fish kill events in subestuaries and

tributaries of North Carolina’s APES between 1991 and 1995, Pfiesteria concentrations reached

concentrations as high as 940±130 cells/ml in the Neuse River and 4,300±2,510 cells/ml in the

Pamlico River (Burkholder, Glasgow, and Hobbs 1995; Burkholder and Glasgow 1997;

Burkholder et al. 1997).

2.4.2. Temperature

Temperature is an important environmental parameter which influences distribution and activity

of marine organisms (Eppley 1972; Lomas et al. 2002; Rose and Caron 2007). Such

physiological processes as uptake of dissolved nutrients and primary and bacterial productivity

8

are temperature-dependent (Caron, Goldman, and Dennet 1986; Nielsen and Tønseth 1991;

Rogers and McCarty 2000).

In general, heterotrophic protists, such as dinoflagellates, are strongly influenced by

temperature and its variations (Chang and Carpenter 1985; Kim et al. 2004; Rose and Caron

2007). Dinoflagellates prefer warm environments and, as numerous studies have demonstrated,

rates of their population growth and bloom development are highest during summer and early

fall (Kremp and Parrow 2006; Hallegraeff, Mooney, and Evans 2009; Granéli et al. 2011).

Temperature is considered the most important environmental parameter controlling germination

of dinoflagellate benthic cysts (Anderson 1980; Binder and Anderson 1987; Rengefors and

Anderson 1998).

Pfiesteria are eurythermal organisms, which means that they can be active across broad temperature range (Burkholder et al. 1993, 1995; Burkholder and Glasgow 1997, 2001). In the majority of water samples collected by Rublee et al. (2004) in coastal waters of nineteen countries, presence of Pfiesteria species was not to be correlated with temperature. Both species can withstand freezing temperatures as low as -20 °C for up to 6 months and heat exposure of

76.5 °C (Gordon et al. 2002). It was determined that the optimal temperature at which Pfiesteria engage their toxic activity is ≥15 °C (Glasgow et al. 2001). On the other hand, both organisms

were capable of producing toxin and killing finfish at temperatures <15 °C (Burkholder,

Glasgow, and Hobbs 1995; Burkholder and Glasgow 1997; Glasgow et al. 2001). Greatest

population growth and peak toxic activity of Pfiesteria species, however, have occurred when

water temperatures were ≥26 °C (Burkholder and Glasgow 1997; Marshall 1999).

2.4.3. Salinity

9

Salinity is considered a “master factor” (Oliver 2005, 10) as it is very important in influencing

biological processes and distribution of organisms in marine systems (Calliari et al. 2006;

Saunders et al. 2007; NOAA 2008).

The optimum salinity for the growth and induction of toxic activity in Pfiesteria piscicida

and P. shumwayae is 15 ppt (Burkholder et al. 1993; Burkholder and Glasgow 1997; Shumway,

Burkholder, and Springer 2006). Nevertheless, similar to other dinoflagellate species (Macedo et

al. 2001; Sullivan and Andersen 2001; Grzebyk et al. 2003), Pfiesteria are euryhaline organisms, which means they can tolerate broad salinity spectrum (Burkholder, Glasgow, and Hobbs 1995;

Sullivan and Andersen 2001; Burkholder and Marshal 2012). Indeed, can bloom in full-strength seawater (Burkholder and Glasgow 2001) as well as in freshwater (Burkholder et al. 1993;

Burkholder and Glasgow 1997, 2001). Springer et al. (2002) found that Pfiesteria demonstrated the most aggressive attack behavior towards bay scallop larvae when salinity of the culture medium was ≈15 ppt. At the same time, Pfiesteria swarmed around the larvae within seconds of exposure at salinities as high as 25 ppt (Springer et al. 2002). Pfiesteria have been cultured cultures in 25-30 ppt salinities (Springer et al. 2002). Bloom formation and toxic activity of

Pfiesteria populations can also occur in waters characterized by salinities lower than the 15-ppt

optimum. During the 1997 Pfiesteria bloom in Maryland, salinities of the Pocomoke, Manokin, and the Chicamacomico Rivers were 7-11, 0.5-10, and <10 ppt respectively (Maryland

Department of Natural Resources 1998 in Stoecker and Gustafson 2002). A number of Pfiesteria toxic events occurred even in freshwater aquaculture facilities (Burkholder et al. 1993;

Burkholder and Glasgow 1997, 2001).

2.4.4. Dissolved Oxygen Concentration

10

Dissolved oxygen (DO) concentration, which is the availability of oxygen dissolved in water

(USGS 2014), is an important environmental parameter in coastal marine environments (NOAA

2008). It is vital to most marine organisms (USGS 2014) and its scarcity or complete absence

can by harmful to them (Diaz and Rosenberg 1995; Peña et al. 2010; USEPA 2012). Numerous

eutrophied coastal systems in the United States experience episodes of hypoxia (low DO

concentration) or anoxia (no DO) (Diaz and Rosenberg 1995; Boynton et al. 1996; Breitburg et

al. 1997). DO concentrations can decline due to a number of naturally and anthropogenically

driven processes (USEPA 2012). Most commonly, DO concentrations decline as algae blooms

senescence and bacteria exhaust its supply to decompose algal remains (Stanley and Nixon 1992;

Pinckney et al. 2000; USGS 2014).

Relevant literature contains only scant information on Pfiesteria in relation to their DO requirements. According to Glasgow et al. (2001) Pfiesteria behaves normally at DO concentrations of ≥3.8 mg/L.

2.4.5. pH pH, which is used to measure acidity/basicity level of aqueous solutions (USEPA 2012a; USGS

2014a), is an important parameter influencing biotic processes in marine ecosystems (Chen and

Durbin 1994; Addy, Green, and Heron 2004; Holmes-Farley 2006). Generally, living organisms exercise control over their internal pH levels so that small changes of external pH do not affect normal progress of their life processes (Holmes-Farley 2006). Substantial declines of water pH, however, can be detrimental to living organisms (Gehl and Colman 1985; Addy, Green, and

Heron 2004; Holmes-Farley 2006).

11

Seawater is well buffered (de Wit et al. 2001; University of Western Australia 2011;

Middelboe and Hansen 2007). This buffering capacity is a result of interplay of a set of

2- - equilibrium chemical reactions which involve carbonate (CO 3 ) and hydrogen carbonate (HCO 3

) ions (University of Western Australia 2011) according to the equations

+  CO 2 + H 2O ↔ H 2CO 3 ↔ H + HCO 3

and

2-  CO 2 + H 2O + CO 3 ↔ 2 HCO 3 ,

+ where CO 2 is carbon dioxide, H 2O is water, H 2CO 3 is carbonic acid, H is hydrogen ion, HCO 3

2- is bicarbonate ion, and CO 3 is carbonate ion. The result of these reactions is that pH values in

coastal marine waters do not fluctuate significantly and generally fall between ≈7.9 and 8.3, a

range which is congenial to marine life (NOAA 2007; University of Western Australia 2011). As

a matter of fact, it has been suggested that pH level of seawater in world’s oceans has been

largely stable for as long as 20 million years (Pearson and Palmer 2000).

Positive growth of dinoflagellate populations has been associated with high pH levels

(Hinga 1992; Macedo et al. 2001). P. minimum populations can grow normally at pH ≈10

(Hansen 2002). marina grew normally at pH ≈10.2 (Pedersen and Hansen 2003).

Blooms of Alexandrium spp. and spp. have also been associated with elevated pH values (Macedo et al. 2001). In fact, dinoflagellates P. minimum and P. micans are two marine

phytoplankton species most commonly associated with elevated pH (Hansen 2002; Pedersen and

Hansen 2003a; Sierra-Beltrán, Cortés-Altamirano, and Cortés-Lara 2005).

Positive growth of Pfiesteria populations and their toxic activity occurs at pH ≥ 6.8

(Glasgow et al. 2001; Burkholder et al. 2004) and ≤ 8.6 (Glasgow et al. 2001) with optimum pH

12 of 7.5 (Simmons, Villareal, and Rublee 2004). At pH outside this range, Pfiesteria populations can grow, albeit at reduced rates (Glasgow et al. 2001).

2.4.6. Turbulent Mixing

Turbulence is one the most important factors controlling population dynamics of planktonic organisms (Furnas, Smayda, Tomas 1989; Margalef 1997; Havskum, Hansen, and Berdalet

2005). Blooms of harmful dinoflagellates have been generally associated with low turbulence conditions (Margalef 1978; Juhl, Velazquez, and Latz 2000; Sullivan and Swift 2003) given that turbulent mixing negatively affects general movement, feeding behavior, cell division rate, and other aspects of dinoflagellate physiology (Juhl, Velazquez, and Latz 2000; Havskum, Hansen, and Bardalet 2005; Hall et al. 2008). According to Weise et al. (2002), low turbulent mixing, which resulted from low wind speeds, was the chief culprit of A. tamarense blooms in St.

Lawrence River Estuary in June 1994. Hallegraeff, McCausland, and Brown (1995) attributed partial responsibility of paralytic shellfish poisoning outbreaks in coastal waters of southern

Tasmania to periods of low wind speed and low-level turbulent mixing.

Laboratory studies have shown that elevated turbulence has a negative effect on population dynamics of Pfiesteria species (Burkholder and Glasgow 1997; Glasgow et al. 2001;

Stoecker et al. 2006). To properly handle water samples containing Pfiesteria zoospores in laboratory setting, Burkholder and Glasgow (1997) and Stoecker et al. (2006) highlighted the importance of drop-pipetting and otherwise gentle handling of Pfiesteria cultures to prevent zoospore cells from encysting and settling out of the water column. In a modeling experiment by

Hood et al. (2006) Pfiesteria population growth was negative as dinoflagellates were exposed to estuarine level turbulent mixing. Turbulence has also been found to affect predatory success of

13

Pfiesteria species. Springer et al. (2002) found oyster larvae kept Pfiesteria zoospores at bay by moving their cilia vigorously generating turbulence to foil the dinoflagellate attack (Springer et al. 2002).

2.4.7. Availability of Dissolved Nutrients

Similar to other heterotrophic dinoflagellates, Pfiesteria can assimilate dissolved nutrients

directly from the water column (Burkholder, Glasgow, and Hobbs 1995; Burkholder and

Glasgow 1997; Glibert et al. 2006a) and assimilate these at rates comparable to rates at which

they feed phagotrophically (Glibert et al. 2006a; Lewitus et al. 1999). In addition to being

efficient consumers of inorganic nutrients (Burkholder, Glasgow, and Hobbs 1995; Burkholder

and Glasgow 1997), similar to other toxic dinoflagellate species (Glibert 1998; Glibert and

Terlizzi 1999; Li, Stoecker, and Coats 2000), Pfiesteria utilize dissolved organic nitrogen (Zhang

et al. 2004; Glibert et al. 2006; Hood et al. 2006; Lewitus et al. 2006), which can constitute a

large proportion of their nutritional requirements (Glibert et al. 2006a). Presented with several

types of nitrogenous compounds, Pfiesteria tend to ingest amino acids and urea followed by

ammonium (NH4 +) and nitrate (NO3 -) (Glibert at al. 2006a).

2.4.8. Availability of Finfish Prey

Toxic dinoflagellates can sicken and kill finfish (Nielsen 1993; Kudela et al. 2008; Tang and

Gobler 2009). Toxic compounds (brevotoxins) released by K. brevis , a notorious finfish killer in

coastal waters of the Gulf of Mexico (Brand and Compton 2007; Glibert et al. 2009), affects all

developmental stages of affected finfish (FFWCC 2014). Individuals exposed to K. brevis swim

erratically, convulse, defecate and regurgitate uncontrollably, and-as their gills fail-suffocate

14

(Magaña, Contreras, and Villareal 2003; FFWCC 2014). Another prolific fish killer, K.

veneficum (Hall et al. 2008; Sheng et al. 2010), narcotizes fish to immobilize them prior to direct physical attack (Sheng et al. 2010); affected fish die as a result of gill inflammation and failure

(Paerl et al. 2009; Park et al. 2011; Place et al. 2012). Nielsen (1993) described Juvenile Atlantic cod ( G. morhua ) dying within two hours of exposure to toxic G. galatheanum . Actively toxic

Pfiesteria have caused disease and mortality in all finfish species tested (Burkholder et al. 1993;

Burkholder and Glasgow 1997; Cancellieri et al. 2001). In fact, it has been postulated that toxic activity of Pfiesteria to be controlled by their access to finfish (Burkholder and Glasgow 1997;

Burkholder et al. 2001a). When deprived of fresh finfish tissue, Pfiesteria may lose their ability to produce toxins (Burkholder and Glasgow 1997; Burkholder et al. 2001a). Pfiesteria affect finfish in two ways: through biotoxins they release and by direct physical attack (Gordon and

Dyer 2005; Glibert and Burkholder 2006). It has been suggested that both TPC members detect chemical cues emitted by traveling schools of finfish (Glasgow et al. 1995; Cancellieri et al.

2001; Saito et al. 2007), which trigger toxic activity of these organisms (Burkholder et al. 1992,

1995; Burkholder and Glasgow 1997). In encystment/excystment experiments conducted by

Seito et al. (2007), Pfiesteria germination rate was approximately five times higher in culture media stocked with fish than in media where finfish were absent. Pfiesteria would swarm around finfish prey, immobilize them with toxins, and feed on their epidermal tissues (Burkholder et al.

1992, 1993; Steidinger et al. 1996). Egerton and Marshall (2006) demonstrated that the P. piscicida actively sought fish blood cells even when these were only 15% of the total food resources available. It has been noted that actively toxic Pfiesteria are able to kill finfish even when physically separated from their prey by dialysis membranes, which prevent direct contact between dinoflagellates and finfish but allow free flow of toxic compounds (Burkholder and

15

Glasgow 1997; Burkholder et al. 2001a; Gordon et al. 2002). Affected finfish appear to be stressed and lose their ability to manage buoyancy and normal locomotory function (Marshall et al. 2000; Burkholder and Glasgow 2001; Gordon, Gordon, and Dyer 2006). Most commonly, finfish affected by Pfiesteria experience gradual deterioration and eventual loss of their mucus coat (Burkholder, Glasgow, and Hobbs 1995; Dykstra and Kane 2000; Vogelbein et al. 2001), a thin epidermal layer responsible for respiration, osmotic regulation, excretion, and disease resistance in fish (Jakowska 1963; Shepard 1994). Even when seemingly insufficiently numerous, Pfiesteria can compromise finfish health and survival. At sublethal levels ( ≈100-250 cells/ml), Pfiesteria can trigger ulcerative diseases (Burkholder et al. 1993; Burkholder and

Glasgow 1997; Noga 2000) as well as affect the neural, immune, and reproductive systems of affected finfish (Burkholder et al. 1993; Burkholder and Glasgow 2001; Lewitus, Glasgow,

Burkholder 1999). Eggs of a number of finfish species, including striped bass ( Roccus saxatillis ;

Burkholder 1999) and Atlantic ( Brevoortia tyrannus ; Burkholder, Glasgow, and

Hobbs 1995), failed to hatch in the presence of actively toxic Pfiesteria cells.

Similar to other heterotrophic dinoflagellates (Hall et al. 2008; Hallegraeff, Mooney, and

Evans 2009; Place et al. 2012), Pfiesteria are mixotrophic organisms, which means they can switch between feeding autotrophically (through photosynthesis) and heterotrophically (through consumption of other organisms) (Granéli et al. 1997; Burkholder and Glasgow 2001). This ability of being able to feed autotrophically through photosynthesis as well as to feed on a wide array of prey organisms is related to Pfiesteria polymorphic nature (Burkholder and Glasgow

1997). Pfiesteria have complex life cycle with at least 24 stages (Burkholder, Glasgow, and

Hobbs 1995; Lewitus, Glasgow, and Burkholder 1999; Burkholder and Glasgow 2001). In this respect, Pfiesteria are “polymorphic and multiphasic” (Steidinger et al. 1996, 159) as they alter

16

their size and shape according to availability and nature of food resources (Burkholder 1999;

Burkholder and Glasgow 1997a.). Its planktonic zoospore stage predominates when flagellated algal prey are abundant. As nonmotile prey becomes more numerous, a larger percentage of

Pfiesteria cells present transform into benthic lobose amoebae (Burkholder et al. 2001; Glasgow et al. 2001). Pfiesteria have demonstrated their preference of finfish tissue (Egerton and Marshall

2006; Glibert and Burkholder 2006), however, when live finfish are scarce or unavailable,

Pfiesteria —as many other toxic dinoflagellates (Li et al. 1996; Berge, Hansen, Moestrup 2008;

Cardoso 2012)—can subsist on a broad spectrum of prey organisms (Burkholder and Glasgow

1995; Burkholder and Glasgow 1997; Glasgow et al. 2001; representants of all marine trophic levels (Burkholder and Glasgow 1997, 2001; Stoecker et al. 2006).

2.4.9. Availability of Algal Prey

Algae are important biotic component of marine ecosystems (Addy and Green 1996; Miller,

Geider, and MacIntyre 1996; Underwood and Kromkamp 1999). Because of their ability to produce organic matter through photosynthesis, algae constitute important component of marine food webs (Pope 1975; ASM 2006).

Similar to other species of heterotrophic dinoflagellates (Li et al. 1996; Li, Stoecker, and

Coats 2000; Kudela et al. 2008), Pfiesteria are consumers of algae (Burkholder and Glasgow

1995; Marshall 1999; Pinckney et al. 2000). Studies have indicated that representants of

Cryptophyta phylum are Pfiesteria preferred algal prey (Burkholder and Glasgow 1995; Lewitus,

Glasgow, and Burkholder 1999; Stoecker and Gustafson 2002), with Rhodomonas and

Cryptomonas genera ingested at highest rates (Egerton and Marshall 2006; Jeong et al. 2006;

Saito et al. 2007). Additionally, similar to a number of other heterotrophic dinoflagellates

17

(Eriksen, Hayes, and Lewitus 2002; Takishita et al. 2002; Nishitani et al. 2012), Pfiesteria can

retain chloroplasts of ingested algal prey (kleptochloroplastidy) and use them to supplement their

food requirements with nutrients generated autotrophically when other sources of nutrients

become limited or unavailable (Burkholder et al. 1993; Burkholder and Glasgow 1995;

Lewitus, Glasgow, and Burkholder 1999).

2.4.10. Interactions with Bacteria

Bacteria are important to the proper functioning of marine ecosystems (Ferguson and Palumbo

1979; Cole, Findlay, and Pace 1988; Cho and Azam 1990). They are predominant decomposers

of organic matter (Ferguson and Palumbo 1979; Cole, Findlay, and Pace 1988) and constitute

food source for numerous marine organisms (Berk, Cowell, and Small 1976; Jürgens, Arndt, and

Rothhaupt 1994). They are most abundant in warm waters (Bjørnsen et al. 1989; Shiah and

Ducklow 1994); hence, the most prolific growth of their populations occurs during summer and early fall (Shiah and Ducklow 1995; Li 1998; Crump et al. 2004). Similar to Pfiesteria , bacterial

populations grow at highest rates while feeding on dissolved organic nutrients (Hagström et al.

1984; Davies et al. 1986; Cole, Findlay, and Pace 1988). Their rapid growth is no match for

metazoans and protists that prey on them; various microbial predators have demonstrated

inability to significantly reduce populations of marine bacteria (Pace 1988). As a consequence,

concentrations of bacterioplankton vary little in the marine realm (Davis and Sieburth 1984; Del

Giorgio et al. 1996; Crump et al. 2004).

Pfiesteria interacts with bacteria in two ways: 1) they can prey on bacterial cells, and 2) they can form symbiotic relationships with specific strains of bacteria, which, as it has been claimed, play an important role in activation of their toxic activity.

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Bacteria are nutritionally valuable (Sherr et al. 1986; Simon and Azam 1989) and

numerous dinoflagellate species are efficient bacterivores (Wikner and Hagström 1988; Kirchner

et al. 1996; Seong et al. 2006). Pfiesteria also prey on bacteria (Burkholder and Glasgow 1995,

1997; Fleming et al. 1999). Pfiesteria zoospores can capture bacteria and concentrate them in

“netlike cytoplasmic extensions” prior to ingestion (Burkholder and Glasgow 1995, 185).

Amoeba stage of Pfiesteria life cycle ingests bacteria prior to asexual reproduction (Burkholder and Glasgow 1995; Burkholder et al. 2001a).

Apart from using bacteria as food source, a number of toxic dinoflagellates can form symbiotic relationships with these organisms (Córdova, Escudero, and Bustamante 2003;

Hasegawa et al. 2007). Most commonly these bacteria belong to the alphaprotobacteria ( α-

Protobacteria) and gammaproteobacteria ( γ-Proteobacteria) groups (Garrity 2001; Alverca et al.

2002), both abundant, and often dominant components of marine food webs (González and

Moran 1997; Piccini et al. 2006). For example, bacteria have been suggested to contribute to production of paralytic shellfish toxins (PST) in Alexandrium spp. (Dantzer and Levin 1997;

Gallacher et al. 1997). Substantiation of the above and similar results of other studies has been inferred from the fact that dinoflagellates capable of producing toxins lose these capabilities in axenic cultures. According Martins et al. (2004), A. lusitanicum can lose its ability to produce toxins when separated of from co-occurring bacterial assemblages. Doucette and Powell (1998) found similar reaction of A. lusitanicum to removal of bacteria. Similarly, certain Pfiesteria strains require the presence of symbiotic bacteria to produce toxins (Tengs et al. 2003; Alavi et al. 2001; Miller and Belas 2004). Alavi et al. (2001) found over 30 bacteria species co-existing with these dinoflagellates. Alavi (2004) found that the presence of Roseobacter sp. was essential for Pfiesteria cultures to grow and produce toxins. When cultured axenically, Pfiesteria can

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produce only limited amounts of toxins (Burkholder et al. 2005; Burkholder and Marshall 2012;

CAAE 2013) and grow at significantly reduced rates (Alavi et al. 2001). Former rates of population growth and toxin production capabilities tend to become restored once bacteria are re- introduced (Burkholder and Marshall 2012).

2.4.11 . Interactions with Microzooplankton

Microzooplankton are abundant and important constituents of marine food webs (Berk, Cowell, and Small 1976; Bernard and Rassoulzadegan 1990; Dolan and Coats 1990). Their populations can grow at very high rates, at times, as fast as algal populations (Hansen, Bjørnsen, and Hansen

1997; Kamiyama and Arima 2001; Gransden and Lewitus 2003). Microzooplankton can feed on a wide range of prey and can clear 100% daily phytoplankton production in a given coastal marine system (Calbet and Landry 2004). By feeding on bacteria (Bernard and Rassoulzadegan

1990; Sherr and Sherr 1987), algae (Montagnes, Berger, and Taylor 1996; Kamiyama and Arima

2001), and other prey, microzooplankton facilitate transfer of energy to higher trophic levels, benefiting larger organisms (such as mesozooplankton) which would be unable to utilize it otherwise (Gifford 1991; Calbet 2001; Godhantaraman and Uye 2003). Microzooplankton can also feed on toxic dinoflagellates and influence population dynamics of these organisms (Jeong et al. 1999; Montagnes and Lessard 1999; Jeong et al. 2004). According to Stoecker, Thessen,

and Gustafson (2008), microzooplankton can potentially clear 100% of daily production of

dinoflagellate cells. Microzooplankton also prey on Pfiesteria . Studies have shown that such

microzooplankton as Tintinnopsis spp., Tintinnidium spp., Nolaclusilis spp., Favella sp .

(Stoecker, Stevens, and Gustafson 2000; Lewitus et al. 2006), Strombidium spp. (Setälä, Autio,

and Kuosa 2005), Stylonichia putrina (Burkholder and Glasgow (1995), woodruffi

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(Gransden and Lewitus 2003), Brachionus plicatilis (Mallin et al. 1995), Mesodinium pulex

(Stoecker, Stevens, and Gustafson 2000) and Strombidinopsis jeokjo (Jeong et al. 2004) prey on these dinoflagellates. According to Vogelbein et al. (2001), microzooplankton can contribute to substantial reduction of Pfiesteria population size. Stoecker, Thessen, and Gustafson (2008) hypothesized that microzooplankton might be able to clear 100% of Pfiesteria biomass.

Stoecker, Stevens, and Gustafson (2000), Stoecker and Gustafson (2002), Stoecker et al. (2002), and Gransden and Lewitus (2003) obtained very similar results stressing that microzooplankton grazing could potentially preclude growth of nontoxic Pfiesteria zoospores in natural settings.

2.4.12. Interactions with Mesozooplankton

Similar to microzooplankton, mesozooplankton are prolific inhabitants of eutrophic coastal marine waters (Reaugh, Roman, and Stoecker 2007; Stoecker, Thessen, and Gustafson 2008), and copepods are the most abundant mesozooplankton group in these types of environments

(Kimmel and Roman 2004; Roman et al. 2005; Calliari et al. 2006). Temperature is the most important environmental variable that influences mesozooplankton population dynamics and, because of their preference for warmer environments (Miller, Johnson, and Heinle 1977;

McLaren and Corkett 1981; Kimmel and Roman 2004), the most dynamic growth of their populations occurs during summer and early fall (Dolan and Coats 1990; Kiørboe and Nielsen

1994). Copepods feed on a wide array of prey (Kleppel 1993; Jones and Flynn 2005; Turner et al. 2005) and are, in turn, nutritional staple for larger zooplankton (Calliari, Britos, and Conde

2009) and numerous invertebrates and finfish (Porter, Pace, and Battey 1979; Sommer et al.

2002; Hirst and Bunker 2003). Mesozooplankton also prey on toxic dinoflagellates (Collumb and

Buskey 2004; Breier and Buskey 2007; Buskey 2008). By ingesting toxic dinoflagellates,

21

mesozooplankton, and mostly copepods, can influence population dynamics of these organisms

and—potentially—suppress development of their blooms (Jeong 1994; Stibor et al. 2004; Breier and Buskey 2007). A number of copepods can prey on Pfiesteria (Burkholder and Glasgow

1995; Mallin et al. 1995; Roman, Reaugh, and Zhang 2006). Roman, Reaugh, and Zhang (2006)

hypothesized that abundant populations of A. tonsa might influence population dynamics of

TOX-B and NON-IND Pfiesteria strains in natural setting and preclude their blooms.

2.4.13. Interactions with Bivalve Mollusks

Similar to a number other species of toxic dinoflagellates (Contreras, Marsden, and Munro 2011;

Saeedi, Kamrani, and Matsuoka 2011), Pfiesteria can: a) intoxicate and kill bivalve mollusks and b) constitute prey of these animals. Studies that investigated interactions of Pfiesteria with bivalve mollusks concentrated on three molluscan species: eastern oyster ( Crassostrea virginica ), hard clam ( Mercenaria mercenaria ), and bay scallop ( Aequipecten irradians ).

2.4.13a. Eastern Oyster ( Crassostrea virginica )

The eastern oyster is a resilient bivalve mollusk, well adapted to fluctuating environmental variables (Sellers and Stanley 1984; Kennedy 1996; Gobler, Peterson, and Wall 2009). It feeds on a wide variety of phytoplankton (Kennedy 1996; Baldwin and Newell 1991; Coen et al. 2007) and, similar to Pfiesteria , is most active during summer and early fall (Sellers and Stanley 1984;

Shumway, Burkholder, and Springer 2006; Gobler, Peterson, and Wall 2009).

Toxic dinoflagellates can negatively affect and even kill early life stages of the eastern oyster (Shumway and Cucci 1987; Mulholland et al. 2009; Tang and Gobler 2009a). Wikfors

(2005) established that larval eastern oyster fail to metamorphose in the presence of actively

22

toxic P. minimum . Tang and Gobler (2009a) demonstrated that C. polykrikoides and K. brevis

killed oyster larvae. In a study by Stoecker et al. (2008), high mortality rates of oyster larvae occurred upon exposure to K. veneficum . Toxic dinoflagellates can also kill oyster juveniles. In

Mulholland, Boneillo, Minor (2004), 20% of oyster juveniles died within 72 hours of exposure to actively toxic C. polykrikoides . Luckenbach et al. (1993) demonstrated 100% mortality of juvenile oysters in the presence of P. minimum . In Smolowitz and Shumway (1997), eastern oyster juveniles died when cultured with G. sanguineum . Pfiesteria can also kill early eastern oyster larvae and juveniles. According to Springer (2000), Springer et al. (2002), and Shumway,

Burkholder, and Springer (2006), actively toxic P. piscicida are drawn to oyster tissue through chemosensory attraction in a fashion analogous to their attraction to finfish tissue. In Springer et al. (2002) oyster larvae and juveniles were affected negatively by toxins released by TOX-B

Pfiesteria cells cultured with no prior access to life finfish, contrary to results of earlier studies which stressed the importance of finfish presence during toxic Pfiesteria outbreaks. According to Burkholder et al. (1995) and Springer (2000), larval eastern oysters can be killed by actively toxic Pfiesteria within minutes of exposure. Shumway, Burkholder, and Springer (2006) found that juvenile and larval eastern oyster oysters can be intoxicated and killed by Pfiesteria toxins while isolated from Pfiesteria cells by a dialysis membrane. According to Springer et al. (2002) and Shumway, Burkholder, and Springer (2006), Pfiesteria might negatively affect eastern oyster early life stages in natural settings.

Unlike larvae and juveniles, adult oysters can ingest toxic dinoflagellates. Hégaret et al.

(2007) noted predatory behavior of eastern oyster toward A. fundyense . In this study, oysters accumulated A. fundyense toxins in their tissues but displayed no signs of intoxication. Leverone,

Shumway, and Blake (2007) found that out of four mollusk species tested, eastern oysters were

23

able to ingest the largest proportion of K. brevis cells while suffering no ill health effects. May et al. (2010) noted no mortalities of adult oysters after 24-hour exposure to A. monilatum . Adult oysters also ingest toxic Pfiesteria cells (Burkholder and Glasgow 1997; Burkholder and

Glasgow 2001). Burkholder, Glasgow, and Hobbs (1995), noted that no mortalities of adult oysters occurred after three weeks of persistent ingestion of actively toxic Pfiesteria cells.

2.4.13b. Hard Clam ( Mercenaria mercenaria )

Similar to eastern oysters, hard clams are well adapted to environmental variability (SSERC

1999a). Temperature is the key factor regulating metabolism and reproductive success of these animals (Lorio and Malone 1995); therefore, their highest activity occurs during summer and early fall (SSERC 1999; USACE 2004; Bricelj 2009). Hard clams feed primarily on phytoplankton (Langdon and Newell 1996; Greenfield et al. 2004) and thrive in eutrophic environments abundant in it (Weiss, Carmichael, and Valiela 2002; Carmichael, Shriver, and

Valiela 2004; Gobler, Peterson, and Wall 2009).

Toxic dinoflagellates can cause Ill health and mortalities of hard clams. Tang and Gobler

(2009a) noted significant mortality rates of larval clams in the presence of C. polykrikoides . In addition to lengthened development time, K. brevis precipitated mortality of larval clams

(Leverone et al. 2006). At times, hard clams burrow in sediment upon exposure to toxic dinoflagellates (Shumway 1990). According to Vigfúsdóttir et al. (2004), M. mercenaria engaged in such behavior of upon exposure to blooming K. micrum . In a number of instances, hard clams suffered no ill effects after ingestion of toxic dinoflagellates. May et al. (2010) found that 100% of their M. mercenaria specimens survived 24-hour exposure to actively toxic A. monilatum . Similar to bay scallops (Springer 2000) and eastern oysters (Springer 2000; Springer

24

et al. 2002; Shumway, Burkholder, and Springer 2006), Pfiesteria are drawn to larval and juvenile M. mercenaria through chemosensory attraction, and attack and ingest these bivalve mollusks (Springer 2000, Shumway, Burkholder, and Springer 2006). Adult hard clams, on the other hand, can ingest these toxic dinoflagellates with no adverse health effects (Burkholder,

Glasgow, and Hobbs 1995; Springer et al. 2002; Shumway, Burkholder, and Springer 2006).

According to Springer et al. (2002), grazing activity of adult hard clams ingesting toxic cells of

P. piscicida was unaffected. Burkholder, Glasgow, and Hobbs (1995) documented no mortalities of adult hard clams after three weeks of continuous feeding on of Pfiesteria TOX-A cells.

2.4.13c. Bay Scallop ( Aequipecten irradians )

Bay scallops inhabit shallow estuaries and near-shore, low-turbulence coastal waters (Fay,

Neves, and Pardue 1983), abundant in dissolved nutrients (Shriver, Carmichael, and Valiela

2002) and phytoplankton prey (Fay, Neves, and Pardue 1983; Kirby-Smith 1970; Shriver,

Carmichael, and Valiela 2002). Adult bay scallops are characterized by a wide temperature (Fay,

Neves, and Pardue 1983) and salinity (Shumway, Burkholder, and Springer 2006) tolerance.

Although their activity can occur at temperatures as low 0 °C, scallops spawn and are most active during summer and early fall (Fay, Neves, and Pardue 1983; Springer et al. 2002;

Shumway, Burkholder, and Springer 2006) coinciding with the highest activity of Pfiesteria species (Springer 2000; Springer et al. 2002; Shumway, Burkholder, and Springer 2006).

Similar to hard clams and eastern oysters, bay scallop adults can ingest toxic dinoflagellates. Hégaret et al. (2007) found that scallop feeding rates increased when fed cells of toxic A. fundyense . Yan et al. (2003) demonstrated that survival of the scallop larvae was not affected when the mollusks were grown with A. tamarense at concentrations of 500–10,000

25 cells/ml for 48 hours. Other dinoflagellates can harm and kill bay scallops (Tang and Gobler

2009; FFWCC 2014). Wikfors and Smolowitz (1993) demonstrated rapid mortality of bay scallops exposed to actively toxic P. minimum . Smolowitz and Shumway (1997) stressed high susceptibility of these bivalve mollusks to G. aureolum . Tang and Gobler (2009) found that C. polykrikoides were able kill 99% of bay scallop larvae in a 10-hour period. Leverone et al.

(2006) demonstrated lengthened development time and reduced survival of scallop larvae exposed to actively toxic K. brevis . In Leverone, Shumway, and Blake (2007), toxins of K. brevis were responsible for a significant reduction of scallop feeding rates. Pfiesteria can also negatively affect bay scallops. According to Springer (2000) Pfiesteria zoospores are drawn to adult scallop tissue through chemosensory attraction, while Springer et al. (2002) demonstrated development of similar dynamic between Pfiesteria and scallop larvae. Burkholder et al. (1993) noted that bay scallops did not stimulate Pfiesteria to emerge from their cysts and initiate their toxic activity. The authors pointed out, however, that after Pfiesteria -related 1992 scallop mortality event at National Marine Fisheries Service (NMFS) facility in Beaufort, Miami, a hypothesis was formulated which states that the presence of finfish is not a prerequisite for the initiation of toxic Pfiesteria activity and their predatory behavior towards bay scallops. Pfiesteria zoospores swarm around bay scallops, attack and consume them (Springer 2000; Springer et al.

2002; Shumway, Burkholder, and Springer 2006). In Burkholder et al. (1992), Pfiesteria killed all adult bay scallops in less than 20 minutes of exposure. Mortalities occurred even when actively toxic Pfiesteria cells were separated from scallop prey by a dialysis membrane (Springer

2000; Springer et al. 2002). Adverse health effects resulting from contact with Pfiesteria cells and their toxins would result even at low concentrations of these dinoflagellates.

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2.4.14. Interactions with Other Toxic Dinoflagellates

Although published literature does not consider possible interactions between Pfiesteria and

other species of toxic dinoflagellates as one of the determinants of Pfiesteria population

dynamics, to determine whether Pfiesteria could acclimatize and establish self-sustaining populations in a particular environment, it seems worthwhile to investigate potential nature of such interactions to establish whether activity of these organisms could preclude establishment of viable Pfiesteria populations. Studies have shown that toxic dinoflagellates can ingest other toxic

dinoflagellate species (Jeong and Latz 1994; Matsuyama, Miyamoto, and Kotani 1999; Jeong et

al. 2001). In Casco Bay, Maine, Turner et al. (2005) found that dinoflagellates contained toxins

of other dinoflagellate species. In Rodríguez, Ochoa, and Alcocer (2005) N. scintillans ingested

toxic G. catenatum . G. catenatum were grazed by P. kofoidii in Matsuyama, Miyamoto, and

Kotani (1998). Nakamura, Suzuki, and Hiromi (1995) demonstrated that G. dominans and G.

spirale were unaffected by toxins of G. mikimotoi ; population sizes of both dinoflagellate grazers

increased while feeding on actively toxic G. mikimotoi . Pfiesteria can co-occur with K.

veneficum ( Burkholder and Marshall 2012). In experiments by Jeong et al. (2006), P. piscicida

ingested all naked mixotrophic dinoflagellates, including the smallest thecate dinoflagellate H.

rotundata . Pfiesteria can also ingest C. polykrikoides , a highly toxic species, which can co-occur

with other dinoflagellate species even at blooming concentrations (Burson 2009).

3. Status of Physicochemical and Biological Determinants of Pfiesteria Population

Dynamics in Great South Bay Prior to Superstorm Sandy

GSB, similar to other embayments of LI’s South Shore (Taylor, Gobler, and Sañudo-Wilhelmy

2006), is—as Paerl (1988) would designate it—“bloom sensitive" (in Nuzzi and Waters 1989,

27

117). Prior to the passage of superstorm Sandy in October 2012, dinoflagellates were an

abundant group of microorganisms in GSB (Lonsdale et al. 2006; Dooley 2013a), constituting

the second main phytoplankton group active in this system (Carpenter, Brinkhuis, and Capone

1991). During spring, a number of streams and rivers emptying into GSB experienced blooms of

H. triquetr a (Carpenter, Brinkhuis, and Capone 1991). With the arrival of summer, blooms of A.

fundyense (Branca and Focazio 2006; Hattenrath, Anderson, and Gobler 2010), G. aureolum

(Chang and Carpenter 1985; Carpenter, Brinkhuis, and Capone 1991), and C. polykrikoides

(Gobler et al. 2008; Kelley 2012; SSERC 2012) affected this system.

3.1. Temperature

Prior to the passage of superstorm Sandy, temperature in GSB was determined by seasonal

variations in solar radiation, rates at which ocean water flows into the bay through Fire Island

Inlet, and the rate of terrestrially derived freshwater flux (USACE 1999). Summer temperatures in this coastal system would generally reach 25-26 °C (Hinga 2005) exhibiting only minimal variation (USACE 1999, 2004a). Because of its proximity to the Atlantic Ocean and its cooling effect, extreme temperature variability in GSB was rare), although not uncommon (Hinga 2005).

Weiss et al. (2007) recorded 29.77 °C during August. One temperature measurements collected in GSB by Sebastiano (2012) during July was ≈30 °C. During the heatwave that affected New

York and other East Coast states in the summer of 1977, daily average temperatures hovered above 32.2 °C for nine consecutive days, while three of these experienced daily high temperatures in excess of 38.8 °C (Greene et al. 1978). On July 21 of that same year temperature climbed to 40 °C (Greene et al. 1978). High summer temperatures have been also common in tributaries emptying into GSB. One of the temperature measurements obtained by O’Malley

28

(2008) in the lower segment of the Carmans River was 30.8 °C. Figure 3.1 depicts thermal suitability of coastal waters of Eastern United States to positive growth of Pfiesteria populations during July while Figure 3.2 demonstrates high summer temperatures in GSB.

Figure 3.1. Average July Sea Surface Temperature Readings along the U.S East Coast (NEFSC 2013).

29

Figure 3.2. Temperature Measurements at Bellport and Blue Point, GSB (Duval 2008).

3.2. Salinity

Prior to the passage of superstorm Sandy GSB salinities were determined by freshwater runoff

and inflow of marine water through the Fire Island inlet (Bokuniewicz an Schubel 1991; Schubel

1991; Wilson, Wong, and Carter 1991; USACE 1999). The result of this arrangement was that

GSB salinity levels were substantially lower than salinity level of the adjacent coastal ocean

(Bokuniewicz and Schubel 1991). According to Moskowitz (1976), the average salinity of GSB was 20% lower than the ocean. GSB and its tributaries were characterized by a relatively wide range of salinities (Ryther 1954), generally falling within 25-30-ppt range (Hinga 2005) while average salinity of GSB was ≈25.9 ppt (Tanski, Bokuniewicz, and Schlenk 2001; USACE 2004a;

30

Hinga 2005); GSB salinity did not demonstrate marked variations (Schubel 1991; Bricelj 2009).

Most of salinity measurements collected by Hinga (2005) fluctuated only 2 ppt of the median salinity of this system. Figure 3.3 demonstrates the pre-Sandy GSB salinity profile.

Figure 3.3. GSB Salinity (USACE 2004a).

3.3. Dissolved Oxygen Concentration

DO concentration standard for marine waters of New York State is 5.0 mg/L, a DO

concentration which has been demonstrated to afford complete safeguard to marine life (SSERC

1999c). Prior to the passage of superstorm Sandy over GSB in October 2012, DO concentrations

in this system were unlikely to decline to levels harmful to its resident and transient organisms

(Redfield 1952; Dennison, Koppelman, and Nuzzi 1991; Bokuniewicz et al. 1993; SSERC 2000;

NYSDOS 2001; USACE 2004; Hinga 2005). Occasional dips would have occurred, albeit rarely

31

(SSERC 1999c). Out of 5487 DO measurements collected by the Suffolk County Department of

Health Services (SCDHS) in Great South Bay and Hempstead Harbor between 1976 and 2004 only 42 are below hypoxic concentration of 4 mg/L (McElroy et al. 2009). In fact, frequently,

waters of GSB would have been nearly saturated with DO and rarely fallen by more than 10% of

that concentration (Hinga 2005). On occasion, GSB streams and rivers may have become

hypoxic (Foehrenbach 1969; Dennison, Koppelman, and Nuzzi 1991); however, low DO

conditions in these environments would have been sporadic (SSERC 1999c). In a number of

GSB tributaries, DO concentrations would be as high as 15.1 mg/L (USACE 1997). DO concentrations in Patchogue, Swan, and Carmans Rivers sampled by Zaikowski (2007) were over 6 ml/L, often approaching complete saturation. DO data collected by the Suffolk County

Department of Health Services (SCDHS) in the Beaver Dam Creek point to a similar condition.

While close to 80% of SCDHS DO measurements in this creek were lower than the New York

State DO freshwater standard, only 5.4% of these fell below the established marine standard

(SCDHS 2008). Figure 3.4 depicts generally high DO concentrations in GSB prior to the passage of superstorm Sandy over this system in October 2012.

32

Figure 3.4. GSB Dissolved Oxygen Concentrations (USACE 2004a).

3.4. pH

According to the United States Army Corps of Engineers (USACE), prior to superstorm Sandy’s

crossing of coastal LI in October 2012, waters of GSB were weakly alkaline, (within the 7.8-8.6

range; USACE 1997). Leigh-Manuell, Hanlon, and Larado (2010) studied relationships between

GSB’s pH levels, DO concentrations, and temperature. It was demonstrates that pH did not

fluctuate significantly during experimental proceedings and was always ≈7 (Figure 3.5 ).

33

Figure 3.5. GSB pH Profile (Leigh-Manuell, Hanlon, and Larado 2010).

pH was also sufficiently high in tributaries emptying into GSB. According to SCDHS, average pH levels did not vary demonstrably between the upriver freshwater reaches of the Beaver Dam

Creek and its downriver tidal section (SCDHS 2008). One of the highest pH measurements collected by O’Malley (2008) in the Carmans River was 9.36.

3.5. Turbulent Mixing

According to a rather limited number of publications which investigated the character of turbulent mixing in GSB, prior to the passage of superstorm Sandy in October 2012, tidal range

(tidal amplitude) in this system was low (Bokuniewicz and Schubel 1991; USACE 1997, 2004a) and, consequently, tides exerted rather limited influence on the degree of turbulent mixing in

GSB (Figure 3.6 and 3.7).

34

Figure 3.6. GSB Tidal Range (USACE 2004a).

Figure 3.7. Tidal Current Velocities in GSB (USACE 2004a). 35

Wind action, on the other hand, was a significant force influencing the magnitude of turbulent

mixing in GSB. Prevailing winds of the summer months are from the west (SSERC 1999a) and the southwest (Greene et al. 1978), and because the long axis of GSB is oriented parallel to the prevailing wind direction, prior to the passage of superstorm Sandy over this system in October

2012, GSB was a well-mixed system (Nuzzi and Waters 1989; SSERC 1999a).

3.6. Availability of Dissolved Nutrients

Despite considerable nutrient retention potential within its subwatersheds (Kinney and Valiela

2011; Sebastiano 2012), prior to the passage of superstorm Sandy, GSB was abundant in nutrients (Carpenter, Brinkhuis, and Capone 1991; Dennison, Koppelman, and Nuzzi 1991;

Schubel 1991; SSERC 1999; Cosper 2001; SSERC 2012) (Table 3.1).

36

Table 3.1. Eutrophication Status of GSB and Other U.S. Coastal Marine Systems (Bricker et al. 1999).

In 1999, LI Comprehensive Waste Treatment Management Plan estimated total daily flux of dissolved nutrients to this system to be 6,600 lbs. (SSERC 1999). The main sources of nutrients

37

to GSB were surface runoff, submarine groundwater discharge, and atmospheric deposition

(SSERC 2000) (Table 3.2).

Table 3.2. Respective Contribution of Main Nitrogen Sources to GSB (Kinney and Valiela 2011).

The largest source was wastewater (Gobler 2010) as only a limited number of households in

Suffolk County had access to centralized sewage treatment (Munster 2004); only ≈30% of these residences were linked up with sewage treatment plants in 2014 (Abbatecola 2014; Schwartz

2014; Dowdy 2013). The majority of unsewered households treated their wastewater in

traditional septic/cesspool systems (Kinney and Valiela 2011; Monti and Scorca 2003; Suffolk

County 2014). Onsite Wastewater Treatment Systems (OWTS) are particularly prevalent in the

eastern portion of the watershed (SSERC 2000; Kelley 2012). Considering high N and P concentrations in human excrement (Weiskel and Howes 1992), N was often highly concentrated in groundwater beneath areas of GSB watershed where large number of OWTS were utilized

38

(Monti and Scorca 2002; Reay 2004). Between 1987 and 2005 concentration of N in the Upper

Glacier aquifer and Magothy aquifer increased by 40% and 70% respectively (Suffolk County

2014). Runoff of excessively applied lawn fertilizers is also a significant source of dissolved

nutrients to GSB (Munster 2004; Portmess and Petrovic 2010; Kinney and Valiela 2011). In a number of instances, as much as 50% of total fertilizer that is applied to turf infiltrated aquifers and polluted groundwater (Flipse et al. 1984 in Portmess and Petrovic 2010). According to

Kinney and Valiela (2011), turf fertilizer is the third major source of nitrogenous nutrients to

GSB. Contribution of stormwater runoff, which can be a significant nutrient source to coastal marine systems (Howarth, Sharpley, and Walker 2002; Hunter 2003; Glasoe and Christy 2004), was high in GSB as well (deQuillfeldt 2013).

Concentrations of dissolved organic nutrients—which, as it have been pointed out earlier, stimulate the most prolific growth of Pfiesteria populations (Zhang et al. 2004; Glibert et al.

2006; Hood et al. 2006)—were also high in GSB (Ryther 1954; Gobler et al. 2005; Lonsdale et al. 2006) before superstorm Sandy affected this system in October 2012. Lomas et al. (2001) found concentrations of organic nutrients in coastal embayments of LI to be up to 20 times higher than the concentrations of inorganic nutrients. According to Lonsdale et al. (2006), organic nitrogenous and phosphorous nutrients comprised 92% and 100% respectively of total

GSB nutrient pool. Furthermore, recurrent blooms of A. anophagefferens and C. polykrikoides in

GSB also attest to ample supply of organic nutrients in this system given the affinity of these organisms for organic forms of nitrogenous nutrients (Caron et al. 2004; Gobler, Lonsdale, and

Boyer 2005).

3.7. Availability of Finfish Prey

39

Despite the overall decline (Carson 1945; USACE 1996; Sagarese 2009; Nuttall 2010), prior to

the passage of superstorm Sandy, finfish populations in GSB were viable (Dowhan et al. 1997;

SSERC 1998; Cashin Associates 2007) and biologically diverse (McElroy et al. 2009); the bay

had been considered an “excellent fish habitat” (Mc Elroy et al. 2009, 18); over 150 species

frequented embayments and tributaries of GSB (Carson 1945; Greene et al. 1978; Shima and

Cowen 1989; SSERC 1998; Cosper 2001; McElroy et al. 2009). Fishing was an important economic activity in GSB (Dowhan et al. 1997; Kelley 2012). Species of economic importance included Atlantic menhaden ( Brevoortia tyrannus ), blackfish ( Tautoga onitis ), black sea bass

(Centropristis striata ), bluefish ( Pomatomus saltatrix ), butterfish ( Peprilus triacanthus ), scup

(Stenotomus chrysops ), striped bass ( Morone saxatilis ), and summer flounder ( Paralichthys dentatus ) (Carson 1945; USACE 1996; SSERC 1998). Kritzer, Hughes, and O’Reilly (2007) noted significant numbers of alewives ( Alosa pseudoharengus ) in the estuarine section of the

Carmans River. Nine of the twelve LI wild brook trout populations occurred in watercourses draining into GSB (Dowhan et al. 1997; Regional Plan Association 2013). Swan River, for example, is known nationwide for its high abundance of this species (Dowhan et al. 1997;

Cashin Associates 2007). Regional significance of wild brook trout recreational fishery has been recognized for the tidal section of the Carman River (Nelson, Pope & Vooris 2008). The U.S.

Fish and Wildlife Service (USFWS) have included the Champlin and Orowoc Creeks, and the

Swan, Connetquot, and Carman Rivers in their Significant Coastal Fish and Wildlife Habitat

(SCFWH) inventory database (USACE 1996; SSERC 1999b; Cashin Associates 2007; Town of

Brookhaven 2013). Populations of forage species such as bay anchovy and Atlantic silverside

(Shima and Cowen 1989; Castro and Cowen 1989; Bokuniewicz et al. 1993; Dowhan et al. 1997;

USACE 1996) were also numerous in GSB prior to the passage of superstorm Sandy in October

40

2012. In a study conducted by Juanes and Conover (1995), Atlantic silversides and bay

anchovies consistently constituted over 50% of sampled finfish individuals in this system. In a study of GSB commercial fisheries, Nuttall (2010) pointed out that as population size of some species declined over the last several decades, abundance of others species increased. For instance, both commercial and recreational catches of weakfish increased through the 2000s

(Nuttall 2010). According to trawl studies conducted in GSB, bluefish yields increased twofold between 1980 and 2000s (Nuttall 2010). Trawl surveys have also demonstrated fivefold increase in scup catches between 1981 and 2007 (Nuttall 2010). Recreational catches of some species

(black sea bass, for instance) were also sizable (Nutall 2010). Figures 3.8 through 3.10 depict viability of selected finfish catches in GSB prior to the passage of superstorm Sandy.

Figure 3.8. Recreational Catches of Black Sea Bass in GSB (Nuttall 2010).

41

Figure 3.9. Commercial and Recreational Catches of Northern Kingfish in GSB (Nuttall 2010).

Figure 3.10. Commercial and Recreational Bluefish Catches in GSB (Nuttall 2010).

3.8. Availability of Algal Prey

42

Rublee et al. (2006) found that the incidence of Pfiesteria cells in GSB was correlated with high concentrations of chlorophyll a. Indeed, prior to superstorm Sandy, abundant nutrients

(Carpenter, Brinkhuis, and Capone 1991; Dennison, Koppelman, and Nuzzi 1991; Schubel 1991;

SSERC 1998, 1999; Cosper 2001) in tandem with high magnitude of photosynthetically active radiation (PAR) (Nuzzi and Waters 1989) rendered GSB abundant in algae (Schubel 1991;

SSERC 1998, 1999a 1999c; Lonsdale et al. 2006; Newell et al. 2009), with the most prolific growth and highest abundance of these organisms occurring during summer and early fall

(Schubel 1991; Hinga 2005). Indeed, according to Lively, Kaufman, and Carpenter (1983) concentrations of algal cells and algae population growth rates in GSB were about three times higher during July and August than during April and May. Annual primary productivity rates in the GSB measured by Carpenter, Brinkhuis, and Capone (1991) in the early 1980s were one of the highest ever recorded in temperate marine environments. Diatoms dominated bay’s planktonic community during winter (Lively, Kaufman, and Carpenter 1983). During summer, smaller chrysophytes and chlorophytes would predominate (Carpenter, Brinkhuis, and Capone

1991; Lonsdale et al. 1996, 2006; Bricelj 2009; Newell et al. 2009) and primary productivity would achieve maximum rates (Bokuniewicz et al. 1993; Boissonneault-Cellineri et al. 2001;

Weiss et al. 2007).

3.9. Interactions with Bacteria

Prior to the severe weather event of October 2012, GSB bacteria (Caron et al. 1989;

Boissonneault-Cellineri 1995; Lonsdale et al. 2006), including alpha- and gammaprotobacteria

(Kelly and Chistoserdov 1997) strains which interact with Pfiesteria symbiotically and facilitate production of their toxins (Tengs et al. 2003; Alavi et al. 2001; Miller and Belas 2004), were

43

abundant in GSB. Majority of these organisms would be transported into its waters with

terrestrial freshwater discharges, namely stormwater runoff (SCDHS 2008; Kelly and

Chistoserdov 2001; deQuillfeldt 2013). Approximately 90% of the total load of bacteria to the

GSB would enter GSB with stormwater runoff (SSERC 1999). Photosynthetic Synechococcus cyanobacteria, which Pfiesteria prey upon (Burkholder and Glasgow 1997), was an important constituent of GSB bacterioplankton (Caron et al. 2004, Sieracki et al. 2004; Bricelj 2009).

Bacterial populations of tributaries draining into GSB were also large prior to the passage of

Sandy. For instance, measurements collected by the Suffolk County Department of Health

Services (SCDHS) between September 2002 and November 2003 shown elevated concentrations of total and fecal coliform bacteria in large sectors of Beaver Dam Creek (SCDHS 2008).

3.10. Interactions with Microzooplankton

Summer microzooplankton populations were abundant (Castro and Cowen 1991; Bokuniewicz et al. 1993; Lonsdale et al. 1996; Gobler, Renaghan, and Buck 2002; Lonsdale et al. 2006) and biologically diverse (USACE 1997; Cosper 2001) in GSB prior to the passage of superstorm

Sandy. It has been suggested that such abundance and diversity was in part due to the general decline of GSB bivalve mollusk populations which, being filter-feeders, controlled the abundance of and other microzooplankton groups (Deonarine et al. 2006; Pan 2010). In their study of A. anophagefferens population dynamics in GSB’s Bay Shore Cove and Patchogue

Bay, Gobler, Renaghan, and Buck (2002) found considerable ingestion of this alga by microzooplankton. Copepod nauplii, which are considered microzooplankton rather than mesozooplankton because of their relative size, also constituted major component of GSB

44

microzooplankton communities in GSB before superstorm Sandy affected this system (Lonsdale

et al. 1996; Castro and Cowen 1991; Smith et al. 2008).

3.11. Interactions with Mesozooplankton

According to a rather limited number of procured publications, prior to the passage of

superstorm Sandy, GSB summer mesozooplankton populations were abundant and typically

dominated by two calanoid copepods, Acartia tonsa and A. hudsonica (Chang and Carpenter

1985; Lonsdale et al. 1996; Cosper 2001; Smith et al. 2008; Bricelj 2009).

3.12. Interactions with Eastern Oysters ( Crassostrea virginica )

GSB eastern oyster populations have suffered substantial losses and low rates of population growth (Bubolo 1985; LoBue and Udelhoven 2013) (Figure 3.11).

Figure 3.11. General Decline of Bivalve Mollusk (Eastern Oyster and Hard Clam) and Menhaden Fisheries in GSB (Nuttall 2011). 45

Commercial catches of this bivalve mollusk in GSB were substantial during the 1800’s but

collapsed in 1893 due to overfishing (Nuttall 2010). Reasonable catches were maintained with

transplanted seeds (Nuttall 2010). Between the early 1800s and 1931, oyster populations in

eastern GSB increased in size, however, due to low salinities that prevailed in the bay during this

period, at significantly reduced rates (Cashin Associates 1993). In GSB’s western parts salinities

were more suitable to oyster growth, however, being located nearer the tidal inlet and the coastal

ocean, transplanted oyster seeds were decimated by oyster drills ( Urosalpinx cinerea ) and other predators (Cashin Associates 1993). In 1931, Nor’easter storm reopened Moriches Inlet (Conley

1999; SSERC 1999a). As salinity of both Moriches Bay and eastern GSB increased, both systems became more hospitable to oyster predators which nearly annihilated remaining o yster

population (Cashin Associates 1993; Tanski, Bokuniewicz, and Schlenk 2001; Nuttall 2010;

Nuttall et al. 2011). Fishery stagnated during the 1930s and 1940s (Nuttall 2010; Nuttall et al.

2011). In 1952 and 1953 picoplanktonic green tide chlorophyte species Nannochloris atomus and

Stichococcus sp. bloomed to extreme concentrations (Ryther 1954; Foehrenbach 1969; Ryther

1989; Newell 2004). Although not considered toxic, monospecific blooms of these two species negatively impacted GSB oyster populations as their low nutritional value and small size (2-4 um in diameter) precluded efficient grazing of these species by oysters (Ryther 1989). Bokuniewicz et al. (1993) pointed out that scarcity of suitable food led to a substantial decline of eastern oyster populations in GSB; the fishery collapsed in mid 1950s (Nuttall 2010; Nuttall et al. 2011). The remaining oyster populations plummeted further in the mid-1980s due to brown tide events

(Bricelj and Lonsdale 1997; Gobler et al. 2008) and an outbreak of H. nelsoni (MSX)—a spore- forming protozoan known to seriously affect oyster health (Burreson, Stokes, and Friedman

46

2000; Hofmann et al. 2001)—in 1998 (Harvell et al. 1999). According to Pan (2010), because of its current environmental standing, viability of eastern oyster populations in GSB has been seriously compromised and its former abundance is unlikely to be restored.

3.13. Interactions with Hard Clams ( Mercenaria mercenaria )

In the past, hard clam was a numerous and economically important species of bivalve mollusk in

GSB (Bubolo 1985; Bricelj and Lonsdale 1997; SSERC 1999a; Greene 2009; Weiss et al. 2009);

the bay used to supply ≈90% of the New York and 45% of the hard clam harvests in the United

States (Flagg and Malouf 1983; Suffolk County 2014). During 1960s and 1970s, GSB was

considered the “clam factory” (Brumbaugh et al. 2006, 17) and the most significant producer of

hard clams in the world (Carter, Wong, and Malouf 1984). Annual commercial harvests averaged

3248 metric tons meat weight between 1970 and 1979, to decline by 97% to an average of 91

metric tons a year between 1996 and 2007 (NYSDEC 2008; Newell et al. 2009). This decline

has resulted because of overfishing (Bubolo, 1985; Bricelj 2009; Bricelj and Lonsdale 1997;

LoBue and Udelhoven 2013), brown tide events (Bricelj and Lonsdale 1997; Greenfield et al.

2004; Greenfield, Lonsdale, and Cerrato 2005), and other causes (Zaikowski et al. 2007).

Furthermore, rates at which remaining GSB hard clam populations would grow was rather low

(Greenfield 2002; Greenfield et al. 2004). Doall et al. (2008), for instance, stocked up a number

of sites in the central part of GSB with individuals transplanted from Long Island Sound. During

the 4 years of the study, value of the condition index (CI), which measures the rate of hard clam

growth (Crosby and Gale 1990), declined during summer months (Doall et al. 2008). Similarly,

Weiss et al. (2007) noted that hard clams transferred to GSB in 2004 and 2005, when brown

tides were absent, were characterized by considerably lower value of the CI index than

47 individuals transplanted to Shinnecock Bay. Individuals collected in the central and eastern parts of the bay by Newell et al. (2009) were significantly smaller than individuals collected in the western part despite high concentrations of organic nutrients and algal prey in these waters.

Newell (2004) and Newell et al. (2009) attributed slow growth and limited reproductive success of these animals to high contribution of picoplankton to the total algal community of GSB.

According to a number of studies, picoplankton have been the most abundant ( ≈50%) phytoplankton group in waters of GSB (Carpenter, Brinkhuis, and Capone 1991; Lonsdale et al.

1996; Boissonneault-Cellineri et al. 2001; Greenfield, Lonsdale, and Cerrato 2005; Lonsdale et al. 2006; Bricelj 2009; Newell et al. 2009). Because picoplankton is too small to be grazed by hard clams effectively, these organisms do not constitute suitable source of nourishment for these mollusks (Bricelj 2009). Although hard clam populations are still sufficiently abundant to support recreational fishery in a number of locations throughout GSB (Nature Conservancy

2011), all restoration efforts of once successful commercial hard clam fishery have failed thus far

(LoBue and Botman 2011). No recovery of these bivalve mollusks has occurred (Carroll, Gobler, and Peterson 2008; Bricelj 2009; Nuttall 2010; Suffolk County 2014) despite limited outbreaks of A. anophagefferens (Bricelj 2009) and significantly reduced harvest pressure (LoBue and

Bortman 2011). Figure 3.12 demonstrates the precipitous decline of commercial hard clam catches in GSB.

48

Figure 3.12. Decline of Commercial Hard Clam Yield in GSB (Kraeuter et al. 2008).

3.14. Interactions with Bay Scallops ( Aequipecten irradians )

Bay scallops, similar to hard clams, used to be numerous in GSB (Rather 2008). According to a number of studies, recurring blooms of A. anophagefferens have been the predominant cause of

the collapse of bay scallop fishery in GSB and a number of other LI embayments (Bricelj and

Lonsdale 1997; Gobler et al. 2005; Suffolk County 2014) (Figure 3.13).

49

Figure 3.13. Decline of Commercial Bay Scallop Catch in GSB (Suffolk County 2014).

Besides devastating effects brown tide events, the decline of GSB bay scallop fishery was linked

to the decline of submerged aquatic vegetation (SAV) beds in this system. Numerous studies

have demonstrated that bay scallops utilize SAV meadows preferentially as habitat (Fay, Neves,

and Pardue 1983; Pohle, Bricelj, and Garcia-Esquivel 1991; Shriver, Carmichael, and Valiela

2002). Blooms of A. anophagefferens decimated GSB SAV beds (Gobler et al. 2005; Gobler

2008), thwarting recruitment of scallop juveniles and, consequently, recovery of bay scallop populations in GSB (Tettelbach and Wenchel 1993; Bricelj and Lonsdale 1997; Bricelj,

MacQuarrie, and Smolowitz 2004). Various projects initiated in GSB to restore these once abundant populations have been unsuccessful thus far (Carroll, Gobler, and Peterson 2008; Pan

2010).

50

According to the simple, inventory-like analysis of the pre-Sandy status of the physicochemical

and biological factors that determine Pfiesteria population dynamics in GSB, prior to Sandy, this system appears to have constituted an environment suitable to Pfiesteria bloom formation

with respect to temperature, salinity, DO concentrations, pH, availability of dissolved nutrients,

co-occurrence of other toxic dinoflagellate species, abundance of finfish, bacterial and algal

prey, as well as interactions with bivalve mollusks. Given stochastic nature of turbulent mixing

in coastal marine environments and the paucity of published studies on a) the effects of turbulent

mixing on food webs in GSB, and b) interactions between zooplankton and toxic dinoflagellates

in this system, examination of the determinants in question—turbulent mixing and interactions

with micro- and mesozooplankton—did not lead to a straightforward answer to the research

question of the current study.

4. Discussion

Because analysis of information compiled in “Status of Physicochemical and Biological

Determinants of Pfiesteria Population Dynamics in GSB Prior to Superstorm Sandy” (Chapter 3) did not facilitate procurement of a straightforward answer to study’s research question, additional research was conducted to complement this inadequacy. The research in question suggests that due to strong influence of geomorphological arrangement on the determinants of

Pfiesteria population dynamics as well as resistance heterotrophic dinoflagellates commonly

express in the face of harsh conditions generally common in coastal marine environments, prior

to October 29, 2012, some sections of GSB appear to have constituted environments where

bloom formation and toxic activity of Pfiesteria species was conceivable.

51

Effective Cell Concentration

According to Rublee (2006), Pfiesteria populations were not sufficiently abundant in GSB to grow beyond background concentrations. It is a common misconception to assume that populations of toxic dinoflagellates (and other undesirable marine microorganisms) must be highly concentrated to negatively affect co-occurring organisms (Smayda 1990; Boesch et al.

1997). Majority of dinoflagellates constitute only a small proportion of total planktonic biota in their non-toxic as well as toxic forms (Örnólfsdóttir, Pinckney, and Tester 2003; Turner and

Borkman 2005; Turner 2006). In fact, HABs can occur when dinoflagellate concentrations are as low as a few hundreds cells per milliliter (Nehring, Hesse, and Tillman 1995; Sournia 1995).

Gosselin, Fortier, and Gagne (1989) noted that actively toxic P. tamarensis negatively affected experimental finfish larvae at the lowest concentration tested ( ≈250 cells/ml). According to

Örnólfsdóttir, Pinckney, and Tester (2003), Texas bloom of K. brevis occurred as the dinoflagellate constituted ≈23% of the extant phytoplanktonic community. Furthermore, aquatic microorganisms are present “discontinuously” (Boero et al. 1996) as their populations can be large at one time and become severely limited or even absent at a later time and vice versa

(Boero et al. 1996; Rengefors and Anderson 1998). Concentrations of so called “hidden flora” can increase relatively quickly and overwhelm marine ecosystems (Anderson et al. 1993, 575).

Despite its generally low abundance, A. anophagefferens reached concentration of ≈1.5×10 6 cells/ml in GSB in 1985 (Bricelj and Lonsdale 1997; Gobler et al. 2004; Nuzzi and Waters

2004). On one occasion, populations of this picoplanktonic alga increased rapidly from background concentrations to account for 97% of total algal biomass in West Neck Bay, an embayment in Peconic Estuary, LI (Gobler and Sañudo-Wilhelmy 2001a). Similarly, concentrations at which Pfiesteria can intoxicate other organisms do not need to be high. The

52

dinoflagellates may constitute mere 10% of the total phytoplankton biomass in its bloom phase

(Burkholder et al. 1992, 1993; Mallin et al. 1995; Lin, Zhang, and Dubois 2006). It has been

demonstrated that concentrations of ≈300 cells/mL were sufficient to stimulate Pfiesteria

populations to intoxicate and kill finfish (Burkholder et al. 1993; Lovko et al. 2003; Rublee et al.

2005). According to Glasgow et al. (1995), these concentrations can be as low as 250 cells/ml.

Moreover, Pfiesteria have come to be nicknamed ‘phantom’ dinoflagellates as they can appear

rapidly in the water column and disappear with comparable swiftness as prey abundance

dwindles or environmental conditions become less hospitable to host robust growth of their

populations (Burkholder et al. 1992; Steidinger et al. 1996; Burkholder and Glasgow 2001).

Temperature

Rublee et al. (2006) concluded that due to suboptimal summer temperatures, pre-Sandy GSB did

not constitute a site which could have hosted positive population growth and toxic outbreaks of

Pfiesteria species. Although all organisms have their own optimal sets of environmental conditions, they are not necessarily inhibited by them (Li, Stoecker, and Coats 2000). Numerous studies have demonstrated that dinoflagellate blooms can occur at temperatures outside their culture-based temperature optimum (Jensen and Moestrup 1997; Cardoso 2012; Yamamoto et al.

2013). C. polykrikoides , for instance, can function properly over a wide temperature range, despite the temperature optimum which characterizes this species (Kim et al. 2004 ; Kudela and

Gobler 2012). Blooms of C. polykrikoides have been occurring predominantly during summer and early fall (Kudela et al. 2008; Mulholland et al. 2009). Nevertheless, this dinoflagellate species can grow at temperatures as low as 10 °C (Imai, Yamagucchi, and Hori 2006). When C. polykrikoides bloomed in coastal California in March 2006, its highest concentrations were most

53

common at 12.7 °C—lowest temperature measurement recorded during the study (Curtiss et al.

2008). Watras, Chisholm, and Anderson (1982) did not find a strong correlation between

temperature and growth rate of Gonyaulax tamarensis ; the dinoflagellate excysted and its populations grew at ≈6 °C while the temperature range for its optimal growth falls between 12

°C and 20 °C.

Although Pfiesteria are eurythermal organisms, their most prolific growth have occurred

during summer and early fall. Consequently, prior to October 29, 2012, the most likely sites of

GSB where positive growth and toxic outbreaks of Pfiesteria species appear to have been most

feasible were located in the easternmost section of this coastal lagoon where water RTs were the

highest and the influence of colder water flowing into GSB from the coastal ocean through the

Fire Island Inlet was the lowest.

Salinity

Because coastal lagoons are characterized by a high degree of separation from coastal ocean and

receive both freshwater and saltwater (Barnes 1980; Kjerfve et al. 1996), salinity in these

systems—similar to a number of other determinants of Pfiesteria population dynamics—varies

according to the distance from tidal inlet(s) and tributaries that drain into them (Debenay et al.

1998; Branco, Kozlowsky-Suzuki, and Esteves 2007; De Vittor et al. 2012) with highest values in the proximity of tidal inlet(s) and lowest in lagoonal extremes ( Sheng, Peene, and Lui 1990;

Smith 1993) (Table 4.2; Figure 3.3). Consequently, salinity in coastal lagoons can vary spatially from hypersaline to fresh (Kjerfve 1994). Prior to the passage of superstorm Sandy, when GSB was characterized by a greater degree of separation from the coastal ocean (Barnes 1980;

Schubel 1991; Kennish and Paerl 2010) than it is currently, its salinity displayed spatial

54

variability (USACE 1999). It generally decreased from west to east with increasing distance

from the Fire Island Inlet and the influence of more saline oceanic water (Wilson, Wong, and

Carter 1991; Bokuniewicz et al. 1993; USACE 1999, 2004; Hinga 2005; Weiss et al. 2007). GSB

salinity would also decrease along the south-north gradient: from the offshore bar to the mouths

of tributaries draining into the northern reaches of this system (Pritchard and Gomez-Reyes

1986; Wilson, Wong, and Carter 1991; Tanski, Bokuniewicz, and Schlenk 2001; SCDHS 2008),

with the highest salinities nearer the Fire Island inlet and the offshore bar (Sebastiano 2012) and

lowest near the mouths of major tributaries (Redfield 1952; Bokuniewicz and Schubel 1991;

USACE 1999). In embayments of this coastal lagoon salinities were commonly 50% of regular seawater (Ryther 1954). USACE (2004a) recorded 19.75 ppt in Bellport Bay. According to

Ryther (1954), a number of larger subestuaries of GSB had salinities less than 10% of the salinity of seawater. Studying bloom dynamics of G. aureolum in the Carmans River Estuary,

Chang and Carpenter (1985) recorded salinity of ≈15 ppt, which, constitutes the optimal salinity

at which Pfiesteria commence their toxic activity (Burkholder et al. 1993; Burkholder and

Glasgow 1997; Springer et al. 2002).

As it is the case in regards to temperature, in general, dinoflagellate populations do not

require their environments to attain optimal salinities to increase in size and engage toxic

activity. C. polykrikoides , for instance, is characterized by a wide range of salinity tolerance,

despite its relatively narrow optimal preference (Kudela and Gobler 2012). Pfiesteria outbreaks

have occurred at salinities other than the 15-ppt optimum. Shumway, Burkholder, Springer

(2006) found that although rates at which actively toxic cells of P. shumwayae killed C. virginica

larvae were the highest at salinities of ≈15 ppt, the dinoflagellates were able to kill these bivalve

mollusks at ≈30 ppt. According to Stoecker and Gustafson (2002), it is possible that the 1997

55

Pfiesteria outbreaks in Maryland’s Pocomoke, Manokin, and Chicamacomico Rivers occurred at salinities lower than the professed 15 ppt optimum due to reduced predation pressure, ample availability of algal and fish prey, or combinations of these and other parameters.

Although Pfiesteria are euryhaline and—theoretically—their populations could grow even in full-strength marine waters beyond Fire Island, the most vigorous toxic activity of these organisms occurs at salinities of ≈15 ppt. Given their apparent preference for more brackish

environments, prior to October 29, 2012, positive population growth of Pfiesteria species and

possible outbreaks of their toxic activity appear to have been most feasible in subestuaries and

lower reaches of watercourses emptying into easternmost section of this coastal lagoon where,

due to terrestrial freshwater flux and diminished influence of the coastal ocean, ≈15 ppt salinities

were more common than at sites located nearer the Fire Island Inlet.

DO Concentration

Bottom-water hypoxic and anoxic conditions are more frequent and of longer duration in deeper

waters than in shallower environments (Stanley and Nixon 1992; SERC 2014) such as coastal

lagoons. In the NRE, where most of the recorded Pfiesteria outbreaks have taken place

(Burkholder and Glasgow 2001; Jochems and Martens 2001), bottom-water hypoxia episodes

(Luettich et al. 2000; Burkholder et al. 2006; Rothenberger, Burkholder, and Wentworth 2009)

and finfish kills associated with low DO concentrations (NCDENR 1999) are relatively common

during summer and early fall. It has been indicated, however, that attribution of approximately

50% of finfish which died in subestuaries and tributaries of APES between 1991 and 1993 to

depressed oxygen concentrations was unfounded as many of the finfish affected had bleeding

sores on them while measured DO concentrations were within acceptable range (Jochems and

56

Martens 2001). Severe hypoxic events occurred in the lower NRE during summer of 1997 and

1998, for example (Baird et al. 2004). Although finfish that died during that time did so in waters often affected by low DO concentrations, mortality events in question were attributed to toxic activity of Pfiesteria species (Burkholder and Glasgow 2001). Eby and Crowder (2002) recorded recurrent hypoxic conditions in the NRE in the summer of 1998; however, these were more common in the upstream, relatively deeper sections of this system. In fact, maximum extent of downstream hypoxic conditions comprised mere 15% of the study area (Eby and Crowder 2002).

Having analyzed EPA STORET Neuse River DO concentration data, Burkholder et al. (1995) noted that out of 7100 DO concentration measurements collected between 1/1/1991 and

12/31/1993 in the lower NRE, only 674 were ≤4 mg/L which is considered to be potentially deleterious to aquatic and marine biota. Adams et al. (2003) pointed out that most of western

Pamlico Sound and the lower NRE is suitable to finfish survival even with low benthic DO concentrations.

Sediments are naturally anoxic (López and Lluch 2000; Middelburg and Levin 2009).

Dinoflagellate cysts can survive buried in such sediments for extended periods of time (Marshall

1999; Pinckney et al. 2000; Bravo and Figueroa 2014). A number of studies have indicated that oxygen is an absolute requirement for germination of most dinoflagellates (Anderson, Taylor, and Armbrust 1987; Rengefors and Anderson 1998). Nevertheless, temporary exposure even to anoxic conditions does not seem to affect ultimate germination success as cysts stored for months and even years in oxygen-deprived sediments can successfully germinate after re- exposure to appropriately concentrated oxygen (Keafer, Buesseler, and Anderson 1992; Nehring

1993; Joyce and Pitcher 2004). According to Kremp and Anderson (2000) cysts of Baltic Sea dinoflagellate species Scrippsiella hangoei can germinate at 15-20% dissolved oxygen

57

concentrations. Similarly, Pfiesteria , which encysts when faced with unfavorable environment

(Anderson, Hood, and Zhang 2003; Anderson, Stoecker, and Hood 2003; Coyne et al. 2006), can germinate while the overlying water column is hypoxic and even anoxic (Saito et al. 2007).

Additionally, Pfiesteria field populations can be active at DO concentrations lower than those

noted in culture studies (Burkholder and Glasgow 1997; Glasgow et al. 2001). When surface or

bottom waters become suboxic, Pfiesteria can simply migrate down or up to evade particularly

low DO concentrations (Glasgow et al. 2001).

Beside adverse consequences that intermittent suboxic conditions can exact on littoral

living resources, episodes of low DO concentrations can initiate processes that can contribute to

creation of conditions congenial to dynamic growth of certain marine microorganisms.

Resuspension of nutrients stored in benthic sediments intensifies as DO concentrations in bottom

waters decline (Middelburg and Levin 2009; Cousins, Stacey, and Drake 2010; Peña et al. 2010);

suboxic episodes severely restrain denitrification processes which normally remove N from a

given waterbody (Cornwell, Kemp, and Kana 1999; Viaroli et al. 2010). As a result, more

nutrients remain in benthic sediments and can be released back into the water column (Kemp et

al. 2005; Middleburg and Levin 2009) where they can be assimilated by living organisms once

again.

Although, overall, GSB was characterized by good standing of its DO concentrations

prior to October 29, 2012, and—in regards to this parameter—GSB appears to have constituted a

congenial habitat for sustained growth of Pfiesteria populations, such growth appears to have

been most plausible in the easternmost section of this coastal lagoon where photosynthetic O2

production was higher than at sites located nearer the Fire Island Inlet.

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pH

According to USACE (1997), to a large extent, the pH of GSB is influenced by

connections of this coastal lagoon to the ocean (as limited as they are). As a result, pH data

collected by the U.S. Army Corps of Engineers are, as it is generally the case in coastal lagoons

(Farjalla et al. 2006), comparable to those of the neighboring coastal ocean (USACE 1997).

Although Leigh-Manuell, Hanlon, and Larado (2010) collected their pH measurements during

winter and spring rather than during summer when Pfiesteria outbreaks have been most common in the NRE (Figure 3.5), given the high photosynthetic activity in GSB during summer (Schubel

1991; Lonsdale et al. 2006; Newell et al. 2009), prior to the passage of superstorm Sandy, pH in this coastal lagoon was high and, therefore, suitable to positive population growth of dinoflagellate populations, including Pfiesteria species.

Pfiesteria and similar species of toxic dinoflagellates can benefit from high pH. Studies have shown that as pH raises, intensity of micro- and mesozooplankton grazing decreases

(Pedersen and Hansen 2003, 2003a; Buskey 2008; Stoecker, Thessen, and Gustafson 2008). As micro- and mesozooplankton grazing intensity decreases, population growth of phytoplankton— normally preyed on by zooplankton—becomes less restricted, and as their concentrations grow, pH level rises further (Pedersen and Hansen 2003a) facilitating algal growth (Pedersen and

Hansen 2003a) and more congenial nutritional environment to Pfiesteria and similar heterotrophic dinoflagellates.

It appears that, prior to the passage of superstorm Sandy over GSB in October 2012, positive growth of Pfiesteria populations in this coastal lagoon was most feasible in its easternmost section where high rates of CO 2 assimilation by abundant algal populations

59 contributed to creation of congenial (high) pH environment and, consequently, relatively low micro- and mesozooplankton predation pressure.

Turbulent Mixing

Influence of tidal action can be quite pronounced in estuaries (Roman et al. 2005; Wetz et al.

2006). Because coastal lagoons are generally shallower (Kjerfve 1986; Miththapala 2013) and less exposed to the influence of the coastal ocean (Kjerfve 1986, 1994; Kennish and Paerl 2010;

Miththapala 2013) than estuaries and other less restricted coastal marine systems, influence of tidal action in these systems can be very limited (microtidal) and—in a number of instances— non-existent (Boynton et al. 1996; Kirk and Lauder 2000; Kennish and Paerl 2010). In this respect, coastal lagoons are more “quiescent” than estuaries and other less restricted coastal marine systems (Kennish and Paerl 2010, 92) (Table 4.1).

Table 4.1. Some Differences between Estuaries and Coastal Lagoons (Miththapala 2013).

Similar to a number of other parameters that influence Pfiesteria population dynamics, tidal range and associated turbulent mixing decreases with increasing distance from tidal inlet(s)

60

(Smith 1990; Merino et al. 1990; Dias, Lopes, and Dekeyser 1999). The tidal amplitude at the

ocean side of the Fire Island Inlet is approximately 1.3 m; it becomes substantially reduced,

however, as tidal wave flows through the inlet and propagates through the bay (USACE 2004a;

Sebastiano 2012; Aretxabaleta, Butman, and Ganju 2014) (Figure 3.6). This physical restriction

results in a tidal range of 0.5 m at this location (USACE 1997, 2004a). As tidal water propagates

through the bay and away from Fire Island Inlet, the tidal range decreases to 0.3 m in the bay’s

easternmost sections (Bokuniewicz and Schubel 1991; USACE 1997, 2004a). While tidal current

velocities at the Fire Island Inlet were ≈1.03 m/s, they were only ≈0.15 m/s in the middle of the

bay and ≈0.10 m/s in its easternmost sections (USACE 2004a; McElroy et al. 2009) (Figure 3.6).

Wind action, on the other hand, is the predominant non-tidal mixing agent in coastal lagoons

(Calliari, Britos, and Conde 2009; Cousins, Stacey, and Drake 2010; Kennish and Paerl 2010;

Mahapatro, Panigrahy, and Panda 2013) as their generally shallow depths enable wind action to influence entire water column (Barnes 1980; Gamito et al. 2005; Pérez-Ruzafa et al. 2005). In the NRE, where the influence of tides is negligible (tidal amplitude ≈7 cm) due to the moderating

effect of the Outer Banks which separate this subestuary from the influence of the coastal ocean

(Stanley and Nixon 1992; Luettich et al. 2000; Hall et al. 2008), wind is the predominant mixing

force (Luettich et al. 2000; Reynolds-Fleming and Luettich 2004; Rothenberger, Burkholder, and

Wentworth 2009); NRE is considered a well-mixed system (Rudek et al. 1991; Wetz and Paerl

2008; Hall and Paerl 2011).

Although laboratory studies determined that Pfiesteria underperforms in turbulent

environments, field studies have shown that they can overcome such conditions. Similar to

numerous dinoflagellate species (Cullen and Horrigan 1981; MacIntyre, Cullen, and Cembella

1997; Kudela et al. 2008), Pfiesteria is a strong swimmer (Burkholder and Glasgow 1997;

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Glasgow et al. 2001). The filose stage of Pfiesteria life cycle, for example, is outfitted with long pseudopodia that facilitate their suspension and movement through the water column

(Burkholder and Glasgow 1997). As turbulence varies over temporal scales (Rothschild and

Osborn 1988; MacKenzie and Leggett 1993; Berdalet and Estrada 2005), Pfiesteria —similar to other species of heterotrophic dinoflagellates (Berdalet et al. 2007; Schaeffer et al. 2009;

Katanao et al. 2011 )—can sink down the water column to evade inhospitable turbulent episodes and ascend back to the surface/near surface when waters become calmer (Burkholder and

Glasgow 1997; Glasgow et al. 2001). During fish kill events, Pfiesteria assume amoeboid form and descend to feed on dead/dying finfish (Morris 1999). In a modeling study of Pfiesteria dynamics by Hood et al. (2006), Pfiesteria population growth did not decline upon exposure to turbulent mixing levels comparable to levels common in shallow environments where these dinoflagellates have been normally found. Even though TOX-A took longer to double in size than the NON-IND functional type did, populations of this functional type prevailed turbulence experiments (Hood et al. 2006). Burkholder and Glasgow (1997) stressed that a Pfiesteria fish kill event occurred in the NRE buffeted by moderate wind. Pfiesteria cells were concentrated at

≈10 3/mL during a fish kill event in the NRE and declined to concentration of 3*10 2/mL as a result of significant mixing associated with passing of a strong low pressure system (Burkholder and Marshall 2012). Nonetheless, Pfiesteria populations did not decline to zero as expected based upon laboratory studies (Burkholder and Marshall 2012). Pfiesteria quickly regenerated their populations which were decimated after Hurricanes Bertha and Fran passed over coastal

North Carolina in 1996 (Glasgow and Burkholder 2000; Glasgow et al. 2001).

During summer months, turbulence-associated mixing can abate substantially as waters become stratified (Paerl 1988; Koseff et al. 1993; Cousins, Stacey, and Drake 2010).

62

Stratification occurs when water column becomes divided into discrete layers, the top mixed

layer and the relatively unmixed bottom layer (Mellard et al. 2011). Stratification can develop as

a result of intense solar heating of the ocean surface (Kjerfve and Magill 1989; Christian, Boyer,

and Stanley 1991; Koseff et al. 1993) or salinity differences associated with high freshwater

delivery following heavy precipitation events (Kjerfve and Magill 1989; Koseff et al. 1993;

Moore, Baird, and Suthers 2006; Hall et al. 2008). In both instances, stratification develops as

denser water is displaced down the water column due to gravitational force (Kjerfve and Magill

1989).

Because of their shallowness and wind-mixed nature, stratification events are considered

to be rare in coastal lagoons (Palma-Silva, Albertoni, and Esteves 2002; Chagras and Suzuki

2005; Murphy and Secor 2006). Nevertheless, stratification may develop (Kirk and Lauder 2000;

Kennish and Paerl 2010) and be a common summer occurrence even in well-mixed systems such as coastal lagoons (Garcia and López 1989; Kjerfve and Magill 1989; Debenay et al. 1998;

Godhantaraman and Uye 2003). Although GSB was considered a well-mixed system (Wilson,

Wong, and Carter 1991; Hinga 2005; Bricelj 2009; Nuttall 2010) prior to the passage of superstorm Sandy in October 2012, stratification events could occur in this coastal lagoon during summer (Hinga 2005; McElroy et al. 2009). It could become particularly pronounced in the lower sectors of watercourses emptying into the easternmost parts of bay, where the influence of tidal currents was severely restricted (USACE 2004a, McElroy et al. 2009).

Positive growth of dinoflagellate populations and their toxic activity occur with higher probability as stratification of the water column develops (Chang and Carpenter 1985; Weise et al. 2002; Hallegraeff, Mooney, and Evans 2009; Place et al. 2012). Studies have shown that availability of sunlight is the most important determinant of phytoplankton population dynamics

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in shallow, nutrient-rich marine systems (Penncock and Sharp 1986; Koseff et al. 1993;

Litchman 2000). When stratification develops, turbidity caused by suspended sediments becomes reduced, availability of ambient sunlight increases (Peperzak 2003, 2005; Thompson et al. 2008;

Emmanuel and Chukwu 2010), and corresponding increase in algal photosynthetic activity occurs (Cloern 1984; Chang and Carpenter 1985; Köhler 1997). The result is that algal cells become concentrated in the upper segment of the water column (Koseff et al. 1993; Moore,

Baird, and Suthers 2003; Thompson et al. 2008) where they can be grazed by Pfiesteria with

greater efficiency. While Chang and Carpenter (1985) noted low concentrations of G. aureolum

in the more open sections of GSB, concentrations of this dinoflagellate were significantly higher

in the stratified waters of the lower Carmans River and its estuary protected from wind action by

wooded areas. Furthermore, despite transitory nature of stratification events in GSB, Pfiesteria

blooms could develop in this system, because dinoflagellate presence in the water column is

“discontinuous” (Boero et al. 1996, 568). Because of their ‘phantom’ nature (Burkholder et al.

1992; Steidinger et al. 1996; Burkholder and Glasgow 2001), Pfiesteria blooms can be as

transitory as stratification events themselves.

Although Pfiesteria can cope with elevated turbulent mixing, positive population growth and toxic outbreaks of these organisms occurred most commonly in stratified, low-turbulence waters. Given such preference, prior to October 29, 2012, positive population growth of

Pfiesteria in GSB were most feasible in the innermost, most sheltered sections of this coastal lagoon away from the Fire Island Inlet and the influence of more turbulent ocean beyond it.

Availability of Dissolved Nutrients

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Although more than 40 streams and rivers drain into GSB (Jones and Schubel 1980; Monti and

Scorca 2003; McElroy et al. 2009) and runoff facilitating impervious materials cover a large proportion of its watershed (Dennison, Koppelman, and Nuzzi 1991; NYSDOS 2001; Kinney and Valiela 2011), in comparison to other sources, surface runoff is generally considered not to be a major source of freshwater (and nutrients it may contain) to this coastal lagoon (SSERC

1999a, 2000; Gobler and Boneillo 2003). The Carmans River—one of the largest rivers emptying into GSB (O’Malley 2008)—delivers approximately 25% of annual surface freshwater flux into this coastal lagoon (Pritchard and Gomez-Reyes 1986). Nevertheless, Carmans River has been determined not to be an important contributor of freshwater to GSB during summer months when precipitation is low and evaporation is high (USACE 1999). Costal nutrient pollution is, however, a “’multi-media problem” (Whitall, Castro, and Driscoll 2004, 26). One such “medium” is groundwater. Groundwater can constitute an important source of freshwater

(and nutrients it contains) to coastal lagoons (Charette, Buesseler, and Andrews 2001; Cañedo-

Argüelles et al. 2011; Pérez-Ruzafa et al. 2012). Besides directly seeping through the bottom, groundwater feeds streams and rivers emptying into GSB (Pluhowski and Kantrowitz 1964;

Town of Brookhaven 2013). According to estimates by Pluhowski and Kantrowitz (1964), groundwater contributes 90-95% of the freshwater flux to these watercourses. Groundwater is considered the second largest source of nutrients to this coastal lagoon (Carpenter, Brinkhuis, and Capone 1991). According to Capone and Slater (1990), groundwater accounts for ≈50% of the total N flux to GSB. Also, as it is the case with other coastal systems where groundwater supplies significant amounts of nitrogenous compounds (Johannes and Hearn 1985; Burkholder et al. 2006; Fear, Paerl, and Braddy 2007), groundwater flux into GSB has been recognized as important to biological processes in this system (Capone and Bautista 1985; Taylor, Gobler, and

65

Sañudo-Wilhelmy 2006) and, more relevantly, stimulation of toxic/noxious microorganisms

(Beck et al. 2008). It has been suggested that blooms of A. anophagefferens have been associated

with nutrient content of groundwater flux to this coastal lagoon (Gobler and Sañudo-Wilhelmy

2001).

Efforts have been initiated to reverse the nutrient overenrichment problem in GSB and

other embayments along LI’s South Shore. A number of studies have demonstrated that

frequently, in regards to HABs, efforts directed at reduction of nutrient loads to coastal waters

and remediation of eutrophication problem have been unsuccessful (Sechi et al. 2001; Paerl et al.

2004; Carstensen et al. 2011). Institution of nutrient reduction measures in Sweden’s Laholm

Bay, for instance, failed to prevent HABs from occurring (Granéli 1987). In LI, sewering

measures had been undertaken in western Suffolk County to address raising nutrient

concentrations in GSB (SSERC 1999b). Since then, considerable reduction in DIN

concentrations has occurred (Nuzzi and Waters 2004; Hinga 2005; Eckhardt, Flipse, and Oksford

1989). According to government of Suffolk County, average concentrations of nitrate, nitrite,

and ammonia declined in this part of the bay by 50% between 1980 and 2005 (Hinga 2005).

Nevertheless, although sewering measures contributed to a decline in DIN concentrations,

corresponding reduction in concentrations of DON have not occurred (Nuzzi and Waters 2004).

According to Gobler, Lonsdale, and Boyer (2005), inorganic nutrient reduction efforts in western

GSB may have contributed to creation of a nutritional environment in which populations of A.

anophagefferens can grow and—because of their ability to utilize organic nutrients more efficiently than inorganic forms—to have an advantage over phytoplankton that do not possess this ability. The eastern parts of GSB have experienced no perceptible reduction in concentrations of N (neither DIN nor DON) (Nuzzi and Waters 2004). Nutrient concentrations

66 have actually increased despite substantial reductions in the flux of nutrient-rich wastes from duck farms formerly numerous in this part of GSB’s watershed (SSERC 2012). The sewage treatment plant on the Patchogue River, for instance, does not provide sufficient treatment of wastewater (Greene 2009). Provided it did, it would not contribute to substantial reduction of nutrient flux to GSB as effluent processed at this facility comprises mere ≈2% of N flux to this coastal lagoon (Koppelman 1978).

Because coastal lagoons are characterized by high degree of separation from the influence of marine waters, nutrient concentrations in these systems—similar to other determinants of Pfiesteria population dynamics—vary spatially according to the distance from tidal inlet(s) and mainland (Barnes 1980; Meyercordt, Gerbersdorf, and Meyer-Reil 1999;

Badylak and Phlips 2004; Kennish and Paerl 2010) (Table 4.2).

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Table 4.2. Status of Some Physicochemical Factors in Celestún Lagoon, Mexico (Medina-Gómez and Herrera-Silveira 2002).

Prior to the passage of superstorm Sandy, such arrangement was also the case in GSB

(Bokuniewicz et al. 1993) (Figure 4.1).

Figure 4.1. Spatial Variability of Dissolved Nitrogen Concentrations in a Typical Coastal Lagoon (GSB; Gobler 2010).

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Nutrient concentrations grew with increasing distance from the Fire Island Inlet and—as

a result—were more concentrated in the northern parts of the bay near tributary mouths and in its

easternmost parts than in its western reaches closer to the Fire Island Inlet (Lively, Kaufman, and

Carpenter 1983; Carpenter, Brinkhuis, and Capone 1991; Dennison, Koppelman, Nuzzi 1991;

Cosper 2001; Bokuniewicz et al. 1993; Rublee et al. 2006). Gobler, Renaghan, and Buck (2002), for instance, found that DIN concentrations in Patchogue Bay were significantly higher than those in Bay Shore Cove. According to the authors, such difference can be explained by the fact that water residence time in Bay Shore Cove—situated nearer the Fire Island Inlet and the influence of the coastal ocean—was lower than in Patchogue Bay situated further from this Inlet

(Gobler, Renaghan, and Buck 2002). Considerable distance from the Fire Island Inlet and influence of the coastal ocean had also contributed to high nutrient concentrations in tributaries draining into the easternmost sections of GSB. Suffolk County Department of Health Services

(SCDHS) found significant concentrations of ammonia in the Beaver Dam Creek (SCDHS

2008). Dissolved nutrients were also abundant in the estuary of the Carmans River (Carpenter and Dunham 1985; Clark, Gobler, and Sañudo-Wilhelmy 2006).

Availability of Finfish Prey

Similar to estuaries, coastal lagoons can be abundant in finfish (Pollard 1994; Poizat et al. 2004;

Machado, Conde, and Rodríguez-Graña 2011), hence, they commonly constitute productive fisheries (Gamito et al. 2005; Kennish and Paerl 2010; Mahapatro, Panigrahy, and Panda 2013).

In Egypt, 50% of total inland finfish yield comes from Lake Manzala (Ramdani et al. 2009).

According to Yañez-Arancibia (1978), ≈80% of finfish populations in coastal Mexico are

associated with coastal lagoons and immediate surroundings functioning of these systems affect

69

(in Muelbert and Weiss 1991). Nearly 50% of annual finfish catch along the coast of Texas

originate from Laguna Madre (Ziegler and Benner 1999). The APES, where Pfiesteria were the predominant cause of finfish mortality during the 1990s (Burkholder and Glasgow 1997, 2001;

Jochems and Martens 2001), constitutes the most important finfish nursery on the Atlantic Coast

(Mallin et al. 2000; Ross 2003; Burkholder et al. 2006) as it provides a nursery habitat for ≈80%

of finfish inhabiting waters of the mid-Atlantic coast (Copeland and Gray 1991). In 1987, the

United States Congress designated the APES an estuary of national importance (NCDENR

2014).

Coastal lagoons are attractive finfish habitats for a number of reasons. First, they are shallow (Emmanuel and Chukwu 2010). Shallow coastal systems are considered to be more attractive to a number of finfish species as predators which prey on these species are less likely to frequent shallow environments (Olney an Boehlert 1988). Second, coastal lagoons are abundant in dissolved nutrients. In general, finfish abundance in coastal waters increases with rising eutrophication level (Schernewski and Schiewer 2002; Lee and Jones 1991; Nixon and

Buckley 2002) as nutrients stimulate algae which stimulate algal predators—zooplankton (Telesh

2004)—which in turn stimulate growth of finfish populations (Schernewski and Schiewer 2002).

Third, due to their shallow depths, moderate magnitude of turbulent mixing, abundance of dissolved nutrients, and the ease with which solar radiation can penetrate the water column, coastal lagoons can be abundant in submerged aquatic vegetation (SAV) (Ménendez and Comín

1989; Lenzi, Palmieri, and Porrello 2003; Menéndez et al. 2003), usually, more extensive than in estuaries and other types of transitory marine systems (Barnes 1980; Knoppers 1994; Pérez-

Ruzafa et al. 2012). SAV meadows are ecologically important as they are generally characterized by higher abundance of finfish than non-vegetated habitats (Castro and Cowen 1991; Franco et

70

al. 2006; Verdiell-Cubedo, Oliva-Paterna and Torralva-Forero 2007). Finfish use these habitats

as feeding grounds, nurseries, and refugee from predators (Pohle, Bricelj, and Garcia-Esquivel

1991; Abecasis, Bentes, and Erzini 2009). In a study by Lazzari (2002) in Casco Bay, Maine, finfish densities were 3.7 times higher in SAV beds than in unvegetated areas. 23 out of 40 finfish species studied by Briggs and O’Connor (1971) in GSB preferred SAV meadows over

other environments. Although, as it has been the case in numerous coastal lagoons around the

world (Castel, Caumette, and Herbert 1996; De Wit et al. 2001; Gamito et al. 2005), there has

been a general decline in SAV abundance in GSB (Bricelj and Lonsdale 1997; Hinga 2005;

Suffolk County 2014), some sections of this system were more successful in keeping their

meadows relatively intact (USACE 2004a; Gobler 2010; Kinney and Valiela 2011). Mouth of the

Carmans River, for instance, is adjacent to 26-acre eelgrass ( Zostera marina ) and widgeon grass

(Ruppia maritima ) beds (Town of Brookhaven 2013).

Rublee et al. (2006) pointed out that finfish populations in GSB have dwindled

precipitously over the past several decades and that they were no longer sufficiently abundant to

stimulate toxic activity of Pfiesteria species. It has to be pointed out that majority of finfish

species which can be encountered in coastal lagoons are not permanent residents of these

systems but seasonal migrants (Barnes 1980; Gamito et al. 2005; Kennish and Paerl 2010). Such

species evolved various mechanisms to cope with fluctuating environmental conditions; they are

able to adjust their blood and internal organs to endure movements between saltwater and

freshwater (Childs et al. 2008). Seasonal migrants use lagoonal tributaries as nursery, spawning,

and feeding grounds (Gamito et al. 2005; Garcia et al. 2007; Abecasis, Bentes, and Erzini 2009).

Such species that frequent GSB as bluefish (Juanes and Conover 1994), butterfish (Carson

1945), silverside (USFWS 1983; Conover et al. 2005), spot (Adams et al. 2003), bay anchovy

71

(Luo 1993), weakfish (Carson 1945; NY Seafood Council 2013), striped bass (NY Seafood

Council 2013), and alewife (NYSDEC 2014) travel in dense schools which Pfiesteria have attacked most frequently in the NRE (Burkholder, Glasgow, and Hobbs 1995; Glasgow et al.

1995; Cancellieri et al. 2001). In the APES, concentrations of TOX-A zoospores have been noted to be at their maximum in late summer to early fall as Atlantic menhaden ( Brevoortia tyrannus ) commence their annual migration out of the lagoon back to coastal ocean (Manooch 1988,

Glasgow et al. 2001). Such species as weakfish (Carson 1945), bluefish (Bokuniewicz et al.

1993), and alewife (Kritzer, Hughes, and O’Reilly 2007; Hughes and O’Reilly 2008; Kelder

2010) migrate into GSB’s tributaries in the spring and back to the coastal ocean during late summer/early fall.

Given the highest probability of Pfiesteria outbreaks during summer when many migratory finfish move back to the Atlantic, as well as propensity of these organisms for attacking schooling finfish, prior to the passage of superstorm Sandy, Pfiesteri a bloom formation in GSB appear to have been most feasible in the SAV beds and lower tributaries emptying in the easternmost section of this coastal lagoon.

Availability of Algal Prey

Algae constitute a significant proportion of planktonic communities of coastal lagoons during summer (Barnes 1980; Badylak and Phlips 2004; Garcia et al. 2007). This abundance results from generally long water residence times (Badylak and Phlips 2004; Mahapatro, Panigrahy, and

Panda 2013), high concentrations of dissolved nutrients (Bricker et al. 1999; Velasco et al. 2006;

Cousins, Stacey, and Drake 2010), and shallow depths (Boynton et al. 1996; McGlathery,

Anderson, and Tyler 2001; McGlathery, Sundbäck, and Anderson 2007), which facilitate

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efficient penetration of water column by solar radiation (Brito et al. 2009; Piccini et al. 2009;

Kennish and Paerl 2010). The result is that c oastal lagoons are usually more abundant in algae than estuaries and other transitory marine systems (Knoppers, Kjerfve, and Carmouze 1991;

Kennish and Paerl 2010) and one of the highest net primary productivity ever recorded in natural ecosystems (Knoppers 1994; Sorokin, Sorokin, and Gnes 1996; Conde et al. 1999). Chlorophyll a concentration measured by Moyà et al. (1987) in Es Cibollar lagoon, Majorca, Spain, was only

2.71 mg/m 3 in February but rose to 23.20 mg/m 3 during August. Flores-Verdugo et al. (1988) found that annual primary production in coastal lagoons of Mexico was ≈6 times higher than in other less tidally restricted marginal marine systems of this country. Studying relative contribution of benthic and pelagic photosynthesizers to primary production in Kirr-Bucht and

Rassower Strom, small coastal lagoons in the southern Baltic Sea, Meyercordt, Gerbersdorf, and

Meyer-Reil (1999) found that concentration of chlorophyll a in more restricted Kirr-Bucht waters was ≈7 times higher than in Rassower Strom. In Thau lagoon (France), Bacher, Bioteau, and

Chapelle (1995) found surface concentrations of diatom-dominated primary producers (70% of all total phytoplankton present) to be about 40 times greater than in the coastal Mediterranean

Sea. Annual chlorophyll a maxima in estuaries of the Neuse and Pamlico Rivers, where most of the recorded Pfiesteria outbreaks have taken place (Burkholder and Glasgow 2001; Jochems and

Martens 2001), typically exceed 40 µg/L (Glasgow and Burkholder 2000; Glasgow et al. 2001).

Barnes (1980) stressed that while the importance of diatom populations in coastal lagoons should not be belittled, chlorophytes and cryptophytes (algal groups preferably consumed by Pfiesteria ) constitute a larger proportion of planktonic communities in coastal lagoons than in estuaries and other less restricted coastal marine systems In addition to diatoms, chlorophytes and

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cryptophytes are abundant in subestuaries of the APES during summer months (Mallin and Paerl

1994; Pinckney et al. 2000; Wetz and Paerl 2008a).

One algal resident of GSB, Aureococcus anophagefferens , deserves particular attention as its activity has been an important determinant of ecosystemic health of this coastal lagoon. A.

Anophagefferens is a pelagophyte alga with a wide distribution along the East Coast of the

United States (Anderson et al. 1993; Gastrich et al. 2004; Glibert et al. 2007). It bloomed in LI

embayments for the first time in 1985 and has been a recurring problem since (Bricelj and

Lonsdale 1997; Gobler and Sunda 2012; Suffolk County 2014).

To author’s best knowledge, there have been no studies (laboratory or field) which

investigated potential interactions between A. anophagefferens and Pfiesteria species.

Nevertheless, populations of other toxic dinoflagellate species have often constituted significant

proportion of planktonic communities during brown tide events in coastal waters of LI (Caron et

al. 1989; Lonsdale et al. 1996; Sieracki et al. 2004) and Maryland (Deonarine et al. 2006). Caron

et al. (2004) and Buskey and Hyatt (1995) found that population of O. marina grew at their normal rate in the presence of highly concentrated A. anophagefferens cells. Buskey and Hyatt

(1995) also noted that although populations of O. marina grew at low rates when ingesting low concentrations of brown tide cells, they grew well on high concentrations of this species.

Because to their different physiological requirements, it is seems unlikely that A. anophagefferens (and other brown tide species) and Pfiesteria would influence their respective population dynamics in GSB. First, brown tides occur predominantly during late spring and early summer within 15-20 °C temperature range (Cosper et al. 1989; Gobler, Lonsdale, and Boyer

2005). Brown tide species underperform at higher temperatures (Cosper et al., 1989; Gobler et al.

2005) and their blooms collapse at temperatures >25 °C (Nuzzi and Waters 1989, 2004).

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Pfiesteria are thermally resistant. As it was discussed in earlier section (1.2.2), the highest rates of Pfiesteria population growth have occurred within the 26-33 °C temperature range

(Burkholder et al. 1993, 1995; Burkholder and Glasgow 1997, 2001; Marshall 1999). Second, studies found that as salinity levels fall to <28 ppt, population growth brown tide species declines significantly (Cosper et al. 1989a), while Pfiesteria thrive at these and lower salinities

(Burkholder, Glasgow, and Hobbs 1995; Sullivan and Andersen 2001; Burkholder and Marshal

2012).

Beside bivalve mollusks, other organisms that influence population dynamics of

Pfiesteria species (bacteria, algae, micro- and mesozooplankton, and finfish) do not seem to be

affected by the presence of abundant brown tide cells. Caron et al. (1989) and Caron et al. (2004)

investigated the effects of A. anophagefferens on planktonic assemblages finding that all culture

populations experienced no inhibitory effects, grew at their normal rates, and achieved high

concentrations during brown tide events. In fact, on a number of occasions, culture populations

achieved their highest respective concentrations when concentrations of A. anophagefferens were

also at their highest. According to Nuzzi and Waters (1989), phytoplankton sized ≤5 um

outnumbered A. anophagefferens cells during 1986 brown tide event in GSB. Bacterial

populations can also thrive in environments dominated by brown tide species (Gobler and

Sañudo-Wilhelmy 2003; Kana et al. 2004; Sieracki et al. 2004) because, as A. anophagefferens

and other brown tide species, bacteria prefer organic nutrients and ingest them at rates much

higher that the rates at which it consumes inorganic nutrients (Gobler and Sañudo-Wilhelmy

2001a). As it was discussed earlier, spatial distribution of bacteria in coastal lagoons reflects

spatial distribution of algae and availability of organic products of algal photosynthesis.

According to Gobler and Sañudo-Wilhelmy (2001), additions of glucose to culture medium

75 containing brown tide cells and bacteria stimulated the growth of both populations. The most prolific growth of microbial populations noted by Ziegler and Benner (1999) in Laguna Madre,

Texas occurred during a brown tide event. Kana et al. (2004) detected no adverse effects in algae exposed to a dense bloom of A. anophagefferens . In addition to finding that total phytoplankton assemblage grew normally in the presence of abundant A. anophagefferens cells, Caron et al.

(2004) found significant mortalities of this brown tide species due to zooplankton grazing.

According to Sieracki et al. (2004), microzooplankton can ingest A. anophagefferens cells.

Having noted high ingestion rates of A. tonsa nauplii feeding on A. anophagefferens , Smith et al.

(2008) hypothesized that copepod nauplii can effectively control A. anophagefferens populations and preclude development of their blooms in natural settings. Microzooplankton are also unaffected by dense brown tides in GSB (Caron et al. 1989; Lonsdale et al. 1996; Sieracki et al.

2004). In fact, as Gobler, Renaghan, and Buck (2002) pointed out, significant A. anophagefferens mortalities in two GSB embayments—Bay Shore Cove and Patchogue Bay— were due to zooplankton grazing. Deonarine et al. (2006) found that some microzooplankton grazers avoided ingestion of brown tide cells, but grazed alternative prey, maintained their normal grazing rates, and suffered no adverse effects in the presence of abundant A. anophagefferens population. Mesozooplankton are also unaffected by the presence of A. anophagefferens ; copepods can graze densely concentrated cells of these organisms (Lonsdale et al. 1996; Deonarine et al. 2006; Lonsdale et al. 2007). Although a number of finfish species simply leaves waters containing high concentrations of A. anophagefferens cells (Siddall 1987), no ill effects in finfish exposed to such dense populations of this alga have been noted (Shima and Cowen 1989; Buskey et al. 1996; Bricelj 2009). In fact, Duguay, Monteleone, and Quaglietta

(1989) found higher densities of finfish eggs and larvae when A. anophagefferens concentrations

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were high than when abundance of this alga was lower. Castro and Cowen (1989) recorded that

growth rates of bay anchovy larvae exposed to a brown tide cells were among the highest ever

recorded in GSB.

Similar to other physicochemical and biological determinants of Pfiesteria population dynamics, spatial variability also pertains to algal biomass in coastal lagoons; chlorophyll a concentrations tend to be higher in lagoonal interiors depending on the distance from inlets and the influence of the coastal ocean (Jarry et al. 1990; Wasmund and Heerkloss 1993; Soria,

Vicente, and Miracle 2002). Prior to October 29, 2012, spatial distribution of algae also reflected lagoonal geomorphology of GSB; east to west and landward to seaward gradients in this distribution were prominent (Lively, Kaufman, and Carpenter 1983; Cosper 2001; Weiss et al.

2009). Because of longer water residence times (Vieira 1989), concentrations of dissolved nutrients and, therefore, algal biomass, were higher in the easternmost sector of this coastal lagoon than in the vicinity of the Fire Island Inlet (Cosper 2001; Gobler, Renaghan, and Buck

2002; Clark, Gobler, and Sañudo-Wilhelmy 2006; Bricelj 2009; Sebastiano 2012). Consequently, long water residence times in the eastern part of GSB would enable a significant percentage of summer phytoplankton to avoid advection to the open ocean (Ryther 1954). Chlorophyll a concentration gradient was also the case along the north-south axis with highest algal biomass in more landward locations closer to bay’s tributaries (Lively, Kaufman, and Carpenter 1983;

Bokuniewicz et al. 1993; Cosper 2001). During the early 1950’s, for instance, green tide chlorophyte species, Nannochloris atomus and Stichococcus spp. , bloomed in GSB (Ryther

1989; Bokuniewicz et al. 1993). The highest concentrations of these organisms occurred in the

eastern portion of the bay near river mouths and high concentrations of nutrients released from

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industrial-scale duck farms operating in the watersheds of streams and rivers emptying into this

section of the bay (Sieracki et al. 2004).

Given the strong correlation between spatial distribution of algal populations and

abundance of dissolved nutrients in coastal lagoons as well as the fact that algae constitute the

predominant nutrient source for nontoxic Pfiesteria stages, prior to the extreme weather event of

October 29, 2012, probability of Pfiesteria bloom formation of in GSB appears to have been the highest in the easternmost reaches of this coastal lagoon where—due to long water RTs and proximity to freshwater sources—concentrations of both dissolved nutrients and algal chlorophyll a were the highest.

Interactions with Bacteria

Bacteria are commonly very abundant in coastal lagoons (Chin-Leo and Benner 1999; Farjalla et al. 2006; Piccini et al. 2006), at times more abundant than algae (Sorokin et al. 1999; Pugnetti et al. 2010). Because coastal lagoons are frequently situated near urban and agricultural areas

(Macedo et al. 2001; Piccini et al. 2006; Kennish and Paerl 2010), terrestrial runoff can be a source nutrients—and bacteria—to these systems (Fiadrino et al. 2003; Glasoe and Christy 2004;

Reay 2004). In the lagoon of Venice, as measured by Sorokin et al. (1996), bacteria constitute

40-60% of total planktonic biomass. According to Farjalla et al. (2006), abundance of bacteria in

Imboassica Lagoon, Brazil is so high that doubling of bacterial biomass in this system would take only three days. Sorokin, Sorokin, and Gnes (1996) found that concentrations of bacteria in

Comacchio lagoonal system (Italy) were higher than ever recorded in marine waters. According

to Baird et al. (2004), the rate of bacterial production in a mesohaline section of the NRE can be

comparable to the rate of phytoplankton production.

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Coastal lagoons constitute environment more congenial to positive growth of bacterial

populations than estuaries and other marginal marine systems. First, since bacterial populations

grow at highest rates while ingesting organic nutrients (Cole and Pace 1995; Donderski, Mudryk,

and Walczak 1998; Piccini et al. 2006) and phytoplankton produce such compounds through

photosynthesis, abundance of bacteria in aquatic environments is correlated with activity of

primary producers (Chin-Leo and Benner 1999; Lonsdale et al. 2006; Pugnetti et al. 2010).

Studies have shown that a large proportion of organic compounds produced by phytoplankton

are assimilated by bacteria (Hagström et al. 1988); as much as 50% of primary production in

coastal marine environments can be assimilated by bacterioplankton ( Azam et al. 1983; Azam

1998; Lonsdale et al. 2006). Second, typically, activity of aquatic bacteria increases as certain

compounds of the DOM pool are exposed to solar radiation (Lindell, Granéli, and Tranvik 1995;

Moran and Zepp 1997; Piccini et al. 2009). As solar radiation breaks down organic matter

(Miller and Moran 1997; Jørgensen et al. 1998), organic substrates become available to bacteria

(Wetzel, Hatcher, and Bianchi 1995; Jørgensen et al. 1998; Piccini et al. 2009). General shallowness of coastal lagoons (Boynton et al. 1996; Kennish and Paerl 2010), efficient penetration of the water column by solar radiation (Brito et al. 2009; Piccini et al. 2009; Kennish and Paerl 2010), and high concentrations of DOM (Sorokin et al. 1996; Mahapatro, Panigrahy, and Panda 2013) make these systems particularly hospitable to efficient breakdown of organic material and robust growth of bacterioplankton biomass.

Owing to the strong association between primary productivity of algal populations and abundance of bacteria, spatial distribution of bacteria in coastal lagoons tends to follow algal biomass gradient. Prior to the passage of superstorm Sandy in October 2012, distribution of bacterial biomass in GSB followed the west-east and landward-seaward concentration gradients

79 according to distance from the Fire Island Inlet. On average, concentrations of bacteria were highest in the lower sections of tributaries emptying into the innermost parts of this system

(Dennison, Koppelman, and Nuzzi 1991; SSERC 1999, 1999c) as well as their estuaries

(Dennison, Koppelman, and Nuzzi 1991).

Given that Pfiesteria prey on bacteria (Burkholder and Glasgow 1995, 1997; Burkholder

1999) and employ bacterial faculties in the production process of their toxins (Tengs et al. 2003;

Alavi et al. 2001; Miller and Belas 2004), locations of GSB where prior to October 29, 2012

Pfiesteria bloom formation and toxin production appear to have been most feasible was the easternmost section of this c coastal lagoon where shallow depths, abundant sunlight, and high nutrient concentrations facilitated efficient photolysis of organic matter produced by algae through photosynthesis and prolific growth of bacterial populations.

Interactions with Zooplankton

Microzooplankton (Sorokin et al. 1996; Lam-Höai, Rougier, and Lasserre 1997; Godhantaraman and Uye 2003) and mesozooplankton (Buskey et al. 1996; Pombo, Elliott, and Rebelo 2005; de

Wit 2011 ) can be abundant and important components of food webs in coastal lagoons (Barnes

1980). Coastal lagoons are excellent sites where robust growth of zooplanktonic populations can occur as they become more abundant with increasing concentration of nutrients (Gaudy 1989), phytoplankton prey (Mallin and Paerl 1994; Godhantaraman 2001; Godhantaraman and Uye

2003) and—to a higher extent in the case of mesozooplankton—residence time (Pace, Findlay, and Lints 1992; Solidoro et al. 2010; McGlaughon 2012). Mallin and Paerl (1994) found that summer populations of A. tonsa in the lower NRE followed abundance of their phytoplanktonic prey. Wetz and Paerl (2008) recorded high abundance of ciliates in western Pamlico Sound and

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the lower NRE following the passage of Tropical Storm Helene in September 2000. In addition to A. tonsa , Mallin (1991) found abundant NRE populations of Paracalanus crassirostris and

Oithona colcarva . Despite its abundant micro- and mesozooplanktonic predators (Mallin and

Paerl 1994; Valdes-Weaver et al. 2006), the NRE has been considered an “epicenter” of toxic

Pfiesteria activity (Burkholder and Glasgow 2001; Jochems and Martens 2001). Bearing in mind its high concentrations of chlorophyll during summer months (Schubel 1991; SSERC 1998;

1999a; Boissonneault-Cellineri et al. 2001; Lonsdale et al. 2006) and long water residence times

(SSERC 1999, 2000; Hinga 2005), GSB also constitutes an excellent habitat for zooplanktonic grazers (Barnes 1980).

Although microzooplankton can prey on toxic dinoflagellates, grazing by these organisms might not be an effective way of precluding development of HABs. First, frequently, microzooplanktonic grazers simply avoid ingesting actively toxic dinoflagellate cells. According to Jeong et al. (1999), Strombidinopsis sp. avoids consuming actively toxic A. carterae cells.

Stoecker, Guillard, and Kavee (1981) demonstrated development of similar relationship between

F. ehrenbergii and toxic A. carterae . According to J. Burkholder and H. Glasgow of the North

Carolina State University’s Center for Applied Aquatic Ecology, microzooplankton populations can refuse ingestion of actively toxic Pfiesteria cells (Stoecker, Stevens, and Gustafson 2000).

Although ciliates can ingest non-toxic Pfiesteria strains (Stoecker, Stevens, and Gustafson 2000;

Stoecker et al 2002; Lewitus et al. 2006), they tend to reject actively toxic strains (Lewitus et al.

2006). Based on these results, it has been hypothesized that while field populations of nontoxic

Pfiesteria strains can be kept in check by grazing activity of microzooplankton, field populations of actively toxic Pfiesteria strains might avoid such fate and increase in abundance even in the presence of microzooplanktonic predators (Stoecker, Stevens, and Gustafson 2000; Stoecker and

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Gustafson 2002; Hood et al. 2006). Second, microzooplankton can ingest toxic dinoflagellates,

albeit at rates lower than nontoxic strains ( Hansen 1989; Buskey and Hyatt 1995; Buskey et al.

2008). Hansen (1989) noted that ingestion of toxic Alexandrium spp. by F. ehrenbergii declined progressively as dinoflagellate cell toxin content increased. Significantly lower microzooplankton grazing rates were noted in the presence of Pfiesteria TOX-A functional type as compared to grazing rates of TOX-B and NON-IND functional types (Stoecker, Stevens, and

Gustafson 2000; Stoecker and Gustafson 2002; Hood et al. 2006; Lewitus et al. 2006). In a modeling experiment by Hood et al. (2006) population of Pfiesteria TOX-A increased despite significant rates of microzooplankton grazing. Stoecker et al. (2002) found that population declined to near zero after 3 hours of ingesting Pfiesteria TOX-A. Third, at times, heterotrophic dinoflagellates can attack and ingest microzooplankton (Hansen 1991; Inoue, Fukuyo, and

Nimura 1993; Li et al. 1996; Kamiyama and Arima 1997). Jacobson and Anderson (1996) detected ciliate remains inside cells of A. ostenfeldii and G. diegensis . Toxins released by A. tamarense effected backward swimming and eventual death of a Favella ehrenbergii in Hansen

(1989). In Bjørnsen and Nielsen (1991) a steep decline in concentration of microzooplanktonic grazers occurred at depths most abundant in toxic G. aureolum cells. Nagai et al. (2008) cultured

D. fortii by feeding it cells of marine ciliate M. rubra ; eventually, D. fortii completely cleared M. rubra cells out of the water column. Hansen (1991) found that although D. rotundata were preyed upon by T. fusus , frequently, D. rotunda ingested their ciliate predators. When actively toxic, Pfiesteria can ingest microzooplankton which normally prey on them (Burkholder and

Glasgow 1995; Stoecker, Stevens, and Gustafson 2000; Stoecker et al. 2002). S. putrina can consume actively toxic cells of P. piscicida with no adverse effects (Burkholder et al. 1992).

However, allowed sufficient time, vegetative Pfiesteria cells would transform into

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which would either swarm around their potential ciliated predators to draw them out from the

immediate area or attack and kill them (Burkholder et al. 1992, 1993; Mallin et al. 1995).

Burkholder and Glasgow (1997) found that none of the microzooplankton species they studies

preyed on these amoebas. Fourth, as eutrophication level increases, alternations of the structure,

composition, and abundance of zooplankton populations may occur (Beaver and Crisman 1982).

Most typically, microzooplankton biomass decreases with increasing eutrophication level (Park

and Marshall 2000). Stoecker and Gustafson (2002) hypothesized that microzooplankton grazing pressure can be reduced significantly in tributaries draining to coastal marine systems due to generally high nutrient concentrations in these systems. In this spirit, the authors proposed that in addition to abundant cryptophyte prey, high nutrient concentrations, and associated reduction of microzooplankton grazing pressure, tributaries draining to coastal systems constitute environments where prolific growth and toxic outbreaks of Pfiesteria species can occur. Lastly,

Stoecker and Gustafson (2002) proposed that in addition to elevated nutrient concentrations in tributaries emptying into coastal systems, relatively lower salinities that characterize these environments can contribute to lower microzooplankton grazing rates and, consequently, less inhibited growth of dinoflagellate populations (Figure 4.2).

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Figure 4.2. Potential Ingestion of Pfiesteria Zoospores by Microzooplankton according to Salinity (Stoecker and Gustafson 2002).

Naumenko (1992) noted that along with water temperature and predation by finfish, salinity is an important determinant of microzooplankton abundance and activity in the Vistula Lagoon

(Poland/Russia). Garcia and López (1989) found that significant declines in microzooplankton abundance and a corresponding increase in the abundance of algae occur in Laguna Joyuda,

Puerto Rico following precipitation events. According to Stoecker and Gustafson (2002), net growth of Pfiesteria populations in the tributaries of Chesapeake Bay during the summer of 1997 was higher at low salinities because mortality rates of these dinoflagellates due to grazing were relatively lower. Furthermore, microzooplankton grazing rates can be comparatively lower in more eutrophied, less saline tributaries and their estuaries due to higher abundance of mesozooplankton, such as copepods, which prey on microzooplankton and, unlike microzooplankton, have demonstrated preference for environments abundant in nutrients

84

(Roman, Holliday, and Sanford 2001; Roman, Reaugh, and Zhang 2006; Calliari et al. 2006). As it will be discussed below, mesozooplankton prey on microzooplankton and ingest these organisms at higher rates than alternative prey available (Hansen et al. 1993; Calbet and Saiz

2005; Liu et al. 2005). Mesozooplankton also ingest Pfiesteria (Burkholder and Glasgow 1995;

Mallin et al. 1995; Roman, Reaugh, and Zhang 2006). It is probable that how interactions

between copepods and actively toxic Pfiesteria play out in natural settings might be different to how these organisms interact in culture studies; it is likely that mesozooplankton will not

constitute a major factor influencing Pfiesteria population dynamics. First, copepods are not considered efficient grazers (Lawrence et al. 2004; Calliari, Britos, and Conde 2009). On average, they can clear as much as ≈22% of extant phytoplankton communities; most commonly, however, this percentage is as low as 6-10% (Calbet 2001). Studies have shown that although copepods can graze on HABs, in comparison to grazing exacted by microzooplankton, their contribution to reduction and eradication of these blooms is minor (Campbell et al. 2005; Turner and Borkman 2005). In a study in the Mediterranean, Calbet et al. (2003) found that while microzooplankton grazed A. minutum at considerable rates (at times surpassing the rate of population growth of this dinoflagellate), copepods contributed little to removal of these dinoflagellates. In the modeling simulation performed by Hood et al. (2006), ingestion of

Pfiesteria cells by A. tonsa —with microzooplanktonic grazers absent—was not a significant determinant of Pfiesteria abundance. According to Roman, Reaugh, and Zhang (2006), to effectively influence Pfiesteria population size, A. tonsa would have to achieve very high concentrations (>10 copepods/L). Mesozooplankton do not constitute considerable grazing force in GSB and it is microzooplankton rather than mesozooplankton that exert a considerable grazing pressure on smaller phytoplankton in this system (Caron et al. 1989; Lonsdale et al.

85

1996; Boissonneault-Cellineri et al. 2001). Second, mesozooplankton often prey on microzooplankton (Jonsson and Tiselius 1990; Buskey and Hyatt 1995; Nielsen and Kiørboe

1994), organisms—which, as it was discussed earlier—may significantly influence Pfiesteria

population dynamics (Burkholder and Glasgow 1995). Copepods do not feed passively but

obtain their nutrients in a selective manner (Turner and Tester 1989; Gismervik and Andersen

1997; Hirst and Bunker 2003). Because they are unable to graze small microorganisms

efficiently (Paffenhöfer 1984; Berggreen, Hansen, and Kiørboe 1988; Richardson and Verheye

1998), when presented with a choice, mesozooplankton tend to concentrate their grazing effort

on larger cells (Berggreen, Hansen, and Kiørboe 1988; Paffenhöfer 1988; Calliari, Britos, and

Conde 2009) ingesting microzooplankton at higher rates than rates at which they feed on other

microorganisms (Hansen et al. 1993; Calbet and Saiz 2005; Liu et al. 2005), including toxic

dinoflagellates (Gill and Harris 1987; Schnetzer and Caron 2005). Mesozooplanktonic

commonly select microzooplankton even if alternative prey is more abundant (Stoecker and

Egloff 1987). Studies have indicated that grazing activity of mesozooplankton might effectively

influence population dynamics of microzooplankton (Kiørboe and Nielsen 1994; Nielsen and

Kiørboe 1994; Calbet and Saiz 2005) and promote population growth of phytoplankton which

microzooplankton would normally ingest at significantly higher rates (Glaus-Porter 1976;

Buskey and Hyatt 1995; Stibor et al. 2004). Schnetzer and Caron (2005) and Merrell and

Stoecker (1998) demonstrated that as mesozooplankton grazed on microzooplankton, diatom

populations flourished. Some mesozooplanktonic grazers behave in a similar fashion with

respect to toxic dinoflagellates. According to Sipura, Lores, and Snyder (2003), grazing activity

of A. tonsa can free heterotrophic dinoflagellates (in part or completely) from both grazing and

competition pressures and effect unperturbed growth of their populations. Sheldon, Nival, and

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Rassoulzadegan (1986), demonstrated that dinoflagellate populations declined as a result of

microzooplankton grazing but remained stable upon introduction of mesozooplankton that

released dinoflagellates from heavy microzooplankton grazing. Considering prior studies that

investigated trophic interactions between dinoflagellates, micro-, and mesozooplankton

(Stoecker and Sanders 1985; Gifford and Dagg 1988; Hansen, Bjørnsen, and Hansen 1994),

Stoecker and Gustafson (2002) proposed that as mesozooplankton ingest microzooplankton at higher rates than they consume dinoflagellates, total grazing pressure on toxic Pfiesteria zoospores might decline significantly. In a modeling study, Hood et al. (2006) introduced A. tonsa into a medium containing cells of Pfiesteria TOX-A functional type and microzooplankton feeding on these actively toxic cells. Upon brief acclimatization period, A. tonsa began to graze microzooplankton while Pfiesteria concentration increased (Hood et al. 2006). According to

Wetz and Paerl (2008), mesozooplankton can control microzooplankton population dynamics in the NRE during summer (M. S. Wetz and J. C. Taylor, unpublished results). Lonsdale et al.

(1996) noted that summer copepod populations in GSB commonly prey on microzooplankton at higher rates than rates at which they ingest other prey and, by doing so, constitute an important control of microzooplankton population dynamics in this coastal lagoon as well. Frequently, ciliate mortality rate would surpass the rate of their population growth rendering abundance of ciliates in GSB relatively low (Lonsdale et al. 1996). Chang and Carpenter (1985) found that the

Carmans River estuary—a major subestuary of GSB—is highly hospitable to blooms of G. aureolum due to low zooplankton grazing pressure. The authors found large copepod population in this section of GSB while, a year earlier, Chang (1984) detected no zooplankton grazing of G. aureolum cells. It appears very likely that low (Chang and Carpenter 1985) and zero (Chang

1984) predation-related loses of G. aureolum in the Carmans River estuary was a result of

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trophic interactions between micro- and mesozooplankton. Third, mesozooplankton ingest actively toxic dinoflagellate cells but at low rates (Turner and Tester 1989; Turner and Borkman

2005; Cohen, Tester, and Forward 2007). Breier and Buskey (2007) found that although A. tonsa ingested toxic K. brevis cells, rates at which this was occurring were considerably lower than rates at which this copepod ingested nontoxic dinoflagellate species available. Turner and

Anderson (1983) found low ingestion rates of A. hudsonica grazing on G. tamarensis . Guisande et al. (2002) noted that copepods fed on actively toxic cells of A. minutum , however, as A. minutum toxin content increased, copepods ingested more toxic cells at a significantly reduced rates. Similarly, Ives (1987) noted progressive reduction of copepod grazing rates in tandem with increasing toxicity level of ingested Alexandrium spp. cells. Dam and Colin (2005) recorded low rates of egg production by A. tonsa when the only prey presented to these copepods was P. mimimum . However, these rates increased significantly when diatom T. weissflogii was introduced into the culture medium, indicating that P. minimum did not constitute suitable food for A. tonsa . Analogous correlation can also develop between mesozooplankton and Pfiesteria .

Hood et al. (2006), for instance, noted “ Acartia ’s reduced preference for grazing on the TOX-A functional type” (Hood et al. 2006, 475). Fourth, although mesozooplankton can ingest actively

toxic dinoflagellate cells and suffer no ill health effects (Frangópulos et al. 2000), frequently,

such effects do develop; at times, toxic dinoflagellates can even kill mesozooplanktonic grazers

(Delgado and Alcaraz 1999; Waggett, Hardison, and Tester 2012). Gill and Harris (1987) noted

mortality of copepods C. helgolandicus and T. longicornis upon ingestion of actively toxic G.

aureolum cells. C. polykrikoides , which has been a recurring problem in GSB, can kill their

mesozooplanktonic predators (Jiang et al. 2009; Jiang, Lonsdale, and Gobler 2010). Some of the

non-lethal effects that can result from interactions between actively toxic dinoflagellates and

88 mesozooplankton include reduction of feeding rates and egg production, and general survivability (Huntley et al. 1986; Turner and Tester 1997; Collumb and Buskey 2004). A. tonsa produced fewer eggs upon persistent ingestion of toxic K. brevis cells (Collumb and Buskey

2004). Jiang et al. 2009, Jiang, Lonsdale, and Gobler (2010) and Jiang, Lonsdale, and Gobler

(2010a) recorded reduced rates of grazing, egg production, and egg hatching in A. tonsa as the copepods consumed actively toxic cells of C. polykrikoides . Gill and Harris (1987) noted that copepods C. helgolandicus and T. longicornis experienced severely reduced rates of egg production when fed actively toxic cells of G. aureolum . In a number of instances, behavioral changes occurred in copepods ingesting toxic dinoflagellates, alterations that affected copepod swimming, photosensitivity, and general survivability (Cohen, Tester, and Forward 2007; Berge et al. 2012; Hong et al. 2012). Sykes and Huntley (1987) noted adverse effects of P. brevis on locomotory faculties of C. pacificus . Huntley et al. (1986) recorded analogous effect in Calanus pacificus exposed to P. brevis . Although there have been no recorded cases of copepod mortalities upon ingestion of actively toxic Pfiesteria cells, behavioral changes have been noted.

Mallin et al. (1995), for instance, noted that although some A. tonsa specimens preyed on toxic

P. piscicida cells, upon their ingestion copepods moved erratically and collided with walls of the culture vessel. Such behavior might render copepods more vulnerable to predation and thus be detrimental to general survivability of these organisms in natural setting (Mallin et al. 1995;

Burkholder and Glasgow 1997). Lastly, mesozooplankton tend to settle for ingestion of the most abundant prey organisms (Turner and Tester 1989; Abreu et al. 1994; Gismervik and Andersen

1997). Similar to other toxic dinoflagellates (Örnólfsdóttir, Pinckney, and Tester 2003; Turner and Borkman 2005; Turner 2006), Pfiesteria do not require to be highly concentrated to become toxic (Burkholder et al. 1993; Lovko et al. 2003; Rublee et al. 2005). Consequently,

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mesozooplankton is less likely to ingest toxic Pfiesteria cells while concentrating their grazing efforts on most numerous prey. Considering low affinity of microzooplankton for low salinity waters as well as low ingestion rates of toxic Pfiesteria cells by micro- and mesozooplankton, prior to the passage of superstorm Sandy, Pfiesteria bloom formation and toxin production appear to have been most feasible in the eastern extremity of this system where more brackish environment was the result of proximity to emptying streams and rivers and long distance from the Fire Island Inlet and the influence of more saline water of the coastal ocean.

Interactions with Bivalve Mollusks

Bivalve mollusks are diverse and ecologically important class of aquatic and marine shellfish

(Ward et al. 1993; NOAA 2013; Beukema and Dekker 2014). Indeed, bivalve mollusks have been recognized for their remedial contribution to water quality as their method of feeding involves filtration and, consequently, removal of toxic or otherwise undesirable microorganisms from the water column (Gainey and Shumway 1991; Philippart et al. 2007; Harke, Gobler, and

Shumway 2011).

Shallow depths, restricted circulation, long residence times (Phlips, Badylak and Grosskopf

2002), and the presence of SAV beds, where bivalve mollusks are more likely to feed, spawn, and seek refuge (Fay, Neves, and Pardue 1983; Dennison, Marshall, and Wigand 1989;

Hernández Cordero et al. 2012), make coastal lagoons excellent bivalve mollusk habitats; their populations can reach very large sizes and bivalve mollusks can dominate benthic animal communities in these systems (Dekker 1989; Kjerfve 1994; Bacher, Bioteau, and Chapelle

1995). Consequently, many coastal lagoons are heavily utilized for shellfish aquaculture (Melià et al. 2003; Loubersac et al. 2007; Torello and Smith 2014). Thau lagoon in southern France is

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inhabited by ≈100 molluscan species (de Wit 2011) while boasting one of the most abundant oyster populations in France (Dupuy et al. 2000). This coastal lagoon is considered to be “the biggest shellfish breeding area in Europe” (Gilbert et al. 1997, 144); shellfish-farming occupies

≈20% of the total area of this system (La Jeunesse and Elliott 2003). Ria Formosa (Portugal) boasts a healthy bivalve mollusk fishery (Chícharo and Chícharo 2001; Caetano, Vale, and

Bebianno 2002; Martins et al. 2003). 90-95% of the total Portuguese yield of these shellfish originate in this system (Chícharo and Chícharo 2000; Santos et al. 2004). Olenin (1997) found that one species bivalve mollusk species—D. polymorpha —occupied ≈23% of the benthic sediment surface in Curonian Lagoon (Russia/Lithuania). As it is with respect to finfish, APES is the most important shellfish nursery along the Atlantic Coast of the United States (Burkholder and Glasgow 1997, 2001). Geomorphological arrangement (shallow depths and restricted circulation) also renders GSB an attractive bivalve mollusk habitat (SSERC 1999a; LoBue and

Bortman 2011). For example, although the size of GSB eastern oyster population has shrunk severely over the years (Sebastiano 2012), eastern oysters are cultivated in aquaculture (Kelly and Chistoserdov 2001; Sebastiano 2012). Judging by limited mortalities of these populations, the physical, chemical, and biological framework of GSB appears to make this coastal lagoon a system where growth and survival of bivalve mollusks can take place (Sebastiano 2012).

The fact that some species of bivalve mollusks can ingest toxic Pfiesteria cells and suffer

no adverse effects might not be an effective mechanism by which these animals control

population dynamics of toxic dinoflagellates (Newell 2004). First, in certain instances, bivalve mollusks simply avoid ingestion of toxic dinoflagellates. Eastern oysters (Shumway and Cucci

1987; Hégaret, Wikfors, and Shumway 2007) and hard clams (Shumway and Cucci 1987;

Hégaret et al. 2007), for example, have been noted to avoid ingestion of some toxic

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dinoflagellate species by closing their shells. Second, although bivalve mollusks can ingest toxic

dinoflagellates, rates at which this ingestion occurs is often significantly lower than ingestion

rates of nontoxic species/strains (Hégaret and Wikfors 2005). May et al. (2010) found that hard clams ingested toxic cells of A. monilatum at lower rates than they ingested Cryptomonas sp. or a nontoxic strain of A. tamarense . According to Shumway and Cucci (1987), a similar dynamics can develop between hard clams and actively toxic P. tamarensis . Wikfors and Smolowitz

(1993) found that when hard clams were exposed to toxic P. minimum , their growth rate experienced significant reductions despite the availability of suitable prey. Springer et al. (2002) found that eastern oysters ingested Pfiesteria TOX-B and NON-IND zoospores at a rate of

≈3000 cells/ml/hr, while the rate at which they grazed on TOX-A zoospores was ≈1700 cells/ml/hr. Similarly, Shumway, Burkholder, and Springer (2006) noted that the rates at which adult bay scallops, eastern oysters, and hard clams grazed toxic strains of P. shumwayae were significantly lower than rates at which these bivalve mollusks ingested nontoxic strains of this dinoflagellate. Burkholder et al. (1993) noted that as bay scallops filtered low concentrations of toxic Pfiesteria zoospores, the rate at which they opened and closed their shells declined significantly. Third, fecal material excreted by bivalve mollusks can contribute significantly to eutrophication of their immediate environment (Bartoli et al. 2001; Izawa and Kobayashi 2006;

Byron and Costa-Pierce 2011) and, as a result, initiation of HABs. Studies demonstrated significant concentrations of organic nutrients in sediments directly below shellfish beds which, under suitable conditions, can be remineralized and reintroduced into the water column (Haven and Morales-Alamo 1966; Doering, Oviatt, and Kelly 1986; Hartstein and Stevens 2005).

According to data collected by De Vries et al. (unpubl.) in Lake Grevelingen, Holland, deposition of fecal material on sediment surface by resident bivalve mollusk populations

92 supplied 3-4 times more nutrients than nutrients delivered via other ways (Hily 1991). In a study by Haven and Morales-Alamo (1966), the rate at which bivalve excrement was deposited in benthic sediments was 7 times higher than the rate at which nutrients generated via other means settled to the bottom. It is important to reference the significance of such nutrient delivery as it might influence population dynamics of phytoplankton and SAVs (Doering, Oviatt, and Kelly

1986; Kautsky and Evans 1987). Peterson and Heck (1999) demonstrated that fecal material excreted by oysters significantly contributed to seagrass productivity in St. Joseph Bay, Florida.

Doering, Oviatt, and Kelly (1986) found that overall production of their mesocosm culture doubled upon addition of hard clams. Souchu et al. (2001) hypothesized that one of the main reasons as to why abundant bivalve mollusk populations in Thau lagoon (southern France) failed to control the growth of bloom-forming organisms was the abundant supply of nutrients derived from fecal material that these animals excreted. Such studies as Doering, Oviatt, and Kelly

(1986) and Bartoli et al. (2001) concluded that nutrients supplied to sediments by bivalve mollusks can stimulate activity of phytoplankton and effectively cancel their overall grazing effort. Fourth , many dinoflagellates can survive passage through digestive tracts of bivalve mollusks that ingest them and regain their motility upon reintroduction into water column following bivalve excretion (Bauder and Cembella 2000; Hégaret et al. 2008; May et al. 2010).

Viable cells of A. fundyense (Hégaret et al. 2008) and A. monilatum (May et al. 2010) have been detected in biodeposits of hard clams. Passage of intact dinoflagellate cysts through digestive tracts of adult bay scallops has been described in Hégaret et al. (2008) for P. minimum . Bioactive cells of such species as P. minimum (Hégaret and Wikfors 2005), A. fundyense (Hégaret et al.

2008), and A. monilatum (May et al. 2010) can emerge from biodeposits of eastern oysters.

Pfiesteria have been also recorded to survive the passage through digestive tracts of bivalve

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mollusks. According to findings of Shumway, Burkholder, and Springer (2006), 23 of 49

Pfiesteria cysts excreted by hard clams in biodeposits gave rise to fully flagellated individuals capable of producing toxins and killing finfish. In a similar fashion, Springer et al. (2002) demonstrated that 50% of Pfiesteria zoospores consumed by eastern oysters excysted and recovered full motility within 6 hours of being excreted.

Given the poor health of bivalve mollusk populations in GSB, as well as general unpalatability, nutritional inadequacy, and—therefore—low rates at which bivalve mollusks ingest toxic Pfiesteria cells, it appears that, in regards to bivalve mollusk predation pressure, prior to superstorm Sandy, GSB constituted an environment where Pfiesteria blooms appear to

have been feasible.

5. Conclusion

Superstorm Sandy was one of the largest storms ever to hit the northeastern United States. Apart

from loss of life and damages to residential neighborhoods and coastal infrastructure, Sandy also

affected coastal aquatic and marine ecosystems (Blain 2012). Eleven billion gallons of untreated

and partially treated sewage contaminated coastal waters as it flowed from flooded sewage

treatment facilities (Kenward, Yawitz, and Raja 2013). In Great South Bay, storm surge cut

through Fire Island creating three new tidal inlets (NPS 2013; NY Sea Grant 2013; Vergano

2013) and the U.S. Army Corps of Engineers had to close two of these shortly after the storm

(Dooley 2013; Vergano 2013), but left the third one—the Old Inlet—open (Aretxabaleta,

Butman, and Ganju 2014; NY Sea Grant 2013).

Following Sandy and the resulting breach in Fire Island, the previously restricted eastern

extremes of GSB where Pfiesteria outbreaks appear to have been most probable, appear to have

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become less likely to host Pfiesteria bloom formation. With Old Inlet connecting to the ocean, residence times in the previously sheltered sections of GSB are much shorter than before Sandy affected this system and, consequently, the physicochemical and biological conditions of these waters are now different. While availability of finfish prey has increased (Dooley 2013; Vergano

2013; Goldberg 2014), parameters such as salinity, turbulent mixing, temperature, and availability of algal prey have changed (Kelley 2012; Chen and Brumer 2013; Flagg and Flood

2013) so that the probability of Pfiesteria bloom formation in GSB appears to have become lower than before Sandy affected this system. One of the conclusions of the current study is that residence time, determined by the degree to which water in a given body of water is renewed from the coastal ocean, is an important hydrological parameter as it influences a number of water quality parameters and, therefore, composition, distribution, and population dynamics of organisms in coastal marine systems (Monsen, Cloern, and Lucas 2002; Crump et al. 2004;

Cucco and Umgiesser 2006). HABs (Garcon, Stolzenbach, and Anderson 1986; Furnas, Smayda, and Tomas 1989; Kudela et al. 2008), including those caused by Pfiesteria (Burkholder,

Glasgow, and Hobbs 1995; Weber and Marshall 1999; Ferreira et al. 2005), have typically occurred in environments characterized by long water residence times. Short residence times seem to preclude Pfiesteria bloom formation. This has also been indicated by studies in coastal waters of New Zealand and South Carolina where blooms of these organisms have not been considered a threat as most of these systems are characterized by short water residence times

(Lewitus et al. 2002; Rhodes et al. 2002). Rublee et al. (2005) pointed out that the substantial depression of residence times in subestuaries of the Albemarle-Pamlico Estuarine System following the passage of hurricanes Dennis, Floyd, and Irene in 1999 was the predominant factor contributing to the absence of Pfiesteria outbreaks in that system since 1999. Pfiesteria , it has

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been suggested, had been flushed out from the Pamlico and Neuse River estuaries with copious

rainfall generated by the 1999 hurricane trio (Jochems and Martens 2001; Burkholder et al. 2004;

Rublee et al. 2005). This study explored suitability of GSB to Pfiesteria blooms prior to the passage of Sandy by analyzing the factors that influence Pfiesteria population dynamics in the physicochemical-biological context of a typical coastal lagoon. It was demonstrated that prior to the passage of Sandy the most tidally restricted sections of GSB appear to have constituted environments congenial to Pfiesteria bloom formation with respect to physicochemical and biological parameters that drive dynamics of Pfiesteria blooms. Included in this analysis was the magnitude of turbulent mixing and interactions of phytoplankton with zooplankton, parameters analysis of which did not facilitate the attainment of a satisfactory answer to the research question—whether GSB constituted an environment where, prior to the passage of Sandy,

Pfiesteria bloom formation and toxic activity were feasible.

Coyne et al. (2001) stressed that an analysis of physicochemical and biological factors that influence Pfiesteria population dynamics is crucial to predict and prevent blooms. However, as much as such analysis demonstrates that occurrence of Pfiesteria outbreaks in GSB appears to have been feasible prior to the passage of Sandy, it fails to explain why Pfiesteria outbreaks have never occurred in GSB despite the presence of these organisms in this system, the ability of these populations to produce toxins, and the apparent alignment of the factors which influence their population dynamics. Some of the possible reasons for the lack of Pfiesteria blooms in GSB prior to the passage of Sandy and after are outlined below.

HABs are complex phenomena (Jewett et al. 2007); their development, persistence, and senescence are unpredictable and their behavior erratic despite our knowledge of physicochemical and biological parameters which influence their population dynamics and toxic

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activity (Smayda 2002; Hall et al. 2008; Smayda and Trainer 2010; Cardoso 2012). One

particularly intense bloom of K. veneficum in the NRE was described as a “surprise dinoflagellate bloom” (Hall et al. 2008, 412). Similarly, identification of the mechanism governing the onset and persistence of A. anophagefferens blooms in GSB and other embayments of LI’s North and South Shores have also been rather difficult (Taylor, Gobler, and

Sañudo-Wilhelmy 2006; Wurch 2011). In fact, development and persistence of these blooms have been so detrimental to the local fishery that the Suffolk County Department of Health

Services intended to formulate a statistical model of brown tide population dynamics, and possibly, to predict their occurrence in GSB (Marks 1994). Outbreaks of Pfiesteria have also been unpredictable and erratic (FDA 1997; Weinhold 2002). As it was discussed in Chapter 4,

Pfiesteria are known as ‘phantom’ dinoflagellates because their blooms can develop unexpectedly and rapidly and disappear with similar stealth and swiftness (Burkholder et al.

1992; Steidinger et al. 1996; Burkholder and Glasgow 2001).

Pfiesteria population dynamics and bloom formation might be influenced by parameters and conditions that have not been discussed in available literature but which are known to influence population dynamics of other toxic dinoflagellates. For instance, it is possible that ecological success of Pfiesteria in GSB has been precluded by dense populations of A. anophagefferens and recurrent brown tide events in this system. As it was discussed in Chapter

4, toxic dinoflagellates often co-occur with highly concentrated Aureococcus anophagefferens and cells comprising significant proportion of planktonic communities during brown tide events (Buskey and Hyatt 1995; Caron et al. 1989; Lonsdale et al. 1996;

Sieracki et al. 2004; Deonarine et al. 2006). The fact that there has been no documentation available dealing with possible interactions between Pfiesteria and brown tide species does not

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rule out the possibility that their blooms have been the causative factor that precluded Pfiesteria

bloom formation in GSB. Another possible reason for the absence of Pfiesteria outbreaks in

GSB is interactions of these dinoflagellates with viruses. Viruses are abundant in marine environment (Fuhrman 1999; Wommack and Colwell 2000; Bibak and Hosseini 2013), and their presence was also observed in GSB (Vaughn 1979; Gobler et al. 2004, 2007). They are ecologically important as apart from constituting prey for larger microorganisms (Fuhrman and

Suttle 1993; Fuhrman 1999) viruses infect and are able to lyse HAB phytoplankton cells and, therefore, influence population dynamics of these organisms and contribute to preclusion of their blooms (Tarutani et al. 2001; Gobler et al. 2007; Bibak and Hosseini 2013).

Although the probability of Pfiesteria outbreaks appears to have become lower in GSB due to shorter residence times in this coastal lagoon following the passage of Sandy and opening of the Old Inlet, presence of these organisms in GSB waters should not be discounted. Causes of the absence of Pfiesteria blooms by factors not yet known which might influence such blooms in the future must be investigated further not only to better understand governing dynamics of

Pfiesteria species but to comprehend population dynamics and toxic activity of similar toxic dinoflagellates. In the concluding remarks of “ And the Water Turned to Blood ”—an account of

Pfiesteria blooms in North Carolina—Rodney Baker states that “we still don’t know whether

Pfiesteria will be a plague upon our waters of Old Testament proportions. Crucial scientific

questions remain to be answered” (Baker 1997, 328). Indeed, there is still much that we need to

learn about Pfiesteria , including their interactions with A. anophagefferens and viruses and the

exact mechanism governing their toxin production. As our knowledge of these organisms grows

we are becoming better equipped to keep their populations in check and mitigate negative effects

of their outbreaks more effectively.

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