Tampa Bay Estuary Program 

Technical Publication # 03-04 

SEAGRASS COMMUNITIES of the GULF 

COAST of FLORIDA: STATUS and ECOLOGY 

FINAL REPORT 

AUGUST 2004 
CLINTON J. DAWES

RONALD C. PHILLIPS

GEROLD MORRISON

August 2004

CLINTON J. DAWES

University of South Florida

Tampa, Florida, USA

RONALD C. PHILLIPS

Institute of Biology of the Southern Seas

Sevastopol, Crimea, Ukraine

GEROLD MORRISON

Environmental Protection Commission of Hillsborough County

Tampa, Florida, USA

August 2004
ii

CITATION

Dawes, C.J., R.C. Phillips, and G. Morrison. 2004. Seagrass Communities of the Gulf Coast of Florida:

Status and Ecology. Florida Fish and Wildlife Conservation Commission Fish and Wildlife Research

Institute and the Tampa Bay Estuary Program. St. Petersburg, FL. iv + 74 pp.

COPIES

This document may be obtained from the following agencies:

Tampa Bay Estuary Program FWC Fish and Wildlife Research Institute

100 8th Avenue SE 100 8th Avenue SE

Mail Station I-1/NEP ATTN: Librarian

St. Petersburg, FL 33701-5020 St. Petersburg, FL 33701-5020

Tel 727-893-2765 Fax 727-893-2767 Tel 727-896-8626 Fax 727-823-0166

www.tbep.org http://research.MyFWC.com

AUTHORS

Clinton J. Dawes, Ph.D.

Distinguished University Research Professor

University of South Florida

Department of Biology

Tampa, FL 33620

cdawes@cas.usf.edu

Ronald C. Phillips, Ph.D.

Associate

Institute of Biology of the Southern Seas

2, Nakhimov Ave.

Sevastopol 99011

Crimea, Ukraine

ronphillips67@hotmail.com

Gerold Morrison, Ph.D.

Director, Environmental Resource Management

Environmental Protection Commission of Hillsborough County

3629 Queen Palm Drive

Tampa, FL 33619

813-272-5960 ext 1025

morrisong@epchc.org
iii

TABLE of CONTENTS

iv Foreword and Acknowledgements

1 Introduction

6 Distribution, Status, and Trends

15 Autecology and Population Genetics

28 Ecological Roles

42 Natural and Anthropogenic Effects

49 Appendix: Taxonomy of Florida Seagrasses

55 References
iv

FOREWORD

The waters along Florida’s Gulf of Mexico coastline, which stretches from the tropical Florida Keys in the

south to the temperate Panhandle in the north, contain the most extensive and diverse seagrass meadows

in the United States. Seagrass meadows rival or exceed most kinds of agriculture in their productivity and

also provide unique aesthetic and recreational opportunities. The importance of seagrasses as food, shelter, and essential nursery habitats for commercial- and recreational-fishery species and for the many other

organisms that live and feed in seagrass beds is well known. A single acre of seagrass can produce over 10

tons of leaves per year and can support as many as 40 thousand fish and 50 million invertebrates. This

high level of production and biodiversity has led to the view that seagrass communities are the marine

equivalent of tropical rainforests.

The importance of seagrasses to society has become fully recognized by government agencies. Seagrasses are now receiving focused attention from environmental managers, who require integrated science

to aid in developing seagrass-protection programs. Studies concerning the ecology, biology, and management of Gulf-coast seagrasses are increasingly diverse and complex; yet a synthesis of this research has not

been prepared since the late 1980s. The need for an up-to-date synthesis has resulted in the production of

this document, which compiles and organizes the many diverse research efforts that have been accomplished for this region since that time.

Holly Greening David W. Crewz

Senior Scientist Research Scientist

Tampa Bay Estuary Program Florida Fish and Wildlife Conservation Commission

Fish and Wildlife Research Institute

August 2004

ACKNOWLEDGEMENTS

Many individuals contributed their time and expertise towards the completion of this document. Foremost among them are the three authors, Clinton J. Dawes (University of South Florida), Ronald C.

Phillips (now at the Institute of Biology of the Southern Seas, Crimea, Ukraine), and Gerold Morrison

(now at the Environmental Protection Commission of Hillsborough County). David W. Crewz (Florida

Fish and Wildlife Conservation Commission, Fish and Wildlife Research Institute [FWC–FWRI]) provided outstanding contributions to content, editorial review, and document design.

A dedicated group of seagrass scientists and resource managers located throughout the Gulf coast of

Florida and elsewhere provided documentation (published and unpublished) and helpful comments on

earlier drafts of this document. We appreciate the comments provided by Diane Altsman, Walt Avery, Seth

Blitch, Catherine Corbett, Frank Courtney, Mike Crane, Tom Cuba, Kellie Dixon, Mark Flock, Mark Fonseca, Tom Frazer, Roger Johansson, Alice Ketron, George Kish, Eric Lesnett, Jeannine Lessman, Graham

Lewis, Robin Lewis, Kevin Madley, Rob Mattson, Gary Raulerson, Bill Sargent, Andy Squires, Betty Staugler, Larinda Tervelt, Dave Tomasko, Tom Ries, Bob Virnstein, and Kim Yates. Llyn French (FWC–FWRI)

provided an invaluable service by designing and composing the document for publication.

This document is a joint product of the Gulf of Mexico Program (Larinda Tervelt, project lead), the

Florida Fish and Wildlife Conservation Commission, Fish and Wildlife Research Institute ([formerly the

Florida Marine Research Institute] Kevin Madley, project lead), the United States Geological Survey

(Jimmy Johnson, project lead), and the Tampa Bay Estuary Program (Holly Greening, project manager).
NTRODUCTION

The goals of this review are to update what has

become known about seagrasses since the publication of Zieman and Zieman’s review in 1989 and to

assess the current status of seagrass habitats on the

Gulf coast of Florida.

Published studies about seagrasses have

increased substantially since the 1960s.

Seagrass losses on Florida’s Gulf coast, documented in the 1970s and 1980s, became the

basis for state and local government involvement in developing management approaches.

Whether continuous or patchy in plant coverage, all seagrass communities should be viewed

as having the same valued functions, such as

serving as nurseries.

At least thirteen ecological roles have been

assigned to seagrass communities, including

roles as primary producers, as habitat for animals and plants, and as support for food webs.

Florida Gulf-coast seagrass communities support multimillion-dollar commercial and recreational industries, especially with regard to

fishing. Turtlegrass, Thalassia testudinum

Ron Phillips photo
2 GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

VISION and SCOPE

During the 1930s, most seagrass studies in the

northern hemisphere focused on the decline and

demise of Zostera marina L. (Eelgrass) on both

sides of the Atlantic Ocean. During this period,

only the work by Setchell addressed the distribution, taxonomy, and ecology of seagrasses (Setchell

1920, 1929, 1934, 1935). In the late 1950s, the

Florida State Board of Conservation (now the Fish

and Wildlife Research Institute [FWRI], a part of

the Florida Fish and Wildlife Conservation Commission) initiated research efforts dealing with the

ecology and distribution of Florida’s seagrasses

and their roles in coastal and

estuarine processes, especially

regarding food webs and habitat

characters supporting recreationally and commercially important

fish and shellfish. Part of this

effort culminated in a publication by Phillips (1960a), one of

the few at that time concerning

seagrass ecology or any aspect of

seagrass biology (Zieman 1987).

In 1978, a bibliography compiled

by the Seagrass Ecosystem Study listed over 1,400

titles worldwide (Bridges et al. 1978), and by 1982,

a community profile of south Florida seagrasses

contained over 550 references (Zieman 1982).

In a summary of seagrass studies published

over a period of 25 years, Zieman (1987) found

that initially nearly all seagrass literature was

descriptive and qualitative. By 1970, most published works were quantitative, and development

of conceptual models had begun. By 1980, increasingly robust models of the mechanisms by which

seagrass systems develop and maintain their productivity were being proposed and used as guides

for developing proposed research (e.g., McMillan

1978, 1980).

By 1982, scientists, resource managers, and

agency personnel monitoring and managing bays

and estuaries on the Atlantic and Gulf of Mexico

coastlines of Florida, such as for the Indian River

Lagoon, Charlotte Harbor, Sarasota Bay, Tampa

Bay, and Pensacola Bay, noted dramatic seagrass

losses. Starting around 1950, those areas experiencing large population increases also experienced

seagrass losses, probably as a result of increasing

development pressure (Figure 1.1). To address seagrass losses in the State of Florida, management

programs were initiated between 1985 and 1995 to

conserve and restore seagrass communities. The

State’s Surface Water Improvement and Management (SWIM) programs, within the Water Management Districts, address seagrass conservation

issues statewide. Federally sponsored National

Estuary Programs (NEPs) were designated for four

specific estuaries: Tampa Bay National Estuary

Program (TBNEP, now TBEP), Sarasota Bay

National Estuary Program (SBNEP), Charlotte

Harbor National Estuary Program (CHNEP), and

Indian River Lagoon Program (IRLP). Since that

time, considerable research, particularly regarding

the light requirements of different seagrass species,

has been conducted. This work was stimulated by

the development of goals and targets established

by the SWIM Districts and NEPs with respect to

reducing eutrophication and nutrient loadings in

Figure 1.1 Satellite

image of Florida,

showing population

centers. Graph

shows decline of

seagrass as 

population

increases. 

LANDSAT 2000 image processed by FWC–FWRI FWC–FWRI
the major bays and estuaries. In several of these

estuaries, the large seagrass losses observed in the

early 1980s have halted, and moderate gains in seagrass extents have been occurring since about 1988

in some areas.

Although many recent studies concerning seagrass ecology and biology along Florida’s Gulf

coast have been and are being conducted, a synthesis of this information has not been compiled since

the 1989 publication of “Ecology of the Seagrass

Meadows of the West Coast of Florida: A Community Profile” by Zieman and Zieman. Several

resource-management and science programs have

identified the need for an updated synthesis of seagrass information:

The FWC Fish and Wildlife Research Institute

has developed a framework for a statewide Seagrass Conservation Plan (Morrison et al.

2003a).

In August 2000, over 70 seagrass scientists and

managers attended a Seagrass Management

Symposium convened by the Tampa Bay Estuary Program (Greening 2002a). Workshop participants identified, as a first and critical step in

seagrass management, the need for a compilation of scientific information on Florida’s seagrasses published since the review by Zieman

and Zieman (1989).

The US Geological Survey’s (USGS) Coastal

and Marine Geology Program initiated a “Gulf

of Mexico Estuaries Assessment” in 2001, using

Tampa Bay as the subject of a pilot study. An

initial element of the study is to develop a “synthesis report” and web-based information bank

that would link directly to the USGS National

Estuaries Assessment.

The Gulf of Mexico Program (GMP) made a

commitment that “By 2004, the GMP will complete development of an updated gulf-wide

characterization of the status and trends of seagrasses and coastal wetlands” (Gulf of Mexico

Program 2003).

Knowledge of seagrass ecology and distribution

within Florida has progressed substantially in the

last 20 years. The objective of this publication is to

summarize available data and information about

seagrass research performed along the Gulf coast

of Florida since 1985. Literature and studies published prior to 1985 are summarized in Zieman

and Zieman (1989), which is available from the

FWC Fish and Wildlife Research Institute Web site

(www.floridamarine.org). The geographical scope

of this document extends from Florida Bay and

the Florida Keys at the southern extreme northward and westward through the Florida Panhandle

to the Alabama border (Figure 1.2). Although the

distance is only about 700 km (435 miles) from

Florida Bay to Apalachicola Bay (extending over

6.5° of latitude), the aquatic climate changes dramatically. In Florida Bay, conditions are tropical,

whereas in the Panhandle region, conditions are

temperate and delimit the northern distribution in

the Gulf for several Florida seagrass species,

including Thalassia testudinum (Turtlegrass; Figure 1.3) and Syringodium filiforme (Manateegrass;

Figure 1.4).

CHAPTER 1 | INTRODUCTION 3

Figure 1.2 Demarcated watersheds of Florida’s Gulf coast.

Figure 1.3 Thalassia testudinum (Turtlegrass)

Figure 1.4 Syringodium filiforme (Manateegrass)

Graphic by FWC–FWRI adapted fromNOAA Coastal Assessment Framework

Ron Phillips Ron Phillips
DEFINITION of SEAGRASS HABITAT

Simply put, habitat is where an organism resides.

Nevertheless, habitat is not only where organisms

live but also includes how they live there. Seagrass

habitat is an ecological function provided by seagrasses. It is the location where certain organisms

can thrive (survive, grow, and reproduce).

Seagrass habitat is defined in this document as

a physical space containing seagrasses in sufficient

quantity and pattern to produce the appropriate

structural and physiological characteristics to support organisms typical of seagrass communities.

These characteristics include food webs based on

organic-matter production, nutrient cycling, detritus production, shelter, and sediment formation.

Continuous-coverage beds as well as patchy

beds of seagrasses provide critical and valued habitat functions. Fonseca et al. (1998) found that scattered or patchy Z. marina beds covered many

thousands of acres of estuarine seafloor in North

Carolina, had shoot densities and primary production equivalent to those of continuous-coverage

beds, had significantly greater below-ground biomass than did continuous-coverage beds, and

often supported densities of economically valuable

animal species, e.g., pink shrimp, similar to those

of continuous-coverage seagrass beds. They concluded that seagrass habitat must be recognized as

indicating not only continuous-coverage seagrass

beds, but also chronically patchy areas, therein

considering the unvegetated spaces between vegetation as seagrass habitat as well (Figure 1.5).

Thus, the value of seagrass habitat should not

be judged on the basis of seagrass densities or patterns, but upon the provided functions. In this

document, any reference to areas covered by seagrass means seagrass habitat, as long as valued

functions are present and measurable. One should

consider that patchy seagrass beds perhaps represent areas in the process of recovering from past

disturbances, or they may be areas held in a patchy

pattern because of the characteristics of the present physical environment. In any event, patchy seagrasses support valued animals and plants and

display typical seagrass functions.

ECOLOGICAL and 

ECONOMIC IMPORTANCE 

of SEAGRASS COMMUNITIES

Seagrasses are a vital component of Florida’s

coastal ecology and economy. They provide nutrition and shelter to animals that are important to

marine fisheries, provide critical habitat for many

other animals (e.g., wading birds, manatees, and

sea turtles), and improve water quality (Thayer et

al. 1997,1999; Livingston 1990; Kenworthy et al.

1988b; McMichael and Peters 1989; Stedman and

Hanson 1997; Valentine et al. 1997). For example,

Heck et al. (2003) found a strong link between seagrass abundance and those of juvenile finfish and

shellfish that was related to habitat structure.

In systems where seagrasses occur, nearly all of

the commercially and recreationally valuable estuarine and marine animals depend on seagrass beds

as refuge or habitat for parts or all of their life

cycles (Kikuchi and Peres 1977; Thayer et al. 1978,

1984; Kikuchi 1980; Ogden 1980; Thayer and

4 GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

Figure 1.5 A mixture of patchy and continuous seagrasses along a mangrove shoreline in Tampa Bay, Florida.

Gandy Photography, 2001
Ustach 1981; Phillips 1984). As reported by

Wingrove (1999) in the Florida Keys, hundreds of

fish species, including many of commercial value,

rely on seagrass habitats during some parts of their

life cycles. Seagrasses help support a thriving, multimillion-dollar recreational fishery including, as

an example, the shallow-water seagrass flats fishery

seeking bonefish and tarpon. In addition, over 30

species of tropical invertebrates that depend on

seagrasses are collected in the Florida Keys annually for the marine aquarium industry.

Short et al. (2000) list ecological services provided by seagrasses (modified here):

Primary production (food for animals and support for fisheries and wildlife)

Canopy structure (habitat, refuge, nursery, settlement and support of fisheries)

Epibenthic and benthic production (support of

food webs and fishery support)

Nutrient and contaminant filtration (improved

water quality, support of adjacent habitats, support of fisheries)

Sediment filtration and trapping (improved

water quality, countered sea-level rise, support

of adjacent habitats)

Epiphytic substratum (support of secondary

production, production of carbonate sediment,

support of fisheries)

Oxygen production (improved water quality,

support of adjacent habitats, support of fisheries)

Organic-matter production and export (support of estuarine and offshore food webs, support of adjacent habitats, support of fisheries)

Nutrient regeneration and recycling (support of

primary production, support of adjacent habitats, support of fisheries)

Organic-matter accumulation (support of food

webs, countered sea-level rise, support of fisheries)

Dampening of waves and currents (prevention

of erosion/resuspension, support of adjacent

habitats, increased sedimentation)

Seed production/vegetative expansion (selfmaintenance of habitat, support of wildlife)

Self-sustaining ecosystem (recreation, education, landscape-level biodiversity)

Costanza et al. (1997) and Costanza (1999)

stated that, for the entire biosphere, the economic

value of all ecosystem services for 16 biomes is in

the range of 16 to 54 trillion US $ y–1, with an average of 33 trillion US $ y–1. They considered this to

be a minimum estimate. The value of coastal environments, including estuaries, coastal wetlands

(mangroves and salt marshes), seagrass beds and

algae, coral reefs, and continental shelves, is of a

disproportionately high value. These communities

cover only 6.4% of the world’s surface, but they are

responsible for 43% of the estimated value of the

world’s ecological services.

In Florida, seagrass beds are directly responsible

for bringing in millions of dollars annually from

out-of-state and resident recreational boaters and

fishermen and commercial fishermen (Bell 1993;

Milon and Thunberg 1993; Virnstein and Morris

1996; Virnstein 1999; Wingrove 1999; Thomas and

Stratis 2001). Seagrass beds on the Gulf coast of

Florida are important not only for the ecological

services they provide, but for the economic health

of the state and region.

DOCUMENT CONTENT

The focus of this review is the biology and ecology

of seagrasses and of seagrass communities on

Florida’s Gulf coast.

Chapter 2 considers distribution of the Florida

Gulf coast seagrasses and reports the trends in

areal extents of seagrass beds, as recorded by monitoring efforts of various local and regional programs. Chapter 3 synthesizes new information

regarding autecology and presents what is known

regarding genetic analyses of Florida seagrasses.

These genetic techniques were not applied to seagrasses prior to the 1990s. Chapter 4 addresses the

ecological roles of seagrass communities, of their

macroalgal components (epiphytic and drift), and

of adjacent coastal communities (mangroves and

salt marshes). Chapter 5 focuses on the natural

and anthropogenic effects on Florida seagrasses.

The Appendix presents keys to the Florida seagrass

species and presents brief taxonomic descriptions

for the families, genera, and species (includes figures).

A companion document entitled “The Florida

Seagrass Manager’s Toolkit” (Morrison et al.

2003b) addresses seagrass management in Florida

and is available at www.floridamarine.org.

CHAPTER 1 | INTRODUCTION 5
ISTRIBUTION, STATUS, 

and TRENDS

Florida’s Gulf coast can be divided into four regions—South Florida, Gulf Peninsula, Big Bend, and Panhandle—for the purpose of assessing near-shore seagrass

community status and trends.

Aerial photography taken during the 1990s revealed that the South Florida region

contained the majority (65%) of the Gulf coast’s seagrass coverage, followed by the

Big Bend (28%), Gulf Peninsula (5%), and Panhandle (2%) regions.

The most abundant seagrass species on the Florida Gulf coast are Thalassia testudinum, Syringodium filiforme, and Halodule wrightii, each of which principally

has a tropical to subtropical distribution. Two other species (Halophila engelmannii and H. decipiens) also occur in the area, in near-shore meadows dominated by

T. testudinum and S. filiforme and in deeper waters where the latter two species

are absent.

A substantial decline in seagrass coverage has occurred in the South Florida

region over the past 15 years, following a dramatic “die-off” that began in Florida

Bay during 1987.

In the Gulf Peninsula region between ca. 1959 and 1982, mapping efforts in

Tampa Bay, Sarasota Bay and Greater Charlotte Harbor revealed reductions in seagrass coverage. However, difficulty in obtaining accurate coverage estimates from

the 1950s-era maps has complicated attempts to quantify these declines, particularly in the Greater Charlotte Harbor area.

Increasing coverage trends have occurred in Tampa Bay and Sarasota Bay since

1982 in response to improved management of nitrogen loadings and increasing

water clarity. Increased rainfall, stormwater runoff, and nutrient loadings associated with the 1997–1998 El Niño event interrupted the trends in seagrass coverage

gains, but they appear to have resumed in recent years.

Several other areas within the Gulf Peninsula region—including Charlotte Harbor

Proper and Lemon Bay—have been mapped approximately biennially since 1988.

No significant seagrass coverage trends have been reported from these recent

mapping efforts.

The Big Bend region is a unique “zero-energy” coastline and contains Florida’s

second-largest near-shore seagrass bed. The region has received relatively little

research and management attention. The only long-term seagrass coverage

changes reported have been localized losses attributed to the effects of an industrial facility that discharges to the Fenholloway River and Apalachee Bay.

In portions of the Panhandle region, which is also poorly studied, seagrass coverage may be increasing in some low-salinity areas and declining in some areas of

higher salinity.

In addition to these near-shore seagrass resources, recent work on the West

Florida Shelf indicates the presence of extensive, seasonal, deep-water Halophila

beds, which may exceed four hundred thousand hectares (one million acres).

Shoalgrass, Halodule wrightii

Ron Phillips photo
CHAPTER 2 | DISTRIBUTION, STATUS, AND TRENDS 7

DISTRIBUTION

Seagrasses are a relatively small group of flowering

plants that have adapted to survive and reproduce

in the marine environment. They are present in all

coastal states of the U.S., with the exception of

Georgia and South Carolina, where a combination

of freshwater inflows, high turbidity, and large

tidal amplitude restricts their occurrence (Thayer

et al. 1997).

The most abundant taxa in Florida’s near-shore

waters are T. testudinum, S. filiforme, and H.

wrightii, each of which principally has a tropical to

subtropical distribution (Zieman and Zieman

1989). Thalassia testudinum (Turtlegrass) is the

largest of these species, with long strap-shaped

leaves and robust rhizomes (see Appendix for taxonomic keys, descriptions, and illustrations).

Extensive seagrass beds are usually dominated by

this species, either alone or in combination with

other species, such as Syringodium filiforme.

Syringodium filiforme (Manateegrass) can be distinguished by its cylindrical (terete) leaves that,

because they are brittle and buoyant, are frequently broken off from the parent plant and dispersed widely by winds and currents. Halodule

wrightii (Shoalgrass) has flat, narrow leaves and a

shallow root system. It is thought to be an early

successional species in the development of seagrass

beds in the Gulf of Mexico and Caribbean Sea.

Three other species, Halophila engelmannii

(Stargrass), H. decipiens (Paddlegrass), and H.

johnsonii (Johnson’s Seagrass), are also found in

Florida’s coastal waters. In the Big Bend region, H.

engelmannii and H. decipiens are scattered

throughout beds dominated by T. testudinum and

S. filiforme but also occur in deeper water where

these latter two species are absent (Iverson and Bittaker 1986). Halophila decipiens has been found in

the Big Bend and Tampa Bay regions and at depths

to 90 m near the Dry Tortugas (Zieman 1982), and

it forms single-species stands in depths of 20 m or

more, beyond the deep edge of the extensive T. testudinum/S. filiforme beds (Zieman and Zieman

1989, Dawes and Lawrence 1990). Halophila johnsonii is a relatively newly described species and is

morphologically similar to H. decipiens (Eiseman

and McMillan 1980). Halophila johnsonii is now

listed as a threatened species by the National

Marine Fisheries Service (NMFS; 2002) and is

apparently an endemic whose range is restricted to

the lagoon systems of Florida’s southeastern

(Atlantic) coast. It has not been documented to

occur on the Gulf coast (NMFS 2002), and recent

evidence suggests it is genetically indistinguishable

from H. ovalis, a species of the Indo-Pacific region

(Waycott et al. 2002).

A seventh species, Ruppia maritima (Widgeongrass), is a euryhaline plant that is often encountered in the waters of Florida’s Gulf coast,

particularly in estuaries such as Homosassa Bay

(Koch and Dawes 1991a, b) and Tampa Bay (Lazar

and Dawes 1991). This species can form dense

beds, as found in upper Tampa Bay (Lazar and

Dawes 1991). In recognition of its broad salinity

tolerance, some authors have suggested that R.

maritima may be thought of as a freshwater species

that is also capable of living in saline environments, rather than a seagrass in the strictest sense

(e.g., Zieman 1982, Kuo and den Hartog 2001).

In addition to seagrasses, drift and attached seaweeds also make up an important component of

the total submerged aquatic vegetation (SAV) in

many areas of Florida (Dawes et al. 1987, Dawes

1986). In the Big Bend region, for example, benthic green algae in the order Caulerpales— including Halimeda incrassata, seven species of Caulerpa,

and two species each of Udotea, Penicillus, and

Codium—are important associates in the region’s

seagrass beds, with standing crops exceeding those

of seagrasses in some areas (Mattson 2000).

Depth-related zonation patterns of Gulf-coast

seagrass beds (Figure 2.1) have been described by

Lewis et al. (1985a), Iverson and Bittaker (1986),

Zieman and Zieman (1989), and Mattson (2000).

As a general rule, H. wrightii and R. maritima tend

to be more abundant in shallow inshore areas

because they tolerate frequent tidal exposure and

low salinities. Thalassia testudinum and S. filiforme

Fig 2.1 An example of an inshore-offshore seagrass zonation

profile on Florida’s Gulf coast (from McNulty et al. 1972).
8

reach their highest abundance and biomass in

slightly deeper areas, often forming dense singleor mixed-species stands (Figure 2.2). In some

areas, H. wrightii exhibits a second abundance

peak along the deep-water edge of T. testudinum/S.

filiforme meadows (Iverson and Bittaker 1986, Zieman and Zieman 1989). Halophila spp. form

sparse beds in deeper waters outside these meadows (Zieman and Zieman 1989, Fonseca et al.

2001). In the Big Bend region and Tampa Bay, H.

engelmannii also occurs in low-salinity areas (<5

ppt) within 1–2 km of river mouths, where an

ability to tolerate low light levels in waters of relatively high color may be an important factor

explaining its persistence (Dawes 1967, Zimmerman and Livingston 1976, Mattson 2000).

When considering the distribution of seagrasses

on Florida’s Gulf coast, dividing the area into four

regions is helpful. Regions defined by Sargent et al.

(1995) are summarized in Table 2.1 and Figure 2.3:

South Florida (Florida Keys and Florida Bay to

Estero Bay)

Gulf Peninsula (Estero Bay to Anclote Key)

Big Bend (Anclote Key to Ochlockonee Point)

Panhandle (Ochlockonee Point to the FloridaAlabama border)

The spatial distribution and areal extent of seagrasses vary substantially between these regions

(Zieman 1982, Iverson and Bittaker 1986, Sargent

et al. 1995). Recent status and trends in seagrass

coverage at this geographic scale are summarized

in the following section.

REGIONAL STATUS and TRENDS

Sargent et al. (1995) estimated that, on a statewide

basis, Florida’s near-shore coastal waters support

approximately 1.1 million ha (2.7 million acres) of

seagrass. This statewide estimate includes 0.8 million ha (1.9 million acres) of dense and relatively

easily mapped seagrasses in state waters where visibility allowed interpretation of bottom communities (within 14.4 km [9 miles] of shore along the

Gulf coast). It also includes, in portions of the

South Florida region, an estimated 0.3 million ha

(0.8 million acres) of sparse and incompletely

mapped seagrass beds that are interspersed with

hard-bottom communities and are thus difficult to

map accurately. This estimate does not include the

sparse beds that occur in deeper waters on portions of the West Florida Shelf (Sargent et al.

1995).

Along the state’s Gulf coast, the coverage of the

sparse deep-water beds of the West Florida

Shelf and the small, patchy mixed-species beds

that occur intermixed with hard bottom outside

the main seagrass beds in Florida Bay remain the

largest question marks in the effort to develop

accurate estimates of overall seagrass coverage. For

example, recent assessments indicate that the total

area of deep-water beds in the Gulf-coast region

may be on the order of 0.4 million ha (1 million

acres), which would place them, on a worldwide

basis, among the most extensive seagrass habitats

currently known (Fonseca et al. 2001).

During the 1990s, aerial photographs were used

to produce digitized maps of seagrass coverage for

GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

Table 2.1 Extent of seagrass coverage and aerial photography dates in four regions of Florida’s Gulf coast

(Madley et al. 2003).

Seagrass Seagrass

Region (hectares) (acres)

Panhandle (1992) 17,474 43,178

Big Bend (1992) 247,598 611,815

Gulf Peninsula (1999) 43,323 107,051

South Florida (1992, 1995) 574,875 1,420,517

Gulf coast total 883,270 2,182,561

Fig 2.2 Dense seagrass beds begin as patches that coalesce into

a larger, more genetically diverse meadow. 
CHAPTER 2 | DISTRIBUTION, STATUS, AND TRENDS 9

each of the regions shown in Figure 2.3. The dates

of these mapping efforts and their results are summarized in Table 2.1. The South Florida region

contained the majority (65%) of the Gulf coast’s

seagrass coverage, followed by the Big Bend (28%),

Gulf Peninsula (5%), and Panhandle (2%) regions.

SOUTH FLORIDA

The South Florida region (Figure 2.3) includes the

coastal waters of Collier, Monroe, and MiamiDade counties (Table 2.1). The portion that lies

immediately south of Cape Romano includes the

Ten Thousand Islands, an area that is dominated

by mangrove islands and tidal channels but that

also contains patches of seagrasses and some large

seagrass beds (e.g., as reported from the Lostmans

River area by Dawes et al. 1995). The Florida Bay

portion, which lies south of Cape Sable and west

of the Florida Keys, is a carbonate-sediment-based

system that supports extensive seagrass beds.

Based on monitoring data collected annually

from 1974 through 1980, Iverson and Bittaker

(1986) noted that, in addition to their greater

extent, the 0.5 million-ha (1.4 million-acre)

Florida Bay seagrass meadows also had about two

to four times the short-shoot densities of T. testudinum and S. filiforme as occurred in the 0.3 million ha (741,000 acre) Big Bend meadows. They

hypothesized that the density differences observed

in the two areas may be a consequence of greater

seasonal variations in solar radiation and water

temperature in the Big Bend, which is at the northern limits of tropical American seagrasses (Iverson

and Bittaker 1986).

Seagrass coverage and condition in the South

Fig 2.3 County-based regions used to describe seagrass distributions along the Gulf coast of Florida (after Sargent et al. 1995)

and containing locations referred to in the text.
10

Florida region have changed since the 1974–1980

period documented by Iverson and Bittaker

(1986). A dramatic decline in coverage began in

western Florida Bay during the summer of 1987

(Hall et al. 1999). Seagrasses in the bay were apparently subjected to decreased light availability

resulting from resuspended sediments and widespread, persistent microalgal and cyanobacterial

blooms. Bay-wide surveys in 1984 and 1994 indicated that the biomass of T. testudinum, S. filiforme, and H. wrightii declined by 28%, 88%, and

92%, respectively, during that 10-year period. The

spatial patterns of seagrass losses suggested that

chronic light reductions, which affected all species,

and “die-off” (rapid, unexplained plant mortality),

which also affected T. testudinum, most likely

caused the overall decline. Although the loss rate

has slowed considerably in recent years, die-off

and regression of seagrasses are still occurring in

parts of the bay (see also Chapter 5).

GULF PENINSULA

The Gulf Peninsula region lies between Estero Bay

and Anclote Key (Figure 2.3) and includes the

coastal waters of Lee, Charlotte, Sarasota, Manatee,

Hillsborough and Pinellas counties. This region is

a moderate-energy coastline, with extensive sand

beaches and barrier islands that enclose two large

estuarine embayments (Tampa Bay and Charlotte

Harbor) and many smaller lagoons (e.g., Estero

Bay, San Carlos Bay, Matlacha Pass, Pine Island

Sound, Lemon Bay, Sarasota Bay, and St. Joseph

Sound) containing the majority of the region’s seagrass beds. Recent seagrass-coverage trends in this

region appear somewhat irregular, apparently

responding to site-specific situations within the

different estuary and lagoon systems.

For Tampa Bay, Sarasota Bay, and the Greater

Charlotte Harbor system, the earliest photography-based seagrass coverage maps for the region

were developed by the FWRI, using aerial photographs taken in the late 1940s and early 1950s and

again in 1982 (Harris et al. 1983, Tampa Bay

Regional Planning Council [TBRPC] 1984, Janicki

et al. 1994). Maps were subsequently prepared by

the Southwest Florida Water Management District

(SWFWMD) for Tampa Bay (for the years 1988,

GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

Figure 2.4 Time series of mapped seagrass coverage within major coastal water bodies of the Gulf Peninsula region (GCH =

Greater Charlotte Harbor; TB = Tampa Bay; CHP = Charlotte Harbor Proper; SJS = St. Joseph Sound; SB = Sarasota Bay; LB

= Lemon Bay) from ca. 1950 through 2002. The 1950 values were developed using mapping methods that differ from those

currently in use and should be viewed as approximations. (Data sources: FWRI, SWFWMD)
CHAPTER 2 | DISTRIBUTION, STATUS, AND TRENDS 11

1990, 1992, 1994, 1996, 1999, and 2002), Charlotte

Harbor (1988, 1992, 1994, 1996, 1999, and 2002),

Sarasota Bay (1988, 1994, 1996, 1999, and 2002),

Lemon Bay (1988, 1994, 1996, 1999, and 2002),

and St. Joseph Sound (1999 and 2002) (TBNEP

1996, Kurz et al. 2000, D. Tomasko pers. comm.).

Time series of seagrass coverages based on these

maps are shown in Figure 2.4.

In Tampa Bay, approximately 46% (7,452 ha or

18,400 acres) of the existing seagrass coverage was

lost between ca. 1950 and 1982 (Figure 2.4) from

the combined effects of dredging and reductions

in water clarity (Haddad 1989). Coverage losses in

Tampa Bay over longer time periods are difficult to

estimate with any accuracy, because of the sparseness of data and absence of aerial photography

prior to ca. 1950. Indirect methods suggest, however, that as much as 81% of predevelopment coverage may have been lost during the years 1879

through 1982 (Lewis et al. 1991).

Between 1982 and 1996 Tampa Bay regained

approximately 2,090 ha (5,160 acres) of seagrass,

apparently in response to management efforts that

led to reduced nutrient loadings and increased

water clarity. Reduced nutrient loadings and

increased seagrass coverage were also observed in

Sarasota Bay during the same period (Tomasko et

al. 1996). Seagrass coverage then declined in both

estuaries, apparently in response to the heavy rainfall and increased stormwater runoff that occurred

during the 1997–1998 El Niño event (Johansson

2002a).

Seagrass coverage increased once again in

Tampa Bay during the 1999–2002 mapping period,

as water clarity improved during the relatively dry

years that occurred following the cessation of the

1997–1998 El Niño event. During this period the

total mapped coverage in the bay increased by 501

ha (1,237 acres), to 10,561 ha (26,078 acres). In

Sarasota Bay, on the other hand, the total mapped

coverage declined slightly between 1999 and 2002,

from 4,799 ha (11,850 acres) to 4,740 ha (11,703

acres). Seagrass coalescence occurred in both estuaries during the 1999–2002 mapping period,

through a net increase in the coverage of continuous (as opposed to patchy) seagrass beds (D.

Tomasko pers. comm.).

The documented increases in seagrass coverage

in Tampa Bay and the slight increase in Sarasota

Bay between 1982 and 2002 apparently occurred in

response to improved management of anthropogenic nitrogen loads to both estuaries (Tomasko

et al. 1996). Through the Grizzle-Figg Act (403.086

Florida Statutes), the Florida Legislature required

that all sewage treatment plants discharging to the

two estuaries and their tributaries must provide

advanced wastewater treatment (AWT) prior to

discharge. The City of Tampa upgraded its sewage

treatment plant to AWT in 1979, greatly reducing

the amount of nitrogen entering Tampa Bay from

that source. The City of St. Petersburg implemented a wastewater reuse program which almost

eliminated its direct wastewater discharges to

Tampa Bay. Similar improvements to sewage treatment plants in Pinellas, Hillsborough, Manatee,

and Sarasota counties also helped improve water

quality in the receiving estuaries. By the early

1990s, water clarity in some of the most degraded

portions of Tampa Bay had already begun to

improve (Johansson 1991, TBNEP 1996).

Improved management of seagrass communities has been identified as a priority issue in Tampa

Bay and Sarasota Bay and is being addressed

through broad-based stakeholder groups. Both

systems are part of the U.S. EPA’s National Estuary

Program, which helps to coordinate the management activities of local, state, and federal agencies

and the private sector. The Tampa Bay Estuary

Program is pursuing a resource-based management strategy that seeks to limit anthropogenic

nitrogen loadings at levels needed to achieve its

seagrass-coverage goal of 15,400 ha (38,000 acre),

representing 95% of the seagrass coverage that was

mapped in the Bay in 1950 (TBNEP 1996, Johansson and Greening 2000). The Sarasota Bay

National Estuary Program has adopted a technology-based strategy that seeks to control nitrogen

loadings through the adoption of “best available

technology” for anthropogenic nitrogen sources in

the contributing watershed (SBNEP 1995).

Nitrogen-management strategies are effective

seagrass restoration tools in Tampa Bay and Sarasota Bay. Because phytoplankton are important

sources of light attenuation in both estuaries, by

controlling nitrogen inputs, managers can reduce

phytoplankton biomass, increase water clarity, and

increase the bay-bottom area that receives sufficient sunlight to support seagrasses (e.g., Johansson and Greening 2000). A similar situation

appears to exist in Lemon Bay, where phytoplank-
12

ton are also a major source of light attenuation

(Tomasko et al. 2001). In Charlotte Harbor, however, light attenuation is affected more by water

“color”—naturally elevated levels of dissolved

organic matter discharged from extensive wetlands

in the Peace and Myakka river systems—than by

phytoplankton abundance (McPherson and Miller

1994). Because of its large watershed, Charlotte

Harbor also experiences large seasonal and annual

variations in fresh-water inflow, producing large

fluctuations in salinity that can be stressful to seagrasses (Tomasko and Hall 1999). The nitrogenbased management strategies developed for Tampa

Bay and Sarasota Bay may thus have limited applicability for Charlotte Harbor (D. Tomasko pers.

comm.).

The time series of mapped seagrass coverage in

portions of the Charlotte Harbor system is shown

in Figure 2.4. Much of the reduction in coverage in

Greater Charlotte Harbor (which includes San Carlos Bay, Matlacha Pass, and Pine Island Sound, in

addition to Charlotte Harbor Proper) between ca.

1950 and 1982 occurred in the southern portion of

the system, particularly in Pine Island Sound and

San Carlos Bay. This reduction has been linked, circumstantially, to a series of large-scale anthropogenic activities that occurred in the area during

the 1960s, including dredging of the Intracoastal

Waterway, construction of the Sanibel causeway,

and installation of dam and lock structures in the

lower Caloosahatchee River (Harris et al. 1983). In

addition to direct destruction of seagrass habitats,

these projects have been postulated to have indirectly altered the water clarity, salinity, hydrodynamics and flushing characteristics of the area in

ways that made it less conducive to seagrass growth

and survival (Harris et al. 1983). Due to difficulties

encountered in interpreting and digitizing the aerial

photographs that were taken in the area during the

ca. 1950 period, however, the coverage values shown

for the Greater Charlotte Harbor system in that

period in Figure 2.4 should be viewed as rough

approximations (Harris et al. 1983).

No geographically and technically consistent

mapping of seagrass coverage throughout the

Greater Charlotte Harbor system has been conducted since 1982. The northern portion of the

system—which falls largely within the SWFWMD

and is referred to locally as Charlotte Harbor

Proper—has been mapped approximately biennially since 1988 by the SWFWMD (Kurz et al.

2000). From 1988 through 2002, seagrass coverage

in this area fluctuated between 7,200 ha (17,800

acres) and 7,800 ha (19,300 acres), with no apparent upward or downward trends (Figure 2.4).

Lemon Bay has been mapped over the same time

period and has shown relatively small fluctuations

around an average value of 1,058 ha (2,600 acres),

with no apparent trends (Figure 2.4).

Mapping data from St. Joseph Sound near

Clearwater are available only for the years 1999

and 2002, when an average of 5,840 ha (14,400

acres) were recorded (Figure 2.4).

BIG BEND

The Big Bend region extends from Anclote Key

northwestward to Ochlockonee Point in the Panhandle region (Figure 2.3) and includes the coastal

waters of Pasco, Hernando, Citrus, Levy, Dixie,

Taylor, Jefferson, and Wakulla counties. Zieman

and Zieman (1989) note that this portion of the

coast is unique in that it is an extensive area, with

no offshore barrier islands, where a number of

rivers, creeks, and marshes discharge directly into

the Gulf of Mexico. It is also one of the few examples of a “zero-energy” coastline, with average

breaker heights of 3–4 cm or less and little littoral

transport of sand (Murali 1982). Factors contributing to the low-energy characteristics of the

area include a wide and gently sloping shelf, divergence of approaching wave trains into a large

coastal concavity, the location of the coast in a

generally upwind direction, and the wave dampening effects of old submerged beaches and seagrass

meadows (Murali 1982).

The region is an environmentally diverse area

that can be divided into five subregions (Mattson

2000). The Springs Coast subregion, which extends

from Anclote Key northward to the mouth of the

Withlacoochee River, is dominated by flows from a

series of short, spring-fed river systems: the Weeki

Wachee, Chassahowitzka, Homosassa, and Crystal

rivers. Concentrations of nitrate nitrogen have

been increasing steadily in these rivers in recent

decades, due to increasing anthropogenic nitrogen

discharges in their highly karstic watersheds and

spring recharge areas (Katz et al. 1997). The limestone bedrock and sediments of this portion of the

coast are rich in carbonates, however, and tend to

GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY
CHAPTER 2 | DISTRIBUTION, STATUS, AND TRENDS 13

bind inorganic phosphorus from the water column. Because inorganic phosphorus is less available in the water column, primary production of

near-shore aquatic ecosystems are tilted from Nlimitation toward P-limitation (Hauxwell et al.

2001).

The four subregions north of the Springs Coast

subregion are Waccasassa Bay, Suwannee Sound

and adjacent coastal waters, Deadman Bay, and

Apalachee Bay (Mattson 2000). Discharges from

river systems in these subregions tend to be high in

color during periods of high flow, a factor that

apparently contributes to relatively low seagrass

coverage in the vicinity of the river mouths (Mattson 2000).

Although the inshore and offshore seagrass

beds of the Big Bend are among the largest in the

eastern Gulf of Mexico (Iverson and Bittaker

1986), the region has received relatively little management attention (Mattson 2000). Several mapping surveys have been conducted, but most have

covered only a limited portion of the region and

have produced highly variable coverage estimates

(Mattson 2000). The most extensive, region-wide

mapping efforts have been carried out by Iverson

and Bittaker (1986) and Sargent et al. (1995), producing coverage estimates of 300,000 ha (741,000

acres) and 334,842 ha (827,000 acres), respectively.

Neither of these estimates includes the sparse,

deep-water seagrass beds that are located offshore.

For the entire West Florida Shelf, Fonseca et al.

(2001) estimated the areal coverage of deep-water

H. decipiens beds at 0.4 million ha (1 million

acres), which would place them among the largest

seagrass communities in the world.

In a general sense, the Big Bend area has been

described as one of the least polluted coastal

regions of the continental United States (Livingston 1990). However, the Fenholloway River—a

tributary to Apalachee Bay—is an exception to this

generalization, receiving discharges from an industrial facility that are high in color and contain elevated levels of sulfate, BOD, suspended solids and

nutrients (Mattson 2000). These discharges have

apparently caused localized reductions in water

clarity and seagrass coverage in portions of

Apalachee Bay (Livingston 1993, Livingston et al.

1998).

PANHANDLE

The Panhandle region extends from Ochlockonee

Point westward to the Florida-Alabama border

(Figure 2.3) and includes the coastal waters of

Franklin, Gulf, Bay, Walton, Okaloosa, Santa Rosa,

and Escambia counties. It resembles the Gulf

Peninsula region in being a moderate-energy

coastline with extensive sand beaches and barrier

islands enclosing protected estuaries and lagoons

(e.g., Apalachicola Bay, St. Joseph Bay, St. Andrew

Bay, Choctawhatchee Bay, Santa Rosa Sound,

Escambia Bay, and Pensacola Bay) that contain the

region’s seagrass beds. Like the Big Bend region, it

is the subject of limited recent research and management activity.

Based on aerial photography taken in

1992–1993, Sargent et al. (1995) estimated that

19,509 ha (48,170 acres) of seagrasses were present

in the coastal waters of the Panhandle region. No

region-wide coverage estimates have apparently

been produced since that time. Lores et al. (2000)

assessed recent coverage trends in seagrasses and

other SAV in the Escambia-Pensacola Bay portion

of the region, and provided the following summary:

SAV in the Escambia-Pensacola Bay System

underwent a substantial decline in the late

1940s through the early 1970s;

Although scientific documentation of SAV distribution since that time is lacking, some observations suggest SAV growth in the oligohaline

regions of estuaries in northwestern Florida has

shown recent improvements (e.g., in Mobile

Bay [Alabama], Escambia Bay, and Perdido

Bay);

Evidence suggests these same areas are also losing SAV in the euryhaline regions;

Improvements in water quality of the upper bay

regions is thought to be leading to recovery of

low-salinity seagrasses; and

Continuing increases in coastal development in

the lower bay region, with resulting increased

nutrient input and sediment loading/resuspension, may be having an adverse impact on the

health and productivity of high-salinity seagrasses such as T. testudinum.
UTECOLOGY and

POPULATION

GENETICS

Flowering is known in all six species of seagrasses on the Gulf

coast of Florida, but Thalassia testudinum produces fewer fruits

at higher latitudes (north of Tampa Bay).

Genetic information for Florida seagrasses exists only for T. testudinum, with lower genetic diversity observed in its northern

populations. This may be a result of the lower level of seed production and, therefore, more dependence on vegetative expansion for reproduction.

Genetic data are needed for more populations and for other seagrass species.

The clonal nature of seagrasses allows transport of soluble carbohydrates, proteins, and nitrogen (as glutamine) to stressed

short-shoot and long-shoot rhizome meristems.

Depth distribution of T. testudinum is influenced by a variety of

factors, including water transparency, epiphyte load of the leaf

blades, and water movement.

Salinity variations affect the local distributions of seagrasses. T.

testudinum does not survive if held in culture for over 6 weeks

in 6 ppt seawater. In the field, other factors will raise that

threshold. More data are needed regarding osmoregulation in

euryhaline species such as Halodule wrightii.

Higher sulfide levels in the sediment are toxic to T. testudinum,

occur in areas of eutrophication in a number of Gulf-coast estuaries, and may play a role in the patchy nature of seagrass communities.

Moderate grazing by invertebrates (especially sea urchins), fish,

sea turtles, and manatees may enhance seagrass-bed development and species diversity.

Stable isotope ratios can be used to trace the flow of organic

components in seagrass beds, although macroalgae may have

similar isotopic signatures.

Carbon fixation by most seagrasses occurs via the C3 pathway.

However, it is debatable whether a modified form of the C4 pathway exists in some species.

Entire-plant carbon budgets of T. testudinum and H. wrightii

show that below-ground components account for over 50% of a

plant’s respiration.

Widgeongrass, Ruppia maritima

Ron Phillips photo
CHAPTER 3 | AUTECOLOGY AND POPULATION GENETICS 15

SEAGRASS MORPHOLOGY 

and ANATOMY

Detailed descriptions of seagrass morphology and

anatomy can be found in den Hartog (1970),

Ancibar (1979), Tomlinson (1980, 1982), and Kuo

and McComb (1989). Also see the Appendix, Taxonomy of Florida Seagrasses, in this document.

The basic construction of most seagrasses (Arber

1920) is an indeterminate horizontal stem (plagiotropic rhizome or long shoot) that periodically

produces determinate erect stems (orthotropic rhizomes or short shoots) having leaves and flowers.

Adventitious roots develop from both types of rhizome. Rhizomes are usually cylindrical and below

the sediment surface in species with robust morphologies. In contrast, rhizomes of species with

more delicate morphologies (e.g., H. decipiens and

H. johnsonii) often occur above the sediment surface. Rhizome growth is either sympodial (e.g., H.

wrightii and R. maritima) or monopodial (e.g., T.

testudinum and S. filiforme). Leaves of the Gulfcoast Florida species differ greatly in morphology,

being long, wide and thick in T. testudinum; long,

narrow, and thin in H. wrightii and R. maritima;

long, rigid, and cylindrical in S. filiforme; and

short, thin, and membranaceous in species of

Halophila.

The rhizomes of Florida seagrasses are herbaceous, with little fiber tissue. In seagrasses similar

to T. testudinum, the vascular stele and fiber bundles in the cortexes of the blades, short shoots, and

rhizomes are poorly lignified (Dawes 1986). Epidermal cells of seagrass blades lack stomata and

associated guard cells, contain most of the blades’

chloroplasts, and have a thick outer cell wall covered by a thin, porous cuticle. As determined by

uptake studies (Larkum et al. 1989), the cuticle

apparently does not prevent absorption (e.g., of

CO2, cadmium, or manganese) by the blade. As

with those of T. testudinum, the epidermal cells of

R. maritima may be involved in osmoregulation.

Epidermal cells of R. maritima blades that are

grown in higher salinities (e.g., 32 ppt) develop

masses of gelatinous polysaccharides and form

cell-wall ingrowths that may be involved in ion

binding and exchange between the seawater and

cytoplasm (Kruzcynski 1994).

SEXUAL REPRODUCTION

Although sexual reproduction is known in all the

seagrass species of Florida, with the possible exception of H. johnsonii (see Appendix), data on flowering and seed production and survival are limited

(Ferguson et al. 1993). Leaf and inflorescence scars

(Figure 3.1) have been used to determine shortshoot age, sex, and frequency of flowering in T. testudinum (Cox and Tomlinson 1988; van

Tussenbroek 1994; Witz and Dawes 1995), with

short shoots that produce female flowers often

having narrower leaves than short shoots that produce male flowers (Durako and Moffler 1985).

Inflorescence scars on short shoots of T. testudinum (Fig 3.1) suggest that abundant flowering

occurs in northern areas along Florida’s Gulf coast

(e.g., St. Joseph Bay and Apalachee Bay) and

throughout the Florida Keys (C. Dawes pers. obs.),

yet flowering and seed production have not been

documented for most T. testudinum beds along the

northern Gulf coast. In contrast, in May or June of

each year, thousands of T. testudinum seeds may

occur in the beach wrack at Mathieson Hammock

near Miami and in the wrack lines in the Florida

Keys (C. Dawes pers. obs.; Lewis and Phillips 1980),

and countless more are eaten by birds (Fishman

and Orth 1996).

Seagrass flowering and fruiting occur annually

between February and August along the Gulf coast

of Florida (Phillips 1960c; Durako and Moffler

1987; Witz and Dawes 1995) and on the Caribbean

coast of Mexico (van Tussenbroek 1994). Photoperiod does not appear to influence the onset of

flowering in T. testudinum (Moffler and Durako

1987), S. filiforme, or H. wrightii (McMillan 1982),

as these three species flowered under continuous

light. Instead, water temperature influences flower

development (Moffler and Durako 1987), and all

species in Florida flower within a temperature

range of 20°–26°C (McMillan 1982). Lower water

temperatures (10° to 18°C) in the winter in the

Figure 3.1 Inflorescence, leaf, and root scars on short shoots

of Thalassia testudinum (modified from Witz 1994).
16 GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

Tampa Bay area may cause T. testudinum to flower

later in the spring than it does in the Florida Keys,

causing immature fruits to dehisce early in

response to rapidly rising water temperatures in

May and June (Witz and Dawes 1995).

Seed germination has not been studied in most

Florida Gulf-coast seagrass species. Ruppia maritima seed germination is influenced by salinity

and temperature but not by photoperiod. Seeds

collected from the sediment at the mouth of the

Weeki Wachee River germinated at all temperatures (17°, 23°, and 39°C) and in 0 and 15 ppt but

not 30 ppt salinities. In contrast, seeds from North

Carolina germinated in all temperature and salinity combinations (Koch and Dawes 1991a). Studies

on fish predation of R. maritima seeds demonstrated that they can pass through the gut and thus

be dispersed (Agami and Waisel 1988). The importance of seed reserves as an early source of carbohydrate energy for seedlings of T. testudinum was

proposed by Durako and Sackett (1993). Seedlings

had a lower carbon isotopic fractionation (δ13C plant

– δ13C source = 15.4 ppt) than did one-year-old plants

(21.0 ppt), possibly indicating a shift from using

carbon reserves in the seed to taking up CO2 from

the water column. However, the one-year-old

plants were cultured under artificial conditions

and were preconditioned, which may have altered

their fractionation level.

ASEXUAL REPRODUCTION 

and CLONAL BIOLOGY

Vegetative growth by the long-shoot rhizome is

thought to be the principal means of expansion for

Gulf-coast seagrasses, in light of the curtailment of

sexual reproduction explained above. Seagrass

short shoots can be regarded as ramets, and the

clonal group of ramets connected by long-shoot

rhizomes derived from a single propagule is considered to be the genet (see Harper 1990). Rhizome extension is the basis for vegetative

expansion in all seagrasses (Tomlinson 1974;

Dawes 1998a; Andorfer and Dawes 2002; Dawes

and Andorfer 2002).

Two possible advantages in being clonal are the

ability to carry out vegetative expansion or foraging (Cain 1994) and the ability to share resources

between ramets of the same genet growing in areas

of stress or low nutrients (Tomasko and Dawes

1989; Wijesinghe and Handel 1994; Andorfer

2000). Short shoots of T. testudinum near Mullet

Key in Tampa Bay were isolated from others by

severing the long-shoot rhizome on each side of

the short-shoots that were then shaded. These

short shoots had significantly lower blade growth

than did shaded short shoots that were connected

to other short shoots (Tomasko and Dawes 1989).

Rhizomes adjacent to shaded short shoots were

depleted of soluble carbohydrates and proteins,

while the rhizome portion beyond the severed

short shoots was not, suggesting that sugars and

amino acids stored in the adjacent rhizomes could

be mobilized.

Laboratory culture studies using the stable

nitrogen isotope 15NO3 demonstrated that nitrogen

in the form of glutamine was translocated from

the fourth-oldest ramet of T. testudinum to the

growing long-shoot rhizome meristem (Andorfer

2000). Further, leaf regrowth, after the four

youngest short shoots were clipped, was supported

by carbon fixed in the fifth short shoot and transported to it via the rhizome. This high degree of

short-shoot integration, over a distance of five

ramets, demonstrates the importance of the physiological integration that allows T. testudinum to

expand through regions of low light or nutrients

and to tolerate periods of intense grazing (Andorfer 2000). These studies support earlier findings

for T. testudinum subjected to weekly clipping of

blades (Dawes and Lawrence 1979) and for S. filiforme in the Indian River Lagoon subjected to

shading of connected and severed short shoots

(Rey and Stephens 1996). In the latter study, soluble-carbohydrate levels declined from 26.5% to

18.4% and then stabilized in rhizomes connected

to shaded short shoots; this suggests movement of

soluble carbohydrate from adjacent non-shaded

short shoots, thus showing physiological integration of the genet.

Production of new long-shoot rhizome

branches by T. testudinum occurs principally at the

long-shoot meristem and rarely occurs through

branching of the older long-shoot rhizome. Dawes

and Andorfer (2002) found that older short shoots

were more likely to produce a new long-shoot rhizome than were younger ones and that initiation

of rhizomes was suppressed unless the existing

long-shoot meristem was removed. The presence

of “inactive” or dormant short shoots and rhizomes on T. testudinum suggests that a meristem

bank may exist in T. testudinum beds, as was found
CHAPTER 3 | AUTECOLOGY AND POPULATION GENETICS 17

on the Caribbean coast of Mexico (van Tussenbroek 1996a). Just what role these inactive short

shoots play in vegetative expansion in a seagrass

bed is yet unknown. Perhaps, under certain conditions, the dormant short shoots will again resume

growth and produce new blades or rhizome meristems. Further, seagrasses differ in rhizome-branch

frequency. For example, H. wrightii rhizomes

branch frequently and grow rapidly, whereas those

of T. testudinum do not. Species that exhibit more

aggressive growth are often chosen for restoration

in Florida. Thus H. wrightii is more often selected

for restoration projects than T. testudinum,

although the latter species forms the dominant

beds in Florida and contains the greatest diversity

of species (Zieman and Zieman 1989). The types

of growth (foraging) strategies of seagrasses are

the same as those of terrestrial clonal plants. Rhizomes of H. wrightii show a “guerrilla”-type

growth by branching frequently and spreading

rapidly. In contrast, T. testudinum rhizomes show a

“phalanx”-type growth, growing linearly and with

few lateral axes (Dawes 1998a).

EVOLUTION and GENETICS

The discovery of fossil seagrasses in the Avon Park

formation in Florida (Lumbert et al. 1984) indicates that species of Thalassodendron and Cymodocea occurred with T. testudinum in the late

Middle Eocene (ca. 40 million ybp). These fossils

were preserved as carbonized imprints within the

bedding planes of a micritic dolomitic limestone

in a rock outcrop of the Ocala Arch in central

Florida, which contains the oldest exposed rocks in

the state. Generic determinations were certain,

except for one of the fossils, and species identifications were questionable for two of the fossils. The

fossils were identified as Thalassodendron auriculaleporis den Hartog, Cymodocea floridana den Hartog, T. testudinum, Cymodocea sp., Halodule sp.,

and an unknown Zosteroid.

Species of Thalassodendron and Cymodocea

presently occur only in the Old World tropics

(Indo-Pacific region), indicating that a more

diverse seagrass flora existed in the Caribbean Sea

in the past. One hypothesis suggests that they disappeared from the Caribbean region when the

Central American isthmus elevated and separated

the Atlantic and Pacific oceans, changing dominant circulation patterns, which caused Caribbean

water temperatures to fall. Evidentially, an extensive and diverse seagrass flora existed in central

Florida in the past, judging by the abundance and

extent of the fossilized seagrasses and by the diversity of less well-preserved animals from seagrass

beds (two families of foraminifera, some bivalves,

a bryozoan, a crab carapace, and possibly an ostracod). The fossil findings support the vicariance

model of Heck and McCoy (1979), who proposed

that the present-day distribution of seagrass

species is a product of continental drift, speciation,

and extinction. Hypotheses regarding the origin of

seagrasses include evolution from salt-tolerant terrestrial shrubs (den Hartog 1970) or from freshwater hydrophilous ancestors (Cox and Humphries

1993).

Comparison of isozymes of various enzymes

and molecular-DNA techniques have been used to

examine genetic diversity within and between present-day seagrass populations and species in Florida

and the Caribbean region. Early isozyme studies

examined genetic differences between Caribbean

Sea and Gulf of Mexico populations of T. testudinum, S. filiforme, and H. wrightii but found little intraspecific variation (McMillan 1980). The

studies supported McMillan’s (1978) earlier culture

experiments, which revealed that leaf width was

influenced by the immediate environment and not

by genetic differences between plants.

Being the dominant seagrass in Florida and the

Caribbean (Zieman and Zieman 1989), T. testudinum has been the principal species to have its

genetic diversity measured using isozymes

(Schlueter and Guttman 1998) and molecularDNA approaches (Kirsten et al. 1998; Davis et al.

1999; Waycott and Barnes 2001). Allozyme loci of

14 enzymes in 18 populations in the lower Florida

Keys indicated that asexual reproduction is probably the basis for the low genetic diversity there and

for a trend towards genetic uniformity (Schlueter

and Guttman 1998). Allozyme and Amplified

Fragment-Length Polymorphism (AFLP) analyses

were used to compare genetic diversities of T. testudinum from two sites in Panama and from

another in Bermuda (Waycott and Barnes 2001).

The authors found high levels of genetic uniformity, and they suggested that it is due to vegetative-fragment dispersal, even over a distance of

2,700 km.

In contrast to isozyme and allozyme studies,

two other molecular DNA studies found a higher
18 GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

level of genetic diversity within populations of T.

testudinum, rather than between geographically

distinct ones. Randomly Amplified Polymorphic

DNA (RAPD) analysis showed high genetic diversity within two proximal Florida Keys populations

(e.g., Fiesta Key and Craig Key) and within another

northern population off the Steinhatchee River (as

Apalachicola Bay in Kirsten et al. 1998). Further,

almost all samples from an outlier Jamaican population and the two Florida Keys sites were distinct

genetic individuals within and between the three

populations. This is in contrast to the northern

population that had the fewest RAPD phenotypes

(Figure 3.2). The lower within-community genetic

diversity of northern populations of T. testudinum

may reflect the limited introduction of new genets

(e.g., drift plants) or a low level of flowering and

seed production, perhaps because of less-thanoptimum water temperatures or water transparencies. A second study on T. testudinum, using

DNA-fingerprinting techniques, concentrated on

clonal variation in populations in Florida Bay and

found discrete beds that were not genetically uniform (Davis et al. 1999), again emphasizing the

role of sexual reproduction in maintaining population genetic variation. Thus, beds of T. testudinum in more tropical regions contain a greater

number of distinct genets than do beds at higher

latitudes. The difference in genetic diversity

between tropical and subtropical beds may reflect

a lower rate of seed production in more northern

sites, as shown for seagrasses in Tampa Bay (Witz

1994; Witz and Dawes 1995). This may explain the

lower genetic diversity for the northern population

of T. testudinum found off the Steinhatchee River

(Kirsten et al. 1998). The concept that sexual

reproduction is less likely to be successful when an

organism encounters less-than-optimal temperatures was described by Gessner (1970). However,

little is known about flowering and the production

of viable seeds on most of the Gulf coast of

Florida, with the exception of some data for

Tampa Bay and the Florida Keys.

Another study (Angel 2002) using RAPD analysis compared three populations of H. wrightii from

Texas (Christmas Bay, Corpus Christi) and from

Florida (Florida Bay). All individuals appeared to

have unique genotypes, with plants from Corpus

Cristi more closely clustered to those of Florida

Bay, which Angel suggested was due to similar

habitats that acted as a selective force. The molecular phylogeny of 11 species of Halophila found that

H. decipiens and H. engelmannii were distinct

species and that there was a 100% overlap between

populations of the former species between populations from Australia, the Caribbean, and Florida

(Waycott et al. 2002). Further, the internal transcribed spacer (ITS) region of the ribosomal DNA

indicated that H. johnsonii from the Indian River

in Florida could not be distinguished from H.

ovalis and needs further study.

PHYSIOLOGICAL ECOLOGY

At least five abiotic properties influence the morphological and anatomical adaptations of seagrasses (Dawes 1998a), including the (1) osmotic

effects of salt water; (2) availability of dissolved

CO2 and nutrients; (3) intensity and quality of illumination; (4) density (greater than air) and

mechanical drag of an aqueous medium; and (5)

effects of an aquatic medium on the dispersal of

pollen and seeds.

The general physical requirements of five

Florida seagrass species were discussed in Phillips

(1960a). Seagrasses show physiological adaptations

to a variety of abiotic and biotic conditions,

including salinity, temperature, water motion,

anoxia, nutrient limitation, epiphytes, irradiance,

infection, and herbivory (Kuo and den Hartog

1989). The most-studied factor has been the relationship between light and the depth distribution

of seagrasses. Nevertheless, other physical, geological, and chemical characters must be examined to

explain the patchy distribution of seagrasses on the

Gulf coast of Florida (Koch 2001). Koch identified

Figure 3.2 Randomly amplified polymorphic DNA (RAPD)

phenotypes within four populations of Thalassia testudinum.

(Modified from Kirsten et al. 1998)
CHAPTER 3 | AUTECOLOGY AND POPULATION GENETICS 19

tidal amplitude (Figure 3.3) and wave energy (Figure 3.4) as two factors that influence seagrass

depth distribution. Tolerances to common abiotic

factors may differ between populations of a species

from different environments, as shown for R. maritima grown from seed (Koch and Dawes 1991b).

Different strains grown from seeds collected in the

Weeki Wachee River estuary and from North Carolina had growth rates that were similar under different combinations of temperature and

photoperiod but differed with regard to salinity.

This was also reported in earlier studies for other

species (McMillan 1979, 1984). In those studies,

different strains of T. testudinum, S. filiforme, and

H. wrightii showed different tolerances to temperature, and these tolerances were often related to

distinct environmental origins. These findings suggest that transplanting from one area to a different

site at greater distances (e.g., seeds from Biscayne

Bay to Tampa Bay) may not be as successful as

when using local plants.

A large global-level literature base shows the

critical effects of water quality on the presence,

growth, and vitality of seagrasses. Adequate light

transmission is vital to the presence of seagrasses

and affects the density and biomass of seagrasses at

all depths (Dixon and Leverone 1991; Duarte

1991; Johansson 1991; Kenworthy and Haunert

1991; Lapointe and Clark 1992; Tomasko 1992,

2002; Lapointe et al. 1994; Fletcher and Fletcher

1995; Kenworthy and Fonseca 1996; Kurz et al.

1999, 2000; Dixon 2000; Neely 2000). Depth determinations (±10 cm) can be obtained with a mapping-grade differential Global Positioning System,

carrier-phase equipment and surveyed benchmarks (Johansson 2002b). In Tampa Bay, the deep

edges for H. wrightii ranged from –0.30 to –0.76 m

below mean low water (MLW), from –1.19 to

–1.96 m for S. filiforme, and from –0.57 to –1.73 m

for T. testudinum. The maximum depth of T. testudinum at Port Manatee in Tampa Bay was –1.6

m, in agreement with the above determinations

carried out before the 1997–1998 El Niño event

(Dixon and Leverone 1995). The shallowest deep

edges for all three species occurred at the northernmost sites, and the deepest deep edges were

near the mouth of Tampa Bay, where more oceanic

(clear water) conditions exist.

Neely (2000) demonstrated an inverse relationship between light reduction and short-shoot numbers for H. wrightii in Tampa Bay. Leaves on plants

exposed to a 43% light reduction were longer than

those on shoots in control plots or in other lightreduction treatments. Her experiments showed

that loss of biomass was a result of winter die-back

and also occurred with a >60% reduction in light.

The data showed that epiphytes and surface scattering of light must be accounted for when calculating

carbon budgets. Durako and Hall (1992) found

that for T. testudinum the carbon supply became

nonlimiting when light was reduced to levels that

limited photosynthetic rates. Tomasko (1992)

demonstrated that H. wrightii, when growing

under a canopy of T. testudinum leaves, had longer

rhizome internodes, less frequent branching, and

lower above-ground to below-ground biomass

ratios than when growing alone. Studies by SandJensen (1977, 1990), Cambridge et al. (1986), and

Tomasko and Lapointe (1991) showed that heavy

epiphyte loads on seagrass leaves are indicative of

nutrient over-loading in the water column and,

therefore, eutrophication, i.e., poor water quality

(see Chapter 5). In addition to the adverse effects of

nutrient-enriched water on photosynthetic functioning of seagrass leaves, increased levels of nutrients can result in high epiphyte loads that shade

and weigh down the leaves, further lowering them

in the water column.

Figure 3.3 Influence of tidal range on seagrass depth distribution. Minimum depth is deeper with greater tidal range (Koch

2001). Reprinted with permission from Estuarine Research Federation.

Figure 3.4 Influence of wave energy on seagrass depth distribution. Greater wave energy and deeper wave-mixing depths

result in deeper minimum depths (Koch 2001). Reprinted with

permission from Estuarine Research Federation.
20 GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

The relationship between light, both quantity

and quality, and the depth at which seagrasses

grow was the subject of two workshops in Florida

in the 1990s, one sponsored by NOAA (Kenworthy

and Haunert 1991) and the other by the Indian

River Lagoon National Estuary Program (Morris

and Tomasko 1993). Even though they can grow

under high irradiences, seagrasses are regarded as

shade plants (Hillman et al. 1989), with compensation values (Ic) of 25 to 50 μmole-photons m–2 s–1

and saturation irradiances (Ik) of <200 μmolephotons m–2 s–1. Studies comparing photosynthetic

responses of seagrasses to different irradiance levels (i.e., Ic, Ik, and Pmax levels) often show a relationship to temperature and salinity, as seen for two

species of Halophila (Dawes et al. 1989). For example, the estuarine, intertidal populations of H.

johnsonii in the Indian River Lagoon tolerate a

wide range of temperatures (10°, 20°, and 30°C)

and salinities (15, 25, and 35 ppt) and did not

become photoinhibited by high photosynthetically

active radiation (PAR; ca. 500 μmole-photons m–2

s–1). In contrast, H. decipiens from deeper, nonestuarine Gulf-coast populations were intolerant of a

10°C temperature, 15 and 25 ppt salinity, and PAR

above 300 μmole-photons m–2 s–1 (Dawes et al.

1989).

Depth distributions of Florida’s seagrasses are

governed by PAR levels in Tampa Bay (Miller and

McPherson 1995; Dixon and Leverone 1995, 1997;

Dixon 2002) and the Indian River Lagoon (Gallegos and Kenworthy 1996). Recent work has

demonstrated more specific light requirements for

the dominant seagrass species (Table 3.1). PAR levels measured at the deep edges of T. testudinum

beds in Tampa Bay were significantly higher than

the required average light level of 11% (of surface

value) proposed for seagrasses by Duarte (1991).

Also, T. testudinum in Florida Bay (Fourqurean

and Zieman 1991) and Caribbean seagrasses

(Onuf 1994) required at least 22% of surface PAR

levels, when whole-plant requirements and average

daily irradiances were considered. Continuous in

situ measurements of underwater PAR levels for

one year at the shallow and deep edges of four T.

testudinum beds in Tampa Bay also indicated the

Table 3.1 Light requirements of Florida seagrasses.

Minimum Light Requirement

Species and Sites (percentage of total surface irradiance) Reference

Thalassia testudinum

Big Bend area, FL 15.3% Iverson and Bittaker (1986), as

reported by Duarte (1991) and 

by Dennison et al. (1993)

Northwest Cuba 23.5% Duarte (1991)

Puerto Rico 24.4% Duarte (1991)

Charlotte Harbor, FL 20% Tomasko and Hall (1999)

Tampa Bay, FL 22.5% Dixon and Leverone (1995)

Tampa Bay, FL 20.1% (annual mean) Dixon (2000)

Corpus Christi Bay, TX 14% (4% survival after 12 months) Czerny and Dunton (1995)

Syringodium filiforme

Florida Bay, FL 18.3% Iverson and Bittaker (1986) as

reported by Duarte (1991)

Indian River Lagoon, FL 17.2% Duarte (1991) pers. comm. from 

J. Kenworthy (1990)

Indian River Lagoon, FL 24%–37% Kenworthy and Fonseca (1996)

(southern region)

Halodule wrightii

Indian River Lagoon, FL 17.2% Dennison et al. (1993) pers. comm.

from J. Kenworthy (1990)

Florida Bay, FL 10%–20% Fourqurean and Zieman (1991)

Laguna Madre, TX 7%–23% Tomasko and Dunton (1991);

Dunton (1994); Czerny and 

Dunton (1995)
CHAPTER 3 | AUTECOLOGY AND POPULATION GENETICS 21

maximum depth limit in the bay occurred at about

20.5% of subsurface PAR (Dixon and Leverone

1995, 1997). Another study on irradiance in the

more oceanic waters at the mouth of Tampa Bay

showed that, using 20% light penetration as the

threshold, sufficient light was available for T. testudinum beds to extend to two meters of depth

(Miller and McPherson 1995). That depth is similar to measured seagrass-bed depth limits, which

ranged from 1.98 to 2.37 m in estuaries (Dixon

and Leverone 1995). Evaluating short-shoot

density, Neely (2000) found that shallow, monospecific beds of H. wrightii showed a direct relationship between density and irradiance level at

Mullet Key in Tampa Bay (Figure 3.5). The same

level of density reduction with shading occurred in

the cool and warm seasons, although summer

short-shoot densities were almost twice those of

the winter densities, indicating the critical effect

that underwater irradiance has on seagrasses. Similar declines in leaf growth, short-shoot density,

biomass and primary production have been caused

by shading T. testudinum (Lee and Dunton 1997;

Kraemer and Hanisak 2000) and H. wrightii

(Czerny and Dunton 1995).

Light attenuation that limits seagrass depth distribution can result from a variety of sources (Kenworthy and Haunert 1991; Morris and Tomasko

1993). One factor governing light effects on seagrass depth distribution is the level of epiphytic

biomass. The average annual attenuation by epiphytes on water-column PAR levels available to T.

testudinum blades was from 32%–47% in Tampa

Bay, 40%–56% in Sarasota Bay, and 21%–44% in

Charlotte Harbor (Dixon and Leverone 1997).

Variation in epiphyte loads on T. testudinum

blades resulted in a 7%–67% seasonal range in

PAR attenuation in Tampa Bay, 13%–99% in Sarasota Bay, and 6%-89% in Charlotte Harbor (Dixon

and Leverone 1997). Thus, in determining seagrass

light requirements, seasonal levels of epiphyte

loads should be considered in addition to the

actual PAR levels (Dixon 2002).

Other factors shown to have critical effects on

light available to seagrass, in a light model developed for the Indian River Lagoon, were total suspended solids (e.g., tripton or non-algal particulate

matter) and chlorophyll a concentration, which

were the dominant causes of light attenuation

affecting seagrass depth distribution (Christian

and Sheng 2002). Attenuation of light by turbidity

is a critical abiotic factor governing availability of

light to seagrasses, as shown by the effects of

dredging in the Laguna Madre of Texas that

resulted in the decline of H. wrightii (Onuf 1994).

Depth distribution, biomass, and photosynthetic

responses were compared for shallow and deep

populations of T. testudinum on the Gulf coast of

Florida (Dawes and Tomasko 1988; Dawes 1998b).

In general, biomass decreased while chlorophyll a

increased with greater depth (Table 3.2), and the

ratio of above- to below-ground biomass

remained similar.

Also, light quality (i.e., wave-length distribution) and PAR affect seagrass growth. Exposure of

H. wrightii, H. engelmannii, and S. filiforme to

UV-A and PAR irradiation resulted in photosynthesis inhibition in the latter species (Trocine et al.

1982). A low ratio of red:far-red wave lengths

reduced branching and node formation in R. maritima, a response that may allow plants shaded by

larger plants to escape by emphasizing horizontal

rhizome elongation (Rose and Durako 1994). A

similar response by H. wrightii, when shaded by T.

testudinum, was noted in the Florida Keys

(Tomasko 1992).

Photosynthetic responses have been used to

evaluate seagrass tolerances to salinity and temperature variations (Dawes et al. 1987, 1989). Estuarine populations of H. engelmannii and H. johnsonii

from the Gulf coast and Indian River Lagoon of

Florida, respectively, had broader tolerances to a

range of temperatures and salinities when compared to Gulf-coast oceanic populations of H.

engelmannii and H. decipiens. Differences in photosynthetic factors (e.g., Pmax, α, Ik) were evident

with freshly collected plants and after culturing

Figure 3.5 Average density (short shoots m–2) of Halodule

wrightii at control and three increasing levels of shading

(30, 50, 80%) in Tampa Bay seagrass beds (± 1 S.D.). (Modified from Neely 2000)
22 GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

plants under common garden conditions for a

number of weeks using various combinations of

temperature and salinity. Salinity tolerance of T.

testudinum from San Carlos Bay in the Caloosahatchee River Estuary in southwest Florida was

assessed using long-shoot rhizomes bearing two to

four short shoots (Doering and Chamberlain

2000). After a four-week acclimatization period in

ambient seawater (35 ppt) in the laboratory, salinity was changed daily in 1.5 ppt increments.

Regardless of other variables, blade length

decreased (Figure 3.6) and blade production

ended after 43 days in the lowest (6 ppt) salinity.

Blade length was positively correlated with salinity

after 43 days of exposure. Also, the number of

blades per short shoot and blade biomass were

similar in 12, 18, 25, and 35 ppt salinities after 43

days of exposure (data not shown).

Salinity affects the local distribution of seagrass

species on the Gulf coast of Florida (Phillips

1960b, c), with euryhaline species such as H.

wrightii able to grow in a broad range of salinities

(e.g., 5 to 55 ppt; Dunton 1996). However, the

effect on seagrass distribution due to salinity alone

is difficult to separate from the effects of other factors, as shown in the general reduction of seagrass

beds throughout Tampa Bay after the 1997–1998

El Niño event. The 1998 rainfall increase resulted

in higher levels of dissolved inorganic nitrogen,

chlorophyll a (phytoplankton), and color content

(tannins) from extensive runoff into the bay

(Johansson 2002a), which in turn reduced PAR

penetration in the water column. In addition, the

increase in rainfall directly lowered salinities, with

Middle Tampa Bay salinities declining from a winter high of 30 ppt to 10 ppt by March and then to

less than 5 ppt over a two-week period in April,

1998 (Dawes, unpublished data). Therefore, comFigure 3.6 Blade length and mean blade elongation rates for

Thalassia testudinum cultured for 43 days in salinities ranging from 6 to 35 ppt. Letters indicate statistical differences

between treatments and means, different letters are statistically different (P < 0.05). (Modified from Doering and Chamberlain 2000)

Table 3.2 Photosynthetic characters (Pmax, alpha, Ic, total chlorophyll), primary production, and leaf growth (± S.D.)

of Thalassia testudinum in deep- and shallow-water populations at two Florida Gulf-coast sites.

Leaf Relative Primary

Growth Rate Production Total Chl

Site (mg dw · g dw–1 · d–1) (g dw · m–2 · d–1) (mg · g dw–1)

Egmont Key

Shallow (–0.4 m) 28.7 (5.4) 11.71 4.86 (0.99)

Deep (–2.5 m) 27.4 (6.3) 8.17 3.92 (1.28)

Anclote Key

Shallow (–1.0 m) 31.1 (9.3) 1.39 3.00 (0.84)

Deep (–2.5 m) 20.5 (0.3)* 1.00 2.39 (0.70)

Pmax Ic Alpha

Site (mg O2 · g dw–1 · h–1) (μE · m–2 · s–1) (Pmax ⁄ Ic)

Egmont Key

Shallow 4.54 (0.70) 39 0.063

Deep 6.31 (1.21) 45 0.069

Anclote Key

Shallow 6.08 (0.92) 32 0.103

Deep 2.60 (0.33)* 24 0.097

Modified from Dawes and Tomasko (1988). * = probability <0.01
CHAPTER 3 | AUTECOLOGY AND POPULATION GENETICS 23

bined with higher nutrient loadings and turbiditycaused light attenuation, lower salinities resulted

in a die-back of T. testudinum in Cockroach Bay by

June 1998 and a reduction in seagrass acreage

throughout west-central Florida coastal areas (see

Chapter 2).

Seagrass osmoregulatory responses to changing

salinities involve structural and physiological

changes. The epidermal cells of T. testudinum and

R. maritima show cell-wall and organelle modifications when grown or collected in salinities above

30 ppt (Jagels and Barnabas 1989). In R. maritima,

clones of a single genetic strain cultured in 32 ppt

had convoluted cell membranes, smaller vacuoles,

and numerous mitochondria associated with the

cell membrane (Table 3.3). Also, plasmodesmata

were not evident between the epidermal and cortical cells, and a significant increase occurred in

acidic mucopolysaccharides associated with the

inner cell wall (Kruzcynski 1994). Clones from the

same genetic strain of R. maritima did not show

any of the above features when cultured at a lower

salinity (ca. 10 ppt). The increase in cell membrane

surface area and associated mitochondria are characteristic of cells active in osmoregulation, and

acidic carbohydrates may be involved in ion transport and binding (Dawes 1998a). In addition,

osmotic roles of the amino acids proline, alanine,

and glutamate have been identified in seagrasses

(Pulich 1986; Adams and Bate 1994). Pulich

(1986) reported that proline functions as an

organic osmoticum in R. maritima, T. testudinum,

and H. wrightii, and alanine serves that role in

H. engelmannii.

Data on temperature effects on seagrasses are

limited to studies examining flowering and seed

germination (see earlier discussion) and from a

study of leaf-blade production near an electricitygenerating plant (Barber and Behrens 1985). In the

latter study, the productivity of S. filiforme

increased in cooler months and decreased in

warmer months in areas where water temperatures

were enhanced by thermal pollution, whereas T.

testudinum showed little or no response.

Water movement influences the spatial distribution of seagrass beds, in addition to the effects on

depth distribution from tidal and wave activity (see

also Chapter 5). Water velocities, degree of exposure to waves, and relative water depths strongly

affect the distribution of H. wrightii and Z. marina

beds near Beaufort, North Carolina (Fonseca and

Bell 1998). Water movement, expressed as bladefriction velocities (u* = cm s–1), reduced boundarylayer effects around the leaf, and photosynthetic

rate increased in T. testudinum, up to a current

velocity of 0.25 cm s–1 (Koch 2001). In addition,

increased water movement causes a decrease in

concentration of sediment sulfide, a phytotoxin.

Further, seedlings of T. testudinum collected in the

Florida Keys showed optimum growth under intermediate flow rates (u* = 0.3 cm s–1), whereas stagnant water (u* = 0.0) contributed to lower biomass

possibly because of increased sulfide levels, and

high flow rates (u* = 1.0 cm s–1) reduced nutrient

availability in the sediment pore water (Koch

1999a). Back-and-forth wave motion increases

nutrient exchange between blades and the water

column, in contrast to unidirectional movement or

currents (skimming) that flow over T. testudinum

meadows in the Florida Keys (Koch and Gust

1999). When the flow rate (u*) equaled 25 cm s–1,

the water current began to erode the sand; this

caused a 50% reduction in seagrass coverage.

Repeated intertidal exposure, and thus the

development of tolerance to desiccation, has been

suggested as a causal factor in the ability of

Table 3.3 Structural features of Ruppia maritima epidermal cells from a single monoculture grown in 10 ppt and 32

ppt salinity.

Component of Mean (± SD) Mean (± SD) Significance

Cell Volume at 10 ppt at 32 ppt Level

Vacuolar volume 28.4% (± 19.8) 9.5% (± 9.3) P = 0.03*

Cell membrane 0.5 μm (± 0.7) 5.0 μm (± 4.2) P = 0.001**

invagination

Polysaccharide volume 6.7% (± 3.5) 25.0% (± 13.4) P = 0.03*

Chloroplast volume 8.6% (± 4.8) 13.3% (± 5.1) P = 0.02*

Mitochondria number 12.2 (± 8.9) 17.3 (± 5.5) P = 0.03*

Modified from Kruzcynski (1994). * = significant; ** = highly significant.
24 GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

H. wrightii to grow intertidally, where it may form

extensive beds, in contrast to the subtidal distribution of T. testudinum (Phillips 1960a). However,

factors other than desiccation tolerance may play a

more critical role in terms of the higher vertical

zonation for H. wrightii, for example, a greater tolerance to high irradiances and benefits from higher

nutrient input from the shore (Björk et al. 1999).

Tolerance of exposure to high light levels in

shallow or intertidal waters suggests that seagrasses

may have evolved photosynthetic mechanisms

(e.g., C4 photosynthesis) that reduce damage by

UV radiation or intense PAR. However, photosynthetic mechanisms of all seagrass studied thus far

appear to involve carbon-fixation reactions characteristic only of the Calvin cycle (C3 plants) (Beer

et al. 1977, 1980; Andrews and Abel 1977; Benedict

et al. 1980; Durako 1993), with the possible exception of the Mediterranean species, Cymodocea

nodosa (Beer et al. 1980). Carbon fixation in C3

plants occurs via the Calvin Cycle, in which the

enzyme ribulose-1,5-bisphosphate carboxylaseoxygenase (rubisco) initially produces a three-carbon (C3) compound, phosphoglycerate (Abel and

Drew 1989; Beer 1996). However, an earlier study

(Benedict and Scott 1976) reported that high levels

of labeled carbon (14C-bicarbonate) were incorporated into malate (30%) and aspartate (33%) in T.

testudinum. These two four-carbon acids are commonly associated with the initial fixation step of

the Hatch-Slack pathway of C4 plants. The process

described by Benedict and Scott (1976) for T. testudinum is similar to Crassulacean Acid Metabolism (CAM), which prevents loss of fixed carbon

(as CO2) resulting from photorespiration suffered

by plants under high light irradiance. Because seagrass lacunae store respired CO2, determining the

level of photorespiration in seagrass blades is difficult, although it appears to be lower than in susceptible terrestrial plants (Abel and Drew 1989).

More studies are needed to determine whether any

of the six Gulf-coast species have modified C4 carbon-fixation pathways. Durako (1993) demonstrated that T. testudinum can utilize both

bicarbonate (HCO3

–

) and carbon dioxide, with relatively efficient HCO3

– utilization, despite an

apparently low affinity for this form of carbon.

Plants with C4 photosynthesis have higher (less

negative) δ13C values than those of plants using

only the C3 pathway. It has been shown that seagrasses have δ13C values similar to those of C4

plants. The similarity in δ13C values may be due to

diffusional restrictions on carbon movement into

the chloroplast in a relatively closed carbon-fixation system rather than reflecting their mode (e.g.,

C3 or C4) of carbon fixation (O’Leary 1988). Overall, average seagrass δ13C values range between –10

and –11 ppt, showing stable-isotope-ratio signatures that are usually less depleted in 13C than they

are in other aquatic primary producers (Hemminga and Mateo 1996). Thus, δ13C values can be

used to trace carbon flow in short-term carbonallocation and production studies. In contrast to

the overall averages for seagrasses given by Hemminga and Mateo, the δ13C values for T. testudinum

leaves in south Florida, when growing adjacent to

mangroves (Figure 3.7), ranged from –7.3 to –16.3

with stable-isotope ratios of carbon that were similar to those of mangrove detritus (Lin et al. 1991).

The difference in δ13C values suggests influence of

mangrove detritus on nearby seagrass beds.

Macroalgae may have nitrogen stable-isotope

ratios similar to those of seagrasses, as shown for

δ15N. Using δ15N values, Dillon et al. (2002) determined the sources of nitrogen in Florida. In Sarasota Bay, macroalgae had δ15N values that averaged

+4 per mil greater than those of seagrasses, the

higher values being positively correlated with δ15N

values of wastewater discharge and being similar to

macroalgae δ15N values near populated areas in the

Florida Keys.

Seagrasses facilitate the transport of inorganic

carbon from seawater to the chloroplasts of their

Figure 3.7 Stable-carbon isotope values of seagrasses are

influenced by adjacent mangroves. 
CHAPTER 3 | AUTECOLOGY AND POPULATION GENETICS 25

epidermal cells to make up for slow CO2 diffusion

rates and low CO2 levels in water, as well as for

the low affinity of rubisco for CO2 when the lacunae have high concentrations of O2 (Beer 1996).

More importantly, seagrasses can use HCO3

– from

seawater via carbonic anhydrase-catalyzed extracellular conversions. Unlike seaweeds, which have

photosynthetic rates that are saturated by present-day CO2 levels, seagrasses continue to be

CO2-limited (Beer and Koch 1996). The authors

suggest that when seagrasses colonized the sea in

the Cretaceous, CO2 levels were higher. With

atmospheric CO2 levels again increasing, the

authors suggest that nearshore beds may expand

(but see Beer et al. 2003). However, CO2 concentration in seawater need not be a limiting factor

to seagrass growth under low irradiance conditions, as shown by Durako and Hall (1992). Using

δ13C values, they reported that CO2 concentration

in natural seawater was not limiting for T. testudinum under low light when photosynthetic

rates were low, a relationship that may be relatively common in turbid water.

Entire-plant carbon budgets have been measured for T. testudinum in Florida Bay (Fourqurean

and Zieman 1991) and for H. wrightii in Tampa

Bay (Neely 2000). Below-ground biomass (roots,

rhizomes, and short shoots) of T. testudinum

accounted for 85% of the standing stock and

57.5% of total-plant respiration (Fourqurean and

Zieman 1991). In contrast, below-ground biomass

of H. wrightii ranged from 94% (winter) to 69%

(summer) in Tampa Bay, with an above-ground

productivity of 0.25 g C m–2 d–1 in September

(Neely 2000).

Tomasko and Dunton (1995) compared four

methods (blade clip-and-reharvest, above-ground

biomass values, leaf-production turnover rates,

and below-ground:above-ground-biomass ratios)

for estimating diel rates of primary production in

H. wrightii populations of Texas. The whole-plant

biomass and leaf-turnover approaches appeared to

be useful tools for estimating productivity in H.

wrightii. The blade clip-and-reharvest method

devised by Virnstein (1982) for determining primary production in narrow-bladed seagrasses such

as H. wrightii, underestimated primary productivity. Further, they concluded that the Hsat model

that calculates total PAR for a photoperiod was

reliable and accurate in the prediction of daily carbon gain. However, they stated that although the in

situ measurement of underwater light levels and P

vs. I parameters, as used in Zimmerman et al.

(1994), is a more rigorous approach for calculating

primary production, it requires a large investment

of time and resources.

Because of their conspicuous roles in productivity and biomass generation, both the organic

composition and the caloric values of seagrasses

have been studied (Dawes 1998a). Annual mean

caloric values per gram of dry weight (g dry wt)

for leaves, short shoots, and rhizomes of T. testudinum were 10–13, 8–11, and 12–14 kJ, respectively. Similar values were obtained for S. filiforme

(10–13, 10–14, and 15–16 kJ, respectively) and H.

wrightii (13–15, 12–13, and 14–16 kJ, respectively).

Another study in seagrass beds between Tarpon

Springs and Cedar Key estimated that available

kilocalories ranged from a 16-month low of 344 to

a high of 1,837 kcal–1 m–2, with the highest SAV

biomass and caloric values occurring in the late

spring and summer (Dawes et al. 1987).

Dawes (1998a) found that lipid levels for T. testudinum, S. filiforme, and H. wrightii were low,

ranging from <0.1% to 6.3% of the dry weight for

various organs. Soluble protein ranged from 5% to

22% of the dry weight, with leaves having the

highest content in all three species. Soluble carbohydrates ranged from 6% to 54% of the dry

weight, with rhizomes having the highest levels.

Soluble protein and soluble carbohydrate levels

show pronounced seasonal and species variations,

as summarized by Zieman and Zieman (1989) and

Dawes (1998a).

Rhizomes, followed by short shoots, are the

principal storage organs for soluble carbohydrates

and proteins in T. testudinum, S. filiforme, and H.

wrightii (Dawes and Lawrence 1980) and in H.

engelmannii (Dawes et al. 1987). Further, rhizomes

are the source of soluble carbohydrates and proteins for regrowth of cropped blades and for initiation of spring blade growth in T. testudinum

(Dawes and Lawrence 1979, 1980; Tomasko and

Dawes 1989). In a review of carbon, nitrogen, and

phosphorus content in the leaves of 27 seagrass

species, Duarte (1990) found that they accounted

for 34%, 2%, and 0.2% of the dry weight, respectively, with a mean C:N:P ratio of 474:24:1.

Although the cell walls of seagrasses contain lignin,

celluloses, and hemicelluloses, they are in lower

concentrations than in terrestrial monocots (Table

3.4). Soluble, nonstructural carbohydrates (hemi-
26 GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

celluloses) can account for 1%–16% of cell-wall

dry weight in seagrass blades, whereas structural

cellulose and lignin account for 18%–40% and

0.2%–5% respectively. By comparison, tropical

grasses such as tall fescue (Festuca arundinacea

Schreb.) and Bermudagrass (Cynodon dactylon (L.)

Pers.) have 37% and 40% hemicellulose, 32% and

27% cellulose, and 10% and 6% lignin per g dry

wt–1, respectively (Dawes 1986). The adaptations of

seagrasses to a hydrophytic environment are seen in

their flexible blades (i.e., few fiber bundles) and low

amounts of lignin. Perhaps the brief life spans of

leaves of many seagrass species, ca. 15 days in T. testudinum (Witz and Dawes 1995; van Tussenbroek

1996b), may not allow time for significant production of lignin to occur, compared to the long leaf

life spans (months) of terrestrial plants, thus making seagrass blades more edible to grazers.

RHIZOSPHERE and 

SEDIMENT DYNAMICS

Seagrasses serve as sediment traps by acting as baffles, causing sediment-laden water to slow and

drop its sediment as shown for seagrasses of

Florida Bay (Prager and Halley 1999). Even species

with diminutive morphologies, such as H. decipiens, can reduce sediment movement (Fonseca

1989). However, studies in T. testudinum beds in

Tampa Bay show that sediment resuspension does

occur because of tidal currents, so these communities are both sinks and sources of suspended matter (Koch 1999b).

Seagrasses modify the anaerobic sediment in

which they grow. Roots and rhizomes release oxygen translocated from the shoot, cause bioturbation of the soil, and add organic matter via

decomposition (Moriarty and Boon 1989). Chemical changes wrought by the below-ground components are critical for plant survival because high

levels of sulfides in anoxic reducing sediments are

toxic to plants (Erskine and Koch 2000; Koch and

Erskine 2001; Carlson et al. 2002). The above-cited

experiments created sediment sulfide concentrations up to 13.0 mM; natural sulfide levels of

healthy seagrass beds in Florida Bay average <2

mM, whereas in areas of seagrass die-off, the phytotoxin averages 4–10 mM (Carlson et al. 2002).

Some evidence exists that die-off of overdeveloped

beds of T. testudinum in Florida Bay may be due

partly to lowering of plant resistance to the parasite Labyrinthula sp. during periods of increased

environmental stresses such as high salinity, high

or low temperatures, hypoxia, and high sediment

sulfides (Durako 1994; Blakesley et al. 2002; see

Chapter 5). Evidence for sulfide toxicity was found

when sulfide addition (6 mM) in tank cultures was

combined with high salinity and temperature over

a period longer than 28 days (Koch and Erskine

2001), thus mimicking die-off conditions in

Florida Bay. Carlson et al. 2002 suggested that the

higher sensitivity of T. testudinum to sulfide levels

than that of H. wrightii may be due to the former

species having a higher below- to above-ground

biomass ratio. They also reported high sedimentsulfide levels in other estuaries, including Sarasota

Bay, which may help explain the patchy distribution of seagrass beds (Figure 3.8). This suggests

that carbonate sediments, as found in Florida Bay,

are not the only type of substrate that contain high

sulfide concentrations and that seagrass losses (but

not die-off) in other areas on Florida’s Gulf coast

may also reflect sediment conditions. The release

of oxygen into its rhizosphere by T. testudinum and

the resultant decline in sulfides (converted to sulfates) was shown in seagrass beds in Laguna

Madre, Texas (Lee and Dunton 2000a).

Table 3.4 Cell-wall constituents of seagrass leaves, including lignin, cellulose, and soluble carbohydrates.

Species Site (Ref.) Lignin Cellulose Sol. Carb.

Thalassia testudinum Florida (1) 0.2–2% 18–32% 1–9%

Syringodium filiforme Florida (1) 1–3% 19–26% 10–16%

Halodule wrightii Florida (1) 2–4% 25–33% 4–14%

H. uninervis Gulf of Acaba (2) <1% 41% –

Halophila stipulacea Gulf of Acaba (2) 2% 36% –

H. ovalis Gulf of Acaba (2) <1% 33% –

Heterozostera tasmanica Philip Bay (3) 5% 20% –

Modified from Dawes (1986). l. Dawes (1986); 2. Baydoun and Brett (1985); 3. Webster and Stone (1994).
CHAPTER 3 | AUTECOLOGY AND POPULATION GENETICS 27

Much of the sulfide in anoxic sediments derives

from the activities of sulfate-reducing microorganisms. Many microbial associations in the bulk soil

and with the plant’s roots can provide a means for

obtaining scarce nutrients. For example, fungal

associations (mycorrhizal) might provide phosphorus or micronutrients. Nielsen et al. (1999)

hypothesized that the nature of seagrass sediments

restricts the development of vesicular-arbuscular

mycorrhizae (VAM) because they could not find

the association in the rhizosphere of T. testudinum,

in contrast to their finding that VAM are associated with submergent and emergent freshwater

angiosperms. Because they did not find any

(VAM) associations with the roots of T. testudinum, Nielson et al. (1999) suggested that formation of the association was restricted by the

combined effects of highly anaerobic nature and

high salinity of seagrass sediments.

Since the classic studies of Patriquin and

Knowles (1972), who used acetylene reduction by

the seagrass rhizosphere to infer nitrogen fixation,

a number of studies have described nitrogen fixation and nutrient uptake by marine angiosperms

in greater detail. Based on acetylene-reduction

experiments, Kenworthy et al. (1987) determined

that bacteria associated with the roots and rhizomes of T. testudinum fixed nitrogen. Nitrogen

fixation occurs in Lyngbya wollei (Phlips et al.

1992), a freshwater cyanobacterium. Perhaps L.

majescula, a common mat and epiphytic species in

seagrass beds of Florida’s Gulf coast, has a similar

capability.

Seagrasses take up nutrients principally from

sediment pore water (e.g., ammonium and phosphorus; Fourqurean et al. 1992) but also from the

water column (e.g., nitrate and ammonium;

Touchette and Burkholder 2000). However, porewater nutrient concentrations in seagrass beds can

vary widely, as shown in Florida Bay, where

median values of 0.34 μM for soluble reactive

phosphorus and 78.6 μM for ammonium occur

(Fourqurean et al. 1992). Because of rapid growth

rates, seagrasses require large amounts of fixed

nitrogen (10 to 450 mg N m–2 d–1), with ammonium being the preferred form (Moriarty and

Boon 1989). By contrast, nitrate and nitrite concentrations are usually low (<5 μM) in anaerobic

sediments, presumably because they are rapidly

used by denitrifying and anaerobic bacteria.

Seagrasses show a variety of responses when

nutrients become limiting. Phosphorus is usually

the most limiting nutrient in seagrass sediments of

Florida Bay, where calcium carbonate rapidly

binds any free PO4

–3 (Powell et al. 1989; Short et al.

1990; Touchette and Burkholder 2000), and its levels have been correlated with seagrass-bed development (Fourqurean et al. 1992). In Florida Bay,

phosphorus apparently controls the successional

sequence between early-stage H. wrightii beds and

late-stage T. testudinum (Fourqurean et al. 1995).

The former species tolerates higher nutrients better than T. testudinum does and will replace it in

areas of natural or anthropogenic eutrophication.

Nevertheless, factors other than nutrient levels

may contribute to the replacement of T. testudinum by H. wrightii. Studies in Florida Bay

(Carlson et al. 2002) reported that higher sulfide

levels in the root zone, resulting from enhanced

seagrass nutrition, can act as a phytotoxin in sediments and limit T. testudinum. By contrast, ammonium deficiency limits growth of T. testudinum in

Laguna Madre, Texas (Lee and Dunton 2000b) and

of S. filiforme in the Indian River Lagoon (Short et

al. 1993).

Figure 3.8 Frequency distribution of pore-water sulfide concentrations in carbonate sediments of Florida Bay (Rabbit Key

Basin) and in siliceous, terrigenous sediments of Sarasota

Bay (Carlson et al. 2002).
Manateegrass,

Syringodium filiforme,

with drift algae

COLOGICAL

ROLES

Biomass production of the three Florida seagrass species with

larger statures is highly variable, with highest values reported

for each in summer and early fall. Below-ground biomass

accounts for 50%–90% of total plant biomass.

Thalassia testudinum produces 10–19 leaf blades per short shoot

per year, with lower turnover rates in more northern latitudes.

Determinations of leaf production rates for other species are

needed.

Although vegetative growth is the primary method for seagrass

expansion, little is known about rhizome extension rates for

most species.

Seagrass communities serve as habitats and nurseries for over

170 species of invertebrates and more than 100 species of fish.

About 60 fish species using seagrass beds as habitat have noncommercial values and have poorly known requirements.

Epiphytic biota are often seasonal, serve as food, and enhance

seagrass community structure but may also negatively influence

seagrass growth by attenuating light.

A positive relationship exists between epiphytic load on seagrass

blades and nutrient enrichment of their environment, whether

the source of the nutrients is natural (e.g., bird rookeries) or

human (e.g., septic tanks or stormwater runoff).

Drift macroalgae deposited on seagrasses can result in seagrass

die-back but can also serve as habitat, as nutrient sinks, as food,

and as a transport vehicle for invertebrates and fish.

Unlike overgrazing, which has been documented twice in

Florida waters, moderate grazing may enhance the productivity

and expansion of T. testudinum. The role of secondary metabolites that may reduce grazing in seagrasses is not known.

Data regarding competition between seagrass species or with

their associated macroalgae are limited.

Animal species using seagrass habitats also often use nearby salt

marshes, mangrove forests, and coral reefs.

The export of inshore seagrass detritus offshore into deep water

is poorly documented for Florida waters but may be an important source of organic matter in deep marine systems.

FWC–FWRI photo
CHAPTER 4 | ECOLOGICAL ROLES 29

SEAGRASS COMMUNITY ECOLOGY

Like Florida’s coastal mangrove and salt-marsh

communities, seagrass communities are important

primary producers, stabilize mobile sediments,

serve as habitats and nurseries, and are direct and

indirect food for diverse fauna (Dawes 1998a).

Further, the characteristics of seagrasses (e.g.

short-shoot densities, presence of inflorescence

scars, levels of storage products) can aid in determining whether seagrass beds are stressed, thereby

contributing to the monitoring of conditions in

coastal and estuarine communities (Dennison et

al. 1993; Dawes 1998a).

In a study of seagrass distributions and conditions on the northeastern Gulf of Mexico, Livingston et al. (1998) listed four significant

predictors of seagrass community health. These

predictors are photic depth, light quality, water

quality (color, dissolved organic carbon, and

chlorophyll a), and sediment characteristics. The

principal predictor is light quantity (duration and

intensity) in relation to seagrass depth distribution, which is related to water transparency, as

shown in the review on seagrass depth limits by

Duarte (1991). In this regard, water transparency

was found to be the principal factor influencing

the depth distributions of T. testudinum in Tampa

Bay (Dixon and Leverone 1995, 1997) and growth

form of H. wrightii below a canopy of T. testudinum in the Florida Keys (Tomasko 1992).

However, light quality influences seagrass morphology as shown by altering the red to far-red

ratios in cultures of Ruppia maritima (Rose and

Durako 1994).

Seagrass communities have been measured for

standing stock biomass (g dry wt m–2) and shortshoot densities (numbers m–2). Biomass data are

commonly limited to above-ground material,

mostly the photosynthetic blade tissue. Such data

are available for the larger-stature seagrass species,

including T. testudinum, H. wrightii, R. maritima,

and S. filiforme (Table 4.1), but data are usually

lacking for the smaller-stature species (i.e., species

of Halophila). In general, the below-ground biomass (roots and rhizomes) constitutes 50% to 90%

of standing stock for the larger-stature species

(Zieman and Zieman 1989). For example, total

biomass estimates (above and below ground) for

T. testudinum in Florida Bay ranged from 195 to

2,254 g dry wt m–2, and 85% of this was below

ground (Fourqurean and Zieman 1991). Beds of S.

filiforme in the Indian River Lagoon have an

annual mean above-ground leaf biomass of 87 g

dry wt m–2 (± 151) or 43% of the total, whereas

below-ground roots and rhizomes accounted for

117 g dry wt m–2 (± 94) or 57% of the total plant

biomass (Short et al. 1993). Dunton (1990) found

that in the Laguna Madre of Texas below-ground

biomass for H. wrightii was 50% to 85% of the

total and for R. maritima was 20% to 70% of the

total biomass. On Florida’s Gulf coast, the greatest

leaf biomass for T. testudinum occurs in spring and

early summer, whereas below-ground biomass

(roots and rhizomes) is greatest in the fall (Dawes

et al. 1985). Unlike below-ground biomass, leaf

biomass in T. testudinum can show a four-fold seasonal fluctuation, with above- to below-ground

ratios of 0.17 to 0.4 in December and 0.4 to 1.0 in

April in Tampa Bay (Dixon and Leverone 1995).

Leaf turnover rates were reviewed by Short and

Duarte (2001), and techniques to measure leaf

growth were compared by Tomasko and Dunton

(1995) and van Tussenbroek (1996b). Leaf production (g dry wt shoot–1 y–1) in T. testudinum was

measured by puncturing leaf clusters at their bases

and measuring new leaf production after a few

days (Tomasko and Dawes 1989). In a review of

leaf turnover rates for T. testudinum, Marbá et al.

(1994) reported that 14 to 19 leaves are produced

annually by each short shoot. Leaf longevity on

short shoots of T. testudinum in the Tampa Bay

area using leaf-scar chronology was 14.6 days at

Sunset Park in Tarpon Springs and 19.2 days in

south Tampa Bay (Witz and Dawes 1995) and 24

days in a lagoon in the Yucatán, Mexico (van

Tussenbroek 1996b). Leaf production generally

ranged between 10 and 19 leaves short shoot–1 y–1

in Florida Bay (Zieman et al. 1989), the Indian

River Lagoon (Gacia 1999), Tampa Bay (Witz and

Dawes 1995), Charlotte Harbor (Tomasko and

Hall 1999), Lower Laguna Madre, Texas (Kaldy et

al. 1999), and the Caribbean coast of Mexico

(Marbá et al. 1994; van Tussenbroek 1996b). Populations growing in cooler waters had lower annual

rates of leaf production (10–15 leaves shoot–1 ) for

the Indian River Lagoon and the Lower Laguna

Madre in Texas compared to those in Tampa Bay.

Determination of leaf turnover in narrow-bladed

seagrasses, such as H. wrightii (Tomasko and Dun-
30

ton 1995) and S. filiforme (Fry and Virnstein

1988), requires using an approach different from

the leaf-punch technique. Leaf productivity by S.

filiforme, determined using leaf clipping and photography, was 1.8 g dry wt m–2d–1 (Virnstein 1982;

see Short and Duarte 2001 for a detailed review of

techniques).

Little is known about rhizome growth rates of

Florida’s seagrasses, yet this is the sole method for

vegetative expansion. Reviews indicate that vegetative growth differs among species, rhizome

elongation rates being higher for smaller-stature

seagrass species (Marbá and Duarte 1998). In

their review, mean rhizome elongation rates were

highest for H. wrightii, the smallest stature of the

larger Caribbean seagrasses (223, range; 81–365

cm yr–1), followed by S. filiforme (123, range:

52–182 cm yr–1) and then the largest stature

species T. testudinum (69, range: 22–152 cm yr–1).

Short et al. (1993) found that S. filiforme rhizome

elongation exceeded 100 cm y–1 in the Indian

River. At Mullet Key in Tampa Bay, rhizome

internode length and increase in seagrass patch

size for H. wrightii were positively correlated; the

larger the internodes of H. wrightii, the greater the

increase in seagrass coverage at the edges of

patches (Jensen and Bell 2001). The authors also

found that a 100-fold increase in phosphorus, but

not nitrogen, resulted in a significant increase in

rhizome internode length. Jensen and Bell (2001)

suggested that the observed patch-size patterns in

seagrass beds (e.g. larger landscape features) may

be determined by small-scale factors (e.g. nutrient

availability).

GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

Table 4.1 Standing stock (biomass) and productivity of southeastern United States seagrasses and closely related

species (congeners) from other areas. Procedures differ greatly and influence the ranges in biomass and productivity.

Biomass Productivity

Species Locality (g dry wt · m–2) (g C · m–2 · d–1)

Halophila

H. ovalis India 48

Western Australia 40–60

H. engelmannii Texas (Gulf coast) 1.6

Thalassia

T. testudinum Florida (general) 20–8,100

Florida (Atlantic coast) 20–1,800 0.9–16.0

Florida (Gulf coast) 75–8,100

SW Florida Bay (Gulf coast) 60–125

Indian River, FL (Atlantic coast) 890

Corpus Christi, TX (Gulf coast) 454–885

Mexico, lagoon (Gulf coast) 420–1,418

Mexico, reef (Gulf coast) 573–811

Cuba (Caribbean Sea) 20–800 9.3–12.5

T. hemprichii Queensland, Australia 70

Syringodium

S. filiforme Florida (general) 15–200

Indian River, FL (Atlantic coast) 27–81

S. isoetifolium New Guinea (East Indies) 327

Halodule

H. wrightii Mullet Key, FL (Gulf coast) 120 0.25

Florida (general) 0.8

North Carolina (Atlantic coast) 105–200 0.5–2.0

Laguna Madre, TX (Gulf coast) 70–300

Mississippi Sound (Gulf coast) 256

H. uninervis New Guinea (East Indies) 150

Ruppia maritima Texas (Gulf coast) 0–200

Modified from McRoy and McMillan (1977), Zieman and Wetzel (1980), and Hillman et al. (1989).
CHAPTER 4 | ECOLOGICAL ROLES 31

STRUCTURE, SHELTER, 

AND PREDATION 

Seagrass beds function both as nurseries (Lewis

and Stoner 1981; Lewis 1984; Virnstein 1987; Zieman and Zieman 1989; Fonseca et al. 1996b; Bell et

al. 2001; Heck et al. 2003) and as habitats for

invertebrates (Greening and Livingston 1982;

Virnstein 1987; Valentine and Heck 1993), fish

(Stoner 1983; Gilmore 1987), sea turtles (Williams

1988a), and manatees (Lefebvre et al. 1989). The

role of seagrass meadows as nurseries was

reviewed by Heck et al. (2003) who reported a

strong link between seagrass abundance and abundances of juvenile fin fish and shell fish. The principal factor in this positive correlation was the

structure that seagrasses added to the habitat.

Invertebrates add structure and serve as food, thus

supporting increased diversity of other forms. For

example, high densities of the mussel Modiolus

americanus in St. Joseph Bay reached 2000 individuals m–2, which enhanced invertebrate diversity

and secondary production (Valentine and Heck

1993). The high variation of animal diversities

among seagrass sites may be due in part to wave

and current activity (Bell et al. 1994). The authors

found that the mean numbers of fish and copepods and mean numbers of fish species were

higher in low-energy seagrass beds (e.g., minimum

exposure to wind direction, speed, and effective

fetch in Tampa Bay and Onslow Bay, North Carolina). Densities of macroinvertebrates in seagrass

beds on Florida’s Gulf coast vary widely among

sites (Table 4.2), ranging from 910 individuals m–2

(Rookery Bay) to 33,485 individuals m–2 (Tampa

Table 4.2 Densities of macroinvertebrates at various sites in Florida, arranged geographically from the Atlantic coast

to the northern Gulf of Mexico. Single-density values are means; ranges generally represent several sampling sites. Several values are derived indirectly.

Mesh Number

Seagrass Faunal Size Density of

Study Site Species* Group (mm) (indiv. m–2) Species Reference

Card Sound Tha total 0.8 1,085 – Brook (1977)

Bahia Honda Tha Amphipoda 1.0 102 5 Nelson (1980)

Florida Bay Tha total 0.8 2,794 52 Brook (1978)

Rookery Bay Hal Amphipoda 1.0 910 3 Nelson (1980)

Tampa Bay Hal Polychaeta 0.62 13,313 44 Santos and Simon (1974)

Tha Polychaeta 0.62 33,485 Santos and Simon (1974)

Anclote Anchorage Syr total 1.0 2,347 68 Mahadevan and Patton (1979)

Syr total 0.5 9,538 58 Mahadevan and Patton (1979)

Tha total 1.0 3,724 63 Mahadevan and Patton (1979)

Tha total 0.5 18,916 50 Mahadevan and Patton (1979)

Apalachee Bay Tha/Syr total 1.0 1,782 to 2,424 72 Lewis and Stoner (1981)

Tha/Syr total 0.5 3,154 to 4,754 86 Lewis and Stoner (1981)

Tha total 0.5 16,108 80 Lewis and Stoner (1983)

Tha/Hal Crustacea 0.5 6,716 46 Lewis (1984)

Tha/Syr/Hph total 0.5 2,827 170 Stoner (1980b)

Apalachicola Bay Hal total 0.5 38,780 58 Sheridan and Livingston (1983)

St. Andrew Bay Tha total 0.7 7,567 89 Saloman et al. (1982)

Hal total 0.7 3,370 67 Saloman et al. (1982)

Tha/Hal total 0.7 4,192 86 Saloman et al. (1982)

Santa Rosa Sound Tha total 0.5 13,260 144 Morton et al. (1986)

Hal, core total 0.5 9,020 41 Stoner et al. (1983)

Hal, suction total 0.5 6,077 37 Stoner et al. (1983)

Modified from Virnstein 1987. *Hal = Halodule wrightii, Tha = Thalassia testudinum, Syr = Syringodium filiforme, Hph = Halophila

engelmannii.
32

Bay). Invertebrate communities of seagrass beds in

Florida are highly regional, and overlap occurs

between subtropical, tropical, and warm-temperate faunas (Virnstein 1987). Holmquist et al.

(1989) found that seagrass beds in Florida Bay play

a secondary role as habitat for decapods and stomatopods compared with their reactions to winter

cold fronts as a function of their positions on mud

banks. Decapod crustaceans, such as caridean

shrimps, usually dominate the larger fauna numerically, similar species of shrimp being widespread

in Florida’s Gulf-coast seagrass beds (Holmquist et

al. 1989). The macrofauna varies widely in species

composition and animal density over distance and

time, including those of amphipods, gastropod

molluscs, and polychaete worms (Virnstein 1987).

Because infaunal, epibenthic, and epiphytic animals of seagrass beds are prey for larger animals,

fish are abundant there (Gilmore 1987). Principally, the larger-sized fish (e.g., spotted seatrout,

Cynoscion nebulosus, and pinfish, Lagodon rhomboides) have been studied. Smaller nonfishery

species (e.g., gobiids, syngnathids), which account

for most of the resident species in seagrass beds,

are less well known but include over 100 species

(Gilmore 1987). As with invertebrates, fish faunas

vary with geographic regions on Florida’s Gulf

coast and include warm-temperate species, eurythermal tropical species, and stenothermal tropical

species (Gilmore 1987). Four subenvironments in

Florida Bay had an overall average density of 11

fish m–2 and a total of 56 species (Sogard et al.

1987). The two-year study found seagrass standing

crop and litter to be the most important determinates of fish numbers.

Monitoring animal densities and diversity in

planted seagrass beds is a useful method for determining habitat restoration success. A study of a

three-year-old planted seagrass bed in Tampa Bay

(Fonseca et al. 1996a, 1996b) found that shrimp,

fish, and invertebrate densities were equivalent to

those of natural communities, although shortshoot density was only one-third those of natural

beds. Another study in Tampa Bay found that

higher numbers of polychaetes occurred in twoyear-old planted sites than in natural H. wrightii

beds (Bell et al. 1993). The rapid increase of animals in planted beds suggests that the threshold

value of habitat structure may be much lower than

previously thought and that the lack of correlation

between patch size and animal abundance suggests

other factors are involved in maintaining animal

populations. Clearly, more studies are needed to

determine animal population relationships in

restored, created, and patchy seagrass beds.

Studies on seasonal animal-population fluctuations and patch formation in seagrass meadows

suggest that they are not static, even under “stable”

environmental conditions. For example, Tampa

Bay seagrass beds experiencing more rapid water

movement have a lower number of fish and

harpacticoid copepods (Bell et al. 2001), which

may also be influenced by organic matter content

in the sediment. In St. Croix, Virgin Islands, tropical seagrass beds fluctuate in response to seasonal

conditions (Williams 1988b). In addition to seasonal changes, severe plant die-off has occurred in

Florida Bay, probably hastened by spatial variations in environmental stresses (Durako 1994; Zieman et al. 1999; see also Chapter 5). Water-column

nutrient increases resulting in eutrophication have

been cited as important environmental stressors

that can shift species dominance in seagrass beds

and influence sulfide production (Carlson et al.

2002). Williams (1987) found that T. testudinum

could shade and replace S. filiforme beds in St.

Croix. She also reported that the successional

sequence from bare sediment began with rhizophytic (psammophytic) algae, then proceeded to S.

filiforme, and ended with T. testudinum, as sediment nutrient concentrations increased over time

(Williams 1990). In this regard, considering nutrient increases emanating from bird perches

installed in Florida Bay, Fourquean et al. (1995)

proposed that H. wrightii might replace T. testudinum in Florida Bay as sediment nutrient levels

rise. However, this study did not consider the effect

of increased sulfide levels in the sediment (Carlson

et al. 2002) or the effect of phosphate and nitrate

on seagrass growth (Powell et al. 1989).

The competition for space or the “space race” is

evident by the numbers of macroalgae, cyanobacteria and microalgae (Hall 1988; Harris 1997) that

grow on seagrass blades and that show strong seasonal changes. At Egmont Key in Tampa Bay, the

five species of cyanobacteria and 20 species of epiphytic macroalgae varied in dominance (by algal

division) during a 15-month study of T. testudinum. Brown algae dominated during the winter and spring months and cyanobacteria in the

GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY
CHAPTER 4 | ECOLOGICAL ROLES 33

summer months (Figure 4.1). Hall (1988) divided

the algae into those that (1) are characteristically

epiphytic on seagrasses; (2) settle and reproduce

on seagrass blades, but are commonly epilithic;

and (3) germinate on blades and then break free

before reproduction. A study of T. testudinum beds

near Leffis Key and Siesta Key in Sarasota Bay

(Table 4.3) found many species of centric and pennate diatoms, six species of cyanobacteria, and 21

species of macroalgae (Harris 1997). Algae epiphytic on H. wrightii and T. testudinum in Tampa

Bay, Sarasota Bay, and Charlotte Harbor (Dixon

and Leverone 1995; Dixon and Kirkpatrick 1995,

1999) and Santa Rosa Sound in Pensacola Bay

(Macauley et al. 1988) change seasonally, species

diversity being negatively correlated with water

temperature. Gacia et al. (1999) found that sheetlike and filamentous species of green-algae genera

(Enteromorpha, Cladophora) were less resistant to

grazing fish than were species of more coarsely

branched red-algae genera (Hypnea, Chondria,

Acanthophora) in the Indian River Lagoon. Algae

epiphytic on seagrass blades can account for 46%

of the total biomass and 60% of the benthic primary production, as found using stable-isotope

analysis, in H. wrightii beds in Mississippi Sound

(Moncreiff and Sullivan 2001).

Studies have shown that epiphytes, in combination with PAR attenuation and seasonal lowgrowth periods, depress blade growth and primary

production and reduce the depth of T. testudinum

in Tampa Bay (Dixon and Leverone 1995, 1997),

Florida Bay (Frankovich and Fourqurean 1997),

the Florida Keys and Keys in Belize, Central America (Tomasko and Lapointe 1991), and Perdido

Key in northwest Florida (Wear et al. 1999).

Tomasko and Lapointe (1991) found water-column dissolved inorganic nitrogen (DIN) concentration was high near Big Pine Key (3.96 μM), an

island in the Florida Keys with over 2,000 septic

tanks. This DIN level is similar to that of a mangrove island in Belize (Man-O-War Key) having 75

pairs of nesting frigate birds (3.80 μM). In contrast, DIN levels were low around Cutoe Key (0.50

μM), a remote mangrove island in the Florida

Keys. The seagrass beds near Big Pine Key and

Man-O-War Key had higher epiphytism, lower

short-shoot densities, lower leaf-area indices, and

lower biomass in contrast to those near Cutoe Key

(Tomasko and Lapointe 1991). Wear et al. (1999)

reported similar results from an experiment using

slow-release fertilizer (Osmocote®) in northwestern Florida at Perdido Key. Nutrification caused

significant increases of diatoms, as well as red and

brown macroalgae, on blades of T. testudinum, H.

wrightii, and S. filiforme over a 12-month period.

Frankovich and Fourqurean (1997) reported that

epiphytic loads on T. testudinum, along a nutrientenrichment gradient in Florida Bay, gave similar

results to the previous studies, and this study indicated that the effect is localized. The authors concluded that epiphyte levels alone are not as

responsive to moderate nutrient enrichment as are

other seagrass characters, such as leaf tissue N:C

ratios. As noted in Chapter 3, epiphytes of seagrasses reduce water-column PAR by as much as

99% seasonally. The question of whether organic

carbon, nitrogen, or phosphorus can be transferred from seagrass blades to the epiphytic community or vice versa is not clear (see Lobban and

Harrison 1994).

Some of the most abundant invertebrates in

Figure 4.1 Seasonal variations in the mean numbers of

cyanobacteria and macroalgae, grouped by divisions, epiphytic

on Thalassia testudinum blades at Egmont Key in Tampa Bay

over a 15-month period (Hall 1988).
34

seagrass beds are the meiofauna, including crustaceans (Lewis 1987) and copepods (Hall 1988).

Harpacticoid copepods are often the most common animal group associated with seagrasses (Hall

and Bell 1988, 1993; Walters and Bell 1986; Walters

1991). One study reported that more than 20,000

harpacticoid copepods migrated h–1 m–2 from the

sediment during the night into the water column,

representing over 50% of the total benthic

harpacticoid community (Walters and Bell 1986).

Further, this postsunset entry into the water column from the sediment involved over 30 species in

15 families (Walters 1991). At Mullet Key in

Tampa Bay, artificial blades (green ribbon) with

artificial epiphytes (cotton fibers) attached had significantly higher densities, compared to green ribbons alone, of copepods, polychaetes, and

nematodes associated with them after three days.

This shows that epiphytic biota enhance structural

complexity of seagrass blades, thereby increasing

habitat values (Hall and Bell 1988). In a similar

study, Hall and Bell (1993) showed that the most

abundant meiofaunal groups on seagrass blades at

Egmont Key were adult harpacticoid copepods,

copepod nauplii, and nematodes. They also determined that meiofauna density was most signifiGULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

Table 4.3 Epiphytic algae on blades of Thalassia testudinum at Leffis Key and Siesta Key in Sarasota Bay (modified from

Harris 1997) and at Egmont Key in the mouth of Tampa Bay (modified from Hall 1988).

Species Leffis Key Siesta Key Egmont Key

Cyanobacteria

Anabaena oscillarioides x xx

Anacystis sp. x x x

Calothrix crustacea x xx

Entophysalis conferta x

Microcoleus lyngbyaceus x xx

Porphyrosiphon notarisii x

Chlorophyta

Boodleopsis pusilla x

Chaetomorpha minima x x

Cladophora spp. x x x

Enteromorpha chaetomorphoides x

Ulva lactuca x

Phaeophyta

Cladosiphon occidentalis x x

Ectocarpus elachistaeformis x

Ectocarpus rallsiae (as Giffordia rallsiae)x x x

Ectocarpus sp. (as E. rhodochortonoides)x x x

Hincksia mitchellae (as Giffordia mitchellae)x x x

Hummia onusta x xx

Myrionema magnusii (as M. orbiculare)x x x

Myriotrichia occidentalis x xx

Rhodophyta

Acrochaetium spp. x x x

Centroceras clavatum x xx

Ceramium cimbricum (as C. fastigiatum)x x x

C. flaccidum (as C. byssoideum)x x x

Chondria dasyphylla x xx

Chondrophycus papillosus (as Laurencia papillosa) x

Erythrotrichia carnea x

Hydrolithon farinosum (as Fosliella farinosa)x x x

Hypnea musciformis x x

Polysiphonia subtilissima x x

Stylonema alsidii (as Goniotrichum alsidii) x
CHAPTER 4 | ECOLOGICAL ROLES 35

cantly correlated with percentage cover of filamentous algae. A 15-month study of T. testudinum

blades at Egmont Key at the mouth of Tampa Bay

(Hall 1988) showed seasonal shifts in the abundance of nematodes (highest in winter) but not

that of harpacticoid copepods and nauplii (abundant throughout the year). Epiphytes trap detritus,

which attracts significantly higher numbers of

copepods that feed on it (Meyer and Bell 1989).

Small grazers (e.g., copepods) can play a role in

epiphyte control (Virnstein 1987), as shown for H.

wrightii plants in the Indian River Lagoon

(Howard and Short 1986).

Along with drift macroalgae, rhizophytic

(psammophytic) algae may equal or surpass seagrass biomass in Florida Gulf-coast seagrass beds

(Dawes et al. 1985). With the exception of some

tropical coenocytic green algae, macroalgae usually

cannot attach to sand or mud. Coenocytic green

algae are a diverse group of siphonaceous (lacking

internal cell walls) algae placed in the order

Caulerpales. They form extensive rhizoid holdfasts

and include species of calcified (Halimeda, Penicillus, Udotea) and noncalcified (Caulerpa, Riphilia,

Avrainvillea) genera. Even though rhizophytic

coenocytic algae may compete for space and light

with seagrasses, they may also serve as pioneer

plants, which then may be replaced by seagrasses,

as shown in a coral-reef lagoon in St. Croix

(Williams 1990).

Over 230 species of unattached macroalgae

(drift algae) are known throughout the world

(Norton and Mathieson 1983). Drift algae are usually noncalcified, remain infertile, and reproduce

vegetatively (Collado-Vides et al. 1994). Two studies, one in Tampa Bay, Boca Ciega Bay, and Anclote

Anchorage (Phillips 1960b) and the second in

Crystal Bay (Phillips 1960c), listed 195 and 46 taxa,

respectively, of algae that could not be distinguished as either drift or attached. Drift algae are

concentrated in seagrass beds where water-current

velocities are reduced and blades will entangle the

algae (Figure 4.2). To some extent, seagrass blade

size may influence entanglement (Bell et al. 1995).

Seagrass beds on the Gulf coast of Florida are

known to contain from 8 to 65 species of drift

algae with the total biomass ranging from 0.002 g

to 930 g dry wt m–2 (Table 4.4). Brown (2001)

found that algal biomass differed for three sites on

the Gulf coast of Florida, both by season and by

site. Annual means of 6.01, 3.83 g, and 2.25 g dry

wt m–2 were obtained for Tampa Bay in Cockroach

Bay, Tarpon Bay, and Sunset Beach respectively.

Although biomass varies with season and site on

the Gulf coast, the types and numbers of species

are similar (Table 4.4).

Drift algae are an important ecological component of seagrass beds because they serve as habitat,

provide transportation, and provide a food source

for invertebrates (Ballantine et al. 1994; Holmquist

1994; Greenway 1995; Knowles and Bell 1998;

Maciá 1999; Brooks and Bell 2001; Rydene and

Matheson 2003). In Florida Bay, Holmquist (1994)

found 61 invertebrate taxa in drift algae mats in

Figure 4.2 Drift algae are entangled by seagrass leaves, as in this Thalassia testudinum bed. 

FWC–FWRI
36

seagrass beds. Average drift rate was 0.5 km d–1,

indicating that drift algae are an effective mode of

transport for animals. Ballantine et al. (1994)

reported that balls of algae in Venezuela contained

12 species of invertebrates. The numbers of

amphipods, isopods, and tanaeids in Venezuela

were significantly higher on two species of red

algae than were associated with blades of S. filiforme, as similarly found in other previous studies

(Knowles and Bell 1998). Virnstein and Howard

(1987) noted that drift algae serve as habitat and

food sources for gammaridean amphipods. Brooks

and Bell (2001) found that drift clumps of the red

alga Hypnea cervicornis, as they moved through

seagrasses and into sandy areas, had significantly

higher amphipod densities; they concluded that

drift algae clumps provide a mobile corridor for

animals. In Tampa Bay, drift algae form an important alternative habitat for 12 of the 20 most abundant juvenile and small adult fish (Rydene and

Matheson 2003). However, drift algae deposited in

seagrass beds can also affect these habitats by causing a seagrass die-back as a consequence of light

and oxygen reduction (Norkko et al. 2000). A

GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

Table 4.4 Number of drift-algae species and dominant species, biomass, and presence/absence (by division) in seagrass beds on the Gulf coast of Florida. Locations include (1) Anclote Estuary near Tarpon Springs (Hamm and Humm

1976), (2) Hillsborough Bay including studies 2a (Kelly 1995) and 2b (Avery 1997), (3) Sunset Beach at Tarpon Springs

(Brown 2001), (4) Cockroach Bay in Tampa Bay (Brown 2001), and (4) Tarpon Bay at Sanibel Island (Brown 2001).

Locations

1 2a 2b 3 4 5

Number of species 65 19 18 9 9 8

Number of dominant species 4 5 9 9 9 8

Mean biomass (g dry wt · m–2) low – – 80 0.07 0.0002 0.14

high – – 930 6.34 9.45 9.32

Cyanophyta

Lyngbya majescula x x

Chlorophyta 

Caulerpa prolifera x

Chaetomorpha crassa x

C. gracilis x

Chaetomorpha sp. x

Enteromorpha spp. x

Enteromorpha clathrata x

Ulva lactuca xx x

Phaeophyta 

Sargassum pteropleuron x

Sargassum spp. x

Rhodophyta

Acanthophora spicifera x xx

Chondrophycus papillosus (as Laurencia papillosa) x

Digenia simplex x

Gracilaria armata x x

G. caudata xxx

G. cornea (as G. debilis) x

Gracilaria spp. x x

Hypnea musciformis xxx

Laurencia intricata (as Laurencia obtusa)x x

L. poiteaui x

Polysiphonia ramentacea x

P. subtilissima x

Solieria filiformis (as Agardhiella tenera) x xxxx

Spyridia filamentosa xx xxx
CHAPTER 4 | ECOLOGICAL ROLES 37

combination of algae deposition and sea-urchin

grazing increased the die-back of T. testudinum in

Florida Bay (Maciá 1999). The amount of driftalgae biomass can also have a negative correlation

with seagrass bed expansion if dissolved nutrients

are high, as shown for Hillsborough Bay (Avery

1997). In a 10-year study, Avery (2000) found that

with improved water quality the 18 species of

dominant drift algae declined in average biomass

from 164 g dry wt m–2 in 1987 to 0.0029 g dry wt

m–2 in 1998 in Tampa Bay. The green coenocytic

alga Caulerpa prolifera had colonized Hillsborough

Bay in 1986 with an initial coverage of almost 220

ha (543 acres) but had declined to zero in 1997. In

contrast, H. wrightii coverage increased from 0.2

ha (0.5 acres) in 1986 to nearly 57 ha (141 acres) in

1998 (Avery 2000).

A few studies exist concerning competition

between seagrasses or between seagrasses and

attached macroalgae, with two reports on T. testudinum, S. filiforme, and H. wrightii and rhizophytic algae in the Caribbean (Williams 1987,

1990). Williams (1987) removed the leaf canopy of

T. testudinum in St. Croix, resulting in increased

irradiance reaching the S. filiforme understory, and

its biomass doubled within nine months. A study

in Tampa Bay indicated that competition occurred

between T. testudinum and H. wrightii (Rose and

Dawes 1999), where T. testudinum had a lower biomass when growing with H. wrightii, suggesting

interspecific competition. Further, leaf biomass

and growth were significantly lower in dense,

monotypic beds of T. testudinum compared to less

dense beds, suggesting light was reduced via shading in dense beds (Rose and Dawes 1999). There

are also reports regarding competition between

H. wrightii and the coenocytic green macroalga

C. prolifera (Bottone and Mattson 1987; Bottone

and Savercool 1993). Although the data are not

conclusive, C. prolifera may act as a weak competitor for space with H. wrightii, coexisting with it as

an opportunistic species under stable environmental conditions in Tampa Bay (Bottone and Savercool 1993). This result is similar to those found in

studies in the Indian River Lagoon (White and

Snodgrass 1990). Further, C. prolifera, although

susceptible to sudden environmental changes (e.g.,

low salinity), can rapidly colonize areas devoid of

vegetation that may occur following seagrass dieback (Bottone and Mattson 1987).

Overgrazing of seagrass beds is a rarity but has

occurred on the Florida Gulf coast. Overgrazing by

Lytechinus variegatus (purple sea urchin) resulted

in large-scale denudation of seagrass beds. Over

20% of an area 26 km by up to 9.5 km were consumed near the Pepperfish Keys south of the Steinhatchee River in the northern peninsula (Camp et

al. 1973). Density at the front of a single aggregate

averaged 636 sea urchins m–2, with individuals

piled upon one another 2–8 individuals deep that

covered the substrate. Dense aggregations of the

same urchin species overgrazed about 0.81 km2 of

T. testudinum in outer Florida Bay; mean densities

ranged from 364 to <1 individual m–2 (Rose et al.

1999). The origins of the urchin feeding fronts are

not understood. In Florida Bay, high urchin densities may reflect unusually high recruitment or a

release from predation pressure because of overharvesting of spiny lobsters and stone crabs (Rose

et al. 1999).

Moderate grazing pressures on T. testudinum

blades were correlated with specific leaf growth

rates (0.024 mg dry wt. d–1) in the Yucatán, Mexico;

this showed that sporadic grazing had little detrimental effect on T. testudinum (Cebrián et al.

1998). Studies on T. testudinum beds in northwest

Florida and a review of the literature indicated that

herbivory plays a major role in stimulating seagrass growth by removing macroalgae (Heck and

Valentine 1995; Valentine and Heck 1999). Also,

long-term effects of grazers have been shown in a

relationship between L. variegatus and T. testudinum in the Gulf of Mexico (Heck and Valentine 1995), where urchins and seagrass coexist in

balance. In these cases, a balance developed among

intensive grazing, loss of habitat, and predation by

fish on the urchins. In the 1995 experiment, protection from predator fish by enclosures resulted in

intense grazing by the urchins, which became most

destructive in winter when T. testudinum could not

recover as rapidly. The more intensively grazed

beds showed significant reductions in above- and

below-ground biomass, which was apparent even

3.5 years after grazing had ceased. Varying the

number of L. variegatus in experimental plots in

St. Joseph Bay, Valentine et al. (1997) found an

increase in shoot density and productivity of T.

testudinum under moderate grazing pressure. In a

later study, Valentine et al. (2000) demonstrated

that the effect of grazing in Florida Bay was related
38

to season and water depth, with grazing having a

stronger effect in the winter or in seagrass beds at

greater depths. Thus, repeating experiments in

environments with differing physical conditions

and in different seasons is critical in order to

determine the effects of blade loss and seagrass

bed maintenance.

Few studies have been carried out on predation

within seagrass beds since the reviews by Orth et

al. (1984) and Zieman and Zieman (1989). Based

on feeding and caging studies, predation, especially by pinfish and pink shrimp, is probably a

major biological interaction influencing invertebrate community structure (Virnstein 1987). In a

study in St. Joseph Bay, increased biomass of a

mussel (Modiolus americanus) in T. testudinum

beds resulted in greater abundance, biomass, and

annual production of other invertebrates (Valentine and Heck 1993).

TROPHIC DYNAMICS

The ultimate fate of plant biomass is mineralization to simple inorganic compounds, with a portion of seagrass leaf decomposition occurring

while it is still attached to the plant. Nevertheless,

most biomass decomposition occurs on and in the

sediment of seagrass meadows (Hemminga and

Duarte 2000).

A long-term study (419 days) in Laguna

Madre, Texas of H. wrightii senescent blades suspended in the water column revealed that decomposition resulted in a loss of 36% of the organic

matter within the first 24 days (Opsahl and Benner 1993). By the end of the study, decomposition

had resulted in a 76% loss of the organic matter.

Neutral sugars were the most abundant in the

starting material and also the most rapidly lost.

In contrast, the most persistent sugars in the

senescent blades were xylose and glucose, reflecting the more stable polymers of cellulose and

xylan from which the sugars are derived. Near St.

Croix, at depths between 14 and 32 m, the

decomposition rates of a more delicate seagrass,

H. decipiens, differed between litter bags that

were buried in the sediment and those tethered at

the surface of the sediment (Kenworthy et al.

1989). After 6.5 days, buried leaves lost 56% of

the original ash-free dry weight compared to only

28% lost by those left on the surface. These findings were similar to two 7-day studies made in

two consecutive years that included photosynthetic measurements in the same area and using

the same species (Josselyn et al. 1986), this earlier

GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

Table 4.5 δ15N and δ13C values of seagrasses and marine macroalgae from various locations.

Plants δ15N δ13C

Seagrasses

Thalassodendron ciliatum (Australia) +3.5 –9.3

Syringodium isoetifolium (Australia) +5.0 –4.0

Thalassia testudinum (Nicaragua) +3.5 –13.2

Halodule wrightii (Texas) +3.9 –10.3

Ruppia maritima (Texas) +3.6 –7.9

Halophila engelmannii (Texas) +3.8 –7.9

Thalassia testudinum (Texas) +3.9 –8.9

Thalassia testudinum (Jamaica) +4.3 –11.1 –x = 3.9 ± 0.5 –9.1 ± 2.7

Marine Macroalgae

Ulva fasciata (Texas) +8.1 –14.6

Ulva lactuca (Texas) +8.1 –14.6

Gelidium crinale (Texas) +7.9 –14.3

Ascophyllum nodosum (Maine) +8.0 –16.9

Fucus vesiculosis (Maine) +8.1 –16.5

Acetabularia kilnori (Australia) +6.5 –9.5 –x = 7.8 ± 0.6 –14.4 ± 2.6

Macko (1981) cited in Fry et al. (1987).
CHAPTER 4 | ECOLOGICAL ROLES 39

study reporting greater than 50% decomposition

of all buried seagrass material.

Seagrass epiphytes appear to be a more significant source of direct food for many invertebrates

than are seagrass detritus or living tissue. Virnstein

(1987) has proposed the following sequence for a

south Florida seagrass food web: (1) the important

primary producers are epiphytic algae, which have

rapid turnover rates; (2) epiphytic algae are preferentially grazed by most species of small invertebrates associated with seagrass blades; (3) small

invertebrates are preyed upon by decapod crustaceans or small fishes that include resident adults

and juvenile seasonal visitors; and (4) decapods

and small fishes are preyed on by larger, mostly

nonresident fishes. Of the 154 grazers listed by

McRoy and Helfferich (1980) that directly feed on

seagrass blades, the most important invertebrates

are echinoderms, molluscs, and crustaceans. For

example, seagrass blades constitute up to 50% of

the total diet for crustaceans (Klumpp et al. 1989).

Crabs, isopods, and sea urchins are the principal

direct grazers on seagrasses, whereas other invertebrates consume seagrass detritus and various algal

epiphytes. Fry and Parker (1979) and Fry (1984),

using stable carbon isotopes (δ13C), showed that

microalgae on the sediment and seagrass blades

are a major source of carbon for grazers. In some

seagrass meadows, micro- and macroalgae predominate in food webs (Virnstein 1987). In a

review of stable-isotope studies, Fry et al. (1987)

concluded that algae are nutritionally more important for consumers than are seagrasses but that

δ13C data alone can be misleading because benthic

algae have values similar to those of seagrasses.

Thus, they recommended using other elements

with stable isotopes, including nitrogen and sulfur

(Tables 4.5 and 4.6).

Table 4.6 δ34S values for sulfur sources and plants, Redfish Bay, Texas, November 1980–February 1981 (from Fry, 1981).

Sources δ34S

Seawater Sulfate +19.7 to +20.0

Sediment Sulfide –23.4

Porewater Sulfate +15.0 to +17.0

Rain Sulfate +6.3a

Leaves Roots

After After

Total Washing Total Washing

Algae, Submerged Plants

Digenia simplex +18.7 +17.2

Gracilaria verrucosa +18.6

Gracilaria debilis +18.9 +17.5

Thalassia epiphytesb +17.4 +15.2

Seagrasses, Submerged Plants

Thalassia testudinum

Sample 1 (November 29)c +15.2 –12.3 –15.9

Sample 2 (November 29) +12.9d

Sample 3 (January 3) 6.9d –17.3

Halodule wrightii +10.4 –9.3

Syringodium filiforme c +11.5 –3.3

Halophila engelmannii +11.2 +11.5

Intertidal, Emergent Plants

Spartina alterniflorae +3.5 –9.4

Avicennia germinans e (as A. nitida) –0.2 –3.2

a

From Jensen and Nakai (1961). b

Epiphytic community collected after freeze drying. Contained some small animals and mud.

c, ePlants marked with the same letter were growing intertwined in the same sediments.

d

Epiphyte-free seagrass.
40

RELATIONSHIPS with 

OTHER COMMUNITIES

Because seagrass beds are mostly found in areas of

low wave energy, they often occur next to tidal-flat,

salt-marsh, and mangrove communities along

Florida’s Gulf coast and are influenced by these

intertidal communities. Salt marshes are estimated

to cover about 170,000 ha (419,900 acres) of

Florida’s coasts; about half of the marsh area

extends from Tampa Bay north to the Alabama

border on the Gulf coast (Montague and Wiegert

1990). The estimated salt-marsh area is substantially less than the estimate of McNulty et al.

(1972) of 214,000 ha (528,580 acres) for the Gulf

coast of Florida. The largest salt marshes on the

Gulf coast of Florida are in the Big Bend area,

from Aripeka to Apalachicola Bay, a coast with low

wave energy and extensive seagrass beds. About

60% of northwest Florida salt marshes are monotypic stands of Juncus roemerianus Scheele (Needle

Rush), which often extend down to 0.2 to 0.5 m

above MLW and thus are adjacent to near-shore

seagrass beds. Above-ground primary production

ranges from 250 to 950 g dry wt m–2 yr–1 for J. roemerianus and 130 to 700 g dry wt m–2 yr–1 for

Spartina alterniflora Loisel. (Smooth Cordgrass).

(Montague and Wiegert 1990), the difference

probably due to the higher density of J. roemerianus culms. The large biomass of salt-marsh plants

and low direct consumption by herbivores (10%)

means that most of the biomass decomposes and

becomes part of the salt-marsh detritus (Montague et al. 1987). Apparently little of the detritus

is exported to near-shore seagrass beds because of

the relatively high elevation of Florida’s salt

marshes and often the presence of a berm that

retains the detritus within the salt marsh (Montague et al. 1987).

In contrast to the low level of detrital export, a

diverse fauna is shared between salt marshes and

adjacent seagrass beds, for example, mullet, spot,

blue crabs, oysters, and penaeid shrimps (Montague and Wiegert 1990). In addition, tarpon

(Megalops atlanticus), snook (Centropomus undecimalis), red drum (Sciaenops ocellatus), seatrout

(Cynoscion spp.), and kingfish (Menticirrhus spp.)

move from near-shore seagrass beds into saltmarsh tidal streams (Lewis et al. 1985b).

The three Caribbean mangrove species, Rhizophora mangle L. (Red Mangrove), Avicennia germinans (L.) L. (Black Mangrove), and

Laguncularia racemosa (L.) C. F. Gaertn. (White

Mangrove) are estimated to cover 189,725 ha

(468,620 acres) along Florida’s coasts (Lewis et al.

1985b). About 90% of mangrove forests (or mangals) occur in the most southern four counties of

Florida (Lee: 14,275 ha, Collier: 29,126 ha, Monroe: 94,810 ha; Dade: 32,931 ha), a region of low

wave energy and consequently abundant seagrass

beds. Because of the lack of severe freezing temperatures since 1989, mangroves extend intermittently northward of Tampa Bay to the western

panhandle on the Gulf coast and north to the

Tomoka River on the Atlantic coast (Odum and

McIvor 1990; D. Crewz pers. obs.). Export of mangrove-leaf detritus is highest in riverine (1.2–2.7 g

carbon m–2 d–1) and fringing forests (0.5–0.7 g carbon m–2 d–1), and these contribute to detrital-based

food webs in near-shore seagrass beds (Odum and

McIvor 1990). Stable carbon isotope levels (δ13C)

of seagrass blades and mollusk shells taken from

seagrass beds near south Florida mangrove forests

had significantly lower mean values (–12.8 and

–2.3 ppt, respectively) than did blades and shells

far from the coastal forests (–8.3 and +0.6). The

lower δ13C values near mangrove forests indicate

release of carbon dioxide by mangrove-detritus

mineralization, conversion to bicarbonate, and

subsequent uptake of bicarbonate by the adjacent

seagrass beds (Lin et al. 1991).

The habitat values of mangals, adjacent seagrass

beds, and nonvegetated open water were compared

for densities of fish, shrimp, and crabs in Rookery

Bay near Naples (Sheridan 1992). Fish densities

were highest in nonvegetated areas (74% of the

total caught), while shrimp and crab densities were

highest in seagrass beds (74% and 47% of the total

capture respectively). However, flooded R. mangle

forests had resident and transient fish and crab

numbers equal to those of other areas on some

occasions. The data suggest that seagrass beds are a

principal habitat for shrimp and crabs, whereas

mangrove prop roots and pneumatophores may

serve an ancillary role for a variety of fish and

invertebrates, contrary to the conclusion drawn in

earlier studies (e.g., Thayer et al. 1978, 1987).

South Florida’s shallow-water coral reefs are

unique in North America. Florida’s reefs are simiGULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY
CHAPTER 4 | ECOLOGICAL ROLES 41

lar in species composition and physiographic features to those in the Caribbean Sea and Bahama

Islands (Dawes 1998a). Occasionally, hard corals

occur as far north as Jupiter Island on the Atlantic

coast (27°N lat.). Coral reefs form extensive threedimensional structures south and west of Cape

Florida and along the Florida Keys archipelago

(Jaap and Hallock 1990). Seagrass beds cover 80%

of the sea bottom between Cape Sable and

throughout Florida Bay, along the Florida Keys,

extending westward to the Dry Tortugas, and

north on the Atlantic coast into Biscayne Bay, thus

co-occurring with the Florida reef tract (Jaap and

Hallock 1990). Aerial photography of Biscayne

National Park showed that moderately dense to

highly dense seagrass beds covered 25,445 ha

(62,849 acres) of the 55,000 ha (135,850 acres) in

the park (Lewis et al. 2002) and were closely associated with coral patch reefs.

The relationship of seagrass beds to patch and

tract reefs is considered to be in danger due to

nutrient-rich runoff from canals in Biscayne Bay

(Lewis et al. 2002) and from septic tanks in the

Florida Keys (Jaap and Hallock 1990). A key feature of coral-reef dependence on adjacent seagrass

beds is their sensitivity to nutrient enrichment,

which enhances macroalgae overgrowth on the

coral structures (Dawes 1998a). Thus, nutrient

removal from the water column by adjacent seagrass beds may play a major role in protecting

coral reefs. In addition, the ecological role of seagrasses as habitat and nursery in the life cycle of

the spiny lobster (Panulirus argus), whose juveniles

settle in algae associated with seagrass beds, has

been noted (Marx and Herrnkind 1985).

Seagrass detritus and vegetative parts are

exported into deep-water habitats, especially on

the Atlantic coast. Deep water (>200 m) is farther

offshore on the Gulf coast of Florida than along

the Atlantic, but seagrass beds are more developed

in the Gulf. The larger-stature seagrass species

extend to about 9 m (Dawes 1974; Iverson and Bittaker 1986). In contrast, the smaller-stature

Halophila engelmannii grows in 1–5 m in Tampa

Bay (Dawes 1967), in 1.4 m off Tarpon Springs

(Phillips 1960b), and to 73 m off the Gulf coast

(Dawes and van Breedveld 1969; Dawes and

Lawrence 1990) and the Dry Tortugas (Taylor

1928). The more diminutive H. decipiens appears

to be limited to deep water on the Gulf coast and

occurs in 20 to 90 m (Dawes and van Breedveld

1969; Zieman 1982; Dawes and Lawrence 1990;

Fonseca et al. 2001; see also Chapter 4). In the

summer and early fall on the Gulf coast, H. decipiens often forms extensive beds in 20–30 m of water

(Dawes and Lawrence 1990). However, no information currently exists regarding the export or

role of detritus from inshore seagrass beds to the

200-m deep continental shelf west of Florida

(Florida Institute of Oceanography 1994).
ATURAL

and 

ANTHROPOGENIC 

EFFECTS

Seagrass regression and die-off observed in Florida Bay in the

late 1980s resulted from a combination of natural and human

causes, including overproduction of seagrasses that resulted in

dense shoots; anoxia that increased sediment sulfide levels;

nutrient inputs that caused algal blooms, reducing light levels;

lack of freshwater inputs that resulted in hypersalinity; and

infection by Labyrinthula, a pathogen of Thalassia testudinum.

The relationship between Labyrinthula infection and die-off of

T. testudinum is not clear, as the pathogen has been found in a

number of Florida Gulf-coast estuaries in seagrass beds that

have not experienced die-off.

Storms appear to have little direct impact on most seagrass communities, except for erosion and deposition of sediment in some

areas, as found following Hurricane Andrew.

The 1997–1998 El Niño event caused a general die-back of seagrass communities because of low salinities and increased nutrient loads and turbidity from runoff.

The importance of long-shore sand bars is not adequately understood, but they may protect seagrass beds from waves and erosion in some locations.

Mechanical damage to seagrass beds continues, with propeller

scarring being common in shallow-water communities.

The effects of small-scale seagrass-bed damage on local animal

populations is not adequately understood; for example, moderately scarred beds show little change in shrimp and pinfish densities.

Nutrient pollution continues to be a serious problem, with 8 of

the 15 Florida Gulf-coast estuaries regarded as being eutrophic.

Nutrient enrichment from natural or human sources results in

increased phytoplankton density and resultant chlorophyll levels, reduction in light penetration, increased seagrass loss,

increased epiphytism, increased sediment hypoxia, and shifts in

seagrass species dominance.

The effects of sediment contaminants on seagrass communities

are not clear. The highest contaminant levels of metals and 

pesticides are associated with fine muds, limited PAR, and low

salinities.

Stargrass, Halophila engelmannii 

Ron Phillips photo
CHAPTER 5 | NATURAL AND ANTHROPOGENIC EFFECTS 43

any studies, including those by resource

management programs of the National

Estuary Programs for Tampa Bay, Sarasota Bay,

and Charlotte Harbor and by the National Estuarine Research Reserves at Apalachicola and Rookery

bays, have emphasized the importance of Florida’s

Gulf-coast seagrass communities. Understanding

of natural- and human-induced seagrass disturbances has increased substantially, as shown in a

review by Durako (1988), Short and Wyllie-Echeverria (1996) and Duarte (2002). Since 1989, many

publications have examined the condition of seagrass communities on Florida’s Gulf coast. These

include studies of the north-central Gulf coast

(Mattson 1995, 2000), Tampa Bay (Lewis et al.

1991, 1998; Johansson and Ries 1997; Kurz et al.

1999; Pribble et al. 1999; Greening 2002a; Johansson 2002a), and Sarasota Bay, Lemon Bay, and

Charlotte Harbor (Kurz et al. 1999; Staugler and

Ott 2000; Tomasko et al. 2001). Declines in seagrass coverage in various estuaries on the Gulf

coast (including Florida Bay) have often been

attributed to interactions between natural and

anthropogenic factors.

In Florida Bay, T. testudinum die-off was noted

in late 1987 by Robblee et al. (1991) and has been

studied by many others since then (Carlson et al.

1994; Durako 1994; Durako and Kuss 1994; Thayer

et al. 1994; Carlson and Yarbro 2001; Fourqurean

and Robblee 1999; Hall et al. 1999; Zieman et al.

1999). Often, natural processes can result in

increased damage to seagrass beds if they had been

previously stressed by anthropogenic impacts, as

appears to have happened regarding the seagrass

die-off in Florida Bay. Despite the acknowledged

interactions of natural and anthropogenic influences in these and many other systems, however,

these categories are addressed separately below.

NATURAL PROCESSES

Natural processes have biological and nonbiological sources and include such influences as storms,

sea urchin population increases, phytoplankton

blooms, and macroalgae overgrowth. Such sources

can cause direct or indirect damage that include

shading, overgrazing, or disease.

Although moderate grazing is common and

may even affect seagrass communities positively,

overgrazing by urchins, and to a lesser extent by

manatees or sea turtles, can result in complete

denudation in some areas. For example, overgrazing by Lytechinus variegatus (purple sea urchin)

resulted in severe seagrass denudation near the

Pepperfish Keys (Camp et al. 1973) and in outer

Florida Bay (Maciá and Lirman 1999; Rose et al.

1999).

As summarized in Chapter 4, epiphytes on seagrass blades have both positive and negative

effects, with most damage being linked to shading

and reductions in photosynthesis. In Tampa Bay,

shading by epiphytes of T. testudinum blades is an

important controlling factor in its depth distribution (Dixon and Leverone 1997).

High levels of drift algal biomass suppress seagrass-bed expansion and, similar to seagrass epiphytism, are usually linked to nutrient enrichment.

In Hillsborough Bay, Avery (2000) reported in a

10-year study (1986–1994) that the 18 species of

dominant drift algae (see Table 4.6 and Chapter 4)

declined from an average biomass of 930 to 80 g

dry wt m–2 and continued to decline through 1997.

In contrast, due to reductions of water-column

nutrients in Hillsborough Bay, H. wrightii beds

expanded from 0.2 ha to 40 ha (Avery 2000). On

the positive side, drift and epiphytic algae probably

serve as nutrient sinks and as food for a variety of

invertebrates and fish (Kharlamenko et al. 2001;

Moncreiff and Sullivan 2001).

The decline in the past 20 years of seagrass

communities in Florida Bay is a conspicuous

example of the effects of natural and possibly

human influences on seagrass populations on the

Gulf coast. In 1987, vast areas of T. testudinum

began dying rapidly. Short-shoot densities

dropped by 22% and standing crop by 28% at 107

sampling stations (Figure 5.1) between 1984 and

1994 (Hall et al. 1999). Losses were highest in

western Florida Bay in areas with high standing

crops. Similar or greater declines in standing crops

were evident for H. wrightii (92%) and S. filiforme

(93%) between 1984 and 1994. Hall et al. (1999)

suggested that the most likely cause was chronic

light reduction due to increased water turbidity.

A pathogen, Labyrinthula, was also suggested as

being involved in the die-off of T. testudinum in

Florida Bay (Porter and Muehlstein 1989; Robblee

et al. 1991). It is similar to one that was implicated

in the wasting disease of Z. marina over the past 50

years in the Atlantic (Burdick et al. 1993). Porter

M
44

and Muehlstein (1989) reported the presence of an

undescribed species of Labyrinthula in blackened,

necrotic lesions on the leaves of T. testudinum collected in seagrass die-off areas of Florida Bay.

Durako and Kuss (1994) found that photosynthesis in necrotic lesions on T. testudium blades was

impaired. They suggested that this may have

reduced oxygen available for transport via the

blade lacunae to belowground organs, resulting in

increased sediment hypoxia and leading to sulfide

toxicity. The pathogen has also been found to be

infecting from 0% to 100% of T. testudinum blades

in at least 10 sites on the Gulf coast (Figure 5.2)

from the Chandeleur Islands (LA) to Florida Bay

(Blakesley et al. 2002). The authors found that

60% to 90% of the sites sampled in Tampa Bay had

infected short shoots of T. testudinum.

Thayer et al. (1994) ascribed the die-off in

Florida Bay to “as yet unknown environmental

stresses.” Studies now suggest that the die-off arises

from a combination of human-derived and natural factors that may collectively stress T. testudinum

and thereby enhance infectivity by the pathogen

(Blakesley et al. 2002). The factors involved include

hypersalinity (Fourqurean and Robblee 1999);

persistent microalgal and cyanobacterial blooms

and resuspended sediments that reduced illumination (Hall et al. 1999); anoxia that increased levels

of sediment sulfide (Carlson et al. 1994, 2002;

Carlson and Yarbro 2001); and enhanced growth

that resulted in high biomass (Zieman et al. 1999).

Hypersaline areas of Florida Bay probably resulted

in part from a lack of flushing in the absence of

periodic hurricanes. This condition may also be

exacerbated by a reduction in freshwater flow from

the Everglades because of surface-water diversion.

Nevertheless, data covering 1989–1995 suggest that

seagrasses were once again growing rapidly in

Florida Bay and that the loss of T. testudinum

appeared to be slowing (Zieman et al. 1999).

Little information is available regarding the

effects of storms on Florida’s Gulf-coast seagrass

communities, except for that following Hurricane

Andrew, which swept across southern Florida on

August 24, 1992, and exited into Florida Bay at

Lostmans River (Ogden 1992; Dawes et al. 1995).

Although coastal mangrove forests were destroyed,

seagrass beds in Lostmans Bay showed no signs of

alteration or depletion of the meiofauna when

sampled two and nine months after the storm

(Dawes et al. 1995). To be sure, during hurricanes

large volumes of water fall simultaneously upon

broad landscapes, leaching large amounts of nutrients and other compounds from soils. Together

with the lowering of salinities, this short-term

pulse of pollutants may alter community dynamics

in seagrass beds.

Also, excessive rainfall over an extended period

can damage seagrass communities. An El Niño

event between December 1997 and March 1998

resulted in over 52 cm (approximately 20 inches)

of rain, which triggered widespread, persistent

phytoplankton blooms, decreasing the light available to seagrasses, along the Gulf coast of Florida

(Carlson et al. 2003). In 1997, 1998, and 1999,

sampling near the Homosassa and Anclote rivers

and in Tampa Bay and Charlotte Harbor found

that short-shoot densities, blade widths and numbers, leaf-area indices, rhizome-apex densities, and

stored carbohydrates of T. testudinum declined

(Carlson et al. 2003) in response to the 1997–1998

El Niño.

Long-shore sand bars often occur seaward of

seagrass communities and have been suggested to

play a role in protecting seagrasses from strong

waves and currents (Lewis 2002). Thus, reductions

in seagrass coverage in some areas of Tampa Bay

may reflect historical loss of these bars, similar to

those in North Carolina (Fonseca and Bell 1998).

In this regard to this idea, Fonseca et al. (2002)

employed a wave-exposure model (Relative ExpoGULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

Fig 5.1 Standing crop (±SD) of Thalassia testudinum in six

ecological regions of Florida Bay 1984 and 1994 (Hall et al.

1999). MA = Mainland, NE = Northeast, EC = East Central,

IN = Interior, AT = Atlantic, GU = Gulf.
CHAPTER 5 | NATURAL AND ANTHROPOGENIC EFFECTS 45

sure Index—REI) to evaluate the influence of longshore sandbars on seagrass coverage in Tampa Bay.

Their results hindcasted that erosion or loss of

long-shore sand bars on the southeastern shore of

the bay, where wave exposure is relatively high, had

caused the loss of seagrass originally located

between the shore and the historical sand bars.

They recommended that no restoration efforts be

conducted in areas of strong wave exposure without including engineering to reduce REIs.

ANTHROPOGENIC EFFECTS

Livingston (1987) summarized anthropogenic

(human-derived) effects on seagrass meadows

(Table also presented in Zieman and Zieman

1989), including direct physical damage and pollution, with an emphasis on nutrient enrichment.

Since that review, various studies have described

anthropogenic effects and also include estimates of

seagrass losses (see Chapter 4). Nutrification and

sediment loading from maintenance dredging of

shipping channels and vessel-generated resuspension are cited as important causes of seagrass

declines (Schoellhamer 1991; Tomasko and

Lapointe 1991; Culter and Leverone 1993; Quammen and Onuf 1993; Lapointe et al. 1994). Dredging, although now generally restricted, continues

in all estuaries on the Gulf coast to maintain channels and create new harbors. Certainly, such activities cause temporary resuspension of

contaminated sediments and contribute to reduction of water transparency (Godcharles 1971;

Schoellhamer 1991; Ailstock et al. 2002).

Mechanical damage is also done to seagrass

beds by erosive effects following boat (and ship)

groundings, by propeller wash, and by ship and

boat wakes in shallow waters. Kenworthy et al.

(1988a) found that boat wakes substantially

increased bottom shear stress along edges of shallow seagrass beds. Even though nets used for baitshrimp trawling in Tampa Bay do not appear to

cause much direct damage to seagrass beds (Meyer

et al. 1999) a large diversity of animals are killed as

by-catch, plus trawling may suspend sediments

which decreases water clarity.

Propeller scarring is another type of physical

damage to seagrass beds (Figures 5.3, 5.4) that

continues to increase in all areas of coastal Florida

(Sargent et al. 1995). Many shallow flats and mud

banks are now severely eroded due to constant

scarring, ship groundings, chronic wave action

from boats, and water-current scouring (Kruer

1994). Culter and Leverone (1993) stated that prop

scarring was visible in almost all seagrass beds in

Sarasota Bay. Sargent et al. (1995) determined that

6.2% (over 70,000 ha or 183,000 acres) of Florida’s

1.1 million ha (2.7 million acres) of seagrass beds

have been scarred by boat propellers (prop-dredging, prop cuts) and by similar causes, principally

in coastal waters less than 2 m deep. They concluded that scarring in seagrass beds has become

acute because of increasing human population

densities, increasing popularities of boating, fishFigure 5.2 Mean percentage of Thalassia testudinum shortshoots infected with a pathogenic species of Labyrinthula at

ten sites along the Gulf coast. Numbers over bars are sample

sizes. NE = not examined. (modified from Blakesley et al. 2002).
46

ing, diving, and other water sports, and increasing

tourism. Scarring level was divided among three

categories: 1) light scarring, in which less than 5%

of the seagrasses in a particular map polygon were

scarred; 2) moderate scarring, in which 5% to 20%

of the seagrasses within the polygon were scarred;

and 3) severe scarring, in which more than 20% of

the seagrasses in the polygon were scarred (Sargent

et al. 1995). The highest level of scarring was

found in Monroe and Citrus counties, which also

contained the highest amount of seagrass coverage

(688,259 ha; 1.6 million acres). Of this total, they

calculated that 23,332 ha (57,630 acres) were

scarred with 17.3% of the seagrasses being in

Monroe County and 15.8% in Citrus County.

The influences of prop scars on seagrass community production have been studied in Tampa

Bay and Charlotte Harbor (Durako et al. 1992;

Clark 1995; Dawes et al. 1997; Bell et al. 2002),

Sarasota Bay (Folit and Morris 1992), and the

Florida Keys (Matthews et al. 1991). Fragmentation of seagrass beds in Tampa Bay due to propeller cuts did not appear to have any consistent

effects on some animal populations over a oneyear period, as long as seagrass patch sizes were

greater than 1 m2 (Bell et al. 2002). The numbers

of pinfish (L. rhomboides), pipefish (Syngnathus

scovelli), and eight species of epibenthic shrimp

were similar in moderately scarred (6% to 31%

loss of the beds) and nonscarred seagrass beds in

Tampa Bay. The results of these studies suggest

that propeller scars that fragment seagrass beds

may enhance certain faunal development caused

by edge effects along the cuts as long as they are

not too severe. Nevertheless, a recent study of scarring in a T. testudinum bed in Puerto Rico revealed

a negative effect of scarring on crabs and molluscs

up to 5 m from the scar. Also, shrimp species

within the scar differed from those in the nonscarred seagrasses. Fish populations did not show

an effect from the scarring. Further studies are

clearly needed to define the effects of moderate

scarring compared to those of severe scarring on

seagrass productivity.

More recent studies have confirmed previous

observations by Jones (1968), Godcharles (1971),

and Zieman (1976) for the long time period

required for scars to heal (Eleuterius 1987; Durako

et al. 1992; Dawes et al. 1997). When seagrass rhizomes and roots are completely removed by scarring, sediment is destabilized and resuspension

occurs, thereby lowering water transparency and

retarding seagrass regrowth into the scar.

Regrowth of T. testudinum into prop scars in

Tampa Bay was estimated to require an average of

3.5 to 4.1 years for existing propeller scars and up

to 7.6 years in newly made ones (Dawes et al.

1997). Shorter recovery periods in existing scars

probably reflect their older ages and shallower

depths. Based on studies at Weedon Island

(Durako et al. 1992) and Cockroach Bay (Dawes et

al. 1997) in Tampa Bay, the authors recommended

that shallow-water seagrass beds be protected from

propeller scaring. Criteria for the protection of

seagrass beds have been proposed and adopted for

Tampa Bay (Clark 1995; Stowers et al. 2002).

Another source of physical impacts that

degrade seagrass communities is docks, which

shade seagrass beds, as shown in Perdido Bay

(Shafer 1999), Charlotte Harbor (Loflin 1995), and

the Indian River Lagoon (Beal and Schmit 2000).

Smith and Mezich (1999) surveyed 200 out of the

3,592 permit-exempt single- and multi-family

docks that were deemed to have the potential to

GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

Figure 5.3 Propeller scarring from recreational boating in a Thalassia testudinum bed near Panama City, Florida.

FWC–FWRI photo
CHAPTER 5 | NATURAL AND ANTHROPOGENIC EFFECTS 47

damage seagrass beds in Palm Beach County. They

found that 16% of the docks were larger than regulations allowed, owing to additions, and that

approximately 1,491 single-family docks in the

county had affected seagrass negatively. They estimated that, in total, docks eliminated 20.4 ha (50.4

acres) of seagrass beds, which is equal to 2.1% of

the seagrass extents in Palm Beach County. Their

recommendations were that all docks in Florida be

considered within the jurisdiction of FDEP permitting programs. Currently, environmental

resource permits are not required within designated Aquatic Preserves if the docks are less than

46.5 m2 (500 ft2

) or if they are less than 92.9 m2

(1,000 ft.2) when outside Aquatic Preserves. Smith

(1998) concluded that strong enforcement of the

regulations protecting sovereign submerged lands

(253.04–05, 380.05, F.S. and Chapter 18-14, F.A.C.)

would act as a powerful deterrent against boat and

dock damage to seagrass systems.

Direct physical damage of seagrasses can result

from changes in freshwater inputs. The importance to watershed management and reduction of

freshwater removal from tributaries flowing into

Gulf-coast estuaries has been reviewed for Tampa

Bay (Zarbock 1991) and for Charlotte Harbor

(Kurz et al. 1999). Estevez (2000) found that

changes in the amount, timing, and location of

freshwater inflow are primary stressors to estuarine and oceanic seagrass communities, with salinity changes often being a first-order stressor

(Montague and Ley 1993). The present estimated

daily freshwater inflows into Tampa Bay (Zarbock

1991) range from 152.0 to 214.6 m3 s–1 (1,792 to

2,530 ft3 s–1) compared to historical estimates of

193.4 m3 s–1 (2,280 ft3 s–1). Although timing and

duration of riverine inflow, along with nutrient

loading and changes in water color, clearly affect

seagrass communities in Charlotte Harbor, the

relationships between these factors is more complex in Tampa Bay and Sarasota Bay (Kurz et al.

1999). Similarly, diversion of fresh water before it

enters the Everglades, coupled with a decrease in

hurricane-induced flushing (Thayer et al. 1994)

and rising salinities and sediment anoxia (Zieman

et al. 1999), has been proposed as influencing factors in the precipitous decline of Florida Bay seagrasses.

Short and Neckles (1999) hypothesized that

global warming will probably influence seagrass

distribution by causing increases in sea level, in

storm frequency and intensity, in disease from

higher water temperatures, and in turbidity from

eutrophication. The authors also predicted an

increase in ultraviolet radiation that will damage

seagrasses, as has been shown for macroalgae

(Larkum and Wood 1993) and terrestrial plants

(Caldwell et al. 1989). In addition, the shallow

coastal and estuarine waters and low elevations on

the Gulf coast of Florida will certainly be affected

by any eustatic sea-level rise. A potential positive

effect on seagrass distribution may occur from an

increase in dissolved inorganic carbon (Ci) composition in seawater, as seagrasses are presently

CO2 limited and have an affinity for Ci (Beer and

Koch 1996).

Nutrient enrichment, probably the most common human effect in Gulf-coast estuaries, can

result in a variety of changes in seagrass communities. A number of these changes have been presented in Chapter 4. These include an increase in

water-column chlorophyll levels (Johansson 1991;

Janicki et al. 1999; Tomasko et al. 1996, 2001; Morrison et al. 1997), in algal epiphytism and shading

Figure 5.4 Propeller scarring from commercial fishing in a seagrass bed in Pine Island Sound, Florida.

FWC–FWRI photo
48

of seagrass blades (Tomasko and Lapointe 1991;

Tomasko et al. 1996; Dixon 2002), in shifts in

dominant seagrass species (Duarte 1995;

Fourqurean et al. 1995), and in sediment hypoxia

(Carlson et al. 2002).

A review of U.S. estuaries (Bricker et al. 1999)

cited 15 estuaries on the Gulf coast of Florida that

had various levels of eutrophic conditions. Eight of

these estuaries (Florida Bay, the southern Ten

Thousand Islands, Caloosahatchee River, Charlotte

Harbor, Sarasota Bay, Tampa Bay, Choctawhatchee

Bay, and Perdido Bay) had high levels of eutrophic

symptoms. More detailed data are available regarding nutrient loadings and damage to seagrasses in

Tampa Bay (Treat and Clark 1991; Zarbock et al.

1994, 1996; Treat 1997; Lewis et al. 1998; Pribble et

al. 1999; Greening 2002b), Sarasota Bay (Haddad

1989; Tomasko et al. 1992, 1996), and Lemon Bay

(Tomasko et al. 2001). Tampa Bay has been the

subject of several reviews (Treat and Clark 1991;

Treat 1997; Greening 2002a) that include data

regarding point- and nonpoint-source pollution,

domestic and industrial effluents, accumulation of

toxic substances in the sediment, and potential

brine damage from desalination facilities.

Many studies worldwide support the contention that seagrass distribution, survival, and

growth are principally determined by water clarity

(Batiuk et al. 1992; Dennison et al. 1993; Duarte

1991; Kenworthy and Haunert 1991; Kenworthy

1993; Koch and Beer 1996; Morris and Tomasko

1993; and Tomasko et al. 2001; also individual

papers contained in the publications of Bortone

2000 and Greening 2002a). However, other factors

such as waves and water currents can affect seagrass distribution as well (Koch 2001; see Chapter

2). Water clarity is influenced by nutrient inputs,

which stimulate the growth of phytoplankton

(particularly by nitrogen), and by total dissolved

and suspended solids. An overabundance of nutrients leads to eutrophication that supports

increased levels of phytoplankton, drift algae biomass, and coverage by epiphytes on seagrass

blades, all of which result in shading of seagrass

leaves (Sand-Jensen 1977, 1990; Cambridge et al.

1986). In all cases a decrease in water clarity over

seagrass beds was correlated with a decline in seagrass standing stock (Orth and Moore 1983; Short

et al. 1996; Short and Wyllie-Echeverria 1996; Fitzpatrick and Kirkman 1995; Lewis et al. 1985a).

The influence of possible sediment contaminants on seagrass beds is not understood. The concentration of metals, nutrients, petroleum,

chlorinated hydrocarbons, and radionuclides correlate with the smaller particle-size sediments

(muds; <63μm in diam.) in Tampa Bay (Brooks

and Doyle 1991) and with sediment-quality factors

(Zarbock et al. 1997). Thus, concentrations of all

contaminants are highest in west-central Hillsborough Bay and Old Tampa Bay (Brooks and Doyle

1991) for total organic matter (to 16%), total

hydrocarbons (>40 μg g–1), total organic nitrogen

(0.1%), and total phosphorus (0.4%). Using the

triad method of weighting sediment chemistry,

toxicity, and benthic samples, Zarbock et al. (1997)

determined that upper and middle Hillsborough

Bay, parts of Old Tampa Bay, Boca Ciega Bay, and

western Middle Tampa Bay had the most contaminated sediments. In a review of sediment quality in

Tampa Bay, Grabe (1999) found that approximately 1% of its sediments were subnominal and

had a high probability of being toxic. However, at

the levels known for Tampa Bay, evidence that the

heavy metals and pesticides directly affect seagrasses does not exist.

In conclusion, it is apparent that a mix of natural and human-induced events has negatively

influenced the seagrass meadows on the Gulf coast

of Florida. Because of limited historical data,

determining the extent of seagrass decline is often

difficult, although detailed historical analyses are

possible in some areas, such as for Tampa Bay

(Greening 2002a). Anthropogenic influences, particularly those that result in a decline in water

quality, are considered the principal causes of seagrass losses throughout the world (Short and Wyllie-Echeverria 1996) and in the estuaries of

Florida’s Gulf coast (Lewis et al. 1985a). The loss of

seagrass communities continues worldwide

(Duarte 2002) as well as on Florida’s Gulf coast

(see Chapter 2). However, restoration management

can result in the expansion and enhancement of

seagrass meadows, as seen in Tampa Bay and Sarasota Bay (Greening 2002b). Future studies are

needed to distinguish secondary and primary

stressors that result in seagrass declines, such as the

modification of watershed-level stormwater runoff

and direct nutrient input.

GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY
ix seagrass species, known throughout the Caribbean Sea and Gulf of Mexico, occur on the Gulf coast of Florida. A seventh species, Halophila johnsonii, is known only from the Atlantic coast of Florida. Worldwide, sixty

species of seagrasses are recognized. They are currently placed by Tomlinson

(1982) in 12 genera and four families in a single class of monocots, the Liliopsida (Kuo and McComb 1989) and one subclass, Helobiae (Tomlinson 1982).

The inclusion of Ruppia maritima as a seagrass (Dawes 1998a; Littler and Littler 2000) brings the number of species to 61 and families to 5. The family

and genus descriptions below are modified from Dawson (1966), Tomlinson

(1982), and Kuo and den Hartog (2001). A key based on vegetative characters

is followed by species descriptions modified from den Hartog (1970), Littler

and Littler (2000), and Kuo and den Hartog (2001).

S

AXONOMY of FLORIDA

SEAGRASSES

APPENDIX

Charles Seaborn photo
50 GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

ARTIFICIAL KEY to SPECIES, BASED on VEGETATIVE FEATURES

(see also Florida Department of Environmental Protection: http://www.fiu.edu/~seagrass/key/seagrasskey.html)

1. Leaf blades cylindrical; terete in cross section . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .Syringodium filiforme.

1. Leaf blades flat . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .2.

2. Leaves paddle-shaped, associated with scale-like leaves . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .3.

2. Leaves strap-like or linear . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .5.

3. Leaves in a pseudo-whorl; each shoot with 2 scales at the base and 2 scales halfway up the shoot . . . . . . .

 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .Halophila engelmannii.

3. Leaves paired, with 2 scales at their base . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .4.

4. Leaves linear-elliptical, margins entire . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .Halophila johnsonii.

4. Leaves oblong to elliptical, margins minutely serrate . . . . . . . . . . . . . . . . . . . . . . . . . . .Halophila decipiens.

5. Leaves greater than 3 mm wide; rhizomes scaley . . . . . . . . . . . . . . . . . . . . . . . . . . . .Thalassia testudinum.

5. Leaves usually less than 3 mm wide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .6.

6. Leaf tips blunt with dentate points; rhizomes usually straight . . . . . . . . . . . . . . . . . . . .Halodule wrightii.

6. Leaf tips with fine points; rhizomes often in a zig-zag pattern . . . . . . . . . . . . . . . . . . . . .Ruppia maritima.

HYDROCHARITACEAE Jussieu

Family of 15 genera, diverse and cosmopolitan, of

which 3 are seagrasses (Enhalus, Halophila, Thalassia). Plants aquatic herbs, leaves either large, linear

to strap-shaped and with sheathing base

(Thalassia) or small, usually nonlinear, with distinct petiole and without sheathing base

(Halophila). Rhizomes with scale-like leaves, and

branching via apical bifurcation. Flowers unisexual

and anthers either sessile or on long slender filaments. Female flowers usually with long hypanthium and inferior ovary. Fruits fleshy and often

indehiscent.

THALASSIA Banks and Solander ex König. 1805.

Ann. Bot. (König et Sims) 2: 96 (Gr. thalassa:

marine, of the sea).

Rhizomes creeping, horizontal, thick, fleshy,

and indeterminate; bearing scale leaves and erect

short-shoots (rhizomes) at irregular intervals that

arise from rhizome meristem. Erect short shoots

determinate, encased in leaf sheaths, producing

foliate leaves and flowers. Roots unbranched,

fleshy, arising at nodes of short shoots. Leaves

strap-shaped, with sheathing bases, and growing

from basal meristem. Short shoots with conspicuous leaf and inflorescence scars. Plants are dioecious. Male flowers on short stalks, in clusters of

1–3. Female flowers solitary. Fruits globose and

opening by irregular splitting, containing several

large angular “seeds” (seedlings).

Thalassia testudinum Banks and Solander ex

König. 1805. Ann. Bot. (König et Sims) 2: 96 (L.

testudinis: a tortoise). Common name: Turtlegrass

(Figure A.1.4).

Rhizomes creeping, 3–6 mm in diameter, 4–7

mm long between nodes, with scale leaves. Roots

single, unbranched, arising at short-shoot nodes,

producing delicate root hairs at tips. Erect short

shoots with clusters of 3–7 leaves; leaves to 2 cm

wide, with persistent colorless basal sheath, 9–17

parallel veins, and lacunae that are continuous

throughout the plant. Plants dioecious. Flowers on

short peduncles, with single whorl of white tepals.

Male flowers with 3 to 12 stamens, pollen grains in

mucilage, forming moniliform chains. Female

flowers bearing one ovary with 6–8 carpels; style

divided into 2 filiform stigmata. Fruits buoyant,

15–20 mm long, rough, fleshy, tips pointed, with

one to few “seeds” (seedlings).

HALOPHILA Du Petit-Thouars. 1806. Nova

Madag. 2 (L. halo: salt + philus: loving).

Rhizomes horizontal, usually on surface and

bearing two scales at each node, producing a lateral shoot and one unbranched root, short erect

shoots either with paired leaves on long petioles,

distichously arranged along the axes, or in a

pseudo-whorl at top of shoot. Leaves oval, ellipti-
APPENDIX | TAXONOMY OF FLORIDA SEAGRASSES 51

cal, lanceolate or linear, with mid and marginal

veins. Plants monoecious or dioecious. Flowers

covered by 2 bracts; sepals small, petals absent.

Male flowers stalked, with 3 stamens having sessile

anthers. Female flowers sessile, with 3–6 styles.

Fruits ovoid capsules, bearing many seeds.

Halophila decipiens Ostenfeld. 1902. Bot.

Tildsskr. 24: 260 (L. decipio: beguiling, deceptive).

Common name: Paddlegrass (Figure A.1.1).

Rhizomes thin, with transparent, hairy scales,

internodes 1.0–4.5 cm long. Petioles 3–15 mm

long, bearing one pair of leaves; leaves oblong to

elliptical, 10–25 mm long, 3–6 mm wide, with

short rigid unicellular hairs on one or both sides of

blades, margins finely serrate. Plants monoecious,

male and female flowers on same stalk, covered by

spathe, stalks to 1.5 mm long. Male flowers with 2

anthers to 1 mm long. Female flowers with 3 styles

1.5–2.5 mm long; single ovary to 1 mm long.

Fruits 2.5 mm long, 1.5 mm wide, elliptical, bearing up to 30 seeds; seeds oval, to 0.2 mm long.

Halophila engelmannii Ascherson in Neumayer.

1875. Anl. Wiss. Beobeibet Reiser. p. 368 (Named

for George Engelmann [1809–1885], an American

physician and German botanist who settled in St.

Louis). Common name: Stargrass (Figure A.1.3).

Rhizomes indeterminate, narrow, with internodes 2–4 cm long. Determinate, erect shoots 20–40

mm tall at each node. Leaves with petioles, 2 scales

at base and another pair half-way up shoot, petioles to 2 mm long, bearing 2–4 pairs of leaves in

pseudo-whorls. Blades oblong to linear-oblong,

10–30 mm long, 3–6 mm wide, tips pointed and

margins serrate. Plants dioecious. Male flowers

with 3 imbricate tepals and 3 stamens. Female

flowers with one sessile, inferior ovary, ovaries 3–4

mm long, ovoid, with 1–3 styles; styles to 30 mm

long. Fruit globose, fleshy capsule, 3–4 mm in

diam. with several minute subspherical seeds.

Halophila johnsonii Eiseman. 1980. In Eiseman

and McMillan, Aquatic Bot. 9: 16 (Named for J.

Seward Johnson, cofounder of Harbor Branch

Oceanographic Institute, where N. Eiseman studied). Common Name: Johnson’s Seagrass. (Figure

A.1.2).

Rhizomes thin, 1 mm in diam., nodes bearing

scale leaves on upper and lower surfaces, internodes 1–2 cm long. Petioles 10–20 mm long, bearing pair of linear to spatulate leaves; leaves slightly

asymmetrical, 5–25 mm long, 1–4 mm wide, margins entire and surfaces glabrous. Plants dioecious

(?), only female plants found. Female flowers sessile, with three styles 4–6 mm long; seeds

unknown. occurring only in Atlantic Florida

(Sebastian Inlet to Biscayne Bay) and classified as a

threatened species. May not be distinct from

H. ovalis (R. Brown) J.D. Hooker f., 1858.

CYMODOCEACEAE N. Taylor

Family of 5 morphologically distinct genera and

about 20 species, all seagrasses. Plants perennial

herbs. Rhizomes creeping, leafy, herbaceous or distinctly woody, scale-bearing. Leaves distichous

with distinct sheath and blade; blades linear, flat or

terete, leaves with numerous tannin cells. Plants

dioecious. Flowers solitary or in cymose inflorescences and usually terminal. Male flowers subsessile or stalked with 2 anthers producing filiform

pollen. Female flowers with 2 free carpels, each

with 1 style. Fruits 1-seeded nuts, indehiscent.

HALODULE Endlicher. 1841. Gen. 1368 (L. halo:

salt + dule: loving).

Rhizomes monopodially branched; each node

producing one to several unbranched roots and

erect short shoot with 1–4 leaves. Leaves narrow,

linear, with 3 longitudinal veins and short basal

sheaths. Plants are dioecious. Flowers solitary, terminal, enclosed by leaf. Male flowers stalked, with

2 anthers. Female flowers with 2 free carpels, with

long undivided styles. Fruits oval to slightly flattened, with a stony pericarp.

Halodule wrightii Ascherson. 1868. Bot. Zeitung

(Berlin) 26: 511 (Named for Charles Wright

[1811–1885], who collected extensively in the

southwest U.S. and Cuba and sent specimens to

Ascherson). Common names: Shoalgrass, Cuban

Shoalgrass, or Shoalweed (Figure A.2.3).

Rhizomes slender, bearing 2–4 roots and a

short shoot at each node; internodes 0.75–3.5 cm

long, with elliptic scale leaves 5–10 mm long. Leaf

sheaths 1.5–6.0 cm long; blades 5–12 (–20) cm

long, 0.50–1.5 (–2.0) mm wide, with 3 parallel

veins; blade tips with 2–3 short horn-like points.

Plants are dioecious. Flowers lack tepals; male

flowers on peduncles 1.5–2.0 cm long, anthers

slender, 3.5–5.0 mm long. Female flowers with one
52 GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

oval to elliptic ovary, 1.5–2 mm wide, with 10–28

mm long style. Fruits about 1.5–2.0 mm in diam.

with stony pericarp and stylar beak.

SYRINGODIUM Kützing. 1860. In Hohenacker,

Alg. Marin. Sicc. 9: 426 (Gr. Syringx: a tube +

odium: similar to; thus like Syringa, a genus of

flowering plants).

Rhizomes indeterminate, with monopodial

branching and scale leaves; each node bearing 1–3

branching roots and erect, unbranched, determinate short shoot. Short shoots producing 2–3 cylindrical leaves with basal sheaths. Plants dioecious.

Flowers in erect cymose infloresences; male flowers

stalked, with 2 anthers. Female flowers with 2 free

carpels, each with short style. Fruits with stony,

smooth pericarp.

Syringodium filiforme Kützing. 1860. In

Hohenacker, Alg. Marin. Sicc. 9: 426 (L. filiforme:

long and thin). Common name: Manateegrass

(Figure A.2.2).

Rhizomes 2–4 mm in diameter, nodes bearing

2–4 roots with abundant root hairs, scale leaves,

and erect short shoot. Short shoots with 2–3

leaves; leaves to 30 cm long, 1–2 mm in diam.,

cylindrical, with extensive basal sheathing, cylindrical. Plants dioecious. Male flowers on short

peduncles, with 2 anthers. Female flowers sessile,

with leaf sheath and 2 free carpels. Fruits oval to

oblique, to 6 mm long, 3–5 mm in diam.

RUPPIACEAE Hutchins

Family of one genus and 2–3 species, all seagrasses,

closely related to the Potomogetonaceae. Plants

glabrous, submerged perennial herbs found in

alkaline lakes and brackish to oceanic waters. Rhizomes creeping and little differentiated from erect

axes, with monopodial branching. Leaves opposite

or alternate, narrowly linear, with single medial

vein and short basal sheath. Plants monoecious.

Flowers perfect, on spikes subtended by 2 subopposite involucral leaves with slightly inflated

sheaths; two flowers per spike, each with 2 stamens

and few to 4 carpels with short, stout to finely

attenuate styles and peltate stigmas. Fruits indehiscent achenes, the outer layers soft, decaying and

inner pericarp stony.

RUPPIA Linnaeus. 1753. Species Plantarum, p.

127 (Named for Heinrich Reinhard Ruppius

[1688–1719], a German botanist who wrote the

Flora of Jena).

Genus with characters of the family.

Ruppia maritima Linnaeus. 1753. Species Plantarum, p. 127 (L. maritima: of the ocean, marine).

Common names: Ditchgrass; Widgeongrass (Figure A.2.1).

Rhizomes thin, 1–3 mm in diameter, profusely

branched, with growth forms; often geniculate.

Leaves produced at nodes, sheath 6–10 mm long.

Reproductive axes on erect stalks 60–90 cm long.

Leaves to 1 dm long, 0.3 mm wide, 1 veined, with

pointed tips and appearing to arise directly from

rhizome. Plants monoecious. Flowers on short (>6

cm) peduncles elongating after anthesis, becoming

loosely coiled spiral. Flowers with 2 sessile anthers

2-celled, elliptical, 0.6–0.7 mm long; carpels oval,

to 2 mm long, style short and stout to attenuate.

Fruitlets (nutlets) 2–3 mm long, on short stipe,

often curved.
APPENDIX | TAXONOMY OF FLORIDA SEAGRASSES 53

Figure A.1 HYDROCHARITACEAE: 1 Halophila decipiens, Paddlegrass (× 2⁄3); 2 H. johnsonii, Johnson’s Seagrass (× 3⁄4); 3 H. engelmannii, Stargrass (× 3⁄4); 4 Thalassia testudinum, Turtlegrass (× 1⁄ 3). Scale approximate. Smaller phenotypes are common.

1

3

4

2
Figure A.2 RUPPIACEAE: 1 Ruppia maritima, Widgeongrass (× 1⁄4). CYMODOCEACEAE: 2 Syringodium filiforme; Manateegrass

(× 1⁄4); 3 Halodule wrightii, Shoalgrass (× 1⁄ 3). Scale approximate. Smaller phenotypes are common.

54 GULF COAST FLORIDA SEAGRASS: STATUS AND ECOLOGY

1

2

3
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