EPA/600/R-09/050 

June 2009

Seagrasses and Protective Criteria: A Review and Assessment of 

Research Status

 

Edited by: 

Walter G. Nelson 

Western Ecology Division 

National Health and Environmental Effects Research Laboratory 

U.S. Environmental Protection Agency 

Newport OR 97365 

Contributors:

Cheryl A. Brown, Bruce L. Boese, Theodore H. DeWitt, Peter M. Eldridge (deceased), Mark G. 

Johnson, James E. Kaldy III, Walter G. Nelson, David R. Young, Robert J. Ozretich, David T. 

Specht 

Western Ecology Division, National Health and Environmental Effects Research Laboratory, 

U.S. Environmental Protection Agency, Newport OR 97365 
ii 

Disclaimer

The information in this document has been funded wholly or in part by the U.S. Environmental 

Protection Agency. It has been subjected to review by the National Health and Environmental 

Effects Research Laboratory and approved for publication. Approval does not signify that the 

contents reflect the views of the Agency, nor does mention of trade names or commercial 

products constitute endorsement or recommendation for use. 

Acknowledgements 

The authors thank the reviewers of this document for their many perceptive and helpful 

comments. Dr. Timothy Nelson provided review comments for all chapters of the report. Dr. 

Robert Virnstein and Ms. Lori Morris kindly provided reviews of multiple chapters, while Dr. 

Evamarie Koch provided a review for the chapter on hydrodynamic impacts on seagrasses. Dr. 

Jim Power provided assistance with the effort to develop an Access database of bibliographic 

entries for seagrass literature. Expert secretarial assistance was provided by Ms. Karen Ebert and 

Ms. Jimmie Cheney. 

This report is dedicated to the memory of Dr. Peter M. Eldridge, outstanding seagrass scientist, 

U.S. EPA colleague, and friend to many involved in this effort. 

Citation 

The appropriate citation for this report is: 

Nelson, Walter G. (ed.), 2009. Seagrasses and Protective Criteria: A Review and Assessment of 

Research Status. Office of Research and Development, National Health and Environmental 

Effects Research Laboratory, EPA/600/R-09/050. 
iii 

Abstract 

This report is a review and summary of the current status of scientific information 

relevant to the establishment of protective criteria for the most widely distributed seagrass 

species of the United States, eelgrass Zostera marina, and turtlegrass Thalassia testudinum. The 

report focuses on scientific information related to major limiting factors for seagrass survival, 

and assesses the degree to which environmental factors may need to be included in the 

development of adequately protective criteria. The review confirmed that there is a great deal of 

scientific information currently available concerning the responses of Zostera marina and 

Thalassia testudinum to a wide range of environmental factors. However, interactive effects 

among factors influencing seagrass survival remain relatively poorly known, especially across 

broader regional scales. This appears true even for such fundamental environmental 

characteristics as salinity and temperature and their interactions in the expression of nutrient or 

sediment impacts on SAV, although research is beginning to fill this gap. The question remains 

as to whether current modeling approaches, whether empirical or mechanistic, are adequate to 

predict the response of seagrasses to even single stressors. A key concern is that there is a high 

level of uncertainty in being able to predict the trophic pathway for expression of nutrient 

impacts on seagrasses. Thus, water quality criteria based on nutrient concentrations may not be 

adequately protective of seagrass resources. Alternate standards based on water clarity or water 

column chlorophyll a criteria may not be adequately protective if the principle expression of 

nutrient impacts occur through the epiphyte or macroalgal pathways. There are also important 

influences on seagrass survival through sediment associated mechanisms that may not be 

adequately captured by water quality criteria alone. There may be advantages to looking for 

integrative, plant based seagrass condition indicators, such as sucrose content, that relate to the 

ability of seagrasses to survive within a temporally varying environment. Such measures may be 

an appropriate method to integrate water column and sediment impacts into single protective 

criteria. 
iv 
v 

Table of Contents

Chapter 1 

Conceptual Framework for a Review of Research Needs for Development of National 

Water Quality Criteria Protective of Seagrasses ...................................................................1.1 

Walter G. Nelson

Chapter 2 

The Role of Light and Sucrose as a Limitation to Zostera marina Growth 

and Distribution ........................................................................................................................2.1 

Robert J. Ozretich

Chapter 3 

Water Column and Sediment Nutrients as Limits to Growth of Zostera marina and 

Thalassia testudinum ..................................................................................................................3.1 

James E. Kaldy III

Chapter 4 

Salinity as a Limiting Factor for the Seagrasses Zostera marina and Thalassia 

testudinum ..................................................................................................................................4.1

Bruce L. Boese

Chapter 5 

The Effects of Hydrodynamic Factors on Seagrasses ............................................................5.1 

Cheryl A. Brown

Chapter 6

Interactions of Zostera marina and Thalassia testudinum with Sediments ..........................6.1 

Peter M. Eldridge, Mark G. Johnson, and David R. Young 

Chapter 7

The Interaction of Epiphytes with Seagrasses under Nutrient Enrichment .......................7.1 

Walter G. Nelson

Chapter 8

Macroalgal Interactions with the Seagrasses Zostera spp. and Thalassia testudinum..........8.1 

David R. Young

Chapter 9 

The Effects of Temperature and Desiccation on the Seagrasses Zostera marina and 

Thalassia testudinum .................................................................................................................9.1 

David T. Specht and Bruce L. Boese
vi 

Chapter 10

The Effects of Bioturbation and Bioirrigation on Seagrasses ............................................10.1 

Theodore H. DeWitt

Chapter 11

Research Gaps in Relation to Setting Protective Criteria ...................................................11.1 

Walter G. Nelson 
vii 

List of Figures 

 Figure 1.1. Example of pathway for establishing and implementing seagrass protection and 

restoration targets. ............................................................................................................1.6 

Figure 1.2. Diagram of the critical path from Goal Statement to Implementation by the 

States/Tribes of numeric standards for nutrients. ............................................................1.7 

Figure 1.3. Diagram of the presumed relationship of effectiveness of a nutrient criterion in 

protecting resources at the site scale versus the scale at which the criterion is set. .........1.8 

Figure 2.1. Photosynthesis vs. Irradiance curve where alpha is the initial slope of the curve. ...2.3

Figure 2.2. An idealized change in incident light intensity throughout a summer and winter day 

at 45 degrees N. ................................................................................................................2.4 

Figure 2.3. Gross maximum productivity, gross Pmax (■), Ik values (▲), and dark respiration 

rates. ...............................................................................................................................2.13 

Figure 2.4. Days to plant death in dark assuming the sole respiratory carbon source is leaf. ..2.13 

Figure 2.5. Photon flux necessary to replace carbon lost to respiration as a function of 

temperature (assuming equal daylight and nocturnal respiration rates). .......................2.18 

Figure 3.1. Schematic diagram of the effects of increased nutrient loading. ..............................3.8 

Figure 5.1. Relationship between configuration of seagrass beds and shoot morphology and 

current regime. . ................................................................................................................5.8 

Figure 5.2. Growth conditions of seagrass as a function of current speed. ...............................5.10 

Figure 6.1. Both the water column and sediment environments influence seagrasses. ..............6.3 

Figure 7.1. Schematic diagram of seagrass plant illustrating typical patterns of distribution of 

epiphytes within and among blades. ................................................................................7.3 

Figure 7.2. Horizontal arrows indicate external factors influencing abundance of seagrass 

epiphytes. .........................................................................................................................7.4 

Figure 8.1. Effects of experimental benthic green macroalgal (Ulva spp.) manipulations on shoot 

density of Zostera marina. ...............................................................................................8.2

Figure 10.1. Illustration of some of the adverse effects of bioturbation on seagrasses. ...........10.3
viii 

Figure 10.2. Effects of bioturbating burrowing shrimp (Callianassa spp.) on turtle grass 

(Thalassia testudinum) on subtidal sand flats on St. Croix island. ................................10.5

Figure 10.3. Results of two laboratory experiments showing decreased survivorship of 

seagrass. .......................................................................................................................10.13
ix 

List of Tables 

Table 1.1. Sections of the Clean Water Act which pertain to the protection of seagrasses. ..... 1.4 

Table 1.2. List of factors potentially limiting seagrass growth and survival, and related chapter 

in this review. .................................................................................................................1.10 

Table 2.1. Observations of the death of Zostera marina plants in the field and laboratory. NR = 

not reported. .....................................................................................................................2.8 

Table 2.2. P vs. I experimental results for Zostera marina L. from various locations and 

measurement conditions. ..................................................................................................2.9 

Table 2.3. Sucrose content of Zostera marina L. tissues. NR = not reported. ..........................2.10 

Table 2.4. Conversion factors from various sources. ................................................................2.12 

Table 2.5. Statistics of variables used for calculations. ............................................................2.12 

Table 3.1. Summary of nitrogen uptake rates for Z. marina. Vmax in units of μM gdw lf-1 h-1 is 

the maximum uptake rate, Ks is the half saturation concentration (μM). .......................3.2 

Table 3.2. C:N:P ratios for Z. marina from the literature. ..........................................................3.3 

Table 3.3. Selected list of literature examining the effect of eutrophication on Z. marina

communities. ....................................................................................................................3.7 

Table 3.4. Water column and sediment nutrient concentrations (FM) associated with selected 

studies of Zostera habitat. ..............................................................................................3.10 

Table 4.1. Reported salinity tolerance ranges for eelgrass (Zostera marina) populations. ........4.2 

Table 4.2. Reported salinity tolerance ranges for turtlegrass (Thalassia testudinum) 

populations. ......................................................................................................................4.2 

Table 6.1. Distribution of Zostera species in relation to sediment texture or grain size. ...........6.4 

Table 6.2. Literature values for dissolved sulfide toxicity to marine/estuarine plants. ..............6.6 

Table 7.1. Studies of light attenuation by epiphytes in coastal marine systems, revised and 

expanded from Table 1, Brush and Nixon (2002). ..........................................................7.6 
x 

Table 7.2. Summary of laboratory microcosm and mesocosm nitrogen and phosphorus 

enrichment assays using seagrasses. ..............................................................................7.12 

Table 8.1. Summary of sources of significant bioturbation impacts for different seagrasses ....8.2 

Table 10.1. Summary of sources of significant bioturbation impacts for different seagrasses 10.3 
1.1

1.0 Conceptual Framework for a Review of Research Needs for 

Development of National Water Quality Criteria Protective of 

Seagrasses

Walter G. Nelson

1.1 Background of EPA Submerged Aquatic Vegetation Research Program

 U.S. coastal waters provide tremendous natural, economic, and public health benefits. 

Excess nutrients, together with excess carbon, cause cultural eutrophication. These nutrients, 

primarily nitrogen in marine systems, tend to stimulate excessive algal growth, which then 

causes a series of events leading to negative ecological effects such as loss of submerged aquatic 

vegetation (SAV), degradation of benthos, and areas of hypoxia/anoxia. Some of these changes 

affect use of the Nation’s aquatic resources, and pose risks to human health and the environment 

(National Research Council 2000; U.S. EPA 2002). It has been estimated that 40% of U. S. 

estuarine surface area shows poor water quality conditions due to eutrophication (Bricker et al. 

1999). Evidence suggests that nutrient inputs from point and nonpoint sources will continue to 

increase for the foreseeable future (Vitousek et al. 1997); therefore, it is important for the EPA to 

conduct research to diagnose, understand, quantify, and predict the risks that excessive nutrients 

pose to the coastal marine environment. While nutrients can be a stressor resulting from human 

activities, they can also have beneficial effects. Therefore, understanding the responses of 

estuarine and coastal water bodies to elevated nutrient loadings, and developing the ability to 

distinguish between positive and negative effects, is a priority of for EPA coastal research. 

 Within EPA, the National Health and Environmental Effects Research Laboratory 

(NHEERL) has undertaken an extensive research program to improve the scientific basis for 

setting ecologically based water quality standards. The research program is described in the 

NHEERL Aquatic Stressors Framework (U.S. EPA 2002). This document defined loss of SAV 

habitat as a major assessment endpoint for nutrient effects research. Seagrasses are one important 

component of the broad category of SAV within estuarine and some near coastal waters. 

Seagrasses typically inhabit mesohaline and polyhaline portions of estuaries, are critical to 

maintaining estuarine “health” and ecosystem function, and are widely regarded as a cornerstone 

of estuarine productivity. As such there is an increased awareness of their importance worldwide 

(Short & Wyllie-Echeverria 1996; Short & Neckles 1999). 

Anthropogenic nutrient loading stress to seagrasses can be manifested in a variety of 

ways including a primary “toxicity type” stress (i.e., high nutrient loading kills plants directly), a 

secondary stress (i.e., light limitation from phytoplankton blooms) or a combination of primary 

and secondary stresses. Increased nutrient loading can result in an accumulation of epiphytic, 

macroalgal and phytoplankton biomass that shades seagrass and results in loss of areal coverage. 

Losses may result from direct impacts such as physical removal from the action of dense algal 

mats, or via indirect effects that result from sulfide toxicity and light reduction. Because 

estuaries are complex ecosystems, it is important to understand both the stressors (e.g., light, 
1.2

temperature, plankton blooms, toxicity, algal mats, etc.) and end-point response variables (i.e. 

seagrass growth, biomass, C:N, carbohydrate reserves, etc.). The consequences of seagrass loss 

may be indirect but still have dramatic and far reaching consequences to the estuarine ecosystem 

(e.g. shifts in food web structure). 

 One aspect of the NHEERL nutrient effects research on SAV has focused on developing 

nutrient load-response relationships for seagrass habitats. This research effort is designed to 

provide a scientific basis for developing nutrient criteria that would help protect seagrass habitats 

from degradation or loss, and to provide management tools that will aid in SAV restoration 

efforts. A principal tool is the development of seagrass stress-response models focused on two 

dominant seagrass species of the U.S., eelgrass Zostera marina and turtlegrass Thalassia 

testudinum. These seagrass stress-response models are designed to deal with multiple stressors, 

such that the effects of nutrients, increased sediments in the water column, and effects on 

seagrasses due to alteration of the sediment biogeochemistry can be evaluated. 

 In the development of the SAV stress-response models, the ultimate goal is to develop 

and couple the stress-response models, which are plant scale and physiologically based, to 

population scale models. The Western Ecology Division (WED) has developed a seagrass 

stress-response model for Z. marina (Kaldy and Eldridge 2006), that has been parameterized 

with data from Yaquina Bay, Oregon, as representative of conditions in the Pacific Northwest. 

In parallel with this effort, Gulf Ecology Division (GED), Gulf Breeze, Florida, has validated a 

similar stress-response model previously developed for Thalassia testudinum, a common 

seagrass of the Gulf of Mexico region. Ultimately, the aim is that this suite of stress-response 

models can be applied to estimate seagrass response to nutrient stress over the range of 

distribution of the target seagrass species. 

 The specific objective of the eelgrass stress-response model development is to provide a 

tool for analysis of future scenarios that would indicate what changes might be expected to occur 

in response to nutrient enhancement in estuaries. The initial eelgrass stress-response model is 

one which predicts response at the individual seagrass plant level, and thus is applicable at the 

patch scale. Model components include seagrass responses to physical influences such as 

reduced light due to anthropogenic nutrient loading, and to anthropogenic influences on sediment 

geochemical cycling, in order to accurately predict seagrass nutrient responses. 

 While it would be desirable to link the patch scale models to population scale models, it 

is possible to use the patch scale model as a stand alone tool to assist in development of nutrient 

criteria protective of seagrasses (Brown et al. 2007). By collecting data on physical parameters 

(light availability, sediment conditions, water temperature) at stations located across the span of 

the estuarine system of interest, the model can be used to estimate seagrass plant responses to 

increased loads of nutrients and/or sediments at these measurement points. Under the 

assumption that stations are at least locally representative, model predictions at these multiple 

locations can be used to examine whether proposed nutrient standards are adequately protective 

across the entire estuarine system, or whether they may need to be varied across an estuarine 

gradient. 
1.3

1.2 Regulatory Background

 The U.S. Environmental Protection Agency is concerned with the protection of 

seagrasses under two sections of the Clean Water Act (33 U.S.C. 1252 et seq.), Section 304(a) 

and Section 404(c). Under section 304(a) (1) the agency is charged with development of water 

quality criteria reflecting the latest scientific knowledge on effects of pollutants on aquatic biota, 

including “plant life”. The agency is further charged (Section 304(a) (2)) with providing timely 

scientific information on factors necessary to maintain the chemical, physical and biological 

integrity of the nations waters. Under section 404 which regulates dredging and dredged 

material disposal under the lead of the Army Corps of Engineers, EPA is authorized to deny 

issuance of dredged material disposal permits where such activity will have “an unacceptable 

adverse effect on municipal water supplies, shellfish beds and fishery areas (including spawning 

and breeding areas), wildlife, or recreational areas.” While seagrasses are not specifically 

named, it is clear that the protection of seagrass habitat is encompassed within both sections of 

the Clean Water Act. 

 In response to the charge to develop water quality criteria for all of the nation’s waters, 

EPA (2001) has published the Nutrient Criteria Technical Guidance Manual: Estuarine and 

Coastal Waters. The document provides an extensive overview of the issues of nutrient over 

enrichment in marine waters. It provides guidance on steps for development of nutrient criteria 

(see section 1.4 below), and provides suggestions for key variables and measurement methods 

for assessing eutrophic condition. Finally it describes management approaches by which nutrient 

criteria can be used to protect water quality. Ultimately, it is the states and tribes that adopt 

water quality criteria and standards, and they would be responsible for adopting any criteria 

specifically protective of seagrasses. 

1.3 Pathways to Water Quality Criteria Setting

 Kenworthy (1992) provides a detailed history of the early evolution of water quality 

criteria relevant to submerged aquatic vegetation, tracing the original guidelines for developing 

state standards back to the Federal Water Quality Act of 1965. His review similarly addressed 

the history and development of the State of Florida water transparency and turbidity standards. 

His assessment at that time was that both the federal guidance and Florida state standard were 

inadequate and unlikely to sufficiently protect seagrass. 

 The Kenworthy review of the history and process of setting one state water quality 

standard, the Florida transparency standard, suggested that in the past, the process of setting 

water quality criteria might propagate uncertainties contained in guidance documents in such a 

way as to lead to flawed criteria. As a specific example, the state standard at the time included a 

numerical definition of a light compensation point derived from phytoplankton which was 

inappropriate for the survival of seagrasses. 
1.4

Table 1.1. Sections of the Clean Water Act which pertain to the protection of seagrasses. 

SEC. 304. (a)(1) The Administrator, after consultation with appropriate Federal and State agencies and other 

interested persons, shall develop and publish, within one year after the date of enactment of 

this title (and from time to time thereafter revise) criteria for water quality accurately 

reflecting the latest scientific knowledge (A) on the kind and extent of all identifiable effects 

on health and welfare including, but not limited to, plankton, fish, shellfish, wildlife, plant 

life, shorelines, beaches, esthetics, and recreation which may be expected from the presence 

of pollutants in any body of water, including ground water; (B) on the concentration and 

dispersal of pollutants, or their byproducts, through biological, physical, and chemical 

processes; and (C) on the effects of pollutants on biological community diversity, 

productivity, and stability, including information on the factors affecting rates of 

eutrophication and rates of organic and inorganic sedimentation for varying types of 

receiving waters.

SEC. 304. (a)(2) The Administrator, after consultation with appropriate Federal and State agencies and other 

interested persons, shall develop and publish, within one year after the date of enactment of 

this title (and from time to time thereafter revise) information (A) on the factors necessary to 

restore and maintain the chemical, physical, and biological integrity of all navigable waters, 

ground waters, waters of the contiguous zone, and the oceans; (B) on the factors necessary for 

the protection and propagation of shellfish, fish, and wildlife for classes and categories of 

receiving waters and to allow recreational activities in and on the water; and (C) on the 

measurement and classification of water quality; and (D) for the purpose of section 303, on 

and the identification of pollutants suitable for maximum daily load measurement correlated 

with the achievement of water quality objectives.

SEC. 404. (c) The Administrator is authorized to prohibit the specification (including the withdrawal of 

specification) of any defined area as a disposal site, and he is authorized to deny or restrict 

the use of any defined area for specification (including the withdrawal of specification) as a 

disposal site, whenever he determines, after notice and opportunity for public hearings, that 

the discharge of such materials into such area will have an unacceptable adverse effect on 

municipal water supplies, shellfish beds and fishery areas (including spawning and breeding 

areas), wildlife, or recreational areas. Before making such determination, the Administrator 

shall consult with the Secretary. The Administrator shall set forth in writing and make public 

his findings and his reasons for making any determination under this subsection.

 The development of water quality criteria is ultimately based on the concept of protection 

of a designated use for a particular body of water. For the example of Florida, there are five 

designated use categories for surface waters of which Class III - Recreation, Propagation and 

Maintenance of a Healthy, Well Balanced Populations of Fish and Wildlife is the Designated 

Use that applies to the protection of seagrasses (Kenworthy 1992). The language of the Florida 

Class III Designated Use is derived directly from the language of the Clean Water Act (e.g. Sec. 

304. (a)(2), Table 1.1). While encompassing seagrasses, in does not specifically identify 

protection and propagation of seagrasses as a Designated Use. 

 The Seagrass Conservation Plan for Texas (1999) indicates that in order to achieve the 

stated objective to “Ensure water and sediment quality beneficial to the seagrass community” a 

series of steps were involved. A first and important strategy in the Texas plan was to propose 

that “Seagrass Habitat” be added to the list of Designated Uses for the state of Texas. With this 
1.5

basis, water quality criteria specifically protective of seagrasses could be developed and added to 

the Texas Surface Water Quality Standards. Finally, the Texas Plan proposed the 

implementation of water-based Best Management Practices. 

 A second example of creating a specific designated use to address seagrass protection 

comes from Chesapeake Bay (EPA 2003). In order to reach the ambitious goal of correcting 

nutrient and sediment problems within the Bay and its tributaries by 2010, regional management 

organizations recognized the need to create new designated uses that would offer better 

protection of biotic resources and would be a better representation of desired water quality 

objectives. A very precisely defined “shallow-water bay grass designated use” was one such 

refined designated use proposed. The designated use applies to tidal waters, from the intertidal 

zone to segment specific depth zones from 0.5 to 2.0 m. Another interesting aspect is that the 

designated use includes a seasonal use component, such that it applies only during the bay grass 

growing season which varied among salinity zones within the Bay. 

 Development of water quality criteria protective of seagrasses may be principally driven 

by the desire to restore a severely degraded resource. One example of a critical path (Figure 1.1) 

for seagrass protection derived from seagrass restoration goals is provided by the plan developed 

by the Tampa Bay National Estuary Program (Johansson and Greening 2000). Within the 

framework of the general designated use, targets for extent of seagrass restoration are first set. 

Setting of quantitative seagrass restoration targets may be based on a variety of lines of evidence. 

Seagrass conservation targets developed for the Indian River Lagoon of Florida (Virnstein and 

Morris 2000, Steward et al. 2005) involves three approaches, development of a potential target 

based on distribution in the best available habitat (a reference condition approach), the use of a 

historically based distribution target, and a “critical minimum” approach which establishes the 

lowest threshold below which seagrass extent should not fall. 

 The site specific strategies developed by the Tampa Bay NEP (Johansson and Greening 

2000), Texas Parks and Wildlife (1999), etc. can be generalized (Figure 1.2) to a critical process 

path (e.g. U.S. EPA 2001; Batelle 2008). Within the U.S. EPA Regional Offices, there are 

Nutrient Coordinators who establish Regional Technical Assistance Groups to assure that the 

best available current information is brought to the criterion development process, and 

inappropriate guidance is weeded out. As suggested by Figure 1.2, a critical early decision is the 

consideration of scale at which to set a water quality criterion. Considerations of estuarine 

system classification and the degree to which ecoregional classifications can be applied are 

required in order to reach beyond a water-body by water-body approach to criteria setting. 

 The challenge for U.S. EPA in developing guidance for water quality criteria protective 

of the seagrass resource on a national basis is presaged by Kenworthy (1992) in his consideration 

of development of a general state standard for Florida. He suggests that because of the broad 

variation in physical systems, and in the sources and causes of water quality impairments, “a 

general standard will fail to adequately protect seagrasses.” Instead, “site or region specific 

standards will be more effective”. The quandary of this trade off between site scale effectiveness 

and spatial scale at which a criterion is set is summarized in Figure 1.3. Variation in physical 
1.6

systems becomes greatly magnified as one moves from state to national scales. The highly 

specific designated use boundary delineation defined for Chesapeake Bay (EPA 2003) is one 

example of the application of site-specific standards across an important regional resource. 

However, at the national scale, factors limiting the ability to achieve broadly protective 

water quality criteria for seagrasses are the economic limitations and constraints imposed by the 

necessity to develop water-body specific criteria. In contrast to states such as Florida, or those 

bordering the Chesapeake Bay, many states have little or no technical capability for monitoring 

seagrass systems, and have far more limited resources for developing and implementing SAV 

criteria. The present report seeks to begin a process of considering how best to develop water 

quality criteria protective of seagrass by first evaluating current knowledge on seagrass ecology 

in the context of developing protective standards for seagrasses. 

Figure 1.1. Example of pathway for establishing and implementing seagrass protection and 

restoration targets. The approach illustrated is for the Tampa Bay National Estuary 

Program (modified from Johansson and Greening 2000). 

Adopt quantitative 

seagrass restoration 

goals

Determine light 

requirements needed 

to meet seagrass 

restoration goals

Determine water 

quality necessary to 

maintain light 

requirements

Determine chl a 

concentration targets 

necessary to maintain 

water clarity 

requirements

Determine nutrient 

loads necessary to 

maintain chl a targets

Define and implement 

nutrient management 

strategies to achieve 

load reduction goals

Measure progress 

toward targets and 

goals. Reassess as 

needed. 
1.7

Figure 1.2. Diagram of the critical path from Goal Statement to Implementation by the States/Tribes of numeric standards for 

nutrients. 
1.8

Figure 1.3. Diagram of the presumed relationship of effectiveness of a nutrient criterion in protecting resources at the site scale versus 

the scale at which the criterion is set.
1.9

1.4 Goals and Limitations: Reviewing the Environmental Requirements of Major U.S. 

Seagrass Species 

 During the 1980's, a series of summary documents on seagrass communities of the U.S. 

were produced by the U.S, Fish and Wildlife Service (Phillips 1984; Thayer et al. 1984; Zieman 

1982; Zieman and Zieman 1989). These broad reviews encompassed aspects of seagrass ecology 

ranging from environmental tolerances to descriptions of the communities of plants and animals 

associated with seagrass habitats. Since that time there has been a tremendous increase in the 

research literature on seagrasses. In preparing for the present review, Western Ecology Division 

scientists established a Microsoft Access bibliographic database of greater than 10,000 entries 

related to seagrass biology. 

 The aim of the present report is not to duplicate the scope of earlier summaries, but 

instead to focus on providing an updated review of scientific knowledge of the most broadly 

distributed U.S. seagrass species that will be relevant to the process of developing protective 

criteria of major estuarine SAV resources of the continental U.S. As originally conceived, the 

review effort was to cover the three most widely distributed species of seagrasses in continental 

U.S. waters: Z. marina (eelgrass), T. testudinum (turtle grass), and H. wrightii (shoalgrass). The 

effort was to include authors from other EPA divisions who could provide regional knowledge 

on seagrass ecology, together with specialized knowledge on aspects of seagrass ecology. 

Unfortunately, this concept was not realized as divisional research priorities shifted and scientists 

were unable to fully participate in the project. Thus, individual chapter authors determined 

which seagrass species were included within the scope of coverage for a given topic area. 

 The principle focus of this review evolved to become the environmental requirements of 

Z. marina, while T. testudinum received more limited coverage depending on topic area, and 

other species are dealt with to a much lesser degree. A second limitation to the report is that 

publication has been repeatedly delayed, again due to shifting priorities, and therefore the most 

recent seagrass literature may not be included. Despite this, the report synthesizes a great deal of 

information that it highly relevant to the management and protection of seagrasses as a critical 

aquatic resource. 

The principle goal of the review is to highlight any critical uncertainties that must be 

addressed by research in order to be able to develop protective criteria. Thus, our review 

examines what is known concerning the species-specific requirements for the range of principal 

factors potentially limiting seagrass growth and survival within the limits described above (Table 

1.2). The review is structured in chapters which describe individual limiting factors. Each 

chapter provides the background for a particular factor or factors related to seagrass ecology, a 

review of relevant research, and an evaluation of whether or not there are significant research 

gaps in relation to providing guidance to the setting of effective protective criteria. 

 The background section describes the mechanisms by which the limiting factor (e.g. 

epiphytic load) may potentially influence the growth, survival, reproduction and distribution of 

seagrasses. The review of research has a narrow focus on work that is directly relevant to the 
1.10

limiting factor and seagrass condition. For example, while a comprehensive evaluation of what 

is known about taxonomic composition of epiphytic cover may be of significant academic 

interest, it would not be a focus of the review unless it provides some essential insight into 

seagrass survival. The focus is instead on whether there is adequate evidence to indicate whether 

for example, epiphytes can limit seagrass growth, and under what conditions and by what 

mechanisms. The section on Research Gaps in Relation to Setting Protective Criteria assesses 

the state of the science for the limiting factor with regard to setting protective criteria for 

seagrasses. Questions addressed in this section are for example, “Should epiphyte loading be 

somehow integrated into a protective criterion?” or “Is there adequate research in place to 

suggest an approach, or even a potential target?” 

Table 1.2. List of factors potentially limiting seagrass growth and survival, and related chapter in 

this review. 

Limiting Factor Subject to Anthropogenic 

Modification

Chapter 

Light Yes 2 

Nutrients Yes 3 

Salinity Yes 4 

Current and Wave Exposure Yes 5 

Sediment Characteristics (Grain size, 

dissolved oxygen, sulfide) 

Yes 6 

Epiphytes Yes, through indirect 

mechanisms 

7 

Macroalgae Yes, through indirect 

mechanisms 

8 

Temperature (Freezing, Heat Stress, 

Desiccation) 

Yes, through both direct and 

indirect mechanisms 

9 

Bioturbation Yes, through indirect 

mechanisms 

10 

Diseases Yes, through indirect 

mechanisms 

10 

Herbivory Yes, through indirect 

mechanisms 

7 for micrograzers 
1.11

1.5 Literature Cited

Batelle. 2008. Conceptual Plan for Nutrient Criteria Development in Maine Coastal Waters. 

Report prepared for EPA Region 1, Maine Dept. of Environmental Protection, and U.S. 

EPA Oceans and Coastal Protection Division. Batelle, Brunswick ME. EPA Contract 

No. 68-C-03-041. pp. 31. 

http://www.maine.gov/dep/blwq/report/2008/nutrient_criteria_report_2008.pdf 

Bricker, S. B., C. G. Clement, D. E. Pirhalla, S. P. Orlando, and D. R. G. Farrow. 1999. National 

Estuarine Eutrophication Assessment: Effects of Nutrient Enrichment in the Nation's 

Estuaries. NOAA, National Ocean Service, Special Projects Office and the National 

Centers for Coastal Ocean Science. pp. 71. 

Brown, C.A., W.G. Nelson, B.L. Boese, T.H. DeWitt, P.M. Eldridge, J.E. Kaldy, H. Lee II, J.H. 

Power, and D.R. Young. 2007. An Approach to Developing Nutrient Criteria for Pacific 

Northwest Estuaries: A Case Study of Yaquina Estuary, Oregon. USEPA Office of 

Research and Development, National Health and Environmental Effects Laboratory, 

Western Ecology Division. EPA/600/R-07/046.

Johansson, J. O. R. and H. S. Greening. 2000. Seagrass restoration in Tampa Bay: A resourcebased approach to estuarine management. pp. 279-294. In: S. A. Bortone (ed.). 

Seagrasses: Monitoring, Ecology, Physiology, and Management. CRC Press. Boca 

Raton. pp. 318. 

Kaldy, J.E. and P.M. Eldridge (eds.). 2006. Development of mechanistic models to guide 

establishment of protective criteria for seagrasses. Report on the Zostera marina and 

Thalassia testudinum stress-response models. Office of Research and Development, 

National Health and Environmental Effects Research Laboratory, EPA Internal Report. 

Kenworthy, W. J. 1992. Do federal water quality criteria and Florida state water quality 

standards protect seagrasses? A feasibility evaluation of alternative criteria and standards 

to monitor, evaluate and regulate water transparency. Appendix II, pp. A-11 to A-29, In: 

A. Hart (ed.) Proceedings and Conclusions for the Submerged Aquatic Vegetation 

Initiative. Indian River Lagoon, National Estuary Program. Melbourne, Florida. (Also as 

pp. A-19 to A-48, In: Morris, L. J. and D. A. Tomasko (eds.), 1993. Proceedings and 

Conclusions of Workshops on: Submerged Aquatic Vegetation and Photosynthetically 

Active Radiation. Special Publication SJ93-SP13. St. Johns River Water Management 

District, Palatka, Florida. 

National Research Council. 2000. Clean coastal waters: understanding and reducing the effects 

of nutrient pollution. National Academy Press, Washington, DC. pp. 405. 

Phillips, R. C. 1984. The ecology of seagrass meadows in the Pacific Northwest: a community 

profile. US Fish and Wildlife Service. FWS/OBS-84/24. pp. 85. 

Short, F. T. and S. Wyllie-Echeverria. 1996. Natural and human-induced disturbance of 

seagrass. Environmental Conservation 23:17-27. 

Short, F. T. and H. A. Neckles. 1999. The effects of global climate change on seagrass. 

Aquatic Botany 63: 169-196. 

Steward, J., R.W. Virnstein, L. J. Morris, and E. F. Lowe. 2005. Setting seagrass depth, 

coverage, and light targets for the Indian River Lagoon system, Florida. Estuaries 28:923-

935. 
1.12

Texas Parks and Wildlife. 1999. Seagrass conservation plan for Texas. Texas Parks and 

Wildlife Department, Resource Protection Division. Austin, Texas. pp. 79. 

Thayer, G. W., W. J. Kenworthy, and M. S. Fonseca. 1984. The ecology of eelgrass meadows of 

the Atlantic coast: a community profile. U.S. Fish and Wildlife Service FWS\OBS84\02. pp. 147. 

U.S. EPA. 2001. Nutrient criteria technical guidance manual: Estuarine and coastal marine 

waters. U.S. EPA, Office of Water. EPA-822-B-01-003. 

U.S. EPA. 2002. Aquatic Stressors Framework and Implementation Plan for Effects Research. 

US EPA, Office of Research and Development, National Health and Environmental 

Effects Laboratory, pp. 179. 

U.S. EPA. 2003. Ambient water quality criteria for dissolved oxygen, water clarity and 

chlorophyll a for the Chesapeake Bay and its tidal tributaries. U.S. EPA, Office of Water 

and Office of Science and Technology. EPA 900-R-03-002. 

Vitousek P. M., J. D. Aber, R. W. Howarth, G. E. Likens, P. A. Matson, D. W. Schindler, W. H. 

Schlesinger, and D. Tilman. 1997. Human alteration of the global nitrogen cycle: sources 

and consequences. Ecological Applications 7:737-750. 

Virnstein, R. W. and L. J. Morris. 2000. Setting seagrass targets for the Indian River Lagoon, 

Florida. pp. 211-218, In S. A. Bortone, ed. Seagrasses: Monitoring, Ecology, Physiology, 

and Management. CRC Press. Boca Raton. 

Zieman, J. C. 1982. The ecology of the seagrasses of south Florida: a community profile. U.S. 

Fish and Wildlife Service FWS\OBS-84\02. pp. 158. 

Zieman, J. C. and R. T. Zieman. 1989. The ecology of the seagrass meadows of the west coast of 

Florida: a community profile. U.S. Fish and Wildlife Service Biol. Rep. 85(7.25). pp. 

155. 
2.1

2.0 The Role of Light and Sucrose as a Limitation to Zostera 

marina Growth and Distribution

Robert J. Ozretich

2.1 Background

 Light is a fundamental requirement for seagrasses. The energy derived from photons is 

used to reduce carbon dioxide and fuel the biosynthesis of carbohydrates that make up the bulk 

of these plants, amino acids and lipids. Without light consisting of a sufficient quantity of 

photons of wavelengths overlapping the absorption spectra of a seagrass’ photosynthetic 

pigments, insufficient carbon dioxide will be fixed to fulfill the plant’s respiratory needs 

resulting in the plant’s death or failure to grow or reproduce. 

 Anthropogenic activities that affect the quality or quantity of light available to seagrasses 

can be direct and indirect. Uncontrolled construction sites within an estuary’s watershed can 

lead to elevated loads of suspended sediments that increase light scattering in the water, thereby 

reducing the amount of light that can potentially reach seagrasses. Similarly, addition of 

nutrients to a watershed from excess run off of nutrients from agricultural and urban sources can 

promote blooms of phytoplankton and seagrass epiphytes that reduce both the quality and 

quantity of light available to seagrasses. 

 Determining the light requirements of seagrasses has been approached both by 

considering the needs of individual plants and from consideration of the maximum depth 

distributions of a given species. Individual plant requirements have focused on the plant’s 

response to light through the experimental determination of P vs. I (photosynthesis versus 

irradiance) curves (Figure 2.1). By comparing rates of photosynthesis with respiration, the daily 

number of hours of light saturating irradiance (Hsat, Figure 2.2) necessary to balance respiratory 

demands has been estimated. Estimates of Hsat are often coupled with monitoring data to 

evaluate the suitability of locations for sustained Zostera marina growth. The maximum depth 

of occurrence of Z. marina, when compared to the local attenuation coefficient, has also been 

used to estimate the minimum light requirement as a percentage of water-incident light. This 

approach lets the plant’s growth patterns account for the astronomically and meteorologicallyinduced variation in time and latitude of the incident light field (Figure 2.2) and the in situ

changes in water composition to which it is exposed. In general, survival at a location over time 

requires that the plant’s carbon balance is positive and that the periods of negative carbon 

balance are short enough that recovery is possible during more favorable conditions. 

 To reach the leaf surface of seagrasses, sunlight has to pass through the atmosphere, the 

air-water interface, through the water column and finally, through a film of epiphytes. 

Application of the current understanding of the spectral and intensity changes in light prior to 

reaching the leaf surface has resulted in bio-optical-physical models of varying complexities. 

Some models have progressed to the point where the spectral dynamics of light through a stand 

of seagrass has been coupled with primary production models to predict in situ growth and 
2.2

distribution. Modeling of intensively and extensively studied systems has led to proposed habitat 

requirements for submerged aquatic vegetation. 

2.2 Review of Relevant Research

2.2.1 Radiative Transfer Theory

 Understanding the spectral and intensity changes in light as it travels from the sun into 

and through natural waters has lead to several books and articles on radiative transfer theory and 

optical oceanography (Preisendorfer 1986; Jerlov 1976; Kirk 1994). This research has resulted 

in a thorough understanding of the interaction of light with dissolved and particulate material in 

water as well as the in-water consequences of meteorological and astronomical changes (Mobley 

1994). This understanding is expressed through the publication of radiative transfer equations 

(RTEs) that, when combined in the numerical radiative transfer model, Hydrolight (Sequoia 

Scientific, Inc., Redmond, WA) represents a mathematically complete rendering of these 

processes that is available to the public. Other solutions of the RTEs have been published that 

produce comparable results in certain applications (Gallegos 1994; 2001). 

 The basic optical properties of natural waters that the RTEs deal with are the absorption 

and scattering of light as it passes through a volume of water; both of these processes are 

wavelength-dependent. This is because the scattering interaction of matter and light is a function 

of matter’s dimensions compared to light’s wavelengths, and because interactions with matter in 

which energy is absorbed are a quantum phenomena they are also a function of wavelength 

(Mobely 1994). 

 The absorption and scattering coefficients of light are considered to be a result of the 

summation of the contributions from each water component. This can be expressed in the 

following way for the spectral (by wavelength) absorption coefficient, a (λ): 

 

 a (λ) = aw (λ) + aDOM (λ) + aphto (λ) + aTSS (λ) Equation 2.1

where aw, aDOM, aphto, and aTSS are the contributions to absorption by water molecules, dissolved 

organic matter (DOM), phytoplankton, and total suspended solids, respectively. Each is 

wavelength-dependent and the last three are also concentration dependent. 

 Absorption and scattering are considered inherent optical properties (IOPs) of water and 

in combination contribute to the spectral, diffusive attenuation coefficient, λ KD which is 

considered an apparent optical property (AOP) because it varies with the in-air geometry and 

intensity of the light source. 

 The consequence of absorption and scattering of light in water is expressed through the 

following, familiar expression of “Beer’s Law”: 

 )*(

0

ZK

Z D eII λ λλ − = Equation 2.2
2.3

where λ

ZI is the irradiance of wavelength, λ, Z meters below the surface, and λ0 I is the surface 

irradiance. 

-0.05

0

0.05

0.1

0.15

0.2

0 50 100 150 200 250 300 350

μmol photons m-2 sec

-1

μmol O2 mg chl a-1 min-1

Pmax

Ik I R c

alpha

Figure 2.1. Photosynthesis vs. Irradiance curve where alpha is the initial slope of the curve, Ik is 

the light intensity at the onset of light-saturated photosynthesis, Pmax; Ic, the 

compensation light intensity, is the light intensity at which dark respiration, R, is equal to 

the rate of photosynthesis. The dark respiration determined graphically from P vs. I 

measurement represents dark respiration in the light but may not be representative of a 

lower nocturnal dark respiration because of the plant’s response to the enhanced postillumination respiration phenomenon (Heichel, 1979; Falkowski and Raven, 1997). 
2.4

0

200

400

600

800

1000

1200

1400

1600

1800

0:00 3:00 6:00 9:00 12:00 15:00 18:00 21:00 0:00

μmol photons m-2 sec-1

Summer

 Hsat = 13 h 

 Hcomp = 14.5 h

Ik = 100

Ic = 25

Winter

 Hsat = 8 h

 Hcomp = 9 h

Figure 2.2. An idealized change in incident light intensity throughout a summer and winter day 

at 45 degrees N latitude where Ik and Ic are from the P vs. I curve in Figure 2.1. Hsat are 

the hours during which Ik is exceeded and Hcomp are the hours during which production 

exceeds respiration and the light intensity exceeds Ic. 

2.2.2 Application of Radiative Transfer Theory to Systems

 Historically, Secchi depth measurements were the only estimates of water column optical 

properties that were available and conversion to KD (PAR) for modern modeling applications has 

been found to be site-specific, especially for estuarine settings (Preisendorfer 1986; Giesen et al. 

1990; Batiuk et al. 1992). These conversions have relied on direct comparisons to the results of 

the more recent, broad bandwidth, PAR sensors (LiCor Environmental, Lincoln, NE; 

Biospherical Instruments, San Diego, CA). The use of photoelectric instruments in permanent 

moorings or deployments from vessels have supplanted the Secchi depth measurement as their 

outputs are directly useful in calculating PAR attenuation coefficients under a wide range of in 

situ conditions. The spectrum of light that is useful for plants is that which overlaps the 

absorption spectra of the pigments involved with photosynthesis. This is light with wavelengths 

between 400 nm and 700 nm which is considered photosynthetically available radiation (PAR). 

Measured at the deepest sites where seagrasses are found, PAR has been used to assess long-term 

light requirements as a percentage of PAR incident on the water’s surface (Duarte 1991; 
2.5

Dennison et al.1993). Multi-variant regression models using TSS (or turbidity), chlorophyll a, 

and DOM concentrations as combined predictors of KD (PAR) were found to be insufficiently 

accurate in comparison to RTE-derived estimates of KD (PAR) for the purposes of establishing 

management options for seagrasses in Chesapeake Bay (Gallegos 2001). The consequences of 

different mitigation strategies could be assessed in achieving water quality goals when the 

contributions of these water column components to light attenuation have been determined. 

 Using parameterized absorption and scattering for two different seagrass habitats 

(including Zostera marina), Zimmerman (2003) used the Hydrolight model in combination with 

his own model to investigate the light environment within the canopy of seagrass meadows as a 

function of leaf allometry, canopy structure and modeling boundary conditions. Combined 

model outputs compared favorably (±15%) with measured, spectral irradiances at various heights 

within the canopies. 

 If only the concentrations of the absorbing and scattering water components varied in 

time and space and not their spectral attributes, concentration-based models would be universally 

applicable. One such model parameterization, KD = 0.32 + 0.016 * chl a (μg l-1) + 0.094 * TSS 

(mg l-1), has been done for Chesapeake Bay (Kemp et al. 2004) where, 0.32 is the contribution to 

attenuation by water molecules. Unfortunately, the spectral characteristics of the light absorbing 

and scattering components of natural waters appear to be at least system dependent, if not 

temporally dependent. For example, terrestrially derived material (abiotic particles and DOM) 

from different locales (Gallegos and Kenworthy 1996) can have different optical characteristics. 

Species succession and decomposition over time can affect the optical characteristics of a 

phytoplankton bloom as it is reduces nutrient concentrations and is consumed. Determination of 

these characteristics of the optically important components of natural waters will therefore be 

required for each studied system (Gallegos 2001; Zimmerman 2003). 

 While solution of the RTEs computes PAR reaching the depth of a plant’s leaves, it is not 

often the PAR that actually reaches the chlorophyll-containing surface of the leaves. This is 

because of the colonization of these surfaces by a variety of organisms and the settling of biotic 

and abiotic particles (Twilley et al. 1985; Losee and Wetzel 1983; Kemp et al. 2000; see Chapter 

7). This epiphytic community, which tends to increase over time, absorbs and scatters light 

spectrally and reduces the amount of PAR reaching the leaf surface (Drake et al. 2003). This 

resulting light is the photosynthetically used radiation, or PUR. Kemp et al. (2000; 2004) 

developed algorithms to model KD (PAR) through the epiphytic layer as a function of both 

growing season nitrogen or phosphorous concentrations and TSS. These algorithms were used to 

reduce the percentage of water-incident irradiance supporting seagrass populations that had been 

estimated from field measurements and RTE-modeled estimates of KD (PAR). 

2.2.3 Instrumentation Used in Support of RTE Parameterizations 

 Validation of modeling results requires in situ measurement of the light fields. Longterm and shipboard cast deployments of broad bandwidth, PAR sensors (LiCor Environmental, 

Lincoln, NE; Biospherical Instruments, San Diego, CA) have historically contributed greatly to 

understanding the temporal and spatial variability of this critical measurement. Field-deployable 
2.6

full spectrum scanning radiometers have also been developed such as LiCor’s LI-1800 and the 

much smaller HydroRad (HOBI Labs Inc., Tucson, AZ). In recent years, instruments have been 

developed that can simultaneously measure attenuation and absorption/scattering of multiple 

narrow bandwidths. Examples of these types of instruments are the ac-9 (WetLabs, Philomath, 

OR) and HOBI Lab’s HydroScat. The ac-9 has been used extensively in estimating water 

column IOPs in situ (Zimmerman 2003; Gallegos and Jordan 2002). However, these field 

measurements require labor intensive complimentary laboratory determinations that involve 

filtering of samples to spectrally characterize the dissolved matter using standard scanning 

spectrophotometers and those with integrating sphere attachments (Drake et al. 2003) for 

particulate matter. 

2.2.4 Failure to Survive in Optically Sufficient Systems

 Although the correlations between the distribution of persistent seagrass meadows and 

seasonally-averaged KD (PAR), optically important water column constituents, and the 

application of radiative transfer theory may be sufficient to describe general conditions that 

support seagrass growth (Dennison et al. 1993; Kemp et al. 2004) short-term events can also 

affect growth and survival of seagrasses. Such circumstances are not captured in annualized 

data. Loss of Zostera marina resulting from light limitation has been observed both during field 

observations (Dennison and Alberte 1985; Moore et al.1996; 1997; Cabello-Pasini et al. 2002) 

and experimentation (Alcoverro et al. 1999; Cabello-Pasini et al. 2002; Thom et al. 2002; Biber 

et al. 2005). The field and experimental conditions, and observations of these losses of Zostera 

marina are found in Table 2.1. 

 Continuous monitoring of PAR during the time intervals over which these plants died 

provided researchers the opportunity to evaluate the light requirements of this species on a 

photon basis rather than on the statistically-based approach of percent of incident radiation 

requirement (Duarte 1991; Dennison et al. 1993). The observations of Zostera death (Table 2.1) 

all appear to share the same circumstances, over a similar, 3-5 week time frame they received 

either no light or severely diminished light exposure. For the plants in the dark (Cabello-Pasini 

et al. 2002; Biber et al.2005, Table 2.1), respiration was the dominant process. While the aim of 

these studies differed, each invoked either, the concept of Hsat (the daily period of light-saturated 

photosynthesis, Figure 2.2), or Hcomp (Dennison and Alberte 1982), the daily period of light 

above Ic, the compensation light intensity (Figure 2.1), in their attempts to interpret their 

observations. It has been hypothesized that Zostera marina needs to experience between 3 and 

10 h of Hsat per day to meet the demands of growth and respiration (Dennison and Alberte 1982; 

Marsh et al.1986; Zimmerman et al.1989). Given the seasonality, frequency and duration of 

meteorological processes (night to day, clear to overcast and fog) that control the incoming light, 

and of episodic river discharges (freshets, localized flash floods, and hurricanes) delivering silts, 

clays, and nutrients to coastal waters, the recommended wide range in the number of hours of 

Hsat or Hcomp is insufficiently proscriptive. 

2.2.5 Laboratory Studies of Zostera marina Photosynthesis

 Laboratory experiments to determine seagrass production in response to light have been 

performed with leaf segments (scraped of epiphytes), or whole plants. These tissues are 
2.7

incubated in varying PAR intensity while monitoring the change of dissolved oxygen (Figure 

2.1) or other constituents. The reported units of P in the P vs. I curve are varied (Table 2.2) and 

are often not easily converted to common units because of the lack of the necessary normalizing 

ratios, e.g., dry weight to wet weight, chlorophyll a (chl a) per g fresh weight (fw), dry weight 

(dw) or area (decimeter, dm2). In addition, the variability of the P vs. I components is usually 

not reported and is often high because of the few data points and the steep initial slope (alpha) of 

the relationship (Zimmerman et al. 1991). The effect of temperature on the P vs. I components 

have also been investigated (Drew 1979; Marsh et al. 1986; Bulthuis 1987; Zimmerman et al. 

1989) (see Chapter 9). Instruments and techniques have been developed to monitor the 

fluorescence of various pigments involved in photosynthesis. These include the photosynthetic 

efficiency analyzer and pulse amplified modulated (PAM) fluorometry which have been used to 

assess various acute stresses (Major and Dunton 2002; Bjork et al. 1999) and light-limited 

chronic stress (Biber et al. 2005). 

Section 2.2.6 The Role of Sucrose in the Biochemistry of Zostera marina

 Sucrose has been identified as the dominant carbohydrate of Zostera marina’s tissues 

(Smith 1989) and is thought to be the primary carbon reserve during periods of light limitation 

(Smith et al. 1988; Smith 1989; Kraemer and Alberte 1995; Zimmerman et al. 1995; Zimmerman 

and Alberte 1996). Sucrose is the carbon source for dark respiration during respiratory or 

fermentative processes. During periods of light, sucrose and oxygen produced in the shoots are 

translocated to the roots and rhizomes. Upon the onset of darkness, the production of these 

products ceases and their translocation stops shortly afterwards (1h) (Zimmerman and Alberte 

1996), at which time the roots become fully anoxic (Smith et al. 1984 and 1988). In the dark, 

roots continue to oxidize sucrose, but do so through a fermentative pathway and no metabolic 

products are transported to the shoots. Translocation of inorganic phosphate from the roots to 

shoots is strongly light dependent (Brix and Lingby 1985), whereas a sucrose gradient between 

rhizomes and shoots may offset the reduced downward translocation of sugars during prolonged 

periods of reduced light (Alcoverro et al. 1999). Under dark conditions, the sucrose content of 

Zostera marina leaves was nearly exhausted (~20 μmol sucrose gfw-1) in 21 days (Cabello-Pasini 

et al. 2002) and the shoots did not survive. Under severely light-limited conditions leaves 

reaching similar concentrations also failed to survive (Alcoverro et al. 1999) while the rhizome 

sucrose content was much less reduced. Although the fresh weight concentration of sucrose in 

shoots, roots, and rhizomes is on average 166, 20, and 240 μmol sucrose gfw-1, respectively 

(Table 2.3), the concentrations tends to vary with the age of the leaf and rhizome segment 

sampled (Kraemer and Alberte 1993; Kraemer et al. 1998). Each variable contributes to the 

range of published concentrations (Table 2.3). 
2.8

Table 2.1. Observations of the death of Zostera marina plants in the field and laboratory. NR = not reported. 

Location Month/year Study type

Temperature 

°C

Irradiance 

mol m-2 day-1

Hsat

hour 

day-1

Time to 

death 

weeks Reference

MA, Great Harbor 6/82 field experiment 15 ~2 ~6 ~4 Dennison and 

Alberte 1985 

MA, Great Harbor 8/82 field experiment 21 ~2 ~6 ~4 Dennison and 

Alberte 1985 

VA, York estuary 4-5/87 field monitoring 21 1.5-2.5 ~0 (Hcomp) 3-4 Moore et al. 1997 Mexico, Baja California, Pacific coast 

2-3/98 field monitoring 17 0-4 NR ~4 Cabello-Pasini et al. 

2002 

CA, Monterey Bay 11-12/93 lab experiment 12 ~0.3 2 ~4 Alcoverro et al. 

1999 

WA, Sequim Bay 11/95 lab experiment 9 1-2 NR ~4 Thom et al. 2002 Mexico, Baja California, Pacific coast 

11/00 lab experiment 16 0 0 3-4 Cabello-Pasini et al. 

2002 

NC, North River 2-3/03 lab experiment 10 to 16 0 0 5-7 Biber et al. 2005 
2.9

Table 2.2. P vs. I experimental results for Zostera marina L. from various locations and measurement conditions 

μmol photons 

 m-2 sec-1 Parameters in published units

Location

Temp 

°C

Ik Ic

alpha 

x1000 Pmax Respiration

Published 

units

mg chl a 

[dm-2 ] or 

gfw-1

Reference

AK 931 mg C h-1 gdw-1 McRoy 1974 NC 0.27-1.8 mg C h-1 gdw-1 Penhale 1977 CA 5-20 5 

(15 °C) 

1 

(15 °C) 

 14.2 

 (15 °C) 

1.4 

(15 °C) 

μg C h-1 cm-2 Drew 1979 

MA shall/deep 22 0.4 0.08 -0.13 μmol O2 min-1mg chl a-1 [1.8]/[2.2] Dennison and Alberte 

1982 

VA 1-28 80-385 4.4-8.9 1.13-2.8 mg C h-1 gdw-1 Wetzel and Penhale 1983 VA 15/25 37 11.4 0.42/0.62 μmol O2 min-1mg chl a-1 1.5 Evans et al. 1986 MA 20 15-25 4.4 0.22 0.080 μmol O2 min-1mg chl a-1 [2.5] Dennison and Alberte 

1985 

MA shall/deep 20 67 18 2.7-3.7 0.09-0.21 0.06-0.09 μmol O2 min-1mg chl a-1 [2.7]/[4] Dennison and Alberte 

1986 

MA 0-35 7-90 1-17 4-9 0.055-0.45 0.0083- 0.13 μmol O2 min-1mg chl a-1

and 

μmol O2 min-1 cm-2

[2.2] Marsh et al. 1986 

MA shall/deep 20 78/71 13/14 4.2/2.4 0.33/0.17 0.053/0.034 μmol O2 min-1mg chl a-1 [2.9]/[4] Dennison et al. 1987 CA-Monterey 10/20 5/8 μmol O2 h-1 gfw-1 1.1/1.5 Zimmerman et al. 1989 CA-SF 15 35 0.74 0.06 μmol O2 min-1mg chl a-1 Zimmerman et al. 1991 VA 25 210 12 3.8 0.84 0.06 μmol O2 min-1 mg chl a-1 Goodman et al. 1995 CA-Monterey 15 0.25 0.02 μmol O2 min-1mg chl a-1 2.2 Zimmerman et al. 1996CA-Monterey 15 0.2 0.085 μmol O2 min-1 gfw-1 1.5 Alcoverro et al. 1999 Mexico-Baja 16-22 1-350 10-60 0.6-2.45 μmol O2 min-1gfw-1 2 Cabello-Pasini et al. 

2002 

Mexico-Baja 12-27 50-140 5-10 0.3-0.6 0.03-0.06 μmol O2 min-1gfw-1 Cabello-Pasini, et al. 

2003
2.10

Table 2.3. Sucrose content of Zostera marina L. tissues. NR = not reported 

Sucrose content 

μmol sucrose gfw-1 

(leaves / roots / 

rhizomes)

Loss rate 

μmol sucrose gfw-1 day-1

(leaves / roots / rhizomes) Reference

220 / 30 / 300 Zimmerman et al. 1989 

NR/ 10 / 250 Kraemer and Alberte 

1993 

130 / 10 / NR 

no significant loss 

 (whole plants in dark over 12 hr at unknown 

temperature) 

Zimmerman et al. 1995 

40 / 15 / 250 Kraemer and Alberte 

1995 

100 / 30 / 200 Zimmerman et al. 1996 

90 / 20 / 200 

2.2 / NR / 3.4 

 (whole plants in ~0.3 mol photons m-2 day-1

for 30 days at 12 oC) 

Alcoverro et al. 1999 

233 /NR /NR Touchette and 

Burkholder 2001 

350 / NR / NR 15±1 / NR/ NR 

 (whole plants in dark at 16 oC over 21 days) 

Cabello-Pasini et al. 

2002 

Section 2.2.7 Case for the Role of Sucrose in Establishing the Light Requirements of 

Zostera marina

 The similar times-to-death found for Zostera marina (Table 2.1) suggests the existence of 

an initial reservoir of a relatively uniform concentration of respirable compounds. Sucrose is the 

likely compound that would sustain Zostera marina plants in the absence of light and would be 

replenished when irradiance exceeded Icomp for a long enough time. When its concentration was 

reduced to near zero, the plants could not recover and would, subsequently, die. 

 In the text that follows, a carbon budget is constructed from experiments where sucrose 

was monitored to support the following hypothesis that will be the basis for estimating the light 

requirements to sustain Zostera marina in an environment: Zostera marina survival depends 

upon the maintenance of a leaf sucrose content that is between a low of ~15 μmol sucrose gfw-1

and an average, healthy plant, concentration of 166 ± 40 μmol sucrose gfw-1 (± 1SE). The 

carbon equivalents of this range are ~180 and 1992 ± 480 μmol C gfw-1. 

 Both Pmax and respiration have been found to be linear functions of temperature (Drew 

1979; Marsh et al. 1986) between 0 °C and 35 °C (Figure 2.3). For this carbon budget the 

temperature study results of Marsh et al. (1986) will be extensively used. Regression of gross 

Pmax and respiration data of Marsh et al. (1986) yield the following relationships: 
2.11

 gross Pmax = (0.036 ± 0.0053 * deg C) + 0.11 ± 0.096 (to 30 oC, n = 7) Equation 2.3

 respiration = (0.0085 ± 0.00147 * deg C) – 0.02 ± 0.031 (to 35 oC, n = 8 Equation 2.4 

where the units are μmol O2 mg dm-2 min-1 and ± is 1 SE of the regression coefficients and 

intercepts, respectively. 

 The similarities between Pmax values (on a leaf segment area basis) from photoacclimated 

plants with different chlorophyll contents (Dennison and Alberte 1982 and 1985) suggest that the 

temperature-dependent dark biochemical reactions control the optimum photosynthetic rate. It is 

for this reason that the area-normalized rather than chlorophyll-normalized rates of Marsh et al. 

have been computed. 

 The initial slope (alpha) of the P vs. I curve (Figure 2.1) represents the quantum yield of 

photosynthesis for this plant (Falkowski and Raven 1997). Mean alpha values from Table 2.2 

were converted to quantum yields, μmol O2 μmol photons-1, using area normalizing relationships 

from within the references, e.g. mg chl a dm-2, or using those of Andersen and Johnson 

(unpublished data) found in Table 2.4. The average quantum yield from the alphas in Table 2.2 

is 0.029 ± 0.0075 μmol O2 μmol photons-1 (n = 9). Quantum yields are used to compute the 

photosynthesis-saturating irradiance, Ik, from gross Pmax (equation 2.12). Relatively invariant 

alphas, including some from likely photoacclimated plants, reflect the consequences of the 

tightly coupled, species dependent, chlorophyll a mediated light reactions (Falkowski and Raven 

1997). A consequence of an invariant alpha, but temperature-dependent Pmax, is that Ik, is also a 

function of temperature (Figure 2.3). 

 Sucrose constitutes approximately 90% of the soluble carbohydrate of Zostera marina

(Drew 1983; Smith 1989; Alcoverro et al. 1999). Under optimal light conditions it appears to be 

maintained at seasonally-constant (Cabello-Pasini et al. 2002) but leaf and rhizome segment 

number-dependent, high concentrations (Table 2.3). Unlike the structural carbohydrates, it can 

be readily respired either aerobically or through fermentation (Smith et al.1988; Zimmerman and 

Alberte 1996; Kraemer et al. 1998). 

 Direct evidence for sucrose as the key for Zostera marina survival can be found in the 

studies of Cabello-Pasini et al. (2002) and Alcoverro et al. (1999) where leaf sucrose was 

monitored under varying light conditions. Cabello-Pasini et al. monitored sucrose in leaves from 

a lagoon and offshore, and leaves that were experimentally exposed to no light. In the field, 

sucrose levels were reduced by ~85% following 3 weeks of limited light (~2 mol photons m-2 ⋅

day) that continued to the next monthly sampling when the outer coast plants were found to have 

died. Sucrose was also monitored in plants transplanted to the laboratory with no light, which 

led to significantly reduced survival after 2 weeks, and death to all shoots at 4 weeks (Tables 2.1 

and 2.3). This dark experiment, run at 16 °C with aeration, would have a corresponding leaf 
2.12

Table 2.4. Conversion factors from various sources. 

Andersen and Johnson1 Evans et 

al. (1986)2 

Nelson3 Chlorophyll a from 

Table 2.2 

Units gfw dm-2 gdw dm-2 gfw gdw-1 gfw gdw-1 gdw dm-2 mg dm-2 mg gfw-1

Mean 2.168 0.350 6.26 4.4 0.25 2.79 1.6 

SD 0.538 0.081 0.99 0.4 0.062 0.82 0.40 

n 72 100 215 8 6 

1

 unpublished 2005, mid leaf (#2, #3, #4) section data from Yaquina Bay, Oregon 

2

 2-3 cm leaf tip data (leaf number not specified) 

3

 unpublished 2004, whole #2 and #4 leaf data from Yaquina Bay, Oregon 

Table 2.5. Statistics of variables used for calculations. 

Production Respiration Sucrose 

 μmol O2 dm-2 min-1 alpha Content Loss rate 

5 oC 20 oC 5 oC 20 oC 

μmol O2

μmol photons-1 μmol gfw-1 μmol gfw-1 day-1

mean 0.290 0.828 -0.027 -0.154 0.0285 166 -15.11 -2.22

SD 0.199 0.157 0.070 0.049 0.0113 107 1.97 0.80 

n 7 7 8 8 9 7 4 4 

1

 Cabello-Pasini, et al. (2002) 

2 Alcoverro, et al. (1999)

respiration rate of 0.12 μmol O2 dm-2 min-1 (from Equation 2.4). The following substitutions 

with the experimentally-determined fresh weight to leaf area conversions allow comparisons of 

the respiratory carbon and sucrose carbon losses: 

6.2 12 * 15 Loss Sucrose 12180

5.2.

1

1 * min1440 *

17.2

*

min

12.0 0.879n Respiratio

2

2

2

2

Equation sucrosemol

Cmol

daygfw

sucrosemol

daygfw

Cmol

Eq Omol

Cmol

daygfw

dm

dm

Omol

daygfw

Cmol

μ

μ μ μ

μ

μ μ μ

− ±−= =

− ±−= =

where the rates and conversion factors are from Tables 2.3 to 2.5. 
2.13

Ik

R2

 = 0.90

gross Pmax

R2 = 0.90

respiration

R2

 = 0.85

0

0.2

0.4

0.6

0.8

1

1.2

0 10 20 30

degrees Celsius

μmol O2 dm-2 min-1

0

40

80

120

160

200

240

μmol photons m-2 sec-1

Figure 2.3. Gross maximum productivity, gross Pmax (■), Ik values (▲), and dark respiration rates 

(●) from Marsh et al. 1986. 

While the dark respiration and the sucrose loss rates are significantly different (ts 6.95, 

t.05[9] = 2.262), the higher sucrose loss rate appears to be sufficiently large to account for the 

respiratory losses in this experiment. 

 

The uncertainties of these calculations were determined following the propagation of 

errors formulation for products and quotients (Beers 1957) and of sums and differences (Sokal 

and Rohlf 1981): 

For the composite variable, v that is a multiplicative function of x, y and z, its fractional 

standard deviation Sv (the standard deviation (SD) of v divided by the computed value of 

v) is the square root of the sum of the squared fractional standard deviations of the 

variables of which it is a combination. For the composite variable that results from the 

sum or difference between variables, the SD is the square root of the sum of the squares 

of the variable SDs plus or minus any correlation between them (rxy, below, is the 

correlation coefficient between x and y). 
2.14

 

nSESD Equation d

SDSDrSDSDSDyxvfor Equation c

SDSDrSDSDSDyxvfor Equation b

Equation a

z

SD

y

SD

x

SD S

v

SD zyxvfor

vv

v x xyy yx

v x xyy yx

z x y

v

v

7.2

, 2 7.2

, 2 7.2

, 7.2

2 2

2 2

2 2 2

∗=

−= ∗∗−+=

+= ∗∗++=

⎟

⎠

⎞ ⎜

⎝

⎛ + ⎟

⎟

⎠

⎞

⎜

⎜

⎝

⎛ ⎟ + ⎠

⎞ ⎜

⎝

⎛ ==∗∗=

To conservatively compute the standard error of the calculated values from SDv, n from 

the variable with the smallest sample size was used (Tables 2.4 and 2.5). 

 The lower sucrose loss rate (2.2 vs. 15 μmole sucrose gfw-1 day-1) in the Alcoverro et al. 

(1999) experiments (Table 2.5) reflects the contributions of newly-fixed carbon during the two 

hours of saturating light the plants were exposed to when light was not completely excluded. 

The carbon balance of this experiment at 12 °C is as follows: 

Carbon equivalent of sucrose loss from leaves during the 30-day experiment:

8.2 12 * 2.2 8.426 Equation sucrosemol

Cmol

daygfw

sucrosemol

daygfw

Cmol

μ

μ μ μ Δ − ±− = 

where ∆ is the leaf sucrose loss rate from Tables 2.3 and 2.5. 

Carbon balance from computed carbon fixation and respiration:

 2 hours of saturating light at 12 °C from Equation 2.3 yields a gross production of: 

9.2

2.1

1 * min120 * 17.2 * min

54.0 6.29.24

2

2

2

2 Equation Omol

Cmol

daygfw

dm

dm

Omol

daygfw

Cmol

μ

μ μ μ

⎟

⎟

⎠

⎞

⎜

⎜

⎝

⎛

± = 

where 1/1.2 is the molar carbon yield per oxygen evolved (Zimmerman et al.1996). 

24 hours of respiration in leaf tissue at 12 °C from Equation 2.4 yields: 
2.15

10.2

1

1 * min1440 * 17.2 * min

086.0 6.657

2

2

2

2 Equation Omol

Cmol

gfw day

dm

dm

Omol

daygfw

Cmol

μ

μ μ μ

⎟

⎟

⎠

⎞

⎜

⎜

⎝

⎛ − ±− = 

 The net production rate under this limited light exposure is -32 ± 9.3 μmol C gfw-1day-1, 

where the SE was computed by propagating the uncertainty from the gross production and 

respiration using Equation 2.7a into their differences while accounting for the correlation of 

production and respiration (Equation 2.7b). They are strongly correlated (rpr=0.847) because 

gross production includes respiration. The carbon loss as sucrose, -26 ± 4.7 μmol C/gfw-day, is 

not significantly different than the net production rate (ts 0.88, t.05[9] = 2.262). 

 The accuracy of these budgets is completely dependent upon the accuracy of the rates and 

conversion factors that were used. It was assumed that while the P vs. I determined respiration 

rate was temperature dependent it was not light dependent. The efficacy of these respiration 

rates in long periods of darkness is questionable given that while these rates were determined in 

P vs. I experiments, the so-called “enhanced post-illumination effect” on respiration (Heichel 

1970; Falkowski and Raven 1997) was likely occurring. The consequences of this effect in 

terrestrial plants and phytoplankton have been shown to be extended periods of decreasing 

respiration rates following the withdrawal of light that asymptotically approach a true, nocturnal 

rate. The author is not aware of a demonstration or quantification of this effect on Zostera 

marina respiration so the respiration rates from Equation 2.4 were used for all times. 

Further inaccuracy may be introduced by the choice and source of conversion factors 

(Table 2.4). Factors for the same conversions may depend on the geographic source of plants 

and the methods used including age of leaves, leaf section measured (Table 2.4). The direct 

conversion of leaf area to fresh weight (2.17 gfw dm-2) derived by Andersen and Johnson 

(unpublished 2005 data) was used to reduce the cumulative uncertainty of Equation 2.7a. The 

sequential conversion (4.4 gfw gdw-1 (Evans et al., 1986) and 0.25 gdryw dm-2 from Nelson 

(unpublished 2004 data) results in a combined factor of 1.1 gfw dm-2. Using this sequential 

conversion would increase the computed values in Equations 2.5, 2.9 and 2.10 by a factor of 2. 

With the sequential conversion, the totally dark respiration and sucrose utilization rates 

(Equations 2.5 and 2.6) become closer and not significantly different (157 and 180 μmol C gfw1

day-1 while the net production with 2 hours of light (Equations 2.9 and 2.10) went up to -64 

μmol C gfw-1day-1 which is significantly larger (ts 3.29, t.05[9] =2.26) than the sucrose loss rate of 

-26 μmol C gfw-1day-1 (Equation 2.8). Given the uncertainty of the nocturnal respiration rate and 

the various conversion factors needed for these calculations, the net production losses attributed 

to sucrose losses are seen to be justified. 

Section 2.2.8 A Strategy for Assessing the In Situ Health Status of Zostera marina

 A consequence of the correspondence between negative net production and sucrose loss 

is that the leaf concentration of sucrose attained in optimal light conditions can be used as a 

benchmark for computing the “health” of this plant. Attaining and maintaining this maximum 

level is an indication of a healthy meadow. Concentrations much lower would be suggestive of 

plants under light stress. 
2.16

 Periodic sampling and analysis of leaf sucrose content could be used to evaluate the 

health status of an existing meadow. The suitability of a location and depth for Zostera marina

transplantation, or the reason for the absence of Zostera marina, could be demonstrated by 

calculations from continuous monitoring of in situ irradiance and temperature. By combining 

these measurements with the Pmax and temperature relationships of Marsh et al. (1986) one can 

compute the in situ “sucrose status”. 

 In the laboratory experiments cited above, when carbon fixation was maximal or zero one 

only needed to compute the gross Pmax and respiration rates at the experimental temperatures to 

compute net production. In the field, production will not always be maximal (for the 

temperature) or zero (darkness), but will range from zero to gross Pmax because photons can 

contribute to carbon-fixation at fluxes down to ~0.3 μmol photons m-2sec-1 (Falkowski and 

Raven 1997). For those fluxes at or less than saturating irradiances, the gross production is 

proportional to the ratio of the measured irradiance to the temperature-dependent, 

photosynthesis-saturating irradiance, Ik. Therefore, net production computed from the in situ

temperature and time-interval averaged irradiances is the following: 

* max RPgross Equation 11.2

I

I

P t t

t

k

i

neti = − 

Where, I

i is the average irradiance over time interval, i, and t is the in situ temperature. 

Ik

t

 is calculated by dividing gross Pmax at temperature t by the quantum yield, alpha. This yields 

Ik

t

 with the following units: 

12.2

sec60

min * 100 *

029.0

*

sec min 2

2

2

2

2

2 Equation

m

dm

Omol

mol photons

dm

Omol

m

mol photons

μ

μ μ μ =

The units for i net P are: 

13.2. min *

17.2

* *

2.1 min

*

min

*

1

1 2

2

2

2

2

2

2 Eq igfw

dm

Omol

Cmol

dm

Omol

Omol

Cmol

dm

Omol

igfw

Cmol ⎥

⎦

⎤ ⎢

⎣

⎡

⎟

⎟

⎠

⎞

⎜

⎜

⎝

⎛

⎟

⎟

⎠

⎞

⎜

⎜

⎝

⎛ ⎥ −

⎦

⎤ ⎢

⎣

⎡ = μ

μ μ

μ

μ μ μ

To meet the expectations of the hypothesis, that sucrose maintained within a range of leaf 

concentrations is indicative of healthy Zostera marina plants, the following must occur when the 

optimal sucrose concentration (Sucrosemax) is 166 μmol sucrose gfw-1 (Table 2.3) or 1992 μmol 

C gfw-1: 

 Sucrosemax ≥ iSucrose = (i-1Sucrose + iPnet/12) ≤ 15 Equation 2.14 
2.17

where 12 is the number of carbon atoms per sucrose molecule. When iPnet is positive, carbon is 

added to the sucrose pool up to Sucrosemax. 

For these computations, plant death occurs at iSucrose ∼ 15 μmol sucrose gfw-1, the levels 

found in dying plant leaves (Alcoverro et al.1999; Cabello-Pasini et al. 2002) that may represent 

the physiological minimum for this tissue. 

 Computation of iSucrose accounts for varying light and temperature conditions across 

time and would indicate when light and temperature conditions persisted long enough to sustain 

the growth of Zostera marina. A specific time under a given light flux, e.g. Hsat, and temperature 

can not be assigned a priori due to the dynamic nature of the sucrose pool these calculations are 

modeling. However, the time to draw down the sucrose pool in complete darkness from its 

maximum level can be estimated from the respiration rate alone as a function of temperature 

(Figure 2.4). 

 The minimum photon flux necessary to sustain a plant (no net production, Sucrosemax

maintained) was calculated in the following: 

15.2 100 *

029.0

* 2.1*

2

2

2

2

2 2 Equation

m

dm

Omol

mol photons

Cmol

Omol

daydm

Cmol

daym

mol photons

μ

μ

μ

− μ μ = 

where μmolC dm-2day-1 is the temperature dependent daily respiration rate that is converted to 

photon flux through the quantum yield and normalizing relationships. The daily photon flux 

necessary to offset a plant’s respiratory needs are found in Figure 2.5. 

The values of Figure 2.5 result from the same logic as is invoked by the concept of a 

minimum number of Hsat, or Hcomp hours. However, by themselves, neither is especially useful 

in regard to contributing to testable hypotheses about minimum light requirements because of the 

dynamic nature of in situ light fluxes. 
2.18

0

10

20

30

40

50

60

70

80

90

100

2.5 5.0 7.5 10.0 12.5 15.0 17.5 20.0 22.5 25.0 27.5

degrees Celsius

days to death in dark

476 days

103 days

Figure 2.4. Days to plant death in dark assuming the sole respiratory carbon source is leaf 

sucrose. Error bars (± 1 SE) are from the computations incorporating the uncertainties of 

the underlying measurements (Tables 2.4 and 2.5). 

0.0

0.5

1.0

1.5

2.0

0 2.5 5 7.5 10 12.5 15 17.5 20 22.5 25 27.5 30

degrees Celsius

mol photons m-2 day-1 

to meet respiratory demands

Figure 2.5. Photon flux necessary to replace carbon lost to respiration as a function of 

temperature (assuming equal daylight and nocturnal respiration rates). Error bars (± 1 

SE) are from the computations incorporating the uncertainties of the underlying 

measurements. 
2.19

 2.3 Research Gaps in Relation to Setting Protective Criteria

 The correlative approach utilizing growing season measurements for setting system 

habitat requirements is incapable of accounting for the varying periods of diminished irradiance 

that have led to loss of Zostera marina populations. The iSucrose model (above), utilizing 

irradiance and temperature values from continuous data sets collected from strategically placed 

locations within a system, would provide the means of assessing the system’s ability to maintain 

the benchmark range of leaf sucrose over time and space. The growing season is not the only 

time that Zostera marina in a system can be vulnerable to sub-optimal light conditions. 

 The studies of Alcoverro et al. 1999 and Cabello-Pasini et al. 2002 were both done with 

intact roots and rhizomes but the leaf sucrose loss rate did not appear to be ameliorated by the 

mobilization of sucrose from the below-ground tissues as has been suggested in Thalassia 

testudinum (Lee and Dunton 1996). The inhibition of sucrose translocation during anaerobic 

periods of very reduced, or no photosynthesis (Zimmerman and Alberte 1996) may be 

responsible for the apparent isolation of the below-ground sucrose stores in Zostera marina

during light stress. Because the utility of leaf sucrose concentrations in assessing the health of 

this plant depends on the isolation of its leaf sucrose pools from other tissues additional 

simultaneous studies of these sucrose pools under low light conditions are warranted. Such 

studies would assess the importance of this mechanism of mitigating low light conditions in this 

species. 

 The need to determine the properties of optically important components for each 

estuarine system should be broadly evaluated using the same laboratory techniques, same water 

samples, and to the extent possible, should incorporate the use of multi-spectral instruments, 

such as the WetLab ac-9. To reduce costs of such surveys, the resulting IOPs could be processed 

by the analytically robust, and freely-available RTE model of Gallegos (1994). This model is 

currently included as a Fortran module (C. Gallegos, personal communication, 3/12/07) of the 

Chesapeake Bay Water Quality Model (Cerco and Moore 2001). The goal would be to find 

regionally-similar absorption and scattering spectra (e.g., abiotic-mineral-derived turbidity, peatderived DOM) and ranges of spectral responses for, say, mono-specific and mixed-species algal 

blooms. In addition, such surveys would demonstrate the advantages of uniform data sets for 

system comparisons across space and time. 

 The use of Secchi disk readings should be discontinued and replaced with measures of 

light intensity using PAR sensors as the conversion to KD (PAR) from Secchi readings have 

introduced unnecessary uncertainty in this critical measurement. 

 Physiological studies of P vs. I, sucrose content etc., should always include a range of 

normalizing measures. The minimum, most critical measurements are chlorophyll per fresh 

weight and surface area per fresh weight of blades, and fresh weight for below ground 

components, if studied. Chlorophyll normalization has been shown to reduce the variability of 

several physiological parameters while normalization to fresh weight may minimize confounding 

photoacclimation issues resulting from varying chlorophyll content. The use of blade surface 
2.20

area normalization is a means to directly couple in situ light fluxes and laboratory-determined 

photosynthetic rates. 

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2.21

Drake, L. A., F. C. Dobbs, and R. C. Zimmerman. 2003. Effects of epiphyte load on optical 

properties and photosynthetic potential of the seagrasses Thalassia testudinum Banks ex 

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Drew, E. A. 1983. Sugars, cyclitols and seagrass phylogeny. Aquatic Botany 15:387-408. 

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Heichel, G. H. 1970. Prior illumination and the respiration of maize leaves in the dark. Plant 

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2.22

Kraemer, G. B. and R. S. Alberte. 1993. Age-related patterns of metabolism and biomass in 

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Kraemer, G. B. and R. S. Alberte. 1995. Impact of daily photosynthetic period on protein 

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survival in light-limited environments. Journal of Experimental Marine Biology and 

Ecology 185:191-202. 

Kraemer, G., P. D. Kohrs, and R. S. Alberte. 1998. Developmental changes of sucrose synthase 

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62:189-198. 

Lee, K. S. and K. H. Dunton. 1996. Production and carbon reserve dynamics of the seagrass 

Thalassia testudinum in Corpus Christi Bay, Texas, USA. Marine Ecology Progress 

Series 146:201-210. 

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89-96. In: R.G. Wetzel (ed.) Periphyton of Freshwater Ecosystems. Dr. W. Junk 

Publishers, The Hague, Netherlands. 

Major, K. M. and K. H. Dunton. 2002 Variations in light-harvesting characteristics of the 

seagrass Thalassia testudinum: evidence for photoacclimation. Journal of Experimental 

Marine Biology and Ecology 275:173-189. 

Marsh, J. A., W. C. Dennison, and R. S. Alberte. 1986. Effects of temperature on photosynthesis 

and respiration in eelgrass (Zostera marina L.). Journal of Experimental Marine Biology 

and Ecology 101:257-267. 

McRoy, C. P. 1974. Seagrass productivity: carbon uptake experiments in eelgrass Zostera 

marina and its epiphytes. Aquaculture 4:131-137. 

Mobley, C. D. 1994. Light and Water Radiative Transfer in Natural Waters. Academic Press, 

New York. 

Moore, K. A., R. L. Wetzel, and R. J. Orth. 1997. Seasonal pulses of turbidity and their relations 

to eelgrass (Zostera marina L.) survival. Journal of Experimental Marine Biology and 

Ecology 215:115-134. 

Moore, K. A., H. Q. Neckles, and R. J. Orth. 1996. Zostera marina (eelgrass) growth and 

survival along a gradient of nutrients and turbidity in the lower Chesapeake Bay. Marine 

Ecology Progress Series 142:247-259. 

Penhale, P. A. 1977. Macrophyte-epiphyte biomass and productivity in an eelgrass (Zostera 

marina L.) community. Journal of Experimental Marine Biology and Ecology 26:211-

224. 

Preisendorfer, R. W. 1986. Secchi disk science: visual optics of natural waters. Limnology and 

Oceanography 31:909-926. 

Sokal, R. R. and F. J. Rohlf. 1981. Biometry, 2nd edition. Freeman and Company, New York. 

Smith, R. D., W. C. Dennison, and R. S. Alberte. 1984. Role of seagrass photosynthesis in root 

aerobic processes. Plant Physiology 74:1055-1058. 

Smith, R. D. 1989. Anaerobic metabolism in the roots of the seagrass Zostera marina L. PhD 

dissertation, The University of Chicago. 
2.23

Smith, R. D., A. M. Pregnall, and R. S. Alberte. 1988. Effects of anaerobiosis on root 

metabolism of Zostera marina (eelgrass): implications for survival in reducing sediments. 

Marine Biology 98:131-141. 

Thom, R., A. Borde, S. Rumrill, D. L. Woodruff, G. D. Williams, J. Southard, and S. L. Blanton. 

2002. Factors influencing the spatial and annual variability in eelgrass (Zostera marina 

L.) meadows in Pacific Northwest, USA, Systems. Chapter 5, In: J.K Parrish and K. 

Litle, eds. Pacific Northwest Coastal Ecosystems Regional Study, 2001 Annual Report. 

K. Litle, Seattle, WA. 

Touchette, B.W.and J. Burkholder. 2001. Nitrate reductase activity in a submersed marine 

angiosperm: controlling influences of environmental and physiological factors. Plant 

Physiology and Biochemistry 39:583-593. 

Twilley, R. R., W. M. Kemp, K. W. Staver, J. C. Stevenson, and W. R. Boynton. 1985. Nutrient 

enrichment of estuarine submersed vascular plant communities: I. Algal growth and 

effects on production of plants and associated communities. Marine Ecology Progress 

Series 23:179-191. 

Wetzel, R. L. and P. A. Penhale. 1983. Production ecology of seagrass communities in the 

lower Chesapeake Bay. Marine Technology Society Journal 17:22-31. 

Zimmerman, R. C. 2003. A biophysical model of irradiance distribution and photosynthesis in 

seagrass canopies. Limnology and Oceanography 48:568-585. 

Zimmerman, R. C., D. G. Kohrs, and R. S. Alberte. 1996. Top-down impact through a bottomup mechanism: the effect of limpet grazing on growth, productivity and carbon allocation 

of Zostera marina L. (eelgrass). Oecologia 107:560-567. 

Zimmerman, R. C. and R. S. Alberte. 1996. Effect of light/dark transition on carbon 

translocation in eelgrass Zostera marina seedlings. Marine Ecology Progress Series 

136:305-309. 

Zimmerman, R. C., D. G. Kohrs, D. L. Steller, and R. S. Alberte. 1995. Carbon partitioning in 

eelgrass. Plant Physiology 108:1665-1671. 

Zimmerman, R. C., J. L. Reguzzoni, S. Wyllie-Echeverria, M. Josselyn, and R. S. Alberte. 1991. 

Assessment of environmental suitability for growth of Zostera marina L. (eelgrass) in 

San Francisco Bay. Aquatic Botany 39:353-366. 

Zimmerman, R. C., R. S. Smith, and R. S. Alberte. 1989. Thermal acclimation and whole-plant 

carbon balance in Zostera marina L. (eelgrass). Journal of Experimental Marine Biology 

and Ecology 130:93-109.

3.1

3.0 Water Column and Sediment Nutrients as Limits to Growth of 

Zostera marina and Thalassia testudinum

James E. Kaldy

3.1 Background

Seagrasses are vascular plants that have returned to the aquatic habit. In contrast to the 

algae (e.g. seaweeds), seagrasses have highly differentiated tissues which form true roots, stems 

and leaves, and thus possess xylem and phloem for the transport of water and photosynthate 

respectively. The vascular system is much reduced but has been experimentally shown to exhibit 

bidirectional transport (Thursby and Harlin 1982; Brix and Lyngby 1985). Seagrasses can 

acquire nutrients through both the leaves and the root and thus access both water column and 

sediment nutrient sources to support growth and production. The relative importance of leaf vs 

root nutrient uptake appears to be dependent on the specific environmental conditions and 

nutrient concentrations. This review is intended to give the reader a current basic understanding 

of seagrass-nutrient interactions and dynamics. Additional comprehensive detail on seagrassnutrient interactions are available (e.g., Butler and Jernakoff 1999; Hemminga and Duarte 2000; 

Short and Coles 2001; Larkum et al. 2006). 

In general algae out-compete seagrasses for water column nutrients (WCN) than 

seagrasses, since they have a higher affinity for nitrogen they can take up WCN more quickly. Z. 

marina nitrogen uptake rates exhibit a wide range of values and are summarized in Table 3.1. 

Relatively few studies have addressed phosphorus uptake by Zostera spp. (McRoy and Barsdate 

1970; McRoy et al. 1972; Penhale and Thayer 1980; Brix and Lyngby 1985; Perez-Llorens et al. 

1993). Touchette and Burkholder (2000a,b) provide excellent reviews of carbon, nitrogen and 

phosphorus metabolism in seagrasses addressing many of the ecophysiological and biochemical 

aspects in detail. Unfortunately, there has been little attempt within the literature to report 

uptake rates with consistent units (Table 3.1). The units for phosphorus uptake are especially 

variable. Consequently, comparison of rates between systems is difficult without first making 

site specific conversions based on biomass or shoot density. The sediments of marine 

ecosystems typically have higher nutrient concentrations than the water column. Paradoxically, 

leaf uptake can account for up to 70% of the total plant N uptake (Hemminga et al. 1994). 

Access to both water column and sediment nutrients is an important adaptation that has permitted 

seagrasses to persist and to out-compete algae (seaweeds) under oligotrophic conditions. 

Additionally, recent work has shown that internal resorption of N and P from senescing leaves 

can meet part of the plant=s nutrient requirements (Hemminga et al. 1999). 
3.2 

Table 3.1. Summary of nitrogen uptake rates for Z. marina. Vmax in units of μM gdw lf-1 h-1 is the maximum uptake rate, Ks is the half 

saturation concentration (μM). Units for uptake rates are denoted by superscripted letter. 

Location Leaf uptake Root uptake Reference 

Vmax Ks Vmax Ks

Zostera marina Rhode Island 20.5 9.2 211 104 Thursby & Harlin 1982  Netherlands 0.03 B 0.19a Pedersen & Borum 1993  Netherlands 0.01 B 2.8b 0.01 B 3.9b Hemminga et al. 1994  Netherlands 0.35c 0.33c Pedersen & Borum 1992  Japan 0.001d 0.011d Iizumi & Hattori 1982  Alaska 1.05 B 3.2b 0.9 B 2.9b Short & McRoy 1984 Thalassia hemprichii Indonesia 32-37 21-60 Stapel et al. 1996. Thalassia testudinum  Texas (NH4+) 8-16.5 5-19 8-73 34-765 Lee & Dunton 1999b  Texas (NO3-) 3.4-6.5 2.2-38 Lee & Dunton 1999b 

a g N m-2 d-1b μM N gdw-1 h-1c mg N gdw-1 d-1d g N (g plant N)-1 d-1
3.3 

Table 3.2. C:N:P ratios for Z. marina from the literature. 

Seagrass Location %C %N %P C:N N:P C:P C:N:P Reference 

Multiple spp. global 33.6 1.92 0.23 20 24 474 474:24:1 Duarte 1990 Zostera marina

various 36 2.5 0.39 17 15 246 255:15:1 Duarte 1990 California 38.4 2.37 0.34 20 16 304 Fourqurean et al. 1997 Oregon 34 2.7 0.4 15 17 255 255:17:1 2Kaldy 2006a Oregon 35 1.3 0.2 34 15 510 510:15:1 3Kaldy 2006a Oregon 29 1.4 0.1 24 24 576 576:24:1 4Kaldy 2006a 

18 27 481 481:27:1 1Pérez-Lloréns et al. 1991 12 10 120 120:10:1 1Pérez-Lloréns et al. 1991 13 6 79 79:6:1 1Pérez-Lloréns et al. 1991 11 15 170 170:15:1 1Pérez-Lloréns et al. 1991 19 3 57 57:3:1 1Pérez-Lloréns et al. 1991 

California 39 7 38 274 274:38:1 Atkinson and Smith 1983 Rhode Island 34 18 27 481 481:27:1 Atkinson and Smith 1983 Virginia 42 14 41 584 584:41:1 Atkinson and Smith 1983 Virginia 35-39 2-3.8 0.2-

0.3 

2Moore and Wetzel 2000 

Virginia 27-35 1-2 0.06-

0.15 

3Moore and Wetzel 2000 

Zostera capricorni

Australia 20 17 349 349:17:1 2 Atkinson and Smith 1983 Australia 58 8 465 465:8:1 3 Atkinson and Smith 1983 Australia 33 9 302 302:9:1 4 Atkinson and Smith 1983 
3.4 

Zostera noltii

Netherlands 12 9 104 104:9:1 2Pérez-Lloréns et al. 1991 

21 6 127 127:6:1 3Pérez-Lloréns et al. 1991 16 7 114 114:7:1 4Pérez-Lloréns et al. 1991 

Thalassia testudinum

Barbados 445:32:1 2 Atkinson and Smith 1983 Barbados 601:20:1 3 Atkinson and Smith 1983 Texas 33-37 1.7-

2.7 

15-25 

2Kaldy and Dunton 2000 

Texas 35-37 0.7-

0.9 

45-62 

3 Kaldy and Dunton 2000 

Texas 34-36 1.7-2 21-

24 

2 Lee and Dunton 1999a 

Texas 36-37 0.7-

0.9 

46-60 

3 Lee and Dunton 199a 

Florida 34.6 2.2 0.095 18.5 40.2 1070 2 Fourqurean et al. 1992 Florida 36.9 1.82 0.113 24.6 40.2 937 2 Fourqurean et al. 2002 

Thalassia hemprichii

Indonesia 1.9 0.14 30 2 Erftemeijer & Middelburg 

1993 

0.48 0.07 15 3 Erftemeijer & Middelburg 

1993 

0.98 0.06 36 4 Erftemeijer & Middelburg 

1993 

 1 Literature values from Table 1 for original citations see Pérez-Lloréns et al. 1991  2 Ratios for leaf material  3 Ratios for rhizome material  4 Ratios for root material 
3.5 

 In temperate systems with silastic mud sediments the general paradigm is that nitrogen is 

the limiting nutrient (Orth 1977; van Lent et al. 1995). However, there is some debate about the 

degree of nutrient limitation in temperate systems. Zimmerman et al. (1987) concluded from a 

modeling study that nitrogen limitation of Z. marina is probably very rare. In tropical systems 

with carbonate sediments, there is substantial evidence for phosphorus or iron limitation (Short et 

al. 1990; Fourqurean and Cai 2001; Fourqurean and Zieman 2002; Duarte et al. 1995). 

 However, recent work has shown that carbonate dissolution from seagrass organic acids 

may meet seagrass P requirements (Jensen et al. 1998; Burdige and Zimmerman 2002). In the 

Pacific Northwest (PNW) there appear to be few peer-reviewed publications on nutrient 

limitation with regard to seagrasses (Williams and Ruckelshaus 1993). However, it seems 

unlikely that nitrogen is limiting in these systems given that coastal upwelling occurs during 

summer bringing nutrient rich ocean water to the surface (10-25 μM nitrate; C. Brown unpubl. 

data). During winter, terrestrial run-off through anthropogenically impacted watersheds also 

results in large N loading. Terrestrial run-off through red alder (Alnus rubra) rich secondary 

growth forest can provide substantial N inputs because of the nitrogen fixing bacteria associated 

with the trees. Consequently, even forested watersheds can have high N loading rates (Compton 

et al. 2003). 

The determination of nutrient limitation for many primary producers is often based on 

examination of nutrient ratios. Seawater typically has a relatively fixed ratio of elements. 

Deviations from these ratios provides preliminary evidence for specific processes controlling 

how much of an element is present but, it can be dangerous to use elemental ratios as the only 

evidence for limitation. When N:P < 16 the system may be nitrogen limited (excess 

phosphorus), N:P > 16 system may be phosphorus limited (excess nitrogen). The Redfield ratio 

may be reflective of unicellular organisms (phyto- and bacterioplankton); however because of 

the structural components associated with macrophytes (seaweeds and seagrasses) the classic 

oceanic Redfield ratio (C:N:P = 106:16:1) is not appropriate. 

Elemental ratios can provide a general rule of thumb for nutrient limitation, but only 

experimental determinations truly indicate the rate limiting steps. Literature reviews indicate 

that the median leaf seagrass C:N:P is about 400:20:1 with considerable variability (Table 3.2). 

Thus, on a carbon basis, seagrasses require about 4 times more C and 4 times less N and P than 

phytoplankton cells (Hemminga and Duarte 2000). The C:N for rhizome tissue is often much 

higher than that for leaf material since the rhizomes store fixed carbon. Inference of nutrient 

limitation from C:N:P ratios is even more tenuous for seagrasses than for phytoplankton, since 

these plants have access to both water column and sediment nutrient pools as well as internal 

transport tissues (i.e. xylem and phloem). Consequently, nutrient ratios alone should not be used 

to infer limitation (Touchette and Burkholder 2000a). Manipulative experiments should be 

conducted to determine limiting factors. Additionally, other nutrients can play a role in 

controlling seagrass production, for example Herman et al (1996) present a case study where 

they suggest that decreases in dissolved silica may have been a factor in seagrass decline. 
3.6 

3.2 Nutrient Enrichment and Eutrophication

AEutrophication@ is frequently used to describe the increased input of nutrients, primarily 

nitrogen and phosphorus to receiving waters. Recent work has suggested that eutrophication be 

redefined as Aan increase in the rate of supply of organic carbon to an ecosystem@ (Nixon 1995). 

Sources of organic carbon supply are from either allochthonous or autochthonous primary 

production. Nixon (1995) goes on to suggest a trophic classification scheme based on organic 

carbon supply. Eutrophication is often incorrectly used to include not only the process of 

increased nutrient status but also the effects (e.g. hypoxia, algal blooms, etc.) of this enrichment 

(Richardson and Jorgensen 1996). 

Eutrophication sensu Nixon is caused by both natural and anthropogenic alterations of 

nutrient supply (Jorgensen and Richardson 1996). Coastal upwelling, typically associated with 

western continental margins, is a wind-driven phenomenon that results in increased primary 

production and often leads to enhanced fisheries production (Thurman 1988). Anthropogenic or 

Acultural@ eutrophication alters the availability of nutrient elements within receiving waters 

which has primary, secondary and tertiary level impacts on biogeochemistry, primary and 

secondary production (Jorgensen and Richardson 1996; Vollenweider et al. 1992; Howarth et al. 

2000; Livingston 2001). 

Numerous studies have investigated the effects of eutrophication on seagrass 

communities (Table 3.3). Additionally, the physiological response of Z. marina to light (Chapter 

2) and nutrients (this chapter) have been intensively studied. Most eutrophication studies have 

examined the community level response in experimental systems ranging from aquaria to 

mesocosms to the natural environment (Table 3.3). For Z. marina much of this work has been 

conducted along the East Coast of North America and has resulted in a general theory of seagrass 

response. Specifically, that enhanced nutrient loading leads to a degradation of Z. marina habitat 

(Figure 3.1) by stimulating algal production (micro- and macroalgae) and shading seagrass 

(Short et al. 1991, 1995; McGlathery 2001; Havens et al. 2001). However, there does not appear 

to be a relationship between nutrient input and the algal type (epiphyte vs. macroalgae vs. 

phytoplankton) supplying primary production or between nutrient input and the amount of 

primary production (Nixon et al. 2001). 
3.7 

Table 3.3. Selected list of literature examining the effect of eutrophication on Z. marina

communities. Abbreviations are as follows: Zm = Zostera marina, SAV = Submerged Aquatic 

Vegetation, Epi = Epiphytes, Phyto = Phytoplankton, Macro= Macroalgae, Algae = epiphytes + 

phytoplankton + macroalgae. Positive (+) or negative (-) response in biomass is denoted. 

Organizational 

level 

Experimental 

system 

Response Location Reference 

Community aquaria +Epi, -Zm Virginia Neckles et al. 1993 

aquaria +Zm, -Zm3 Netherlands van Katwijk et al. 1999 

aquaria + Epi Washington Williams & Ruckleshaus 

1993 

aquaria +Epi, -Zm2 Virginia Moore & Wetzel 2000 

field +Zm, +Macro Rhode Island Harlin & Thorne-Miller 

1981 

field +Zm Netherlands van Lent et al. 1995 

field -- Maryland Stevenson et al. 1993 

field +Phyto, 

+Macro, -Zm

Mass. Valiela et al. 1992 

field - Epi, -Zm* Washington Williams & Ruckleshaus 

1993 

field +Macro, -Zm4 Finland Bostrom et al. 2002 

field +Macro, -Zm Mass. Hauxwell et al. 2003 

mesocosm +Epi, -SAV Maryland Twilley et al. 1985 

mesocosm -Zm1 North 

Carolina 

Burkholder et al. 1992, 

1994 

mesocosm +algae, -Zm New 

Hampshire 

Short et al. 1995 

mesocosm +Phyto, -Zm, 

-Macro, -Epi 

Rhode Island Taylor et al. 1995 

Lit. review -- Worm et al. 2000 

Lit. review -- Nixon et al. 2001 

*Suggested nutrient limitation of Z. marina

1

Suggested nitrate toxicity of Z. marina

2

Concluded light dominant factor, only +Epi and -Zm at highest light level. 

3

 Positive and negative effects were dependent on source of seagrass and salinity. 

4

 Conclusion based on inference. 
3.8 

Figure 3.1. Schematic diagram of the effects of increased nutrient loading, which decreases 

seagrass density and biomass as well as promoting a shift toward phytoplankton, epiphyte 

or macroalgal community dominance. Adapted from Short et al. 1991. 

This paradigm is based on several characteristics of East Coast systems and consequently 

may not be applicable to Pacific Northwest estuaries. Specifically, this paradigm applies to 

systems with relatively long residence times (months to years) and to places where nutrient 

inputs are dominated by atmosphere and freshwater. Most PNW estuaries exhibit very short 

residence times (1-30 days) as a result of large freshwater inputs and tidal exchange (Brown and 

Lee 2006). Nutrient inputs from the atmosphere are generally considered low in the PNW (Fenn 

et al. 2004; http://nadp.sws.uiuc.edu/isopleths) and dominant nutrient sources shift seasonally 

between river input during winter and ocean input during summer. The application of the 

nutrient loading/algal response paradigm to the PNW requires further scientific investigation. 
3.9 

The responses of lakes to nutrient loading and eutrophication abatement have been 

examined in a number of studies. Eutrophication abatement has also been an important research 

topic for the aquaculture industry (e.g. Kraufvelin et al. 2001; Porrello et al. 2003a, b) since 

effluent from mariculture activities is often high in nutrients. However, relatively few studies 

have examined the response of temperate estuaries to reduced nutrient loading. Although not 

directly linked to seagrass, O=Shea and Brosnan (2000) conclude that despite reductions in 

municipal and industrial waste water discharges to Long Island Sound, water quality, especially 

since bottom water hypoxia continues to be a problem. Thus nutrient loading and environmental 

problems may not exhibit linear responses, depending on the type of biological community 

dominating the system (Driscoll et al. 2003). 

3.3 Relevant Research

Seagrass nitrogen metabolism and biochemistry has recently been reviewed in detail 

(Touchette & Burkholder 2000a). N acquisition occurs through the leaves and the roots and 

appears to be partially concentration dependent (Thursby and Harlin 1982). Nitrogen uptake 

rates from the literature are summarized in Table 3.1. Ammonium is the preferred N source, 

since it is the reduced form (Touchette and Burkholder 2000a). Leaf N acquisition can account 

for up to 70% of the total nitrogen required by the plant (Hemminga et al 1994; Pedersen and 

Borum 1992, 1993). Additionally, there is internal recycling of nitrogen from senescent tissues 

(Pedersen and Borum 1992; Stapel and Hemminga 1997; Hemminga et al. 1999). Nutrient 

uptake experiments are typically conducted in chambers with isolated root and leaf 

compartments. Changes in concentration through time, as well as radioactive (32P, 14C) and 

stable (15N) isotopes have been used in these experimental systems. Ambient nutrient 

concentrations (Table 3.4) in temperate estuaries are typically sufficient to support seagrass 

production. 

Seagrasses were initially believed to be a phosphorus pump, making sediment bound P 

available to water column organisms through translocation (McRoy and Barsdate 1970). 

However, more recent work suggest that excretion of phosphate is of minor importance (Brix 

and Lyngby 1985; Perez-Llorens et al. 1993). Ambient phosphorus concentrations in the 

environment are reported in Table 3.4. 

Experimental manipulations of nutrient supply suggested that seagrasses can be nutrient 

limited. Some of the first field enrichment experiments added fertilizer to the sediments and 

showed increased leaf length, biomass and shoot density (Orth 1977). Additional field and 

mesocosm research also suggests that Z. marina can be nutrient limited in some situations, 

particularly in sediments with low organic content (Short 1983b, 1987). This implies that 

sediment and water column nutrient conditions need to be considered with regard to setting 

nutrient criteria protective of seagrass. While P is not considered a limiting nutrient in temperate 

systems, there is evidence for P and Fe limitation in tropical carbonate sediments. 
3.10 

Table 3.4. Water column and sediment nutrient concentrations (FM) associated with selected studies of Zostera habitat. 

Seagrass Location Water Sediment Reference 

NH4+ NO3-/NO2= P NH4+ NO3-

/NO2=

P 

Zostera marina

Alaska 20-147 4-28 Short 1983a Netherlands 3-15 1-18 5-17 Herman et al. 1996 Rhode Island 0.1-1.5 0.1-0.8 0-1 Harlin & Thorne-Miller 

1981 

Denmark 1-5.5 0.5-6 250-1100 Pedersen & Borum 1993Alaska 10-1500 Iizumi et al. 1982 Mass. 270-1550 Dennison et al. 1987 Alaska 2-7 5-75 Mc Roy et al. 1972 North Sea 1-14 0-9 50-130 0-2 Hemminga et al. 1994 Alaska 1-15 Iizumi et al. 1980 

Zostera japonica

Oregon 2.2-2.9 7-35 0.6-1.1 60-170 0.8-2.0 3.3-5.6 Larned 2003 Oregon 2-5 2-26 0.5-2 600-2400 0-30 7-40 Kaldy 2006b. 
3.11 

Research has clearly shown that seagrasses are affected by nutrients. Various processes, 

both natural and anthropogenic influence nutrient loading to estuaries. It is also apparent that 

nutrient loading processes are regionally variable and that this variability needs to be accounted 

for in the development of protective criteria. Driscoll et al. (2003) outlined many N sources in 

the northeastern US including atmospheric deposition, anthropogenic point and non-point 

sources (i.e. sewage treatment outfalls, septic system failure, ground-water inputs, etc.), and 

landscape practices (e.g. agriculture). However, this list ignores important N sources in the 

PNW that may influence seagrass including coastal ocean upwelling and forestry practices 

(Compton et al. 2003). The importance of the linkage between land use practices in watersheds 

and the resulting nutrient loads to estuaries is becoming more evident with continued research 

(Short and Burdick 1996; McClelland and Valiela 1998; Correll et al.1992; Valiela and Bowen 

2002; Hauxwell et al. 2003; Compton et al. 2003). Protective nutrient criteria may need to be 

tailored to account for regionally important processes (e.g. ocean currents). 

Nitrogen supply is frequently described as either Anew@ or Aold@ nitrogen. New nitrogen 

is typically inorganic (e.g. N2, NO3

-

, NO2

=

, NOx) and is generally made available through 

processes such as upwelling and nitrogen fixation. Old nitrogen is N made available through 

biological excretion and recycling and is typically composed of more reduced compounds such 

as amines, NH4

+

, urea, etc. Seagrasses prefer NH4

+

 since it requires less energy for 

incorporation, although they can also utilize NO3

-

 (Touchette and Burkholder 2000a). 

Anthropogenic inputs of N to estuarine ecosystems tend to be in the form of new nitrogen, which 

favors algal uptake and production. 

As a result, the impact of enhanced nutrient supply to estuarine seagrass communities is 

typically observed through indirect effects. The shift in nutrient supply favors the development 

of algal communities which reduce underwater light and shade out seagrass. This general trend 

has been observed worldwide in both natural and experimental systems (Nixon et al. 2001). 

Thus, water quality criteria protective of seagrass will probably have a negative impact on algal 

production. Direct toxic effects from nutrients have been reported in the literature (Burkholder 

et al. 1992, 1994; van Katwijk et al. 1997). For example, Harlin and Thorne-Miller (1981) 

performed field nutrient addition experiments in Rhode Island, adding ammonium, nitrate, or 

phosphate to the water column rather than to the sediment. The nitrate supplements, while not 

causing a noticeable change in growth of the above-ground plant, did inhibit the root-rhizome 

fraction of the seagrass. They suggested that this might indicate toxicity of the nitrate addition to 

the test plants. Consistent with this hypothesis was the observation that the eelgrass plants 

sometimes disappeared within a half meter of the nitrate dispenser. 

This observation was supported by the results of Burkholder et al. (1992), who reported 

from mesocosm experiments in North Carolina that nitrate enrichment of the water column 

caused declines of eelgrass, especially at higher temperatures, and that this was a direct 

physiological effect independent of shading by macroalgae. They attributed the effect to internal 

imbalances in nutrient ratios from sustained nitrate uptake through the leaf tissue. However, the 

evidence to support direct toxic effects is still considered somewhat tenuous and is a topic of 

debate in the literature (Moore and Wetzel, 2000). For example, Zostera marina thrives in at 
3.12 

least one Oregon estuary (Yaquina Bay) where both ambient water column and sediment 

nitrogen (nitrate and ammonium) concentrations are 3 to 10 times higher than the levels used in 

any of the experiments that exhibit toxic effects (Brown and Kaldy, US EPA, unpublished data). 

Touchette and Burkholder (2002) outline a physiological mechanism to support the nitrate 

toxicity hypothesis. However, it is important to note that temperature stress may be a 

confounding factor leading to seagrass decline. Bintz et al. (2003) observed that the negative 

effect of elevated water temperature on eelgrass was significantly increased when inorganic 

nutrient concentrations also were increased. 

Experimental and observational evidence indicate that increased organic loading (i.e. 

eutrophication) in estuarine ecosystems results in a shift toward algal dominated production and 

degradation of seagrass habitat, thus altering the organization of trophic levels and the flow of 

energy through the system (See Chapter 8). 

3.4 Gaps in Knowledge

While much of the physiology of nutrient uptake in Z. marina has been examined in a 

crude manner, the details of allocation within the plant and the dominant sources under given 

conditions remain somewhat ambiguous. The issue of direct nitrate or ammonium toxicity 

remains unresolved. This is of particular interest in the Pacific Northwest where water column 

and sediment nitrogen levels can be 3 to 10 times higher than reported toxic levels for long 

periods (days to months) as a result of coastal upwelling and riverine loading with no apparent 

negative impact to Z. marina. Relatively cold water temperatures (annual average 10 EC) may 

help to ameliorate stress. In comparison, areas that appear to show direct nutrient toxicity, e.g. 

Chesapeake Bay and Rhode Island coastal ponds, tend to have warmer summer water 

temperatures (up to 30 EC). As a result temperature and nutrient effects may be confounded and 

work synergistically. Detailed physiological work utilizing newly available technology (e.g. 

compound specific nitrogen isotope analyses, microelectrodes/optodes, mechanistic models, etc.) 

will provide valuable insight on how ramets respond to nutrient stress. Additionally, mesocosm 

and field experiments have shown that Z. marina communities are degraded in response to 

enhanced nutrient loading; however, few studies (if any) have addressed the recovery of these 

systems after abatement of loading. Furthermore, there appears to be little relationship between 

nutrient inputs and the rate and dominant primary producers (Nixon et al. 2001). 

The absolute and relative contributions of different primary producers to net ecosystem 

primary production may provide an important metric for determining the degree of 

eutrophication and/or the potential for eutrophication related impairment of resources. In the 

PNW, the production ecology of benthic marine macrophytes (seagrasses and macroalgae) is 

poorly understood. The production ecology of microalgae is even less well known. It may be 

better to set water quality criteria protective of seagrass based on standards which minimize the 

response of algae to anthropogenic N inputs, since the primary response to nutrient loading is 

mediated through algal blooms which in turn smother seagrass (Nelson and Lee 2001). 

Additionally, whether nutrient criteria protective of seagrass will be equally protective of other 
3.13 

types of habitat (e.g. marshes), as well as economically and recreationally important fisheries, is 

poorly known. 

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3.14 

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3.15 

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3.16 

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4.1 

4.0 Salinity as a Limiting Factor for the Seagrasses Zostera marina

and Thalassia testudinum

Bruce L. Boese

4.1 Background

 Eelgrass, Zostera marina is a euryhaline species (Phillips 1984) found world wide in 

coastal waters that vary from mildly hypersaline to nearly fresh (Table 4.1). However, it appears 

to grow best in estuarine waters with salinities in the range from approximately 5 to just below 

that of normal seawater. In contrast, subtropical turtle grass (Thalassia testudinum), which is the 

dominate seagrass species in Florida and the Caribbean (Phillips 1960) is more stenohaline 

(Table 4.2) and does not tolerate extreme salinity fluxes nor prolonged exposures to fresh water 

(Doering and Chamberlain 2000; Montague and Ley 1993). 

 Alterations in estuarine hydrology due to fresh water diversion, dredging and filling may 

result in alterations of estuarine salinity structure. Areas of high salinity and to a lesser extent 

variable salinities may displace brackish water areas further upstream where higher temperatures 

and more nutrients may be prevalent. Higher or less variable salinities may also encourage the 

growth of the slime mold, Labyrinthula sp. which has been implicated in the historic wasting 

disease die-offs that occurred in eastern and western Atlantic eelgrass populations in the early 

1930’s (den Hartog 1987; Giesen et al. 1990). Geographically isolated Z. marina populations 

may have genetically narrower salinity optima than is suggested by the euryhaline nature of the 

species taken as a whole. This makes it more difficult to predict the effect of local salinity 

alterations and may complicate restoration activities. 

4.2 Salinity Ranges

 The wide range of salinity tolerated in Z. marina (Table 4.1) appears to be related to the 

ability to adapt to changes in salinity by osmotic regulation of cellular solutes via salt excretion 

by epidermal cells (Jagels 1983) and the accumulation of the amino acid proline in hypersaline 

environments (van Diggelen et al. 1987). Although Z. marina is able to survive for a time in 

fresh water, net leaf photosynthesis decreases in waters below 5 and totally ceases in completely 

fresh water (Hellbom and Björk 1999; Biebl and McRoy 1971). Sand-Jensen and Borum (1983) 

noted that a die-off of Z. marina occurred in Danish coastal waters during the winter when plants 

were exposed to salinities below 2. In contrast turtlegrass, T. testudinum is found in a narrower 

salinity range (Table 4.2) with an optimal salinity ranges reported as 25 to 38.5 (Phillips 1960), 

and 17-36 Zimmermann and Livingston 1976). 
4.2 

Table 4.1. Reported salinity tolerance ranges for eelgrass (Zostera marina) populations. 

Location Salinity 

Range 

Citation 

World Wide 0-35 Thayer et al. 1984 

World Wide 0-42 Phillips 1984 

Northern Hemisphere 5-35 den Hartog 1970 

Denmark 13-31 Pinnerup 1980 

Denmark 12-31 Sand-Jensen and Borum 1983 

Denmark (estuaries) 9-23 Wium-Anderson and Borum 1984 

Baltic Sea 6-12 Hellblom and Björk 1999 

Netherlands (marine environments) ~30 van Katwijk et al. 1999 

Netherlands (estuaries) 15-25 van Katwijk et al. 1999 

Netherlands (Lake Grevelingen) 22-32 Kamermans et al. 1999 

France (Thau) 27-41 Rigollet et al. 1998 

Italy (Venice) 25-33 Rigollet et al. 1998 

Chesapeake Bay 14-22 Wetzel and Penhale 1983 

Yaquina Bay, OR 25-33 Kentula and DeWitt 2003 

Table 4.2. Reported salinity tolerance ranges for turtlegrass (Thalassia testudinum) populations. 

Location Salinity 

Range 

Citation 

Florida 28-36 Phillips 1960 

Florida 17-36 Zimmermann and Livingston 1976* 

Florida 24-35 Zieman and Zieman 1989* 

Zieman 1982* 

Texas 30-40 Adair et al. 1994* 

Florida 22-36 Doering and Chamberlain 2000 

Dry Tortugas 35-38.5 Phillips 1960 

Everglades National Park 28-48 Phillips 1960 

Florida (west coast) 25-34 Phillips 1960 

*As cited in Doering and Chamberlain 2000. 
4.3 

4.3 Growth/Shoot/Competition Effects

 While Z. marina is tolerant of some exposure to fresh and hypersaline waters, optimal 

growth rates vary with populations. Zostera marina population increases and declines were 

correlated with relatively slight alteration in the salinity structure of Lake Grevelingen in The 

Netherlands, within the range of 22 to 32 (Kamermans et al. 1999). Laboratory experiments 

supported this interpretation showing optimal growth rates were higher at 22 than 32. However, 

Z. marina collected from an area of relatively greater salinity (Roscoff and Bay of Archchon, 

France) did not exhibit differences in growth rates when subjected to the same salinity regimes 

(Kamermans et al. 1999). Optimal salinities for Pacific Northwest (PNW) Z. marina populations 

have not been determined, but Z. marina appears to grow best at 20-32 in Puget Sound (Philips 

1984). Laboratory experiments on PNW Z. marina suggested that higher growth rates occurred 

at 30 than 10 and that the highest densities of Z. marina in the field were in areas of high 

salinities and lowest temperature (Thom et al. 2001, 2003). This apparent relationship between 

plant densities (as estimated from percent cover) and high salinities combined with low 

temperature was also evident in the Yaquina estuary, Newport, OR (Kentula and DeWitt 2003). 

Similar studies have been conducted on T. testudinum. In six week laboratory mesocosm 

experiments (Doering and Chamberlain 2000), T. testudinum was exposed to a range of salinities 

from 6 to 35, with a variety of simple plant metrics (blades per shoot, growth, biomass, tissue 

nitrogen) determined. The results of this experiment suggested that T. testudinum was adversely 

affected by prolonged exposures to salinities less that 12 (Doering and Chamberlain 2000). The 

results of that study were similar to that of a study (Lirman and Cropper Jr. 2003) in which T. 

testudinum was exposed to short-term (14 day) exposure to salinities which ranged from 5 to 45. 

In that study the maximum growth rate for T. testudinum was observed between 30 and 40 with 

reduced rates at 5 and 45. 

Field populations of T. testudinum appear to occupy a narrower range of salinities than 

those they have been shown to tolerate in the laboratory. This is likely due to competition from 

other seagrass species (Halodule wrightii, Syringodium filiforme) which often co-occur with T. 

testudinum and appear to replace it in hyper- and hyposaline conditions. Greenawalt-Boswell et 

al. (2006) found that in hydrologic regions of Charlotte Harbor (southwest Florida) characterized 

by highly variable salinity regimes, overall seagrass biomass was reduced and that H. wrightii

was likely to replace T. testudinum as the dominant species. In contrast, the construction of the 

Gulf Intracoastal Waterway tended to reduce and moderate hypersalinity events in Laguna 

Madre (southeast Texas). This resulted in a decrease in H. wrightii with its replacement by S. 

filiforme and T. testudinum (Quammen and Onuf 1993). 

 Hypersaline conditions have been observed to occur in Florida Bay and have been 

suggested as a contributing cause to a die off of T. testudinum which began in 1987 (Zieman et 

al. 1999). However, more recent laboratory studies suggest that T. testudinum is highly tolerant 

of salinities as great as 60 (Koch et al. 2007) and field studies which examined the responses of a 

T. testudinum meadow to brine discharges from a desalination plant show no apparent adverse 

effects (Tomasko et al. 2000). The most likely cause of this die off was an interaction between 
4.4 

high salinities, high temperatures (Florida Bay Science Plan 1994). High temperatures and 

salinities also may affect photosynthetic oxygen production, resulting in an increase in sediment 

sulfides which in combination with high temperatures has been implicated in seagrass die-offs 

(Koch and Erskin 2001). Hypersalinity was also implicated in seagrass losses in a Yucatan 

(Mexico) coastal lagoon (Herrera-Silveira et al. 2000). In that study mean salinity of 42 resulted 

in an overall reduction in seagrass coverage which was remediated by increasing freshwater 

inputs to the lagoon which reduced mean to 35 resulting in an increase in seagrass coverage and 

a change in dominate seagrass species which included the appearance of some patches of T. 

testudinum (Herrera-Silveira et al. 2000). 

 However, diverting of fresh water into estuaries has been implicated in several long-term 

seagrass declines (Estevez 2000). For example, Zostera hornemanniana meadows in the Etang 

de Berre, a French Mediterranean lagoon, began to degrade after the completion of a 

hydroelectric dam in 1996 which erratically diverts fresh water into the lagoon (Stora and 

Arnoux 1983). Within a few years most of the Zostera meadows had disappeared with a 

concurrent change in the benthos to a degraded euryhaline community (Stora and Arnoux 1983). 

Montegue and Ley (1993) noted that overall seagrass biomass in Florida Bay decreased with 

increasing variation in salinity, and that these decreases were greatest where salinity was lowest. 

Their work suggests that even when mean salinity changes of water diversion projects are small, 

a small increase in salinity variance may have drastic effects on seagrass populations. Similarly 

fresh water diversions which reduce salinities below the optimal range may adversely affect T. 

testudinum populations, especially recruitment from seedlings. In mesocosm experiments, Kahn 

and Durako (2006) found that while turtle grass seedlings tolerated a reduction in salinity of ~10 

below their optimal range (30-40), they were less adaptable than mature shoots. In addition 

ammonium toxicity tended to increase at these lowered salinities which imply that fresh water 

inputs with high nutrients may be detrimental to recruit survival (Kahn and Duranko 2006). 

 In Z. marina there appears to be a significant interaction between tolerance to higher 

salinity and nutrients. Laboratory experiments in which both salinity and nutrients were 

manipulated suggest that Z. marina subjected to high salinities (30) responded adversely to 

nutrient additions (van Katwijk et al. 1999). Van Katwijk et al. (1999) went on to speculate that 

the world wide decline in Z. marina may be related to nutrient increases in high salinity coastal 

environments. In a more recent paper describing a conceptual model for habitat suitability for Z. 

marina transplants (van Katwijk et al. 2000), this idea was further refined by the suggestion that 

stress resulting from high salinity would adversely affect the plants ability to cope with any 

additional stressor. However, the responses of Z. marina populations to changes in salinity also 

appear to be related to the ambient salinities from which these populations originated, suggesting 

genotypic differences in salinity tolerances (Kamermans et al. 1999; van Katwijk et al. 1999, 

2000). Genotypic differences in eelgrass populations have also been implicated in the apparent 

geographic differences in leaf widths (McMillan 1978). 
4.5 

4.4 Wasting Disease

 The slime mold (Labyrinthula sp.) is a secondary decomposer of seagrasses and algae, 

and may have been the cause of the “wasting disease” decline in Z. marina that occurred in 

Eastern North American Atlantic waters in 1932 (den Hartog 1987) and in the Wadden Sea 

during the same time period (van Katwijk et al. 2000). The co-occurrence of drought and 

associated high salinity waters with this die off has led to speculation that high salinity with high 

temperatures favored this infective agent (Martin 1954). Further support for this idea is that 

Labyrinthula, appears not to be pathogenic below 12-15 (Giesen et al. 1990) and that Z. marina 

populations found in brackish water appeared not be affected by the disease (den Hartog, 1987; 

Vergeer et al. 1995). Whether Labyrinthula was the underlying cause or merely acted as a 

decomposer of eelgrass stressed by some other agent is debatable (den Hartog 1987; Vergeer et 

al. 1995). 

 Wasting disease has also been implicated in the die off of T. testudinum in Florida Bay 

(Blakesly et al. 2002). As with Labyrinthula infestations in eelgrass, it primarily has affects high 

density turtle grass meadows when salinities are high (Blakesly et al. 2002). However, it is 

likely not the major cause of the seagrass die-off noted in Florida Bay (Boesch et al. 1993; 

Zieman et al. 1999; Blakesly et al. 2002). However, as Labyrinthula prefers more saline waters, 

fresh water diversions and possible droughts resulting from global climate change will have a 

positive influence on its growth thus making it more likely to adversely affect seagrasses. 

4.5 Summary/Research Gaps

Generally both turtle and eelgrass have wide salinity tolerances with turtle grass being the 

more stenohaline species. In areas of suboptimal or highly variable salinities both species are 

often precluded by complex interactions with other stressors such as high temperatures, 

suboptimal lighting conditions, wasting disease, and competition with species more tolerant of 

hypo- and hypersaline conditions. 

With the exception of van Katwijk et al. (1999) studies have not addressed the 

interrelationship between nutrients and salinity. Based on that study it is at least possible that 

nutrient additions may have greater impact on eelgrass populations in high salinity areas, 

possibly requiring greater regulatory controls. By the same logic, water diversion projects which 

increase salinity may result in greater potential stress from nutrients and wasting disease. Erratic 

fresh water discharges into bays and estuaries have been shown to have deleterious effects on 

tropical and subtropical seagrasses beyond which would have been predicted by the mean change 

in salinity. However experimental work on the effect of increasing the variance in salinity on Z. 

marina and T. testudinum appears to be lacking. Studies that address these research gaps should 

be conducted on a variety of seagrass genets to assess effects on localized populations. 
4.6 

4.6 Literature Cited

Adair, S. E., J. L. Moore, and C. P. Onuf. 1994. Distribution and status of submerged vegetation 

in estuaries of the Upper Texas Coast. Wetlands 14:110-121. 

Biebl, R. and C. P. McRoy. 1971. Plasmatic resistance and rate of respiration and 

photosynthesis in the seagrass Zostera marina at different salinities and temperatures. 

Marine Biology 8:48-56. 

Blakesley, B. A., D. M. Berns, M. F.Merello, M. O. Hall, and J. Hynirva. 2002. The dynamics 

and distribution of the slime mold Labyrinthula sp. and its potential impacts on 

Thallassia testudinum populations in Florida. pp. 199-207. In: H.S. Greening (ed). 

Seagrass Management: Its Not Just Nutrients! Aug 22-24, 2000. St. Petersburg, FL. 

Tampa Bay Estuary Program. 

Boesch, D. F., N. E. Armstrong, C. F. D’Elia, N. B. Maynard, H. W. Paerl, and S. L. Williams. 

1993. Deterioration of the Florida Bay Ecosystem: An evaluation of the scientific 

evidence. Report to the Interagency Working Group on Florida Bay. Sept. 15, 1993. 

pp.30. 

denHartog, C. 1970. The Seagrasses of the World. North-Holland Publication Amsterdam. pp. 

275. 

denHartog, C. 1987. 'Wasting disease' and other dynamic phenomena in Zostera beds. Aquatic 

Botany 27:3-14. 

Doering P. H. and R. H. Chamberlain. 2000. Experimental studies on the salinity tolerance of 

turtle grass, Thalassia testudinum. pp. 81-98 In: S. A. Bortone (ed)., Seagrasses: 

Monitoring, Ecology, Physiology and Management. CRC Press, Boca Raton. pp. 318. 

Estevez, E. D. 2000. Matching Salinity Metrics to Estuarine Seagrasses for Freshwater Inflow 

Management. pp. 295-307. In: S. A. Bortone (ed)., Seagrasses: Monitoring, Ecology, 

Physiology and Management. CRC Press, Boca Raton. pp. 318. 

Florida Bay Science Plan. 1994. Florida Bay Interagency Working Group. 

Giesen, W. B. J. T., M. M. van Katwijk, and C. den Hartog. 1990. Temperature, salinity, 

insolation and wasting disease of eelgrass (Zostera marina L.) in the Dutch Wadden Sea 

in the 1930's. Netherlands Journal of Sea Research 25:395-404. 

Greenawalt-Boswell, J. M., J. A. Hale, K. S. Fuhr, and J. A. Ott. 2006. Seagrass species 

composition and distribution trends in relation to salinity fluctuations in Charlotte 

Harbor, Florida. Florida Scientist 69: 24-35. 

Hellblom, F. and M. Björk. 1999. Photosynthetic responses in Zostera marina to decreasing 

salinity, inorganic carbon content and osmolality. Aquatic Botany 65:97-104. 

Herrera-Silveira J. A., J. Ramírez-Ramírez, N.Gómez, and A. Zaldivar-Jimenez. 2000. Seagrass 

bed recovery after hydrological restoration in a coastal lagoon with groundwater 

discharges in the north of Yucatan (southeastern Mexico). pp 219-229. In: S. A. Bortone 

(ed)., Seagrasses: Monitoring, Ecology, Physiology and Management. CRC Press, Boca 

Raton. pp. 318 

Jagels, R. 1983. Further evidence for osmoregulation in epidermal leaf cells of seagrasses. 

American Journal of Botany 70:327-333. 
4.7 

Kahn, A. E. and M. J. Durako. 2006. Thalassia testudinum seedling response to changes in 

salinity and nitrogen levels. Journal of Experimental Marine Biology and Ecology 335: 

1-12. 

Kamermans, P., M. A. Hemminga, and D. J. de Jong. 1999. Significance of salinity and silicon 

levels for growth of a formerly estuarine eelgrass (Zostera marina) population (Lake 

Grevelingen, The Netherlands). Marine Biology 133:527-539. 

Kentula, M. E and T. H. DeWitt. 2003. Abundance of seagrass (Zostera marina L.) and 

macroalgae in relation to the salinity-temperature gradient in Yaquina Bay, Oregon, 

USA. Estuaries 48:1130-1141. 

Koch, M. S. and J. M. Erskine. 2001. Sulfide as a phytotoxin in the tropical seagrass Thalassia 

testudinum: interations with light, salinity and temperature. Journal of Experimental 

Marine Biology and Ecology 266:81-95. 

Koch, M. S., S. A. Schopmeyer, C. Kyhn-Hansen, C. J. Madden, and J.S. Peters. 2007. Tropical 

seagrass species tolerance to hypersalinity stress. Aquatic Botany 86:14-24.. 

Lirman, D. and W. P. Cropper Jr. 2003. The influence of salinity on seagrass growth, 

survivorship, and distribution within Biscayne Bay, Florida: Field, experimental, and 

modeling studies. Estuaries 26:131-141. 

Martin, A. C. 1954. A clue to the eelgrass mystery. Transactions of the 19th North American 

Wildlife Conference, Washington, D.C 441-449. 

McMillan, C. 1978. Morphogeograhic variation under controlled conditions in five seagrasses, 

Thalassia testudinum, Halodule wrightii, Syringodium filiforme, Halophila engelmannii, 

and Zostera marina. Aquatic Botany 4:169-189. 

Montagure, C. L. and J. A. Ley. 1993. A possible effect of salinity fluctuation on abundance of 

benthic vegetation and associated fauna in northeastern Florida Bay. Estuaries 16:703-

717. 

Phillips R. C. 1960. Observations on the ecology and distribution of the Florida seagrasses. 

Professional papers series, No 2, Contribution No. 44, Florida Sate Board of 

Conservation Marine Laboratory, St. Petersburg. FL, 72pp. 

Phillips, R. C. 1984. The ecology of eelgrass meadows in the Pacific Northwest: A community 

profile. U.S. Fish and Wildlife Service 85 pp. 

Pinnerup, S. P. 1980. Leaf production of Zostera marina L. at different salinities. Ophelia 

(Supplement) 1:219-224. 

Quammen, M. L. and C. P. Onuf. 1993. Laguan Madre: Seagrass changes continue decades 

after salinity reduction. Estuaries 2:302-310. 

Rigollet, V., T. Laugier, M. L. De Casabianca, A. Sfriso, and A. Marcomini. 1998. Seasonal 

biomass and nutrient dynamics of Zostera marina L. in two Mediterranean lagoons: Thau 

(France) and Venice (Italy). Botanica Marina 41:167-179. 

Sand-Jensen, K. and J. Borum. 1983. Regulation of growth of eelgrass (Zostera marina L.) in 

Danish coastal waters. Marine Technology Society Journal 17:15-21. 

Stora, G. and A. Arnoux, 1983. Effects of large freshwater diversions on benthos of a 

Mediterranean lagoon. Estuaries 6:115-125. 

Thayer, G. W., W. J. Kenworthy, and M. S. Fonseca. 1984. The ecology of eelgrass meadows 

of the Atlantic coast: A community profile. U.S. Fish and Wildlife Service. FWS/OBS84/02. 147 pp 
4.8 

Thom R. M., A. B. Borde, S. Rumrill, D. L. Woodruff, G. D. Williams, J. Southard, and S. L. 

Blanton. 2001. Factors influencing the spatial and annual variability in eelgrass (Zostera 

marina L.) in Pacific Northwest, USA, System (Chapter 5). PINCERS 2001 Annual 

Report. 

Thom R. M., A. B. Borde, S. Rumrill, D. L. Woodruff, G. D. Williams, J. A. Southard, and S. L. 

Sargeant. 2003. Factors influencing spatial and annual variability in eelgrass (Zostera 

marina L.) meadows in Willapa Bay, Washington, and Coos Bay, Oregon. Estuaries 26: 

1117-1129. 

Tomasko, D. A., N. J. Blake, C. W. Dye, and M. A. Hammond. 2000. Effects of the disposal of 

reverse osmosis seawater desalination discharges on a seagrass meadow (Thalassia 

testudinum) offshore of Antigua West Indies. pp. 99-112. In: S. A. Bortone (ed)., 

Seagrasses: Monitoring, Ecology, Physiology and Management. CRC Press, Boca Raton. 

pp. 318. 

van Diggelen, J., J. Rozema, and R. Broekman. 1987. Mineral composition of and proline 

accumulation by Zostera marina L. in response to environmental salinity. Aquatic 

Botany 27:169-176. 

van Katwijk, M. M., G. H. W. Schmitz, A. P. Gasseling, and P. H. van Avesaath. 1999. Effects 

of salinity and nutrient load and their interaction on Zostera marina. Marine Ecology 

Progress Series 193:155-165. 

van Katwijk, M. M., D. C. R. Hermus, D. J. de Jong, R. M. Asmus, and V. N. de Jonge. 2000. 

Habitat suitability of the Wadden Sea for restoration of Zostera marina beds. Helgoland 

Marine Research 54:117-128. 

Vergeer L. H. T., T. L. Aarts, and J. D. de Groot. 1995. The ‘wasting disease’ and the effect of 

abiotic factors (light intensity, temperature, salinity) and infection with Labyrinthula 

zosterae on the phenolic content of Zostera marina shoots. Aquatic Botany 52: 35-44. 

Wetzel, R. L. and P. A. Penhale. 1983. Production ecology of seagrass communities in the 

lower Chesapeake Bay. Marine Technology Society Journal 17:22-31. 

Wium-Andersen, S. and J. Borum. 1984. Biomass variations and autotrophic production of an 

epiphyte-macrophyte community in a coastal Danish area: I. Eelgrass (Zostera marina

L.) biomass and net production. Ophelia 23:33-46. 

Zieman, J. C. 1982. The ecology of the seagrasses of South Florida: a community profile. U.S. 

Fish and Wildlife Services, Office of Biological Services, Washington, D.C., FWS/OBS82/25, pp. 158. 

Zieman, J. C., J. W. Fourqurean, and T. A. Frankovich. 1999. Seagrass die-off in Florida Bay: 

Long-term trends in abundance and growth of turtle grass, Thalassia testudinum. 

Estuaries 22:460-470. 

Zieman, J. C. and R. T. Zieman. 1989. The ecology of the seagrass meadows of the west coast of 

Florida: a community profile, U.S. Fish and Wildlife Service Biological Report 85(7.25), 

155 pp. 

Zimmerman, M. S. and R. J. Livingston. 1976. Seasonality and physico-chemical ranges of 

benthic macrophytes from a North Florida estuary (Apalachee Bay). Contributions in 

Marine Science 20:33-45. 
5.1 

5.0 The Effects of Hydrodynamic Factors on Seagrasses

Cheryl A. Brown

5.1 Background

 Most biological, geological, and chemical processes in seagrass beds are influenced by 

water motions (Koch 2001). In addition, the presence of seagrasses influences the motions of the 

water. Seagrasses reduce water flow (Worcester 1995; Fonseca et al. 1982; Peterson et al. 2004), 

attenuate waves (Fonseca and Cahalan 1992), modify turbulent mixing (Worcester 1995), 

generate vertical secondary flows (Nepf and Koch 1999), induce coherent, canopy-scale eddies 

(Ackerman and Okubo 1993), and retain water during low tides (Powell and Schaffner 1991). 

Water motions have the potential to affect the growth, survival, and distribution of seagrasses 

through both direct and indirect mechanisms. 

 This chapter reviews the effect of hydrodynamic stressors on seagrass growth, survival, 

and distribution, including the effects of current motions, waves and turbulence. Currents are 

relatively steady uni-directional flows, while wave-induced flows are oscillatory. Turbulence is 

temporally and spatially irregular water motions that are superimposed over the larger flow 

pattern, such as unidirectional current or oscillatory wave action (Koch and Verduin 2001). The 

reader is referred to Koch et al. (2006a) for a review of the fundamentals of fluid flow, 

particularly as it relates to seagrass. Hydrodynamic stressors can influence seagrass through 

numerous mechanisms, including direct damage or uprooting of plants due to the hydrodynamic 

forces, erosion of sediment surrounding the plant, and by limiting productivity through either 

shading or diffusional boundary layer limitation. In addition, water motions influence seagrass 

populations through dispersal of pollen (Ackerman 2002), seeds (Orth et al. 1994), and 

reproductive shoots (Harwell and Orth 2002). 

 Anthropogenic activities have the potential to modify the current and wave exposure of 

seagrass communities, which may, in turn, affect their growth and distribution. Hydraulic and 

hydrodynamic modifications of coastal ecosystems, such as freshwater diversions, 

channelization, and damming, have been identified as one of the major environmental issues of 

coastal regions (National Research Council 1994). Hydrological and hydrodynamic alterations 

influence salinity patterns, tidal dynamics, circulation patterns, as well as the supply of nutrients, 

toxics, and sediments to coastal ecosystems, all of which can impact seagrass. There has also 

been an increase in boat and ship traffic in coastal regions, which produces currents and waves 

that can potentially influence seagrass. In addition to anthropogenic factors, natural extreme 

hydrodynamic events such as storms and hurricanes have influenced seagrass distribution and 

survival. 

 Most of the research on the interaction of hydrodynamics and seagrass focuses on the 

effect of seagrass on the hydrodynamics with relatively few studies of the impact of the 
5.2 

hydrodynamics on seagrass productivity and distribution. In addition, most of the studies on the 

effects on hydrodynamic stressors on seagrass are correlative rather than causative. 

5.2 Effect of Seagrass on Currents, Waves, and Turbulence

 The effect of seagrass on currents, waves and turbulence is complicated and depends on 

the hydrodynamic regime as well as specifics of the seagrass habitat (e.g., seagrass type and 

morphology, orientation of leaves and canopy, location within the seagrass bed and depth, shoot 

density, presence of epiphytes and macroalgae). There have been field, laboratory, and modeling 

studies on the interaction of seagrasses and currents, waves and turbulent mixing. Koch et al. 

(2002, 2006a, 2006b) present an excellent review on this topic. 

 Hydrodynamics influence the architecture of seagrass meadows. Seagrass blades are 

easily bent by currents (Fonseca et al. 1982). When bent over, the blades form a dense layer that 

redirects the water flow over and under it. As the current increases, the canopy becomes more 

compressed with maximum canopy bending occurring at water speeds of about 40-50 cm s-1 

(Fonseca et al. 1982). Seagrass blades also flap or flutter when subjected to water flow. The 

range of movement of the seagrass blades may depend upon the amount of epiphyte cover on the 

seagrass blades and heavy epiphyte cover may result in the formation of a closed canopy with 

strong flows above the canopy and little flow within the canopy (Koch 1996). Canopy bending 

and flapping has implications for light availability for seagrass photosynthesis (Zimmerman 

2003) as well as the transport of particulate and dissolved materials between the water column 

and the canopy. 

 The water flow inside seagrass beds is often reduced by a factor of 2 to 10 compared to 

adjacent bare areas (e.g., Scoffin 1970; Worcester 1995; Komatsu 1996; Koch 1996; Koch et al. 

2002; Peterson et al. 2004). Numerous field and laboratory studies have found a reduction of 

flow inside the seagrass canopies resulting from drag associated with the seagrass blades 

(Fonseca et al. 1982; Fonseca et al. 1983; Fonseca and Fisher 1986; Gambi et al. 1990; 

Ackerman and Okubo 1993; Komatsu 1996; Nepf and Vivoni 2000; van Keulen and Borowitzka 

2000; Peterson et al. 2004); however, the degree of reduction is variable. Studies have found 

that the degree of flow reduction is dependent upon the vertical distribution of plant material 

(Ackerman and Okubo 1993; van Keulen and Borowitzka 2000), the vegetation density (Fonseca 

et al. 1982; Peterson et al. 2004), the seagrass morphology (van Keulen and Borowitzka 2000; 

Fonseca and Fisher 1986), the water depth relative to canopy height (Fonseca and Fisher 1986), 

the degree of bending of the seagrass blades (Thomas et al. 2000), the magnitude of the water 

velocity (Fonseca et al. 1982; Gambi et al. 1990), and the distance from the leading edge of the 

seagrass bed (Fonseca et al.1983; Gambi et al.1990; Peterson et al. 2004). Friction associated 

with the seagrass canopy may also act to retain water in seagrass beds. Dense seagrass beds in 

shallow regions have been shown to retain a thin layer (< 20 cm) of water during falling tides 

and this water trapping ability appears to be related to the morphology of the seagrass and 

seagrass density (Powell and Schaffner 1991). This water trapping ability may be important in 
5.3 

preventing desiccation (see Chapter 10) of the seagrass and other organisms that live in the 

habitat. 

 Generally, water flow is reduced with increasing seagrass density and a recent study 

found that the vertically integrated flow varied inversely with the square root of vegetation 

density (Peterson et al. 2004); however, some studies have found that the degree of flow 

reduction is independent of vegetation density (Fonseca et al. 1983; Fonseca and Fisher 1986). 

Current speeds decrease with increasing distance from the leading edge of seagrass meadow with 

maximum reduction occurring within 25 to 50 cm into the bed (Gambi et al.1990). The width of 

this region of decelerating flow is dependent upon the seagrass density (Peterson et al. 2004) and 

within this region the transport of material is dominated by advection from the leading edge of 

the canopy (Nepf and Vivoni 2000). The flow reducing capability of seagrass is negatively 

correlated with the degree of bending of the seagrass canopy (Thomas et al. 2000). When the 

seagrass blades are upright the friction is at a maximum and as the blades bend over when 

subjected to faster current speeds the friction coefficient decreases. There is often accelerated 

flow above the seagrass canopy (Gambi et al. 1990; Worcester 1995) which, when combined 

with the reduced flow within the canopy, results in a high shear stress layer at the canopy-water 

interface (Gambi et al. 1990; Nepf and Vivoni 2000). The presence of flow acceleration over the 

canopy may depend upon whether the seagrass occurs in patches or as a continuous meadow 

(Worcester 1995). In addition, some studies have found that there is a local velocity maximum 

near the bed within the canopy due to a decrease in vegetation density in the sheath region of the 

canopy (Fonseca and Kenworthy 1987; Ackerman and Okubo 1993; Koch 1996, Nepf and 

Vivoni 2000). This near bottom increase in flow speed may result in elevated sediment 

resuspension (Koch 1999b). Some have suggested that there is acceleration around seagrass 

beds due to flow being deflected above and around the seagrass bed (Gambi et al.1990). Granata 

et al. (2001) found that the interaction of hydrodynamics and seagrass distribution (i.e., spatial 

variation in seagrass density, and presence of edges and gaps) results in three-dimensional 

circulation patterns, including increase in current above meadows, upward flow at the edge of the 

canopy, and recirculation patterns at gaps. 

 The interaction of current flow and the eelgrass blades can result in large-amplitude 

synchronous waving of the blades (Fonseca and Kenworthy 1987; Ackerman and Okubo 1993; 

Grizzle et al. 1996), which has been termed “monami” by Ackerman and Okubo (1993). At low 

current speeds, eelgrass blades gently undulate with low-amplitude motions and gentle flapping. 

When the above-canopy water velocity exceeds 10 cm s-1 monami occurs, with maximum 

amplitude motions occurring at current velocities of about 30 cm s-1 (Grizzle et al. 1996). The 

turbulent vertical transfer of momentum is enhanced during monami resulting in more vertical 

exchange between overlying water column and interior of the canopy (Ghisalberti and Nepf 

2002). These coherent eddies have implications for scalar fluxes that govern gas and nutrient 

exchange, seed dispersal, sediment deposition, and chemical reactions in submerged plant 

canopies (Ackerman and Okubo 1993; Grizzle et al. 1996; Ghisalberti and Nepf 2002). Nepf 

and Koch (1999) demonstrated that submerged plant-like arrays exposed to gradients in 

longitudinal velocity in the laboratory produced vertical pressure gradients that drove vertical 
5.4 

secondary flows. These vertical secondary flows can reach up to 15% of local longitudinal 

velocity and may affect the exchange of nutrients between the sediment and water column. 

However, these vertical velocities are not expected under conditions of extreme bending 

(skimming flow) where the mean current is almost entirely deflected over the top of the canopy. 

 Numerous studies have documented a reduction of wave energy (which is proportional to 

wave height squared) in seagrass habitat (Fonseca and Cahalan 1992; Koch 1996; Granata et al. 

2001; Newell and Koch 2004; Koch et al. 2006b). Oscillatory orbital wave motion is reduced 

with depth inside the seagrass canopies (Koch and Gust 1999). Generally, near bed orbital wave 

velocities are lower in seagrass beds and decrease with increasing plant density (Granata et al. 

2001). Fonseca and Cahalan (1992) used a wave tank to examine the effect of H. wrightii, S. 

filiforme, T. testudinum, and Z. marina on wave energy under various combinations of shoot 

density and water depth to leaf length ratio. When the length of the seagrass blades was similar 

to the water depth, the wave energy reduction per meter of seagrass bed was about 40%. Despite 

differences in morphologies, the four species of seagrass had a similar effect on wave energy 

reduction. As the water depths increased (relative to blade length), wave attenuation was 

reduced. Negligible wave energy reduction occurred at water depths greater than 2 times the 

mean leaf length. For S. filiforme, there was a significant increase in wave energy reduction with 

increasing shoot density, though this effect was not seen in the other three species. Oscillatory 

flow generated by waves results in the seagrass blades’ flapping back and forth at the frequency 

of the waves (Koch and Gust 1999), resulting in an opening and closing of the seagrass canopy 

enhancing exchange between the water column and seagrass canopy. Field measurements of 

waves in Ruppia maritima beds showed that the degree of wave attenuation was dependent upon 

water depth and characteristics of the seagrass bed (Newell and Koch 2004, Koch et al. 2006b). 

The degree of wave attenuation varied with tidal stage with maximum observed attenuations of 

50% observed during low tides. Wave attenuation was only observed at shoot density > 1000 

shoots m-2, and highest wave attenuation was observed when the plants were reproductive, 

occupying the entire water column. 

 The interaction of seagrass with hydrodynamics can have various effects on turbulence 

and turbulent mixing between the water column and the seagrass canopy. The reduction of water 

flow caused by drag associated with the seagrass blades and canopies results in the conversion of 

kinetic energy of the mean flow into turbulent kinetic energy (Gambi et al. 1990). Typically, 

there is a maximum of turbulent intensity near the top of the seagrass canopy (Gambi et al. 1990; 

Nepf and Vivoni 2000), which is generated in part by the large shear stress in this region. There 

is a vertical reduction in turbulence with depth inside the canopy (Koch and Gust 1999). Some 

studies have documented a reduction of turbulence (Koch 1996; Granata et al. 2001) and 

turbulent mixing within seagrass canopies (Ackerman and Okubo 1993; Ackerman 2002), while 

others have found that there is an increase in turbulence inside the canopy relative to adjacent 

upstream bare regions (Gambi et al. 1990). Some studies have shown that the presence of 

seagrass blades results in the production of turbulent flow with mean turbulent intensity and 

amount of the water column influenced by the presence of the canopy increasing with distance 

from the leading edge of the grass bed (Gambi et al. 1990). The conflicting effect of seagrass on 

turbulence may be related to the flow dynamics and the configuration of the seagrass bed (e.g., 
5.5 

continuous versus patchy distribution, location within the meadow and canopy, presence of 

epiphytes, vegetation density). Worcester (1995) found that at low seagrass density and low 

current flow, the presence of eelgrass had no effect on turbulent mixing compared to adjacent 

bare areas, while at sites with continuous eelgrass cover there was an increase in turbulent 

mixing above the seagrass canopy relative to adjacent bare areas. Granata et al. (2001) found an 

increase in turbulence at the edge of seagrass meadows. The seagrass blades and canopies may 

also rescale turbulent energy by attenuating low frequency energy and generating high frequency 

energy (Koch 1996). Presence of epiphytes has been found to result in elevated turbulence 

within seagrass canopies (Koch 1996). 

 Koch and Gust (1999) found that the effect of seagrass on hydrodynamics, including the 

effect on mean flow, turbulence and mixing, may depend upon the hydrodynamic conditions at 

the site. At tide-dominated sites, the current bends the seagrass blades producing a skimming 

flow over the seagrass canopy. This closure of the canopy results in reduction of turbulence 

inside the canopy and reduced mixing between the overlying water column and the seagrass 

blades. In contrast, at wave-dominated sites oscillatory wave action causes seagrass blades to 

flap back and forth, and the canopy is repeatedly opened and closed increasing the water 

exchange between the water column and canopy. Using a model combined with observations, 

Abdelrhman (2003) found that the vertical distribution of a constituent in the water column 

determines whether the canopy will enhance or reduce the transport of the constituent. For 

example, a constituent with a vertical profile with maximum concentrations at the surface would 

have transport enhanced by 20%, whereas a constituent with maximum concentrations at the 

bottom would have transport reduced by 30%. 

5.3 Effects of Water Velocity on Seagrass Growth and Distribution

 The reduction of current velocities by seagrasses has positive and negative effects on 

their growth. Advantages of reduced flow include reduced self-shading, reduced sediment 

resuspension, increased settlement of organic and inorganic particles, and high water residence 

time increasing potential for nutrient uptake (Koch 2001). Detrimental effects of reduced water 

velocity include increased phytotoxin concentrations in the sediment and an increase in the 

thickness of the diffusional boundary layer, which may limit photosynthesis (Koch 2001; Koch 

et al. 2002). 

5.3.1. Direct Damage to Seagrass resulting from Currents 

 Scoffin (1970) conducted flume experiments to examine the effect of unidirectional 

currents on erosion of T. testudinum. At current velocities of about 70 cm s-1 flapping of 

seagrass blades occurred, dislodging attached epiphytes and sometimes causing breakage of the 

blades. Sediment removal around the base of the shoots was dependent upon current speed and 

density of the seagrass bed. Extensive sediment removal around the rhizomes and roots occurred 

at current speed of 50 cm s-1 in a sparse grass bed, 100 cm s-1 in a medium density bed, and at 

150 cm s-1 in a dense grass bed (current velocity measured just above the blades). Fonseca et al. 

(1983) proposed that the maximum current velocity that Z. marina can tolerate is 120 to 150 cm 

s

-1. 
5.6 

 To determine the susceptibility of seagrasses to hydrodynamic stressors, one can measure 

the biomechanical properties of seagrass, such as the breaking stress, breaking strain, elastic 

modulus, and toughness. Comparison of these properties to the hydrodynamic forces 

encountered in different environments allows prediction of the probability of damage (Patterson 

et al. 2001). There have been three studies that characterized the biomechanical properties of 

eelgrass (Kopp 1999; Patterson et al. 2001; Fonseca et al. 2007). Patterson et al. (2001) found 

that in natural populations of Z. marina there are always a few strong reproductive shoots that 

would be resistant to extreme hydrodynamic events (such as hurricanes and tropical storms), 

which may ensure the survival of the population. Fonseca et al. (2007) measured the force on Z. 

marina blades in a flume under unidirectional and oscillatory flow and these forces were 

compared to blade tensile strength. Fonseca et al. (2007) conclude that these seagrass blades 

may be damaged or broken under frequently observed storm conditions, and that damage was 

more likely when the seagrass blades were subjected to oscillatory versus unidirectional flow. 

 Previous studies have demonstrated that the morphology of eelgrass leaves is dependent 

upon the nutrient availability (e.g., Short 1983) and there is some evidence that structural 

components of plant tissues (e.g., C:N ratio and cellulose content) may be influenced by the 

nutrient conditions. Kopp (1999) proposed that these changes in morphology and structural 

composition of the leaves may affect the tensile strength of the leaves and the ability of the 

shoots to withstand current and wave energy. Kopp (1999) conducted a set of mesocosm 

experiments to examine the effect of nutrient enrichment on the biomechanical properties of Z. 

marina. This study found that there was a reduction in tensile forces that leaves could withstand 

when subjected to nitrate enrichment for nine weeks. In addition, field measurements of the 

tensile strength of Z. marina leaves revealed that leaves at low nutrient locations could withstand 

26% more force than leaves from high nutrient sites. 

5.3.2 Effect of Currents on Plant Morphology and Configuration of Seagrass Beds 

 Fonseca et al. (1983) found that the physical configuration of Z. marina meadow (the 

ratio of the height to length) was positively correlated with current velocity and that the 

continuity of seagrass cover is inversely related to current speeds (Figure 5.1). There was 

increased mounding of the substratum with increasing current velocity. In regions of unidirectional flow, seagrass is often observed growing in rows perpendicular to the axis of the flow 

(Fonseca et al., 2007; Figure 5.1). Schanz and Asmus (2003) found that hydrodynamics 

influenced the morphology of Z. noltii in the Wadden Sea. Their study included field surveys of 

morphology of Z. noltii in exposed and sheltered locations, cross transplantation experiments, 

and flume experiments to manipulate the current environment. These studies revealed that under 

higher flow conditions the density of the seagrass beds declined, and the seagrass leaf and shoot 

lengths became shorter. A significant decline in shoot morphology and density was observed at 

current velocities > 8 cm s-1. Flume experiments conducted by Peralta et al. (2006) demonstrated 

the growth rates and morphometry of Z. noltii was dependent upon flow conditions. At high 

current velocities (35 cm s-1), the root system enlarged, the cross-sections of the rhizomes and 

leaves increased, and the ratio of above to below ground biomass decreased. Similar changes in 

morphometry (decreased leaf width and length) were observed in natural populations of Z. noltii
5.7 

exposed to more energetic environment associated with the opening of a new tidal inlet (Peralta 

et al., 2005). 

 

5.3.3 Effects of Current on Photosynthesis and Growth 

 Some studies have found that seagrass photosynthesis and growth is related to current 

speed. Conover (1964 and 1968) found that oxygen evolution from individual shoots of Z. 

marina increased as current velocity increased for current speeds of up to 40 cm s-1. Conover 

(1968) reported that the standing stock of Z. marina decreased dramatically in currents greater 

than 50 cm s-1. Nixon and Oviatt (1972) found maximum oxygen evolution of an eelgrass 

meadow at about 16 cm s-1. Using a flume with Z. marina plants, Fonseca and Kenworthy 

(1987) found increased leaf production with increasing current velocity (up to 34 cm s-1). Peralta 

et al. (2006) found a similar increase in growth rates for Z. noltii plants in flume for current 

velocities ranging from 1 – 35 cm s-1. Koch (1994) found that the rate of T. testudinum

photosynthesis increased with increasing flow only at low current speeds with saturation 

occurring at about 0.25 cm s-1 (expressed as blade friction velocities, u*) and photosynthesis was 

inhibited under stagnant conditions. The presence of epiphytes resulted in temporal and spatial 

variability in the boundary layer thickness, thereby reducing the potential for boundary layer 

limitation. Field measurements of blade friction velocities revealed that boundary layer 

limitation should only occur during extremely calm conditions and for very short time periods 

(i.e., fractions of second). Even during relatively quiescent conditions, the boundary layer 

thickness oscillates between non-limiting and limiting at high frequency, suggesting that 

diffusional boundary layer limitation is a transient phenomenon in seagrass environments. Koch 

(1994) suggested that there may be a threshold effect for epiphyte cover. It was hypothesized 

that if the epiphyte cover exceeds a certain thickness, then water will flow over the epiphyte 

cover rather than through it, increasing the thickness of the boundary layer and result in further 

limiting conditions (Koch 1994). Enríquez and Rodrígues-Román (2006) found that the 

photosynthetic electron transport rates of T. testudinum was reduced under low flow conditions 

(< 5.4 cm s-1); however, this seagrass also appears to be able acclimate to flow conditions 

reducing their sensitivity to low flow by about 64%. 

 

 Zimmerman (2003) developed a bio-optical model of irradiance distribution and 

photosynthesis in Z. marina and T. testudinum canopies. This model incorporated the effects of 

canopy architecture on light availability for photosynthesis. The biomass-specific 

photosynthesis of the seagrass canopy responded non-linearly to leaf bending angle. When 

seagrass blades are erect with bending angles less than 10°, photosynthesis is limited by the leaf 

orientation. Bending angles greater than 20° limit photosynthesis because a larger fraction of 

light is absorbed by the upper layers of the seagrass canopy where photosynthesis is already light 

saturated. As discussed previously, bending angle is a function of current velocity therefore selfshading increases with increasing current. Zimmerman (2003) proposed that the productionenhancing aspects of flow (e.g., reduction in diffusive boundary layer thickness) may be offset 

by increased self-shading as leaves bend in response to the flow. 
5.8 

At high flow, there is increased 

mounding of the substratum.

(Fonseca et al.,1983)

Low

Medium

High

There is a reduction in seagrass cover 

under high flow conditions (Fonseca et 

al., 1983) and when subjected to unidirectional flow seagrass grow in rows 

(Fonseca et al, 2007). Black and white 

areas indicate seagrass and bare 

regions, respectively 

Morphometry of the plant is 

dependent upon flow 

conditions. Under high flow 

conditions, the leaves are 

shorter and wider and the 

rhizomes increased.

(Peralta et al., 2006)

Cross-Sectional View Aerial View Individual

Figure 5.1. Relationship between configuration of seagrass beds and shoot morphology and 

current regime. At higher current velocities, there is increased mounding of the beds and 

reduction in seagrass cover (Fonseca et al., 1983). Under high flow conditions, the 

shoots become shorter and wider and the rhizomes increase (Peralta et al., 2006). 

5.3.4 Effect of Currents on Nutrient Uptake

 Thomas et al. (2000) conducted flume experiments in natural communities of Halodule 

wrightii and T. testudinum over a range of uni-directional current speeds. They found that 

ammonium uptake by the seagrass community was dependent upon water velocity and was 

influenced by canopy morphology. Thomas et al. (2000) found that there was a decline in 

efficiency ammonium uptake at higher flow velocities when the seagrass bend. Cornelisen and 

Thomas (2002) used flume experiments combined with isotopically-labeled ammonium to isolate 

the effect of water flow on an individual component of a seagrass community. Ammonium 

uptake of both epiphytes and seagrass leaves were positively correlated with current velocity. 

There was a stronger relationship between the epiphytes and flow than with seagrass leaves and 

flow. Seagrass leaves contributed less than epiphytes to the total uptake of the seagrass 

community. 

5.3.5 Effect of Currents on Sediment Geochemistry and Seedling Survival

 Koch (1999a) found that the interaction of currents and porewater chemistry may 

influence the growth and development of T. testudinum seedlings. In this study, seedling 

mortality was higher for stagnant conditions compared to medium and high current velocities. 
5.9 

Seedling morphology was influenced by current velocity with those seedlings exposed to 

stagnant conditions having reduced biomass with shorter blades and roots, while those exposed 

to medium velocities had the longest blades, largest blade area, and greatest biomass and number 

of roots. In addition, nitrogen, phosphorous and sulfide concentrations in the porewater were the 

highest under stagnant conditions. This study suggests that an intermediate level of porewater 

flux may be beneficial to seagrass growth, while stagnant and high flow conditions may 

contribute to lower biomass through increased phytotoxin and reduced porewater nutrient 

concentrations, respectively. 

 The presence of seagrass shoots in current flow can influence the flux of materials 

between the water column and sediments in permeable sediments. In flume experiments, Koch 

and Huettel (2000) found that the presence of T. testudinum shoots enhanced advective 

porewater exchange. At current velocities of 10 cm s-1 there was enhanced flux of porewater 

upstream and downstream of the shoots. The region of sediment influenced by the shoot-flow 

interaction was dependent upon salinity and orientation of the seagrass shoots. 

5.3.6 Effects of Current on Deposition on Leaves

 Currents can affect the amount of suspended solids deposited on leaves, thus influencing 

photosynthesis (Tamaki et al. 2002). Tamaki et al. (2002) performed experiments to determine 

the effect of deposition of suspended solids on Z. marina leaves on light availability and the role 

of current velocity in removing deposited sediments from leaves. They found that the presence 

of deposited suspended solids on eelgrass leaves (at 3 mg cm-2) reduced the light availability by 

as much as 36%. Based on flume experiments, suspended solids were removed from leaves at 

current velocities greater than 8 cm s-1. Field experiments indicated that suspended solids were 

deposited on transplanted eelgrass in the field at levels that would be sufficient to inhibit 

photosynthesis. 

5.3.7 Effects of Currents on Dispersal and Expansion of Population

 Hydrodynamics have the potential to influence the expansion of seagrasses to 

unvegetated regions through dispersal of seeds and reproductive shoots. Orth et al. (1994) 

conducted field and flume experiments to examine the role of currents in the dispersal of Z. 

marina seeds. They found that there was limited seed dispersion with 80-93% of the seeds that 

germinated remaining inside the 5-m plot. The maximum dispersion distance ranged from 4 to 

14 m. Based on flume experiments, the authors suggest that the limited dispersal of the seeds 

resulted from small-scale topographic features on the bottom (such as burrows, pits, mounds, and 

ripples) that shield the seeds from the flow. Although there is limited seed dispersal, Harwell 

and Orth (2002) found that reproductive shoots with mature seeds were positively buoyant for up 

to 2 weeks during which they can be transported relatively large distances by currents. Based on 

field observations, they found that currents can transport reproductive shoots up to 34 km from 

natural beds. 

5.3.8 Effects of Currents on Epiphyte Coverage

 Schanz et al. (2003) found that epiphyte biomass on Z. noltii was highest at seagrass sites 

exposed to water movement (average current speed of 26 cm s-1), while at sheltered (average 
5.10 

current speed of 10 cm s-1) sites epiphyte coverage was negligible. Results of crosstransplantation and enclosure experiments showed that the epiphyte grazer, Hydrobia ulvae, was 

washed off the seagrass blades at exposed sites. Flume studies revealed that grazer density was 

negatively correlated with current speed and epiphyte biomass was positively correlated with 

current speed. 

5.3.9 Summary 

 In summary, seagrass can be limited by both low and high current velocities (Koch 

2001). The minimum and maximum current velocity for Z. marina growth and occurrence based 

on physiological and mechanical limits are a minimum flow of 3 cm s-1and a maximum current 

of 50-180 cm s-1 (Koch 2001). The growth conditions of seagrass as a function of current speed 

reviewed in this chapter are summarized in Figure 5.2. 

Deposition of

Sediment on 

Leaves

 < 8 cm s-1 (3)

Growth Conditions

Current Speed, cm s-1

Optimal

Diffusional

Limitation

< 5 cm s-1(1)

< 5.4 cm s-1(2)

Physical Damage 

to Plants

70 cm s-1 (9)

Erosion of sediment

resulting in exposure of 

roots & rhizomes:

50-150 cm s-1 (9)

Threshold for damage:

 33-85 cm s-1 (10)

Severe damage:

85-110 cm s-1 (10)

Self Shading

 (11)

Increasing Productivity &

Growth with Flow

Flow range of observations

10 to 40 cm s-1 (4,5)

0 to 16 cm s-1 (6)

2 to 34 cm s-1 (7)

1 to 35 cm s-1 (8)

(1) Koch 1994; (2) Enríquez and Rodrígues-Román 2006; (3) Tamaki et al. 2002; (4) Conover 1964;

(5) Conover 1968; (6) Nixon & Oviatt 1972; (7) Fonseca & Kenworthy 1987; (8) Peralta et al. 2006; 

(9) Scoffin 1970; (10) Thom et al. 1996; (11) Zimmerman 2003 

Figure 5.2. Growth conditions of seagrass as a function of current speed. 
5.11 

5.4 Effects of Wave Exposure on Seagrass Growth and Distribution

 Wave exposure can influence seagrass growth and distribution in several ways, including 

plant breakage, erosion of edges of seagrass beds, influencing plant morphology, modifying 

water exchange between canopy and overlying water column, increasing flux between porewater 

and the overlying water column, and causing sediment resuspension which can result in reduced 

light availability. In a recent review, Koch (2001) suggested that seagrass distribution and 

growth appear to be limited by high but not low wave energy. Koch et al. (2002) proposed that 

the upper limit of seagrass beds can be shifted into deeper waters due to wave energy. In high 

wave energy regions, the shallow region is unvegetated due to continuous sand movement (Koch 

et al. 2006a). In the deeper regions, seagrass typically establish below the maximum wave 

penetration depth. 

 Waves can impact seagrasses by influencing the concentration of suspended particulates, 

which may alter the light environment and availability of nutrients. Suspended particulate 

concentrations are usually higher in unvegetated areas compared to adjacent seagrass beds (Ward 

et al. 1984; Koch 1999b; Granata et al. 2001). The amount of suspended particulates is usually 

reduced in seagrass beds due to reduced resuspension and enhanced settling resulting from 

reduced current velocities in the canopy, reduced turbulence energy within the canopy and at the 

sediment surface, attenuation of wave-induced currents, as well as binding of sediments by the 

roots and rhizomes. Sediment resuspension has the ability to influence not only the light 

environment, but the water column nutrient concentration as well. Nutrients associated with the 

sediments may be released to the water column when sediment is resuspended (Morin and Morse 

1999). 

 Fonseca and Bell (1998) examined the effect of physical setting on the distribution and 

abundance of Z. marina and H. wrightii in North Carolina. Their correlative analysis revealed 

that tidal current speeds, exposure to waves and relative water depths influenced landscape-scale 

features of the seagrass beds. Wave exposure was estimated using a relative exposure index 

(REI), which is based on the maximum wind speeds, percent frequency of wind direction, and 

the effective fetch of the site. Percent cover of seagrass, seagrass perimeter to area ratio, 

sediment organic content, and percent silt-clay declined with increasing REI and current speed. 

There were species-specific differences in the effect of physical setting on seagrass distribution 

and abundance. Along a wave exposure gradient, Z. marina was more likely to be found in 

sheltered rather than exposed areas, while the opposite was true for H. wrightii. There was 

increased flowering and belowground biomass of Z. marina in areas with higher wave exposure. 

Fonseca and Bell (1998) also found that a 50% cover corresponded to a transition level for loss: 

Beds with >50% cover survived in chronic and acute storm events; while beds below this 

threshold did not. 

 Using a multiple logistic regression analysis, Krause-Jensen et al. (2003) found that light, 

salinity, and relative wave exposure (REI) were the main factors influencing eelgrass cover in 

Danish coastal waters. Eelgrass cover was inversely related to REI in water depths of 0 to 4 m. 

However, there was a high occurrence (~50% of observations) of absence of eelgrass in areas 
5.12 

that the regression model predicted as suitable habitat. Krause-Jensen et al. (2003) suggested 

that these discrepancies may be associated with stochastic phenomena, such as extreme storms, 

that are not adequately described by average conditions. Recovery of seagrass populations after 

extreme events may take several years. 

 Frederiksen et al. (2004) found that Z. marina distribution in Danish estuaries was related 

to the amount of exposure to wave dynamics. Eelgrass beds at the sheltered sites were 

continuous, while at exposed sites they formed elongated patches. Patches were more complex 

in wave-dominated regions and the large fluctuations in spatial coverage over interannual time 

scales occurred at exposed sites. Shallow eelgrass populations form characteristic landscapes 

with configuration that is related to degree of physical exposure. Aggregated populations may 

be more resistant to physical disturbances than patchy populations, due to the stabilizing effect of 

the root/rhizome matrix and the reduced patch edge exposed to damage from waves or tidal 

currents. Numerous studies have found that the variability of eelgrass populations (i.e., 

variability in shoot density and biomass) is highest in shallow regions where they are subjected 

to frequent disturbance, due to hydrodynamics and other environmental factors (Middleboe et al. 

2003; Krause-Jensen et al. 2000; Krause-Jensen et al. 2003). In addition, the wave environment 

may influence the plant morphology. Krause-Jensen et al. (2000) found that the shoot to 

rhizome ratio increases with depth (more allocation to photosynthetic tissue) as light and wave 

exposure decrease. 

 High wave energy may prevent seagrass from becoming established and reduce survival 

of seagrass transplants (van Keulen et al. 2003; Paling et al. 2003). Most North American 

seagrasses occur in low energy environments, while along the Australia coast there is more 

exposure to wind and swell waves. In Australia, the limited success of seagrass transplantation 

has been attributed to high wave energy conditions. In the North Sea, van Katwijk and Hermus 

(2000) found a negative relationship between Z. marina transplant success and tidal depths, and 

they hypothesized that this was due to water dynamics (waves) and sediment resuspension and 

movement. At a high energy transplant site, where maximum orbital velocity at the sediment 

surface frequently exceeded 60 cm s-1, none of the transplanted plants survived. While at two 

lower energy transplant sites, with mean orbital velocity of 40 cm s-1, survival was related to 

water depth. They proposed that wave action is too severe at water depths deeper than –0.20 m 

mean sea level (MSL) to support establishment of Z. marina. To re-establish Z. marina beds the 

authors recommend providing shelter from wave action (e.g., use of biodegradable dam-like 

structures). 

 Several recent studies have found that the presence of waves substantially influences the 

porewater flux between the sediment and overlying water column (Precht and Huettel 2003; 

Precht and Huettel 2004), which has the potential to influence seagrass productivity. In 

laboratory flume experiments, Precht and Huettel (2003) found that shallow water waves can 

increase the fluid exchange between sandy sediments (no vegetation) and the overlying water 

column as much as 50-fold relative to molecular diffusion rates. They identified two 

mechanisms which increase the flux of porewater, hydrostatic pressure induced wave pumping 

and topography-related filtering. Oldham and Lavery (1999) found an increase in water column 
5.13 

ammonium that they attributed to increased porewater fluxes associated with interactions 

between hydrodynamics (increased currents, waves and turbulence) and the sediment surface. 

 Thomas and Cornelisen (2003) conducted flume experiments in the field in T. testudinum

habitat to examine the effect of unidirectional and oscillatory flow on ammonium uptake rates. 

The uptake of ammonium from the water column by the seagrass community was 1.5 times 

higher in oscillatory flow than in unidirectional flow. Uptake rates were positively dependent on 

water velocity and turbulent energy in the water column. 

5.5 Effects of Turbulence on Seagrass Growth and Distribution

 Increased turbulence can be both beneficial and detrimental to seagrass growth. Benefits 

associated with enhanced turbulence include faster removal of undesirable substances, and 

enhanced transport of nutrients and carbon through the blade boundary layers (Koch 1996). On 

the negative side, increased turbulence can also lead to more sediment resuspension, resulting in 

reduced light availability (Koch 1996). Seagrasses can benefit from increased turbulence 

through enhanced supply of carbon and nutrients across the blade boundary layer and enhanced 

removal of undesirable substances. There have been no studies that show direct linkage of 

turbulence on seagrass growth and distribution. 

5.6 Anthropogenic Modification of Hydrodynamic Stressors

There is the potential for anthropogenic activities to influence the hydrodynamic 

environment, which may influence seagrass distribution and productivity. There has been an 

increase in recreational and commercial boat traffic in coastal waters. Most of the research on 

the impact of boat activities has focused on direct impacts, including propeller scarring and 

vessel groundings. However, seagrasses may also be impacted by hydrodynamic stressors 

associated with boating activity. Boating activities can modify the hydrodynamics through the 

generation of wakes and currents. Potential impacts associated with boat wakes include 

increased sediment resuspension, release of sediment nutrients into the water column, and 

reduced light. Koch (2002) examined the effect of small-boat wakes on environmental 

conditions in a low wave energy region inside Ruppia maritima habitat and found that the 

potential negative impacts were small compared to natural fluctuations in seagrass habitat. There 

was an increase in water column ammonium associated with increased porewater pumping 

associated with increased wave height. A potential benefit of increased wave activity was the 

dislodgement of epiphytes and particulate matter on leaves. 

 Thom et al. (1996) conducted flume experiments to assess the impact of propeller washes 

on Z. marina. Flume experiments were conducted on intact patches of eelgrass subjected to 

current velocities ranging from 0 to 3.25 m s-1. The lower threshold for plant damage, including 

loss of plants and exposure of rhizome and roots, occurred at current velocities between 33 and 

85 cm s-1. Severe damage occurred when the current speeds were between 85 and 110 cm s-1. 
5.14 

 Eleuterius (1987) proposed that seagrass damaged by motor vessel impacts, such as 

propeller scars and vessel landings, are more susceptible to further erosion and scour due to 

hydrodynamic stressors. Whitfield et al. (2002) found that regions of seagrass that had been 

damaged by vessel landings are more vulnerable to storm events. They found that regions of 

seagrass damage increased in size after the passage of a class 2 hurricane, thereby hindering the 

recovery process. In contrast, healthy intact seagrass beds were undamaged by the passage of 

the hurricane. In addition, it has been proposed that climate change may result in increases in the 

impact of hydrodynamic stressors on seagrass communities, due to increased tidal range, 

currents, and storm activity (Short and Neckles 1999).

5.7 Large-Scale Alterations in Estuarine Hydrodynamics

 Large-scale engineering projects that alter estuarine hydrodynamics have the potential to 

affect salinity patterns which in turn can influence the distribution of seagrasses. Nienhuis et al. 

(1996) suggested that alterations in salinity patterns associated with the emplacement of dikes 

may be responsible for some of the observed changes in Z. marina distribution in the 

Netherlands. Eleuterius (1987) proposed that the alteration of freshwater discharges associated 

with the Mississippi River caused the decline of S. filiforme and elimination of H. engelmannii in 

Mississippi Sound. Quammen and Onuf (1993) proposed that a species shift of the seagrasses in 

Laguna Madre, Texas resulted from moderation of the salinity associated with dredging of the 

Gulf Intracoastal Waterway. For a review of salinity effects on seagrasses, see Chapter 4. It has 

been postulated that one of the reasons why Z. marina has not re-established in the Dutch 

Wadden Sea after its decimation by the wasting disease has been due to hydrodynamic stressors 

(De Jonge and De Jong 1992). Hydrologic modifications in the region resulted in an increase in 

tidal range of 15 to 30 cm and increased current velocities by as much as a factor of 3 in some 

regions. Analysis by De Jonge and De Jong (1992) demonstrated that the reduction in light 

associated with increased tidal range (i.e., increased water depth) is not responsible for the 

changes in the underwater light regime. More recently, hydrodynamics has been proposed as a 

contributing factor to the widespread die off of T. testudinum in Florida Bay, Florida that began 

in 1987. It has been postulated that chronic hypersalinity resulting from freshwater diversions 

and alterations in exchange between the bay and the Atlantic and infilling of the bay due to a 

lack of severe storms may be contributing factors to this decline (Fourqurean and Robblee 1999). 

 The natural variability of seagrass populations is large in shallow water where the 

populations are disturbed by wave action and other physical parameters (Krause-Jensen et al. 

2000; Middleboe et al. 2003; Krause-Jensen et al. 2003). The high frequency of perturbation in 

shallow water is expected to cause a wide range of developmental stages (Krause-Jensen et al. 

2000). Because eutrophication has caused a shift in seagrass distributions to shallower 

environments due to light limitation, there may be an increase in the occurrence of 

hydrodynamic stressors influencing the distribution of seagrass since hydrodynamic stressors are 

elevated in shallow environments (Patterson et al. 2001; Middleboe et al. 2003). 
5.15 

5.8 Natural Hydrodynamic Stressor Events 

 Extreme events, such as storms, hurricanes, and floods, have been reported to damage 

seagrass beds (e.g., Eleuterius and Miller 1976; Preen et al. 1995; Aioi and Komatsu 1996). The 

effects of hurricanes on seagrasses vary widely from increased growth (Oppenheimer 1963), no 

visible effect (Tilmant et al. 1994), removal of shoots (van Tussenbroek 1994), massive loss of 

leaf material (Thomas et al. 1961), to complete destruction of the beds (Preen et al. 1995). 

During extreme events, seagrass beds may be damaged by physical destruction of above ground 

biomass, removal of plants by wave action or sand abrasion, and by smothering due to burial by 

sediment deposits. If storm-related seagrass loss occurs, it is usually rapid and localized; 

however, Preen et al. (1995) documented that 1000 km2

 of seagrass was lost from Hervey Bay, 

Australia, following two major floods and a cyclone within a three week interval. The effect of 

hurricanes on seagrass can be localized with some regions showing no effect, while other regions 

are damaged (Van Tussenbroek 1994). In addition to the physical stresses associated with 

storms, there is also usually modification to the environment including reduced salinity and 

elevated suspended particulate concentrations. The occurrence of Tropical Storm Agnes has 

been postulated as a contributing factor to the decline of seagrass in Chesapeake Bay due to 

reduced salinities and increased suspended sediments (Orth and Moore 1983). After a major 

storm disturbance, eelgrass populations may exhibit extensive growth and increased survival of 

new shoots (Krause-Jensen et al. 2000; Aioi and Komatsu 1996). Fonseca et al. (2000) proposed 

that wave-exposed seagrass habitat may be particularly vulnerable to the effects of extreme 

storm events. 

5.9 Research Gaps in Relation to Setting Protective Criteria

 In a recent review of the physical, chemical and geological factors influencing seagrasses, 

Koch (2001) concluded that more data are needed to define current velocity and wave criteria for 

setting protective criteria. In addition, Koch (2001) suggested that the life stage of the plant 

needs to be taken into account when setting protective criteria. Seagrasses may be more 

susceptible to hydrodynamic stressors when they are seedlings versus when they are in mature 

intact beds. It may be difficult to separate out the effect of the hydrodynamics stressors from 

other environmental stressors due to their interactions. Alterations to the hydrodynamics often 

results in changes in the environment (e.g., salinity, turbidity, nutrients, and sedimentation), 

which may in turn control the distribution of the seagrass. Hydrodynamic stressors may be a 

contributing factor to seagrass decline rather than a limiting factor. 

 In order to develop effective protective criteria, we need more research on the effects of 

hydrodynamics on the productivity, survival and distribution of seagrass. One difficulty in 

setting protective criteria for hydrodynamic stressors is that many factors influence the 

interaction between hydrodynamics and the seagrass. Much of the research reviewed in this 

chapter has demonstrated that the presence of seagrass can have conflicting effects on the 

hydrodynamics (particularly on turbulence). In order to further our understanding of the effect 

of hydrodynamic stressors, we need to be sure to collect sufficient ancillary data for 

interpretation of study results. When collecting hydrodynamic data in seagrass habitat, 
5.16

information also needs to be collected on water depth, location of measurements (both within the 

bed and elevation), canopy height, percentage of water column occupied by seagrass canopy, 

seagrass density, size and patchiness of seagrass bed, wind intensity and direction, tides, 

obstructions to flow (such as macroalgae, gorgonians, clams, and epiphytes), and observations on 

interaction of flow and seagrass (e.g., occurrence of monami, skimming flow, and blade 

flapping) (Koch and Verduin 2001). 

 Koch et al. (2006b) suggested that numerical models of the interaction of waves and 

seagrasses are useful for developing testable hypotheses, explaining observations, designing 

observational studies or restoration efforts, and interpolating sparse data in space and time. 

Hydrodynamic and sediment resuspension and transport models have been developed for 

estuarine systems (e.g., Teeter et al. 2001); however, they often require extensive site-specific 

information for model formulation and calibration. Teeter et al. (2001) reviewed that status of 

hydrodynamic and sediment transport modeling in shallow, vegetated regions and recommended 

that more quantitative information was needed on the effect of atmospheric friction and shear 

stress in shallow seagrass regions and more detailed laboratory and field measurements are 

needed to improve model formulations for sediment resuspension within seagrass beds. 

 Although more research is needed on the effect of hydrodynamic stressors on seagrasses, 

several studies include the effect of hydrodynamic stressors in models/indices to predict the 

success of seagrass restoration projects and provide guidance for which areas are suitable for 

restoration. De Jonge et al. (2000) presented a restoration strategy for Z. marina in the Dutch 

Wadden Sea, which included a GIS-based site selection tool for transplantation sites that 

included hydrodynamic stressors in the selection procedure. Selection of suitable transplant sites 

were based on sediment composition, emersion time, current velocity and wave action. Kelly et 

al. (2001) take into account hydrodynamics, in determining which regions you would expect to 

have high probability of restoration success. Short et al. (2002) presented a site-selection tool for 

transplantation of Z. marina which included wave exposure. 
5.17

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5.19

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5.20

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6.1

6.0 Interactions of Zostera marina and Thalassia testudinum with 

Sediments

Peter M. Eldridge, Mark G. Johnson, David R. Young

6.1 Background 

 As rooted aquatic plants, seagrasses are influenced by the sedimentary environment in 

which they grow. Sediment characteristics such as grain size, mineral composition and organic 

matter content may influence the overall biogeochemical environment of the root zone. Human 

actions can also alter sediment grain size distribution (i.e., dredging activities, soil erosion, 

production of excessive fine sediments through jetty and dike construction, prop wash and wakes 

from boats) which in turn may affect the sedimentary geochemical environment and the rooting 

environment. Further, human actions can affect sediment organic matter content through watercolumn nutrient elevations and the subsequent development of phytoplankton, epiphytes and 

macroalgal blooms. The resulting sedimentation of organic matter (OM) to the benthos leads to 

changes in sediment mineralization rates and redox conditions. Of particular concern are 

generation of phytotoxic substances in anaerobic sediments (e.g., elevated sediment ammonia, 

sulfide, metal ions, and other reduced chemical species) (Figure 6.1). Here we review and 

consider the effects of sediment characteristics on the establishment, survival or growth of 

Zostera marina and Thalassia testudinum. 

 Seagrasses interact physically, biologically and geochemically with both the water 

column and the sediment in which they are rooted. These interactions are often complex and 

depend on local conditions (Thayer et al. 1984; Koch 2001). Probably the most important 

physical process affecting seagrass is light. Low-light stress is manifested in a cascade of effects 

that are intimately related to the sediment geochemistry, especially that of sulfide production. A 

key component controlling sediment geochemistry is the input of organic matter that can become 

available for mineralization. Complex hydrodynamic interactions (see Chapter 2) can cause 

seagrass beds to be either sources or sinks for particulate matter (Nepf and Koch 1999) (Figure 

6.1). Because of mineralization of organic matter input to the root zone through burial, DOM 

release from seagrass (Kaldy et al. 2006) and particle retention of irrigating infauna (Eldridge et 

al. 2004), the seagrass roots and rhizomes are nearly always surrounded by anaerobic sediments 

(Eldridge and Morse 2000; Hebert and Morse 2003; Eldridge et al. 2004). However, if anaerobic 

metabolites are not transformed by secondary redox processes to non-toxic forms, phytotoxic 

substances can concentrate around the roots and rhizomes at levels sufficient to kill the seagrass 

(Carlson et al. 1994; Koch 2001). 

6.2 Review of Research 

6.2.1 Interactions of Sediment Grain Size and Zostera

 The majority of information on the relationship of Zostera to sediment grain size results 

is from field surveys (Table 6.1), and thus is correlative in nature. Zostera spp. are typically 
6.2

found growing in substrata of fine or muddy sand, although this genus is also found growing 

over a wide range of sediment size classes. In the Pacific Northwest, Boese and Young 

(unpublished data) found Z. marina growing in a range of sediment texture classes in Yaquina 

Bay, OR. Although a majority of the Z. marina habitat was in the lower estuary associated with 

sediments characterized by 75-100% sand (up to 25% silt/clay), Z. marina meadows also 

occurred up estuary in sediments containing up to 75% silt/clay. Unfortunately, it is not possible 

to determine if there is an optimum grain size composition for Zostera growth and survival with 

these data. In an experimental study of eelgrass using mesocosms, Short (1987) showed that leaf 

biomass, weight, and shoot height were significantly greater in plants growing on fine grained 

mud than on coarse grained sand although the differences between treatments in this case may 

have been due to nutrient limitation. Thom et al. (2001) cultivated Z. marina for 13 weeks in 

sediment types typically occupied by this species in the Pacific Northwest, as well as coarse, 

organic-poor sand and gravel, which typically are not inhabited by Z. marina (Phillips 1984). 

The greatest growth was observed in the finer grained sediments containing organic matter, and 

lowest growth was measured in the gravel substratum. Coarse-grained sand and a sand/gravel 

mixture produced intermediate growth rates. These experimental results are consistent with the 

Z. marina distribution data summarized in Table 6.1, but as in previous experiments, may be 

confounded by differences in nutrient availability or mineralogy. 

6.2.2 Interactions of Sediment Grain Size and Thalassia

 Sediments found in Thalassia beds typically fall in the “medium sand” size class, and 

range from fine to coarse sand (Table 6.1). As is the case for Zostera, grain size distributions 

have been observed to shift toward finer grain sizes in Thalassia beds (Orth 1977). However, 

values for silt and clay typically are low, ranging from 1 to 34%, and the overall median value 

(taking the mid-point of the range for a given study) is 10% (Table 6.1). Terrados et al. (1998), 

reporting from a study of seagrasses in southeastern Asia (including T. hemprichii), observed 

that seagrass species richness and community leaf biomass declined sharply when the silt and 

clay content of the sediment exceeded 15%. Thus, Thalassia would appear to be more 

vulnerable to damage from siltation than is Zostera. 

6.2.3 Interactions of Sediment with Seeds

Zostera seed germination is surprisingly robust with efficiencies often between 80 to > 

90% under a wide range of oxygen and salinity conditions (Brenchley and Probert 1998; Moore 

et al. 1993). Moore et al. (1993) reported that Z. marina seed germination was often triggered by 

anoxic conditions (in either sediment or water). Such conditions typically occur seasonally in 

finer, organic-rich sediments where the exchange between overlying oxygenated water and the 

sediment is restricted (Koch 2001). Seeds buried to 5 mm in the sediments showed lower 

germination success in the autumn than seed buried at 15 to 25 mm, but in the winter there was 

no apparent effect of burial depth in the sediment. In general, germination of buried Z. marina

seeds began when water temperature dropped below 15 degrees C. Moore et al. (1993) also 

pointed out that there was only about a 1-2 week delay between germination in the sediments and 

the emergence of sprouting seedlings. Furthermore, they found no ungerminated viable seeds in 

any of their test treatments after March. That is, seed germination was >90% with about 80% of 

these forming seedlings. This high success rate might be partially due to the fact that only seeds 
6.3

characterized by an intact, hard seed coat of dark color, and a fully developed embryo were used 

in the Moore et al. (1993) experiment. Harwell and Orth (1999) found high (41-56%) seedling 

survival in burlap bag treatments but only 5-15% in bare sediment treatment due to predation, 

burial, or lateral transport. 

Current

Light

Chlorophyll a TSS

DIN

DIP

Epiphytes SOM

TSS

Waves

SO4

- O2 Root O2 Flux MeOx

HSAVS Mex

NO3

- Oxidized

NH4

+ Reduced

Figure 6.1. Both the water column and sediment environments influence seagrasses. While the 

physical and light attenuating water-column stressor may be the most important to the 

survival of the plant, the sediment geochemical processes, stimulated by sedimentation of 

reactive organic material from the water column, can affect seagrass health. Highly 

reduced sulfidic environments (HS-

) can reduce seagrass production and at times can 

become lethal to seagrass due to root death. HS-

 and metals (Mex) combine to produce 

non-toxic acid volatile sulfides (AVS). Thus the combination of high metal (Fe and Mn) 

in the sediments and diffusion of oxygen from the seagrass roots can reduce the toxic 

effects of sulfide in the near root environment (i.e., the rhizosphere) even under 

conditions of relatively high organic matter input to the sediments. (DIN – dissolved 

inorganic nitrogen, DIP – dissolved inorganic phosphorus, SOM – settling organic 

matter) 
6.4

 The typical range of salinity (~15-30) generally found in estuaries also has minimal 

effects on seed germination (Brenchley and Probert 1998), although under aerobic conditions 

and at salinities > 30 there was a small reduction in germination success (60% versus 80%) for 

lower salinities under aerobic conditions. Anaerobic conditions seem to promote greater 

germination success and shorter germination periods (Brenchley and Probert 1998). 

Table 6.1. Distribution of Zostera species in relation to sediment texture or grain size. Texture 

was available in some publications while grain size was reported in others. 

Seagrass Organic 

Matter (%) Size Class Median Grain Size (mm) 

Silt+Clay

(%) Source 

Zostera marina 1.2 14 Marshall and Lucas (1970) 

 1.2 very fine - 

fine sand2 0.100 - 0.1461 Orth (1977) 

 medium - 

muddy sand 0.12 - 0.50 Nienhuis and DeBree (1977) 

2.2 - 2.7 11 - 13 Kenworthy el al. (1982) 

1.7 6 - 24 Peterson et al. (1984) 

0.4 - 1.0 medium-fine 

sand2 0.17 - 0.34 Kenworthy and Fonseca (1992) 

6.4 - 16 Short et al. (1983) 

 fine sand - 

silty clay Lalumiere et al. (1994) 

0.4 - 3.5 5 - 35 Townsend and Fonseca (1998) 

0.4 – 1.4 Dan et al. (1998) 

 medium - 

fine sand2 0.14 - 0.271 8 - 13 Webster et al. (1998) 

 fine sand2 0.17 - 0.201 4 - 6 Frost et al. (1999) 

0.4- 12 2 - 56 Koch (2001) 

 

medium sand 

- fine 

sand/silt 

 Thom et al. (2001) 

 

medium - 

very fine 

sand2 

0.06 - 0.353 Lee et al. (2002) 

Zostera 

 japonica 1.7 - 3.8 fine sand2 0.13 - 0.231,

3 Posey (1988) 

Zostera 

 noltii fine sand2 0.13 - 0.23 40 - 70 van Lent et al. (1991) 

 very fine 

sand2 0.07 Sprung (1994) 

Zostera 

 muelleri 0.8 - 7.3 1 - 72 Edgar and Shaw (1995) 

Zostera 

 novelandica fine sand 0.193 Heiss et al. (2000) 

Thalassia 

 testudinum 3.5 – 4.9 5 Wood et al. (1969) 
6.5

 15 Scoffin (1970) 

 1 – 34 Burrell and Schubel (1977) 

medium 

sand2 0.35 - 0.361 Orth (1977) 

 0.8 medium 

sand2 0.241,

3 3 Grady (1981) 

 fine sand2 0.15 - 0.20 Wanless (1981) 

 12 – 34 Hoskin (1983) 

 22 Lee and Dunton (1996) 

0.5 - 0.6 McGlathery et al. (1994) 

 medium - 

coarse sand2 0.24 - 0.593 Kuenen and Debrot (1995) 

 fine - coarse 

sand2 0.19 - 0.501 Kalbfleisch and Jones (1998) 

 2 - 9 Livingston et al. (1998) 

0.5 – 2.3 2 - 17 Koch (2001) 

 5 - 12 Terrados et al. (1998) 

 23 - 35 Kaldy and Dunton (2000) 

Thalassia 

 hemprichii 3.5 - 10 medium 

sand2 0.24 - 0.371,

3 Paula et al. (2001) 

1. Calculated from phi (ø) values: (ø) = -log2(mm) (Krumbein and Pettijohn 1938) 

2. Wentworth size class obtained from grain size (Wentworth 1922; Percival and Lindsay 1997) 

3. Median grain size 

6.2.4 Seagrasses and sediment sulfides

 Sulfate reduction is quantitatively the most important diagenetic process in anoxic marine 

waters (Blaabjerg et al.1998). The metabolites from sulfate reduction (H2S and HS-

) may inhibit 

seagrass photosynthesis, growth, and survival through sulfide toxicity (Goodman et al. 1995; 

Terrados et al. 1999). The high concentration of sulfate (2700 ppm) in the seawater and within 

the sediment profile insures that sulfate is nearly always available for reduction in seagrass roots 

and rhizomes. Other oxidants (e.g., nitrate and oxy-hydroxy-metals) although more chemically 

energetic than sulfate are often reduced within the first few centimeters below the sediment 

surface (Berner 1980). Because sulfate is abundant deep within the sediments, sulfide 

metabolites may persist in the sediments for several months to years after mineralization. Hence 

the sediment sulfide can provide a chemical record from phytoplankton blooms, dredging or 

other natural or anthropogenic sedimentation (Eldridge et al. 2004). 

To counteract sulfide accumulation in the rhizosphere, seagrasses transport 

photosynthetically produced O2 through lacunae to the roots (Smith et al. 1988; Caffrey and 

Kemp 1991; Kraemer and Alberte 1993). This transport mechanism is probably an adaptation to 

support aerobic root respiration, but excess O2 diffusing from the roots into the rhizosphere has 

the added benefit of oxidizing sulfide to non-toxic sulfate (Caffrey and Kemp 1991). In addition 

to the photosynthetically produced O2, dissolved O2 in the water column can diffuse through Z. 

marina and T. testudinum lacunal systems, thereby reducing the exposure of seagrass to sulfides 

at night when photosynthetic processes are not active (Koch and Erskine 2001). Binzer et al. 
6.6

(2005) showed that in the dark the degree of O2 saturation in water, and flow velocity were the 

primary determinates of internal oxygen conditions in Cymodocea nodosa. Zostera marina may 

also exhibit some anatomical plasticity in the development of lacunae and shoot size that allows 

the plant to adapt to low light conditions (Penhale and Wetzel 1983; Abal et al. 1994) 

Table 6.2. Literature values for dissolved sulfide toxicity to marine/estuarine plants. 

Pore 

Water 

(PW) or 

Water (W) 

Exposure 

Marine/Estuarine 

Plant

Effect 

Conc. 

(μM)

Duration Response Variable Citation 

PW Zostera marina 2300 24 hr root respiration not reduced Penhale and 

Wetzel (1983) 

PW Zostera marina 400 - 800 21 day decreased photosynthesis Goodman et al. 

(1995) 

PW Zostera marina 70 2 mo decreased leaf elongation rate; 

shoot density unaffected 

Terrados et al. 

(1999) 

W Zostera marina 100-1000 1-3 weeks 

in water 

photosynthesis stopped 

leaf elongation rate stopped 

leaves/shoot decreased 37% 

root non-structural carbohydrate decreased 

81% 

above-ground biomass (shoot/root ) 

decreased 55% 

Holmer and 

Bondgaard (2001) 

PW Thalassia testudinum ~ 1,500 4-6 weeks sulfide suggested as synergistic secondary 

stressor 

Carlson et al. 

(1994) 

PW Thalassia testudinum 80 6 mo seedling mortality 100% Koch (1999) 

PW Thalassia testudinum 2000-6000 48 hr 

leaf elongation rate decreased 43% 

root energy charge decreased 22% 

root ATP production decreased 

no visual signs of acute toxicity 

Erskine and Koch 

(2000) 

PW Thalassia testudinum 6000 14 day 

high salinity (S), 50% mortality 

high temperature (T), 33% mortality 

high S and T, 100% mortality 

Koch and Erskine 

(2001) 

PW Thalassia testudinum 5500 38 day net shoot loss 65% at 34-35°c 

no new leaf emergence at 34-35°c Koch et al. (2007) 

 In marine environments, there is a constant composition of major ions relative to salinity 

so that in high salinity waters there is, as discussed above, an abundance of sulfate to support 

anaerobic diagenesis (Stumm and Morgan 1981). Owing to the potential for sulfide formation, 

plants in euryhaline and mesohaline waters may require sediments which are more oxygenated 

(i.e., coarser sediments) and have higher rates of pore water exchange (Koch 2001) than is 

needed in brackish waters to reduce the effect of sulfide toxicity. 
6.7

 Although sulfate competes poorly as an oxidant, its abundance in seawater ensures that 

sulfate reduction is a dominant geochemical process in near shore sediments (Berner 1980). 

However, the extrapolation of the general trends in near shore sediment diagenesis to seagrass 

sediments is not direct. While Holmer and Bongaard (2001) found that the depth distribution of 

sulfate reduction in a tropical seagrass (Cymodocea rotundata) was positively correlated with 

below-ground biomass, they also found that sulfate reduction was not the major diagenetic 

process leading to nutrient re-mineralization, possibly because of the ability of seagrasses to 

inject O2 and other oxidants into the root zone. 

The ability of seagrasses to promote mineralization of organic material to release NH4

+

while regulating sulfate reduction in the root zone may be an important mechanism to insure 

seagrass survival (Eldridge et al. 2004). The regulation of sulfate reducing bacteria by 

seagrasses may have other benefits. Sulfate reducing bacteria often fix nitrogen (Welsh et al. 

2001), which may lead to increased nitrogen availability and which can also be beneficial to the 

growth of seagrasses. By regulating sulfate reduction the plant can maximize the benefit of 

sulfate reduction while minimizing its toxic effects. Other complicating factors that may 

mitigate the effects of sulfide concentration include the presence of iron (Fe(II)) and manganese 

(Mn2+) in the rhizosphere which may also reduce sulfide toxicity by the formation of insoluble 

metal sulfides (Erskine and Koch 2000). This might support the hypothesis that accumulation of 

sulfides could be mitigated by fertilization of seagrass habitats with Fe(III) compounds. An 

advantage of this proposition would be competitive anaerobic mineralization of organic matter 

by Fe(III) instead of sulfate. The products of iron reduction are not thought to be toxic to Z. 

marina. 

 In terrestrial systems, root exudates (i.e., plant-derived organic compounds added to the 

rhizosphere) can bind with toxic metal ions or other compounds to provide protection under 

chemically adverse soil conditions (Höberg and Jensén 1994). It is possible that seagrasses have 

similar mechanisms to protect against toxic chemicals in sediments; however, little is known 

about root exudation, root growth, and root longevity and mortality in seagrasses. 

 Several studies report on the tolerance of seagrass to soluble sulfides (Table 6.2). While 

there is some evidence that 1 to 2 mmol H2S in the water-column may inhibit seagrass growth or 

cause death of Z. marina (Goodman et al. 1995; Smith et al. 1988), there are few definitive 

studies that show the dose-response of seagrass in long-term controlled-environment studies of 

exposure to sulfides. Holmer and Bondgaard (2001) showed that photosynthesis stopped after 6 

days at water column sulfide concentrations between 100 and 1000 μmol. This study was done 

at saturating irradiance (400-500 μmol photon m-2 s-1). Koch and Erskine (2001) determined the 

response of Thalassia testudinum to sulfides under varying conditions of light, salinity, and 

temperature. T. testudinum showed no response to sulfides until the dose reached 6 mM 

concentration, and a response occurred only in the high temperature treatment (35oC). In 

contrast, longer-term studies by Carlson et al. (1994) showed T. testudinum died after being 

exposed to ~1.5 mM sulfide concentrations. However, Carlson et al. (1994) note that the high 

sulfides at their research sites are probably just one of several factors that contribute to die-off 

episodes rather than the primary cause of death. Other factors such as hyperthermia, 
6.8

hypersalinity and microbial pathogens may act in synergy with high sulfides to induce mortality 

in seagrass (Carlson et al. 1994). 

 There are several internal plant mechanisms through which sulfides can affect seagrass 

survival. Sulfides bind to metal ion cofactors of proteins, inhibiting their activities. Sulfides 

bind the Fe(III) in the heme moiety of the mitochondrial enzyme cytochrome a3 oxidase, 

blocking the terminal step in the electron transport system. Sulfides can also bind Zn(II) in 

carbonic anhydrase replacing the bound hydroxyl necessary for the inter-conversion of CO2 and 

water to bicarbonate (Beauchamp et al. 1984). Currently, there is little information on how 

various plant organelles individually respond to sulfides. In particular, more information is 

needed to determine the effect of sulfides on seagrass plant meristems since this is the site of 

most anabolic processes. Further we have found no studies that address pH effects on sulfide 

toxicity. The pH effect may be important since it affects the speciation of the sulfide. HS-

 has 

been shown to be more toxic to some faunal species than the other sulfide species (Stumm and 

Morgan 1981). New microelectrode methods are now available to measure sulfide concentration 

in the roots, rhizomes, and meristem (Pedersen et al. 2004) making possible toxicological studies 

of seagrass response to sulfides in the meristem and other tissues. 

6.2.5 The Role of the Infaunal Irrigators

 The benefit of seagrass as a source of nutrition and refuge for infaunal communities has 

been demonstrated in numerous studies (e.g., Bostrom et al. 2002; Mattila et al. 1999; Webster et 

al. 1998). Fewer studies, however, have demonstrated the benefits that seagrasses derive from 

the presence and activity of infaunal organisms. Peterson and Heck (2001) found a positive 

relationship between infaunal nutrient cycling and seagrass productivity. Based on modeling 

studies of Thalassia testudinum, Eldridge et al. (2004) suggest that seagrasses derive additional 

benefits from irrigating infauna through the introduction of oxidants from the water column into 

the root zone. The additional oxidants in the rhizosphere help maintain low levels of sulfides 

and other reduced toxicants. A more detailed discussion of the interactions of infaunal irrigators 

and seagrasses is found in Chapter 11. 

6.2.6 Trapping of Particles by Seagrasses

 Seagrasses produce more organic matter than can be consumed by water-column, 

epifaunal, and infaunal organisms (Kaldy et al. 2006; Jones et al. 2003). Much of this organic 

matter becomes sequestered in the sediments (Eldridge and Morse 2000). Particulate deposition 

is further enhanced by the capacity of seagrass to directly retain sestonic particles. The three 

dimensional structure of the seagrass canopy buffers the effects of current velocity and 

turbulence within the canopy (Koch 2001) (Figure 6.1) thereby reducing sediment resuspension, 

total suspended solids concentrations and increasing water clarity. In addition to the settling of 

particles due to decreased turbulence and water flow (Koch 2001), particles physically adhere to 

seagrass leaf surfaces or are trapped by protozoa and possibly other epiphytes that reside on 

leaves. These trapping mechanisms may be the dominant particle sequestration mechanism in 

seagrass canopies, and add significantly to the high rates of organic carbon input into seagrass 

sediments (Agawin and Duarte 2002).
6.9

 Field studies have shown that surficial sediments within seagrass meadows tend to have a 

higher percentage of “fine” material (silt/clay), and a higher organic carbon content, than the 

surrounding non-vegetated (or lightly vegetated) sediments. For example, Marshall and Lukas 

(1970) found that surficial (0-1 cm) sediments from an eelgrass bed in Rhode Island averaged 14 

+ 6% silt/clay, compared to 5 + 3% in unvegetated sediments. Corresponding values for organic 

carbon in eelgrass and unvegetated sediments were 1.89 + 0.34% and 0.88 + 0.17%. Orth (1977) 

reported that the median sediment diameter for sediment cores collected within a Chesapeake 

Bay eelgrass bed was about 0.10 mm (3.3 phi units), compared to about 0.17 mm (2.6 phi units) 

at the edges of the bed. Corresponding values for percent total organic matter were 1.4 and 

0.5%. In another study, Peterson et al. (1984) reported that silt/clay comprised 14-18% of 

sediments from a Z. marina meadow in North Carolina versus 2-3% from the adjacent control 

sand flat. Variations in sediment texture generally reflect differences in physical processes 

related to waves and currents (Burrell and Schubel 1977). Seagrasses often impede currents, 

reduce the flow velocity (Ginsburg and Lowenstam 1958), and increase the accumulation of 

silt/clay fraction affecting the sorting and skewness of the grain size distribution (Wood et al. 

1969; Burrell and Schubel 1977; Fonseca 1981; Kenworthy et al. 1982; Peterson et al.1984; 

Gambi et al.1990; Ackerman and Okubo 1993).

 Fonseca et al. (1983) found no predictable distribution of silt/clay in low surface current 

regimes (up to 53 cm/sec) in Z. marina meadows in North Carolina, while in high currents (up to 

94 cm/sec) there was an inverse relationship between fine sediment content and shear velocity. 

Expanding on this work, Fonseca and Bell (1998) found that sediment composition in seagrass 

habitats was highly variable below a current velocity of about 25 cm/sec, while this variance was 

much reduced at higher current speeds. They suggested that the initiation of motion of sediment 

for the fine sand, characteristic of the North Carolina site, occurs at unidirectional current speeds 

of about 25 cm/sec, and that this speed constitutes a disturbance threshold for silt/clay and 

organic content there. Further, they suggested that at velocities greater than 25 cm/sec there was 

a decreased accumulation of fine sediment that could reduce vegetative spreading and inhibit 

seedling colonization of eelgrass. 

 Seasonality in temperate regions may be important to both the flushing of organic matter 

out of the seagrass beds (Hemminga and Duarte 2000) during winter and the accumulation of 

seagrass or macroalgal biomass in the autumn (Carlson et al.1994). These disturbances may be 

important for the long term survival of seagrass by preventing the deterioration of sediment 

condition (Hemminga and Duarte 2000) or by providing organic material which, upon 

mineralization, provides sufficient NH4

+

 to avoid a nitrogen limitation (Zimmerman et al. 1987).

6.2.7 Sediment Nutrient Effects

 Seagrasses assimilate nutrients from both the water column and sediments. There is 

some debate as to which nutrient source is more important. Active nutrient uptake from 

sediments occurs during daylight hours when the rhizosphere is aerobic. However, during the 

dark period when roots and rhizomes are often subject to fermentation, active below-ground 

uptake is reduced and shoot uptake may become more important. The ability of seagrasses to 

regulate uptake processes depends on a number of factors including the rate of photosynthesis 
6.10

and utilization of stored carbohydrates, the availability of NH4

+

 and NO3

-

 in the water-column, 

and the rate of organic sediments mineralization. (Pedersen et al. 1997; Koch 2001; Eldridge et 

al. 2004; Kaldy et al. 2006). 

 Mineralization is an important source of nitrogen and phosphorous to plants (Short 1987; 

Perez et al. 1991; Perez et al. 1994; Pederson et al. 1997). Mineralization rates may be regulated 

to some degree by the seagrass plant. In sediments with low organic matter content (less than 

about 2.5%), seagrasses receiving high irradiance at the canopy can regulate the sediment redox 

environment with lacunal O2 diffusion from the roots and rhizomes. Maintaining healthy redox 

conditions in the sediments allows the plant to maintain normal physiological nutrient uptake and 

photosynthetic capacity (Koch 2001 and references within; Eldridge et al. 2004). Sulfide 

concentrations (Hebert and Morse 2003) and sulfate reduction rates (Blaabjerg et al. 1998) 

showed diel cycles with greater H2S concentrations during the night and, surprisingly, higher 

sulfide production rates during the day. High sulfate reduction rates during periods of maximum 

photosynthesis suggests that excess carbon produced by primary production is released as DOM 

from the roots and stimulates sulfate reduction. Simultaneously, lacunal O2 oxidizes this 

released sulfide. Radioactive and stable isotope data from several studies suggests that there is a 

strong linkage between seagrass production and DOM release (Holmer and Laursen 2002; Kaldy 

et al. 2006). Using tracer experiments, Kaldy et al. (2006) showed a direct link from seagrass to 

DOM, and then to sediment bacteria, and Holmer and Laursen (2002) showed a positive 

relationship between seagrass photosynthesis and sulfate reduction. A conclusion we reach from 

these result, is that diel variations in seagrass photosynthesis produce pulses of DOM exudates 

from the seagrass roots and rhizomes that simulate daily cycles in sediment sulfate reduction. 

 The interaction between seagrass production and organic matter in the sediment is highly 

variable, and there are studies showing that healthy seagrass can occur in highly enriched organic 

sediments (Koch 2001) (Table 6.1). We assume that either the organic matter in these sediments 

is relatively unreactive or that infaunal irrigation (as discussed earlier) and lacunal O2 release 

aerates the sediments in the vicinity of the seagrass. We note, however, that both infaunal 

irrigation and lacunal O2 release also involves excretion of labile DOC. The literature is unclear 

as to how these metabolites alter the sediment geochemistry (Holmer and Laursen 2002; Kaldy et 

al. 2006). 

 Ammonium, the preferred form of nitrogen for eelgrass, is taken up from pore water 

through the roots (Short 1987). Dennison et al. (1987) established an upper limit for this uptake 

and utilization, finding in mesocosm experiments that interstitial water concentrations of NH4

+

above 100 FM saturate the growth response of Z. marina. In contrast, Van Katwijk et al. (1997) 

reported that an NH4

+

 concentration of 125 FM was toxic to eelgrass, and suggested that the 

toxicity threshold was as low as 25 FM. However, Hebert et al. (2007) reported up to 2500 FM 

concentration of NH4

+

 in the root zone of healthy Z. marina in Yaquina Bay, OR. 

 Maier and Pregnall (1990) showed that Z. marina also utilizes nitrate-rich groundwater 

flowing through a permeable sand layer sandwiched between fine-grained sediments into the 

near shore waters from sandy beaches in Massachusetts. The nitrate-rich groundwater induced 
6.11

nitrate reductase activity in eelgrass plants. This finding suggests the importance of high-porosity 

sandy sediments as a conduit for nutrient-rich ground water to near shore aquatic vegetation, 

including seagrass and macroalgae. The interlayering of permeable sandy sediments with finegrained sediments in a Georgia estuary provided conduits for advective transport of pore water 

constituents out of the sediments (Jahnke et al. 2003). They concluded that such fluxes are 

concentrated into small layers, and as such may constitute a significant input of nutrients to the 

estuary even if permeable, sandy layers comprise a very small proportion of the total bottom 

area. Again, these findings indicate a relationship between sediment grain size and the rate at 

which nutrient-rich ground water can seep into near shore waters, stimulating the growth of 

seagrass and algae. The relationship between aqueous concentrations of nutrients and the 

abundance and health of Zostera is discussed in detail in Chapter 3. 

 Reduction in surficial sediment grain size can decrease the rate of exchange of pore water 

nutrients with overlying water. For example, Short (1983) investigated the accumulation of 

NH4

+

 in interstitial water and onto sediment particles. He found that in highly-reduced eelgrass 

sediments, NH4

+

 is lost from the interstitial pool by diffusion into the thin, oxidized sediment 

surface layer, by adsorption onto sediment particles, and by uptake into bacterial cells and 

eelgrass roots. However, part of this pore water NH4

+

 is termed “exchangeable” because it is 

easily released from the particle surfaces by ion exchange. Short noted that the amount of 

exchangeable NH4

+

 is dependent upon texture, mineralogy, and organic content of the sediment. 

6.3 Research Gaps in Relation to Setting Protective Criteria 

 Both field and laboratory results indicate that Zostera marina and other seagrass species 

are most abundant or productive in fine or muddy sand containing substantial organic matter. 

This type of sediment can contain elevated pore water concentrations of substances such as NH4

+

or dissolved sulfides. Whether or not these constituents act as nutrients or toxins to eelgrass 

plants may depend upon the pore water concentrations, other characteristics of the sediment, and 

the physiology of the exposed plants. Thus, although grain size of the substratum does appear to 

influence the distribution and health of Zostera marina, relatively little is known of the specific 

processes involved in such effects. The percent organic matter in sediments is related to the 

sulfate reduction potential, and hence to the sulfide concentrations in the sediment. The ability 

of seagrass to protect itself from high levels of sulfide in the root zone will be directly dependent 

on availability of light to drive photosynthesis, and may be indirectly dependent on irrigating 

infaunal associates or to the presence of Fe or Mn minerals that detoxify the sulfides in seagrass 

sediments. Quantifying bioirrigation effects and better definition of the relationship between 

available light levels and sulfide concentrations would be helpful in insuring that protective 

criteria based on light levels will be adequate. Additionally, the presence of metal minerals (as 

detoxifying agents) could be used as an evaluation factor in seagrass protective criteria. The 

reported sediment pore water concentrations of NH4

+

 tolerated by Z. marina range over two 

orders of magnitude. Again a better understanding of the relationship between dissolved 

inorganic nutrients and Z. marina physiology is needed. 
6.12

6.4 Literature Cited

Abal, E. G.,N. R. Loneragan, P. Bowen, C. J. Perry,, J. W. Udy, and W. C. Dennison. 1994. 

Physiological and morphological responses of the seagrass Zostera capricorni Ashers. to 

light intensity. Journal of Experimental Marine Biology and Ecology 178: 113-129. 

Ackerman, J. D. and A. Okubo. 1993. Reduced mixing in a marine macrophyte canopy. 

Functional Ecology 7:305-309. 

Agawin, N. S. R and C. M. Duarte. 2002. Evidence of direct particle trapping by a tropical 

seagrass meadow. Estuaries 25:1205-1209. 

Beauchamp, R. O., Jr, J. S. Bus, J. A. Popp, C. J. Boveiko, and D. A. Andjelkovich. 1984. A 

critical review of the literature on hydrogen sulfide toxicity. CRC Critical Reviews in 

Toxicology 13:25-97. 

Berner, R. A. 1980. Early Diagenesis: A Theoretical Approach. Princeton University Press. 

Princeton, N.J. 

Binzer, T., J. Borum, and O. Pedersen. 2005. Flow velocity affects internal oxygen conditions 

in the seagrass Cymodocea nodosa. Aquatic Botany, 83: 239-247. 

Blaabjerg, V., K. N. Mouritsen, and K. Finster. 1998. Diel cycles of sulphate reduction rates in 

sediments of a Zostera marina bed (Denmark). Aquatic Microbial Ecology 15: 97-102. 

Bostrom C., E. Bonsdorff, P. Kangas, and A. Norkko. 2002. Long-term changes of a brackishwater eelgrass (Zostera marina L.) community indicate effects of coastal eutrophication. 

Estuarine Coastal and Shelf Science 55:795-804. 

Brenchley, J. L. and R. J. Probert. 1998. Seed germination responses to some environmental 

factors in the seagrass Zostera capricorni from eastern Australia. Aquatic Botany 63:177-

188. 

Burrell, D. C. and J. R. Schubel. 1977. Seagrass ecosystem oceanography. Pp. 195-232. In: C. 

P. McRoy and C. Helfferich (eds.), Seagrass Ecosystems. Marcel Dekker, New York, 

NY. 

Caffrey, J. M. and W. M. Kemp. 1991. Seasonal and spatial patterns of oxygen production, 

respiration and root-rhizome release in Potamogeton perfoliatus L. and Zostera marina

L. Aquatic Botany 40:109-128. 

Carlson, P. R., Jr., L. A. Yarbro, and T. R. Barber. 1994. Relationship of sediment sulfide to 

mortality of Thalassia testudinum in Florida Bay. Bulletin of Marine Science 54:733-

746. 

Dan, A, A. Moriguchi, K. Mitsuhashi, and T. Terawaki. 1998. Relationship between Zostera 

marina and bottom sediments, wave action offshore in Naruto, Southern Japan. Fisheries 

Engineering 34: 229-204. 

Dennison, W. C., R. C. Aller, and R. S. Alberte. 1987. Sediment ammonium availability and 

eelgrass (Zostera marina) growth. Marine Biology 94:469-477. 

Edgar, G. .J. and C. Shaw. 1995. The production and trophic ecology of shallow-water fish 

assemblages in southern Australia. III. General relationships between sediments, 

seagrasses, invertebrates and fishes. Journal of Experimental Marine Biology and 

Ecology 194:107-131. 
6.13

Eldridge, P. M., J. E. Kaldy, and A. Burd. 2004. Stress response model for the tropical seagrass 

Thalassia testudinum: The interactions of light, temperature, sedimentation, and 

geochemistry. Estuaries, 6: 923-937. 

Eldridge, P. M. and J. W. Morse. 2000. A diagenetic model for sediment-seagrass interactions, 

Marine Chemistry 70:89-103. 

Erskine, J. M. and M. S. Koch. 2000. Sulfide effects on Thalassia testudinum carbon balance 

and adenylate energy charge. Aquatic Botany 67:275-285. 

Fonseca, M. S. and S. S. Bell. 1998. Influence of physical setting on seagrass landscapes near 

Beaufort, North Carolina, USA. Marine Ecology Progress Series 171:109-121. 

Fonseca, M. S., J. C. Zieman, G. W. Thayer, and J. S. Fisher. 1983. The role of current velocity 

in structuring eelgrass (Zostera marina L.) meadows. Estuarine, Coastal and Shelf 

Science 17:367-380. 

Fonseca, M. S. 1981. The interaction of a seagrass, Zostera marina L. with current flow. M.Sc. 

Thesis, University of Virginia, Charlottesville, VA, pp. 35. 

Frost, M. T., A. A. Rowden, and M. J. Attrill. 1999. Effect of habitat fragmentation on the 

macroinvertebrate infaunal communities associated with the seagrass Zostera marina L. 

Aquatic Conservation: Marine and Freshwater Ecosystems 9:255-263. 

Gambi, M. C., A. R. M. Nowell, and P. A. Jumars. 1990. Flume observations on flow dynamics 

in Zostera marina (eelgrass) beds. Marine Ecology Progress Series 61:159-169. 

Ginsburg, R. N. and H. A. Lowenstam. 1958. The influence of marine bottom communities on 

the depositional environment of sediments. Journal of Geology 66:310-318. 

Goodman, J. L., K. A. Moore, and W. C. Dennison. 1995. Photosynthetic responses of eelgrass 

(Zostera marina L.) to light and sediment sulfide in a shallow barrier island lagoon, 

Aquatic Botany 50:37-47. 

Grady, J. R. 1981 Properties of sea grass and sand flat sediments from the intertidal zone of St. 

Andrew Bay, Florida. Estuaries 4:335-344. 

Harwell, M. C. and R. J. Orth. 1999. Eelgrass (Zostera marina L.) seed protection for field 

experiments and implications for large-scale restoration. Aquatic Botany, 64:51-61. 

Hebert, A. B., and J. W. Morse. 2003. Microscale effects of light on H2S and Fe2+ in vegetated 

(Zostera marina) sediments. Marine Chemistry 81:1-9. 

Hebert, A. B., J. W. Morse, and P. M. Eldridge. 2007. Small-scale heterogeneity in the 

geochemistry of seagrass vegetated and non-vegetated estuarine sediments: causes and 

consequences. Aquatic Geochemistry, 13:19-39. 

Heiss, W. M., A. M. Smith, and P. K. Probert. 2000. Influence of the small intertidal seagrass 

Zostera novazelandica on linear water flow and sediment texture. New Zealand Journal 

of Marine and Freshwater Research 34:689-694. 

Hemminga, M. A. and C. M. Duarte. 2000. Seagrass Ecology. Cambridge University Press, 

Cambridge, UK. pp. 298. 

Höberg, P. and P. Jensén. 1994. Aluminum and uptake of base cations by tree roots: a critique of 

the model proposed by Sverdrup et al. Water Air and Soil Pollution 75:121-125. 

Holmer, M. and L. Laursen. 2002. Effect of shading of Zostera marina (eelgrass) on sulfur 

cycling in sediments with contrasting organic matter and sulfide pools. Journal of 

Experimental Marine Biology and Ecology 270:25-37. 
6.14

Holmer, M. and E. J. Bondgaard. 2001. Photosynthetic and growth response of eelgrass to low 

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6.16

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U.S. EPA.

7.1

7.0 The Interaction of Epiphytes with Seagrasses under Nutrient 

 Enrichment 

Walter G. Nelson

7.1 Background

 The surfaces of seagrass blades are often covered by a variable layer of epiphytes, which 

may respond in complex ways to variation in environmental nutrient concentrations and light 

availability. This chapter reviews the observational and experimental literature through 

approximately 2004 on the effects of nutrient enrichment on seagrass epiphytes. The goal was to 

determine whether there is an adequate understanding of the effects of epiphytes on seagrass 

growth and survival to allow incorporation of this factor in criteria protective of seagrasses. 

 Seagrass epiphytes are a diverse, mixed assemblage (Harlin 1980) of macroalgae 

(Phillips 1960; Humm 1964; Ballantine and Humm 1975; Hall and Eiseman 1981; Hall 1988), 

microalgae (e.g. Sand-Jensen 1977; Sand-Jensen and Borum 1984), and a variety of sessile 

animal groups including polychaetes, bryozoans, hydroids and tunicates (Kita and Harada 1962; 

Nagle 1968; Lewis and Hollingworth 1982). The term "epiphytes" will be used to refer to the 

combined microalgal, macroalgal, animal and inorganic components covering seagrass blades. 

Major reviews of seagrass epiphytes include Harlin (1980) and Borowitzka and Lethbridge 

(1989), while van Montfrans et al. (1984), Orth and van Montfrans (1984) and Jernakoff et al. 

(1996) summarized knowledge on epiphyte-seagrass interactions with a particular emphasis on 

micrograzing interactions. 

 While epiphytic algae may have some beneficial effects on seagrasses (Orth and van 

Montfrans 1984; Brandt and Koch 2003), negative impacts appear to predominate (Borowitzka 

and Lethbridge 1989). These effects include: 1) reduction in light available for photosynthesis, 

2) a reduction in the rate of diffusion of materials such as CO2 across the seagrass blade surface, 

and 3) an increase in physical drag, resulting in increased loss of leaves or plants. It has been 

suggested that seagrass leaves with heavy epiphyte cover may become more brittle and break off 

(Borowitzka and Lethbridge 1989), although quantitative data supporting this effect are limited 

(e.g. Heijs 1985). Harlin (1975) suggested that epiphytes may compete with seagrass for water 

column nutrients, but the magnitude of any effect should be minor relative the main effects listed 

above. Suggested benefits of epiphytes include serving as a UV-B filter, which might be most 

important in tropical, oligotrophic waters (Trocine et al. 1981; Brandt and Koch 2003), and as a 

factor potentially limiting desiccation damage for plants in the upper intertidal zone (Penhale and 

Smith 1977, and see Chapter 11). 

 Epiphytes are patchily distributed on seagrass blades (Figure 7.1), and are typically more 

abundant on the distal portions of all blades (Figure 7.2), and most abundant on the oldest blades 

within a plant. The mean life span of leaves will influence the degree to which epiphyte biomass 

can build up on seagrass blades. Typical life spans are Z. marina (27 - 63 d), H. wrightii (34 d), 

and T. testudinum (24 - 50 d) (Borowitzka and Lethbridge 1989). Rates of blade turnover for 
7.2

seagrass can vary widely across seasons, and for Z. marina may range from 50-70 days in 

summer to a maximum leaf age of 200 days in winter (Borum et al.1984). In some cases, low 

blade turnover rates can result in higher epiphytic biomass accumulation, and hence in increased 

light attenuation. Dixon and Kirkpatrick (1995) observed that light attenuation by epiphytes was 

highest during winter months in Sarasota Bay when leaf turnover rates were reduced. However, 

at higher latitudes where winter light may be limiting, winter may be a period of low epiphytic 

biomass in spite of lower blade turnover rate (Williams and Ruckelshaus 1993; Nelson and 

Waaland 1997). 

 Although there have been suggestions that seagrasses such as Z. marina contain 

compounds in the leaf tissue that may inhibit settlement by epiphytes, Borowitzka and 

Lethbridge (1989) concluded that there is little evidence that seagrasses have any means of 

inhibiting epiphyte colonization and growth. 

 Borowitzka and Lethbridge (1989) reviewed the evidence that translocation of nutrients 

from seagrass to epiphytes may occur. They suggested that the work by Brix and Lyngby (1985) 

showed that earlier reports of high rates of nutrient release from seagrass leaves was probably the 

result of methodological problems. Rates of nutrient release reported from seagrasses are too 

low to support levels of epiphyte growth generally observed, so the major source of nitrogen and 

phosphorus appears to be from the water column (Borowitzka and Lethbridge 1989). Thus 

seagrasses function primarily as a substratum supporting the growth of epiphytes, not as a 

primary source of nutrients. 

7.2 Review of Relevant Research

7.2.1 Epiphyte Loads and Limitation of Seagrass Growth and Distribution

 The dominant effect of heavy epiphytic cover appears to be decreased seagrass growth 

and a reduced potential for survival caused by reduced light availability (Sand-Jensen 1977; 

Borum and Wium-Anderson 1980, Bulthius and Woelkerling 1983; Cambridge et al. 1986; 

Silberstein et al. 1986; Sand-Jensen and Revsbach 1987). Epiphytic shading has been suggested 

to be particularly important at lower ambient light levels (Morgan and Kitting 1984; Twilley et 

al. 1985; Wetzel and Neckles 1986). 

 Cambridge et al. (1986), working with seagrasses from Cockburn Sound, Australia, 

provided one of the first suggestions that high epiphytic loads resulting from eutrophic 

conditions directly causes loss of seagrasses. Transplantation experiments with Posidonia 

sinuosa seedlings resulted in leaf area ~60% lower for plants growing in Cockburn Sound after 

48 d. Seedlings were heavily covered with macroepiphytes. Differences in leaf area may have 

been partly due to senescence and breakage of leaf tissue during a storm rather than strictly an 

effect of growth reduction, however. The study noted but did not quantify the presence of 

macroalgal blankets up to 1 m thick in Cockburn Sound, and the relative role of these 

macroalgae versus epiphytes in seagrass loss is unclear. 
7.3

Figure 7.1. Schematic diagram of seagrass plant illustrating typical patterns of distribution of 

epiphytes within and among blades. Epiphyte cover increases from base to tip and from 

youngest to oldest blades. 

 For the freshwater macrophyte (Littorella uniflora) in Danish lakes, Sand-Jensen (1990) 

found a relation of epiphytic load, nutrient load, and depth distribution. Maximum depth of 

macrophyte distribution was inversely related to epiphytic load, and was 0.2 m in the eutrophic 

lake, versus 2.2 m in the most oligotrophic lake. Mean light attenuation due to epiphytes was 

82% (range 64-100%) in the eutrophic lake, and this factor accounted for 62% of the total (water 

column + epiphyte) light attenuation at the leaf surface, versus only 5% at the most oligotrophic 

lake. 

 Studies of light attenuation by epiphytes accumulating on seagrasses or on seagrass 

mimics (glass slides or plastic strips) indicate that the presence of epiphytes can lead to nearly 

100% attenuation of incident light to the surface of individual seagrass blades (Table 1). When 

averaged over entire plants to account for heterogeneity of epiphyte distribution on blades of 

different ages, average light reduction to a seagrass plant is more typically 50-60% (Harden 

1994). Annual averages of light reduction, resulting from seasonal variation in epiphytic load, 
7.4

Figure 7.2. Horizontal arrows indicate external factors influencing abundance of seagrass 

epiphytes. The vertical arrow indicates the variation in important factors associated with 

epiphytes along the transition from the basal to distal portions of seagrass blades. 

were in the range of 32-36% for Thalassia testudinum at the deep edge of its distribution in 

Tampa Bay, Florida (Dixon and Kirkpatrick 1995; Dixon 2000). Monthly averages of light 

reduction ranged from 17-59%, while shallower sites where light was not limiting tended to have 

greater light attenuation by epiphytes (Dixon 2000). 

 A review of seagrass light attenuation studies (Brush and Nixon 2002), suggests that 

studies which have estimated light attenuation using epiphyte suspensions (see Table 7.1) may 

overestimate the attenuation of light for high epiphytic loads. They observed that highest 

epiphytic biomass tended to be generated by larger arborescent algae which tended to float away 

from the seagrass blade when submersed, thus allowing more light to reach the blade than would 

be predicted by the suspension approach. 

 There are few direct measurements of the effect of epiphytes on photosynthesis rates. 

Sand-Jensen (1977) demonstrated that diatom epiphytes reduced the photosynthetic rate of Z. 

marina both by acting as a barrier to carbon uptake and by reducing light intensity. At constant 

illumination, reduction of leaf photosynthesis was a function of bicarbonate concentration, with a 

maximum reduction of 45% at low concentrations, and no reduction at higher concentrations. 

Results indicated that epiphytes are a barrier to carbon uptake because they change the initial 

slope of the photosynthesis (P/I) curve. The inhibition of photosynthesis was proportionally 
7.5

greater (58% vs. 31%) at lower light intensities, again suggesting the impact of epiphytes may be 

most severe when ambient light is low. 

 Using in situ incubations of clipped blades of Halodule wrightii, Morgan and Kitting 

(1984) examined the relative productivity of the seagrass and its epiphytes over a range of light 

intensities. At light levels ≥50% of surface illumination, the epiphytic loads did not appear to 

have a shading effect. This assessment was based on what was deemed a “limited” increase of 

Halodule productivity with higher light levels, although the magnitude of the production increase 

was nearly 50%. In the absence of direct measurements of production with and without 

epiphytes, the conclusion of no shading effect at high light levels appears tenuous. 

 Extrapolation of results from productivity measurements on clipped blades to prediction 

of impacts on rooted plants under field conditions are difficult. Because epiphytes are also 

affected by ambient light levels, the net effect of epiphyte shading on seagrass survival is likely 

to be determined by complex interactions. For example, Neverauskas (1988) experimentally 

reduced light to Posidonia spp. (P. sinuosa and P. angustifolia) by 50% using shade cloth over 

small experimental quadrats (0.0625 m2

) with rhizome connections severed. Epiphyte biomass 

decreased from an initial value of approximately 30 g m-2 to near 0 after nine months of shading. 

The shoot density of Posidonia did not show a decrease until after 9 months, although leaf 

density began to decline after 3 months. Rhizome reserves were sufficient to sustain the plants 

beyond the point in time where low light levels had greatly reduced epiphytic loads. 

 In spite of multiple uncertainties, the guidance document for management of seagrasses 

for Chesapeake Bay (U.S. EPA 2003) incorporates the concept of a percent of surface 

illumination available at the leaf surface to account for the additional attenuation of light 

resulting from the presence of epiphytes. The percent light-at-the-leaf (PLL) can be used to 

establish a minimum light level required for persistence of seagrass in the face of combined light 

attenuation in the water column, and that due to epiphytes. PLL values are calculated by the 

formula: 

PLL = 100 [e – (Kd)(Z)][e – (Ke

) (Be)

] 

where Kd is the light extinction coefficient in the water column, Z is the depth, Ke is the epiphyte 

biomass specific PAR attenuation coefficient, and Be is the epiphyte biomass per unit seagrass 

biomass. Specific equations for determination of Ke and Be are given in U.S. EPA (2003 , Table 

VII-1). Data for the tidal-fresh, oligohaline, mesohaline, and polyhaline salinity zones of 

Chesapeake Bay (Appendix J , U.S. EPA 2003), indicated that the additional reduction in light 

intensity at the leaf surface due to epiphytes was 20-60 % in the lower salinity regions, and 10-

50% in the mesohaline and polyhaline regions. Using an average of 30% additional light 

attenuation due to epiphytes leads to the calculation that the PLL requirement to sustain 

seagrasses is 9% of surface illumination for tidal fresh and oligohaline areas, and 15% for 

mesohaline and polyhaline areas. 
7.6

Table 7.1. Studies of light attenuation by epiphytes in coastal marine systems, revised and expanded from Table 1, Brush and Nixon 

(2002). Maximum epiphyte density reported here is the highest measured in the study, and not necessarily the highest recorded in 

light absorbance measurements, and may differ from that reported by Brush and Nixon (2002). A: artificial light; N: natural light. 

Macrophyte 

Study area (Reference) 

Method Light 

source 

Max epiphyte 

density 

(mg cm-2 leaf) 

Max % light 

absorbance 

(mean) 

Comments, Reference 

Zostera marina Øresund, Denmark Suspension A 1.9 (DW) 99.7 Borum & Wium-Andersen 1980  Øresund, Denmark Suspension A 9.5 (DW) 99.7 Same absorbance data as 1980; Borum et al. 1984  Roskilde, Denmark Suspension A .0039 (chl a) 76 May underestimate total absorption, Borum 1987  Mesocosms, VA Suspension A 20(DW) Measured spectral absorption, Neckles (1993 and 

unpubl. data), cited in Brush & Nixon 2002 

 Mesocosms, RI Scraped and unscraped 

leaves 

N 99 (DW) 90 Effect varied with type of epiphyte, Brush & Nixon 

2002 

 Elkhorn Slough, CA Scraped and unscraped 

leaves 

A 60 Measured spectral absorption, underestimated 

epiphyte biomass by use of phospholipid measure only; Drake et al. 2003 

Thalassia testudinum Terminos Lagoon, Mexico Suspension N 17 (DW) Kemp et al. (1988), cited in Brush & Nixon 2002  Tampa Bay, FL Suspension A 81(51) Excluded necrotic blade tips, underestimates total 

absorbance; Dixon & Kirkpatrick 1995 

 Indian River Lagoon, FL Scraped and unscraped 

leaves 

A 18 (DW) 85 (61) Harden 1994 

 Bahamas Scraped and unscraped 

leaves 

A 36 Measured spectral absorption, underestimated 

epiphyte biomass by use of phospholipid measure only; Drake et al. 2003 

Halodule wrightii Tampa Bay, FL Suspension A 99(48) Excluded necrotic blade tips, underestimates total 

absorbance; Dixon & Kirkpatrick 1995 

 Indian River Lagoon, FL Scraped and unscraped 

leaves 

A 30 (DW) 90 (60) Harden 1994 
7.7

Syringodium filiforme Indian River Lagoon, FL Scraped and unscraped 

leaves 

A 22 (DW) 97 (61) Harden 1994 

Posidonia oceanica Mediterranean Sea, Spain Collection on GF/F filters A 71 (DW) 95(~70) Measured only pigment effect, underestimated total 

light absorbance, Cebrián et al. 1999 

Heterozostera tasmanica Victoria, Australia Scraped and unscraped 

leaves 

A 3.3 (DW) 75 Bulthuis & Woelkerling 1983 

Artificial substrates  Cockburn Sound, Australia Growth on glass slides A 0.004 (chl a) 96(63) Silberstein et al. 1986  Success Bank, Australia Growth on plastic slides A 4.2 (AFDW) Burt et al. (1995), cited in Brush & Nixon 2002  Banc d’Arguin, Mauritania Growth on glass slides N 3 (DW) Hootsmans et al. (1993), cited in Brush & Nixon 

2002 

 Delaware Inland Bays Scraped and unscraped 

mylar strips 

N 17 (DW) Glazer (1999), cited in Brush & Nixon 2002 

 Virginia, Maryland Bays Scraped and unscraped 

plastic strips 

A 4 (DW) 96 Measured spectral absorption, Brandt & Koch 2003 
7.8

7.3 Role of Nutrient Loads in Determining Epiphytic Load

7.3.1. Field Observations

 There are a number of correlative studies of field distribution of epiphyte load versus 

nutrient concentrations which suggest nutrient enrichment increases epiphyte loading. Orth and 

Van Montfrans (1984) reviewed a number of studies from freshwater systems that indicated a 

relation between nutrient enrichment, epiphyte increases, and decrease or loss of macrophytes. 

Borum (1985) observed that the biomass of epiphytes on Zostera marina increased exponentially 

with increasing total N concentration in the water column along a transect in Roskilde Fjord, 

Denmark. The relative increase of epiphytes was 10 times greater than the phytoplankton 

response. 

 Neverauskas (1987) studied the impact of sewage sludge outfalls on the seagrasses 

Posidonia and Amphibolis near Adelaide, Australia. Near the outfall, total loss of seagrass was 

observed within 4 years after discharge was initiated, while partially affected areas at greater 

distances from the outfalls showed high epiphytic loads. Experiments placing artificial substrata 

along a distance gradient from the outfall showed highest epiphytic recruitment at sites closest to 

outfall. 

 Examination of accumulation rates of epiphytes on Heterozostera tasmanica in Victoria, 

Australia showed that highest epiphyte levels were found at the site with the highest nutrient 

input (Bulthius and Woelkerling 1983). During conditions of peak epiphytic growth, biomass 

accumulated at rates that were estimated to diminish light below the compensation point within 

36 days, about half the mean life span of the seagrass leaves. However, the expression of 

impacts on the seagrass tended to be site specific. 

 Silberstein et al. (1986) compared characteristics of the seagrass Posidonia australis and 

its epiphytes between two sites, one of which was near a sewage dispersal line. Epiphytic loads, 

measured as chlorophyll per unit leaf area, were higher at the sewage site which also had lower 

seagrass standing stock, shoot density, flowers, leaf production, and growth. However, the study 

did not sample prior to the introduction of sewage, and it is possible some differences were 

present before the sampling. Also, grazer densities were not compared at the sites. 

 In a comparison of epiphyte and seagrass (Thalassia, Syringodium, Halodule) production 

rates from three sites in Florida and the Bahamas, Jensen and Gibson (1986) found that the 

highest epiphytic biomass were at sites with the highest concentrations of phosphorus and 

silicate. Both Tomasko and Lapointe (1991) and Lapointe et al. (1994) also found relationships 

between total nitrogen concentrations and seagrass and epiphyte response patterns in the Florida 

Keys and Caribbean for Thalassia testudinum and Halodule wrightii. Sites with highest nutrients 

were found offshore both from a populated island with septic tanks and a bird rookery island, 

and were associated with higher epiphytic loads, and low shoot density and biomass (Tomasko 

and Lapointe 1991). In a further comparative study, nutrient concentration zones at sites in the 

Florida Keys were defined as hypereutrophic, eutrophic, mesotrophic, or oligotrophic 
7.9

corresponding to total N concentrations in winter of 38.8, 30.3, 21.6, and 12.8 μM, respectively. 

Seagrasses from the oligotrophic zone typically had lowest epiphyte levels, while those in the 

hypereutrophic and eutrophic zones had high levels of epiphytes and mat-forming macroalgae, 

and low shoot densities and productivity. 

 Epiphyte loads on Thalassia testudinum were also measured along nutrient gradients of 

differing scales in Florida Bay by Frankovich and Fourquean (1997). At a fine scale, they found 

significantly higher epiphyte loads nearest a bird rookery island, with the enhancement effect 

decreasing at between 15 and 30 m from the island. Across Florida Bay as a whole, they found 

that epiphyte load was weakly correlated with total phosphorus, and concluded that epiphyte 

levels are not very sensitive to moderate nutrient enrichment. They suggested that epiphytes 

played no role in generating the Florida Bay seagrass die off, but were instead stimulated by 

nutrients released by dying seagrass. They concluded that epiphyte load may be only a late 

response to nutrient enrichment, and thus not a sensitive nutrient condition indicator. 

 Emphasizing the spatial variability in the expression of epiphyte standing stock in 

response to nutrients loads, Tomasko et al. (1996) found that although the greatest epiphyte 

biomass at sites within Sarasota Bay, Florida was found on one date at the site with highest total 

nitrogen loading, the general pattern of epiphyte biomass showed little relation to the pattern of 

nitrogen loadings. Relative grazer densities, which might have affected the results, were not 

reported, however. 

 Results of an intensive data collection effort in multispecies seagrass beds of the Indian 

River Lagoon, Florida both confirm the importance of spatial variability and also emphasize the 

complexity of interpretations possible from field data (Hanisak 2001). In spite of spatial 

differences in nutrient concentrations at study sites, there were no consistent spatial patterns in 

epiphyte load. Seasonality was a much stronger effect on epiphyte levels than location. 

However, when all data were combined to a single mean per site, above-ground seagrass biomass 

at a site decreased as a function of increased epiphyte biomass. Mean epiphyte load in turn 

decreased as grazer abundance increased. Mean grazer abundance showed a positive 

relationship to above-ground seagrass biomass. As Hanisak (2001) points out, these results may 

be interpreted in two ways: 1) that decreased grazing rates allow increased epiphytes, which 

reduce seagrass biomass, or 2) that increased seagrass biomass increases the amount of grazers 

which leads to reduced epiphytes. 

 The number of in situ experimental nutrient additions in seagrass beds was relatively 

limited until recently. Harlin and Thorne-Miller (1981) performed single nutrient addition 

experiments to Z. marina beds in Rhode Island over a 2-3 month period. While responses of 

seagrass and macroalgae were noted, epiphyte biomass was not quantified, and the nutrient 

additions did not alter the species composition of epiphytic algae. Williams and Ruckelshaus 

(1993) conducted in situ, short term (15 d) ammonium enrichment experiments of both water and 

sediments in Z. marina beds in Puget Sound, WA. There was no significant increase in epiphyte 
7.10

biomass, and in fact epiphyte biomass was significantly reduced by water column ammonium 

enrichment. 

7.3.2 Mesocosm nutrient addition experiments 

 Conditions of various mesocosm experiments conducted to examine the effects of 

nutrient additions on seagrasses are summarized in Table 7.2. Unfortunately, as noted by 

Murray et al. (2000), the extreme variation in mesocosm size, flow rates, type of nutrient 

addition (pulse versus continuous addition), and presence of absence of grazers within 

experimental systems make generalizations difficult. 

 Large mesocosm ponds were used by Twilley et al. (1985) to expose Potamogeton 

perfoliatus and Ruppia maritima to pulse additions of nitrogen and phosphorus (Table 7.2). 

Concentrations of N greater than 60 μM caused declines in the vascular plants in the ponds in 6-

10 weeks. At the highest nutrient addition levels, phytoplankton blooms in the ponds were 

evident. Dense epiphytes developed on plants in all nutrient addition treatments. Epiphytic 

cover was estimated to decrease >80 % of light at the leaf surface at highest nutrient 

concentrations. Epiphytic cover was shown to decrease macrophyte photosynthesis, but the 

authors concluded that the negative effect of epiphytes was insufficient to eliminate macrophytes 

without the additional effect of light attenuation by water column phytoplankton. 

 The effect of elevated ammonium levels in the water column on Z. marina were 

examined by Williams and Ruckelshaus (1993) who used experimental nutrient diffuser systems 

in small laboratory aquaria (Table 7.2). Epiphyte biomass was positively correlated with 

ammonium concentration, and eelgrass growth also decreased significantly with increased 

epiphyte biomass. Epiphyte loads on the order of 75mg/shoot decreased eelgrass growth rates, 

and at loads above 100 mg/shoot, growth rates were reduced by 50%. 

 Using mesocosm tanks of 800 l capacity with high water turnover rates (200 % per day), 

Short et al. (1995) applied bags of slow release fertilizer to achieve a continuous release of 

nutrients. Tanks included mud snails which may have done some grazing on seagrass. Nutrient 

additions decreased eelgrass shoot densities by >50%. There were no interaction effects between 

nutrient addition and light reduction treatments, suggesting that the mechanism causing plant 

decrease in nutrient treatments was light limitation. With full light levels, large increases of 

phytoplankton, macroalgae and epiphytes occurred relative to controls. These results also show 

that under nutrient addition, the magnitude of response of epiphytes can be limited if water 

column light levels are reduced. Kopp (1999) also used continuous nitrate additions in 

combination with a 45% light reduction treatment. There was no statistically significant effect of 

nitrate treatment, while the reduced light treatment significantly reduced epiphytic load. 

However, while not significantly different, the highest epiphytic loads observed were under high 

light and high nutrient conditions. 

 Several studies with large mesocosms have used pulse additions of nutrients to examine 

responses of Zostera marina. The experiment reported by Lin et al. (1996) had limited water 
7.11

turnover (5% per day), while all tanks included a variety of fish and invertebrates, but without 

significant numbers of grazers. There was no increase in epiphyte biomass on eelgrass leaves in 

any treatment. The lack of epiphyte response was attributed to the fact that the mesocosm 

contained multiple trophic pathways, with the primary nutrient response having occurred in the 

form of phytoplankton blooms. These results led the authors to suggest that epiphyte biomass 

may be strongly regulated by light limitation, which is consistent with other mesocosm study 

results (Short et al., 1995; Moore and Wetzel, 2000). The authors concluded that elevated 

nutrient levels are not necessarily predictive of an epiphyte increase. 

 Taylor et al. (1999) used the same experimental systems and similar conditions to those 

used by Lin et al. (1996), but applied a wider range of nutrient treatments. No significant effects 

of nutrient treatments were observed on epiphytes, Z. marina, or drift algae in the mesocosms. 

In these experiments, a brown tide bloom occurred in the control tanks which may have made 

differences with nutrient enrichment treatments more difficult to detect. Using the same 

experimental systems, but switching to a continuous addition of nutrients, Bintz et al. (2003) 

examined the interactive effects of water temperature and nutrient addition. Epiphytic levels 

were shown to significantly increase in warm, nutrient addition treatments as compared to 

unenriched warm or mean temperature treatments, or cool temperature treatments, either ambient 

or nutrient enriched. 

 Experiments by Burkholder et al. (1992) using mesocosms of 1570 l with pulse additions 

of nutrients and limited water turnover showed no effect of nutrient additions on epiphytic loads 

on Z. marina, measured as cell counts rather than biomass. Grazer densities in both nutrient 

additions and controls tended to be quite high in these experiments, and may explain the lack of 

epiphyte response. 

 Moore and Wetzel (2000) used replicated 110 l aquaria to test effects of elevated 

nutrients and reduced light on Zostera marina. The experiment used 16 turnovers per day with 

continuous flow nutrient additions, and tanks included moderate densities of gastropod grazers. 

Epiphyte responses were highly depended on treatment, with only the spring experiment at high 

light levels showing a major (10 times) elevation in epiphyte biomass, principally due to 

macroepiphytes rather than microepiphytes. Both above-ground and below-ground biomass 

showed reductions in apparent response to an epiphyte load of ~ 16 g g-1 of eelgrass. The 

experimental design did not examine the response of seagrass and epiphytes under the 

combination of low light, high nutrients, and low grazers. 

 Mesocosm studies of nutrient impacts are more limited for Thalassia testudinum and

Halodule wrightii. Tomasko and Lapointe (1991) and Lapointe et al. (1994) conducted a series 

of mesocosm experiments (Table 7.2) examining effects of added nutrients alone or in 

combination with light reduction. Using daily pulsed addition of nutrients, both N and P 

additions resulted in increases in epiphyte biomass (as a percent of seagrass biomass versus 

controls). There was also a decrease in rhizome growth rates with nutrient additions in both 
7.12

Table 7.2. Summary of laboratory microcosm and mesocosm nitrogen and phosphorus enrichment assays using seagrasses. Treatment Abbreviations: C = control (ambient conditions), +N = NO3 addition, +P = PO4 addition, +NP = NO3 and PO4 addition, +A 

= NH4 addition, +AN = NH4 and NO3 addition, +AP = NH4 and PO4 addition, +NAP = NO3 and NH4 and PO4 addition, +G = grazer 

addition, D = depleted NH4, ? = unable to determine from information in reference. All ratios presented in Treatment Variables are 

relative nutrient concentrations. 

Species: 

Location 

Treatment 

Variables 

Nutrient 

Addition 

Method 

Mesocosm 

Volume: 

Turnover 

rate 

Shoots 

per 

mesocosm: 

m-2

Algae Grazers Experimental Results and 

Comments 

Zostera marinaNetherlands: van Katwijk et al.1997 

Temperature: 15, 20 EC Sediment: Mud, Sand Nutrients: +AN (9:3;25:25;25:50;25:125;75:75;125:25 :M) 

Continuous feed from stock solutions 

19.4 l : 1.2 d-1 40 Epiphytes 

not apparently measured 

No +A - toxic at 125 μM, probably 

toxic as low as 25 :M +N - no effect Toxicity 8with 8temperature. 

Washington: Williams & Ruckelshaus 1993 

Sediment: C, D, +A (5, 20 :M) 

Water: C, +A (10, 33 :M) 

Continuous feed from stock solutions 

75 l: 61 d-1 26-62 Epiphyte 

biomass measured 

Amphipods possibly present 

Epiphyte biomass 8with 8water 

column NH4. Eelgrass leaf growth 9with 8epiphyte 

biomass. 

Rhode Island: Lin et al. 1996 

Nutrients: C,+P(.7:M),+N(7.6:

M), +NP(7.6,.7:M), 

+A(7.6,.7:M) 

Once per day, dissolved 

4554 l: .05 d-1 518: 

250 m-2

Epiphyte biomass measured 

Added inverts & fish, no epiphyte grazers 

No 8 in epiphytes in nutrient 

additions. Phytoplankton blooms may have limited epiphyte response. 
7.13

Rhode Island: Taylor et al. 1999 

Nutrients: C, +AP (.9,.1;1.8,.2,3.6,.3;7.3,.6 :M) 

Once per day, dissolved 

4554 l: .05 d-1 518: 

250 m-2

Epiphyte biomass measured: added 4 macroalgae 

Added inverts & fish, bottom grazing snails @100 

No significant treatment effects on seagrass, epiphytes or drift algae. 

Rhode Island: Kopp 1999 

Nutrients: C, +N (50, 75, 100, 150 :M) 

Light: C, 55% 

Continuous feed from stock header tanks 

150 l: 1 d-1 50: 200 

m-2

Epiphyte biomass measured after 1 month 

No No sig. effect of nitrate treatment 

on epiphytes (although trend for 

8 for high light, high nutrient), 

low light level 9epiphyte load 

significantly. 

Rhode Island: Bintz et al. 2003 

Nutrients: C, +NP (6.5,.5 :M) Temperature: 3 levels (9 yr mean, 4E above, 4E below) 

Continuous feed from drip bags 

4554 l: .1 d-1 580: 

365 m-2

Epiphyte biomass measure5d: added 2 macroalgal species 

Grass shrimp @ 8 per tank 

Epiphytes 8in warm, nutrient 

enriched versus cool treatments. 

Virginia: Neckles et al. 1993 

Nutrients: C, +AN (16.4-37.8 :M) P (1.8-

3.3 :M) Grazers: C, +G 

Continuous feed from stock solutions 

110 l: 16 d-1 92 

(15/pot): 1500 m-2

Epiphyte biomass measured 

Snails, isopods, amphipods 9-11.4 x103 m-2

Seasonally varied grazer densities and nutrients. Epiphyte biomass 8with 

8nutrients. 

Virginia: Moore & Wetzel 2000 

Light: C, 9,28,42% Nutrients: C, +ANP (7-16.3; 6.6-7.7; 1.4-1.8 :M) 

Continuous feed from stock solutions 

110 l: 16 d-1 92 

(15/pot): 1500 m-2

Epiphyte biomass measured 

Snails; 5,200 m-1

Shading inhibited epiphytes and eelgrass. Epiphytes 8 only with 

high light and nutrients in Spring. 

North Carolina: Burkholder et al. 1992 

Nutrients: C,+N (3.5,7, 35 :M) 

Pulse, daily 1570 l: 

 0.5-0.1 d-1

over a 2 hr period only 

~2000: 700 m-2

Epiphytes assessed as cell counts only 

Amphipods + isopods @ 25-58 x103 

m-2

No significant treatment effects on epiphytes. Variable macroalgal response. Grazer density was high in spring experiment. 
7.14

North Carolina: Burkholder et al. 1994 

Nutrients: Spring C,+N (5 :M) Fall C, +N (10 :M) 

Pulse, daily 1570 l: 

 0.5-0.1 d-1

over a 2 hr period only 

~2500: 800 m-2

Epiphytes assessed as cell counts only 

Amphipods @ ~7-80 x103 

m-2

Epiphyte biomass was “negligible”. Grazer densities very high in both spring and fall experiments 

New Hampshire: Short et al. 1995 

A)Light:11,21,41,61, 94% of ambient B) Nutrients: C,+AP @ 6x C) Light:11,41,94% x Nutrients: C,+AP (7-13 :M) (1.4-2.4 :M) 

Continuous, from bags of slow release fertilizer 

800 l, 2 d-1 200: 

133 m-2

Macroalgae and epiphyte biomass assessed 

Added amphipod predators and mud snails 

Epiphytes 8compared to control 

in high light treatment. Variable responses to nutrients at lower light levels. 

Thalassia testudinumFlorida: Tomasko and Lapointe 1991 

Light: C, -33% Nutrients: C, +ANP (10,1,10 :M) 

Pulse daily, from solutions 

120 l: (?50 d-1) 

static for 14 hr d-1

? 9 Epiphyte 

biomass measured 

Added mixed grazers, density? 

+ANP 8epiphytes and 9rhizome 

growth. Shading 9epiphyte effect. 

Florida: LaPointe et. al 1994 

Nutrients: C, +A (10 :M) , +P (1 :M)), +AP (10,1 :M) 

Pulse daily, from solutions 

120 l: 50 d-1

static for 14 hr d-1

? Epiphyte 

biomass measured 

Added mixed grazers, density ? 

Both +A and +P 8epiphytes and 

9rhizome growth. 

Halodule wrightiiFlorida: LaPointe et. al 1994 

Nutrients: C, +A, +P, +AP 

Pulse daily, from solutions 

120 l: 50 d-1

static for 14 hr d-1

? Epiphytes 

assessed 

Added mixed grazers, density? 

Both +A and +P 8epiphytes; 

only +P 9 rhizome growth. 

North Carolina: Burkholder et al. 1994 

Nutrients: Fall C, +N (10 :M) 

Pulse, daily 1570 l: 

 0.5-0.1 d-1

over a 2 hr period only 

~3768: 1200 m-2

Epiphytes assessed as cell counts only 

Amphipods @ ~7-80 x103 

m-2

Epiphyte biomass was “negligible”. Grazer densities very high in both spring and fall experiments 
7.15

Ruppia maritimaMaryland: Twilley et al. 1985 

Nutrients: C, +NAP @ 3 levels 

Pulse, 6 @10 d intervals, fertilizer

340,000 l: pulse turnover @8-9 d 

? 50-80% cover 

Epiphytes measured 

? not specified but likely 

+N at >60 :M caused seagrass 

9. 

Epiphytes 8. 

North Carolina: Burkholder et al. 1994 

Nutrients: Fall C, +N (10 :M) 

Pulse, daily 1570 l: 

 0.5-0.1 d-1

over a 2 hr period only 

~3768: 1200 m-2

Epiphytes assessed as cell counts only 

Amphipods @ ~7-80 x103 

m-2

Epiphyte biomass was “negligible”. Grazer densities very high in both spring and fall experiments 

Potamogeton perfoliatusMaryland: Murray et al. 2000 

A) C, +NP as Pulse, +NP as Continuous (both 38 :M N, 3.8 

:M P) B) Nutrients: D(2-4 :M), +N (20-24 :M) 

Water exchange: 1,3,6,12 d-1

C) Nutrients: +N (<10 :mol l-1 N, >30 :mol 

l-1 N) 

Grazers: C, +G 

A) Pulse and continuous B, C) Continuous 

A) 10 l B) 100 l: see Treatment Variables C) 100 l: 1 d-1

? Epiphyte 

biomass measured 

A) No B) No C) 0.25 gm-2, 

amphipods, density ? 

B, C) Used sequential runs for different treatments, therefore interpretation of results is difficult because of uncontrolled time effect. See text for effects. 
 7.16

 Halodule and Thalassia. When light reduction was combined with nutrient addition, the 

shading reduced the relative increase in epiphyte biomass. 

7.4 Role of Grazers in Control of Epiphyte Loads

 There is a variety of evidence emerging that complex interactions among trophic 

components in seagrass systems may determine the ultimate extent of the effect of epiphytes on 

their seagrass substrate. 

 A range of studies summarized in reviews by van Montfrans et al. (1984), Orth and van 

Montfrans (1984) and Jernakoff et al. (1996) have indicated either that decreased epiphytic load 

caused by grazer removal tends to improve seagrass growth (e.g. Hootsman and Vermaat 1985; 

Howard and Short 1986; Phillipart 1995) or that there was inhibition of seagrass growth caused 

by increased epiphytic loads. These impacts were observed for a wide variety of seagrass 

species including Zostera marina and Thalassia testudinum, and Halodule wrightii. An 

important issue that remains is whether there is evidence that grazing can control epiphyte 

growth under high nutrient loading conditions, and whether there is any specific nutrient 

elevation level where grazer controls will likely be overwhelmed. Controlled experiments 

combining both nutrient elevation and grazer density manipulations remain relatively limited. 

 Although they examined both the effects of grazers and nutrient additions on seagrass 

and epiphyte production, Williams and Ruckelshaus (1993) did not conduct simultaneous 

experiments with both factors. Mesocosm treatments without the isopod grazer Idotea resecata

had epiphyte biomass almost 300% higher than in tanks with grazers at field densities. However, 

the growth response of Z. marina after about two months was not significantly different, 

presumably because total epiphyte load in the ungrazed treatment was still below the critical 

threshold for seagrass growth impacts. 

 In a series of mesocosm experiments, Neckles et al. (1993) varied nutrient concentrations 

and epiphyte grazer densities to examine the relative effects of these factors on growth of Z. 

marina. Experiments were conducted in early and late summer, fall and spring, with nutrient 

enrichment levels and grazer density and relative composition all varying with season to better 

reflect natural conditions. Experimental outcomes also varied with season. Epiphyte biomass 

increased with nutrient enrichment in all experiments, although marginally so in the fall 

experiment. Grazer impacts were greatest in the two summer treatments, which also 

corresponded to the treatments with the highest density of grazers. The late summer experiment, 

with a grazer density of 11,400 m-2, showed a 592% increase in epiphyte biomass in ungrazed 

treatments. The authors noted that seagrass blades in ungrazed treatments at this time of year 

had dense tunicate populations, suggesting that the grazers inhibited tunicate recruitment as well 

as presumably affecting algal epiphytes. The effects of nutrient enrichment were never large 

enough to overwhelm the impact of grazers. In all experiments the ungrazed, ambient nutrient 

level treatments always possessed epiphyte biomass greater than or equal to the grazed, nutrient 

enriched treatments. As the authors point out, the fact that absolute nutrient level, water 

temperature, grazer density, and grazer composition all varied simultaneously among the 
 7.17

experiments makes it impossible to identify the causes of the differential seasonal responses of 

epiphytes to grazers that were observed. 

 In addition to the factors identified by Neckles et al. (1993), mesocosm experiments have 

demonstrated that grazing impacts on epiphytes may be influenced by a variety of hydrodynamic 

factors. Murray et al. (2000) conducted a series of experiments with the brackish water 

macrophyte Potamogeton perfoliatus to examine the relative impacts of the frequency and timing 

of nutrient additions, the residence time of water within experimental systems, and relative 

trophic complexity of food chains in the mesocosms. By scaling macrophyte and epiphyte 

responses relative to the controls in each experiment, the magnitudes of response to treatments 

could be assessed. Grazing had the largest relative effect on macrophyte growth, except under 

high nutrient loads. With high levels of nutrients, changes from pulsed to continuous nutrient 

addition and from high to low water exchange rates, both led to larger relative responses in 

macrophyte growth. 

 In terms of the ability of grazers to control epiphyte biomass under elevated nutrient 

loads, amphipod grazers were able to decrease epiphytic biomass by 56% relative to controls 

under low nutrient addition conditions, which was associated with a 43% increase in macrophyte 

biomass (Murray et al. 2000). However, while the grazers reduced the magnitude of the epiphyte 

response under high nutrient loads (+63% for grazed versus +112% for ungrazed), the grazing 

impact had no ameliorating effect on the macrophyte response, which was -88% for macrophyte 

growth in both grazed and ungrazed treatments relative to controls. 

 Any factor that influences either the densities of grazers, such as predation, or the feeding 

efficiency of the grazers on epiphytes, such as hydrodynamics, may determine the level of 

impact that epiphytes may have on seagrasses. Schanz et al. (2002) observed that biomass of 

epiphytes on Z. marina was highest in sites exposed to water movement, and that there was little 

epiphyte coverage on seagrass in sheltered areas where abundance of the grazing snail Hydrobia 

ulvae was extremely high (151 x 103

 m-2). In situ flume experiments showed that snail density 

was negatively correlated with current velocity, while epiphyte cover was positively correlated 

with velocity. The authors propose a trophic cascade effect caused by hydrodynamics, where 

fast currents remove or inhibit feeding of micrograzers, thus releasing epiphytes from grazing 

pressure. However, Caine (1980) found an opposite pattern, with epiphyte biomass and 

abundance of the grazing amphipods Caprella laeviscula both being higher in quiet water sites 

versus sites with active wave action. On individual seagrass blades, grazer biomass and epiphyte 

biomass were also positively correlated. These differences suggest that species specific 

differences in the dominant grazers in differing locales may determine the influence of 

hydrodynamics. Caprella, which is adapted to a clinging existence, may be far less subject to 

high current speeds than small grazing snails. Given the wide range of consumption rates for 

different epiphyte grazers (reviewed by Jernakoff et al. 1996, Table 3), grazer community 

composition will clearly be critical to determining the ultimate level of effect on seagrass 

epiphytes. 
 7.18

7.5 Research Gaps

 Epiphyte biomass appears to be a major response variable in determining the ability of 

seagrasses to grow and survive under in situ conditions. In early studies, determination of 

photosynthesis/ irradiance relationships for seagrasses were typically done in the laboratory with 

epiphytes removed from the seagrass shoots (e.g. Williams and McRoy 1982; Rice et al. 1983). 

More recent studies have shown that not only does the presence of epiphytes affect light 

quantity, but also spectral light quality can be altered. Effects of epiphytes on the light available 

to seagrasses is thus an important factor which must be quantified in order to be able to develop 

accurate seagrass-stressor response models for evaluating overall impacts of nutrients to seagrass 

systems. 

 However, the question of whether high epiphyte loads alone can directly result in 

seagrass loss does not yet appear to have been conclusively answered. Responses of seagrass 

ecosystems to nutrient enrichment typically involve multiple trophic pathways, with relative 

responses in phytoplankton, macroalgae and epiphytic algae all potentially occurring. High 

loadings of epiphytes clearly can substantially reduce available illumination to seagrass plants. 

In some systems, light reduction to seagrasses from epiphyte load may reach 60 to 80%, at least 

seasonally (Harden 1994; Dixon 2000). Epiphyte grazing studies generally show that removal of 

epiphytes enhances seagrass growth. These observations are strongly suggestive that persistent 

heavy epiphyte cover will lead to seagrass loss, but to date there is no experimental evidence that 

would identify epiphyte load as the single causative factor responsible for seagrass loss under 

high nutrient loads. 

 The role of epiphytic cover in affecting light availability, and hence seagrass distribution, 

may be an essential element to include in development of management criteria for protection of 

coastal seagrass beds. Tomasko (pers. comm.) has found that, for Thalassia testudinum in 

several southwest Florida embayments, there is considerable variation from bay to bay in the 

minimum light requirements for the species even over this relatively limited geographic region. 

Part of this variation appears to be spatial variation in the typical epiphytic load. The role of 

grazers in determining the ultimate impact of epiphyte increases in response to eutrophication 

still represents a significant source of uncertainty in the development of protective nutrient 

criteria for seagrasses. The technical guidance for ambient water quality criteria for Chesapeake 

Bay (U.S. EPA 2003) provides an explicit formulation for including epiphyte loads in estimating 

light available at the seagrass leaf surface. This is an important step forward, but the model 

formulations do not yet appear to have been extensively validated for other systems. 

 Mesocosm research has shown that different algal components may dominate in the 

biomass response to nutrient enrichment in spite of similar initial conditions, and that 

temperature, nutrient exposure regime and other factors such as grazing intensity may all 

influence the outcome of nutrient enrichment. The laboratory results help support observations 

that suggest that seagrass losses, even in different regions of the same limited system, may be 

caused by competition between seagrasses and different algal components (Short and Burdick 

1996). Such results together suggest that if seagrass stress-response models are to be used to 
 7.19

evaluate whether sets of environmental conditions are adequately protective of seagrasses, such 

models must account for impacts from multiple pathways, and must account for effects of 

trophic cascades. 

7.6 Literature Cited

Ballantine, D. and H. J. Humm. 1975. Benthic algae of the Anclote estuary I. Epiphytes of 

seagrass leaves. Florida Scientist 38:148-162. 

Bintz, J. C., S. W. Nixon, B. A. Buckley, and S. L. Granger. 2003. Impacts of temperature and 

nutrients on coastal lagoon plant communities. Estuaries 26:765-776. 

Borowitzka, M. A. and R. C. Lethbridge. 1989. Seagrass Epiphytes. pp. 458-499. In A. W. D. 

Larkum, A. J. McComb and S. A. Shepherd (eds). Biology of Seagrasses: A treatise on 

the biology of seagrasses with special reference to the Australian region. Elsevier, 

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 8.1 

8.0 Macroalgal Interactions with the Seagrasses Zostera spp. and 

Thalassia testudinum

David R. Young

8.1 Background 

The declining distributions of seagrasses, including Zostera and Thalassia, often are attributed to 

excessive accumulations of macroalgae, which in turn are attributed to anthropogenic nutrient 

loadings (Chapter 2). In this chapter we address the following questions: 

1. What is the evidence from field surveys suggesting negative effects of macroalgae on 

Zostera and Thalassia? 

2. Is there corresponding evidence from field manipulations or laboratory experiments? 

3. Is there evidence of positive effects of macroalgae on Zostera and Thalassia? 

Where available, information on the mechanisms involved in the effect(s) also is presented. 

8.2 Field Evidence Suggesting Macroalgal Impacts on Zostera

There have been numerous reports in recent decades of an inverse relationship between 

the abundance of macroalgae (principally green macroalgae such as Ulva, Chaetomorpha, and

Cladophora species) and the distribution and abundance of Zostera species and other seagrasses 

(Table 8.1). The general conclusion is that Zostera distributions have been impacted by 

increases in macroalgal abundance resulting from eutrophication of near shore marine/estuarine 

waters. The comprehensive summary Seagrass Ecology (Hemminga and Duarte 2000) states: 

“The wealth of reports on seagrass decline following eutrophication renders the negative effect 

of marine eutrophication on seagrass stands an indisputable fact, and indicates that it most likely 

is the main cause of seagrass decline worldwide.” Although there is some evidence that high 

levels of nitrate and ammonium can be directly toxic to seagrasses (Burkholder et al. 1992, 1994; 

Van Katwijk et al. 1997), most researchers have attributed this decline to light reduction caused 

by excess nutrient stimulation of phytoplankton, epiphytes, and/or macroalgae. 

In one of the first reports of macroalgal effects, den Hartog and Polderman (1975) noted 

that in the Dutch Waddenzee, “Ulva especially can form thick deposits on the mud flats, 

suffocating underlying Zostera stands.” Two processes that could cause this “suffocation” are 

prevention of oxygenation of the bottom sediment by direct contact with the water column, and 

reduction in delivery of oxygen via the lacunae to the rhizosphere from reduced photosynthesis 

by the shaded eelgrass (Penhale and Wetzel 1983). Both conditions could lead to prolonged 

periods of sediment anoxia, with the excessive organic loadings from the Ulva mats yielding 

high concentrations of dissolved sulfide, a known phytotoxin, in the pore water (Goodman et al. 

1995). 
 8.2 

Table 8.1. Relationships between macroalgal abundance and the distribution or abundance of 

Zostera marina, Z. noltii, and Thalassia testudinum, in chronological order.

Algal Taxon Effect Location Reference 

Zostera marina 

 Ulva Zostera suffocated Netherlands Coast den Hartog and Polderman (1975) 

 Chaetomorpha macroalgae replaced Zostera North Sea Nienhuis (1983) 

 macroalgae Zostera declined Coast of Poland Plinski and Florczyk (1984)

 macroalgae no effect Coast of Ireland Whelan and Cullinane (1985)

 Ulva Zostera uprooted, buried Northeast Pacific Kentula and McIntire (1986)

 Ulva macroalgae replaced Zostera Venice Lagoon Sfriso et al. (1989) 

 macroalgae macroalgae replaced Zostera Coast of Denmark Funen County Council (1991) 

 macroalgae Zostera declined Coast of Poland Kruk-Dowgiallo (1991) 

 macroalgae Zostera declined Coast of Poland Ciszewski et al. (1992) 

 macroalgae macroalgae replaced Zostera Northwest Atlantic Valiela et al. (1992) 

 Ulva Zostera declined Northwest Atlantic Short et al. (1993) 

 Cladophora Zostera declined Coast of Denmark Thybo-Christesen et al. (1993) 

 Ulva Zostera bed suffocated SW Coast of England den Hartog (1994) 

 macroalgae macroalgae replaced Zostera Northwest Atlantic Lyons et al. (1995) 

Ulva, 

 Chaetomorpha 

Zostera declined North Sea Nienhuis (1996) 

 macroalgae macroalgae replaced Zostera Coast of Portugal Oliveira and Cabecadas (1996) 

 macroalgae macroalgae replaced Zostera Baltic Sea Schramm (1996) 

 Cladophora, 

 Gracilaria 

decreased meadow area Northwest Atlantic Short and Burdick (1996)

Ulva shading killed Zostera Venice Lagoon Coffaro and Bocci (1997) 

Ulva macroalgae replaced Zostera Coast of Portugal Flindt et al. (1997) 

Cladophora, 

 Gracilaria 

shading killed Zostera Northwest Atlantic Hauxwell et al. (1998) 

macroalgae macroalgae replaced Zostera Temperate Zone Raffaelli et al. (1998) 

macroalgae loss of seagrass Northwest Atlantic Bricker et al. (1999) 

macroalgae macroalgae replaced Zostera Northwest Atlantic Valiela et al. (2000a) 

macroalgae macroalgae replaced Zostera Northwest Atlantic Bowen and Valiela (2001) 

Cladophora, 

 Gracilaria 

removal increased Zostera Northwest Atlantic Deegan et al. (2002) 

macroalgae Zostera declined Northwest Atlantic Hughes et al. (2002) 

macroalgae macroalgae replaced Zostera Coast of Denmark Nielsen et al. (2002) 

Cladophora, 

 Gracilaria 

macroalgae replaced Zostera Northeast Atlantic Hauxwell et al. (2003) 

 macroalgae no displacement Northeast Pacific Kentula and DeWitt (2003) 

 Ulva no effect Northeast Pacific Thom et al. (2003) 

 macroalgae macroalgae replaced Zostera Adriatic Coast of Italy Curiel et al. (2004) 
 8.3 

Zostera noltii 

 Ulva Zostera suffocated Netherlands Coast den Hartog and Polderman (1975) 

 Ulva macroalgae replaced Zostera Venice Lagoon Sfriso et al. (1989) 

 macroalgae little effect Dutch Wadden Sea Philippart and Dijkema (1995) 

 Ulva macroalgae replaced Zostera Spanish Atlantic Coast Niell et al. (1996) 

 Ulva, 

Chaetomorpha 

Zostera declined North Sea Nienhuis (1996) 

 Ulva macroalgae replaced Zostera Portuguese Atlantic Oliveira and Cabecadas (1996) 

 Ulva macroalgae replaced Zostera Northwest Atlantic Flindt et al. (1997) 

 Ulva macroalgae replaced Zostera Portuguese Atlantic Pardal et al. (2000) 

 Ulva macroalgae replaced Zostera Portuguese Atlantic Cardoso et al. (2004) 

 Ulva macroalgae replaced Zostera Portuguese Atlantic Patricio et al. (2004) 

Thalassia testudinum 

 macroalgae decreased productivity Gulf of Mexico Cowper (1978) 

 Chaetomorpha, 

 Acetabularia

uprooted; stopped 

recolonization 

Mexican Caribbean Merino et al. (1992) 

 Eucheuma possible competition Gulf of Mexico Perez-Enriquez (1996) 

 macroalgae no effect Gulf of Mexico Bell and Hall (1997) 

 macroalgae possible competitor Columbian Caribbean Angel and Polania (2001) 

 macroalgae potential threat Gulf of Mexico Kopecky and Dunton (2006) 

Nienhuis (1983) observed that eelgrass abundance in a southwest Netherlands’ estuary 

dropped by about 50% between 1978 and 1980. He listed a number of possible causes including 

macroalgal competition, and suggested that the most plausible explanation for the decrease was 

an increase in organic matter deposition on the bottom following increased nitrogen loadings to 

the estuary, which in turn caused rapid anoxia of the sediments. This increase in organic matter 

deposition could have produced “a surplus of toxic substances,” killing the root and rhizome 

system. Two potential phytotoxins which could have been produced in high concentrations 

under such circumstances are dissolved sulfide (Goodman et al. 1995) and ammonium (Van 

Katwijk et al. 1997; 2000). 

A mixed bed of Z. marina and Z. noltii in Langstone Harbor, England was completely 

destroyed by a thick blanket of Ulva spp., due to prolonged anaerobic conditions under the algal 

mat with corresponding high levels of sulfide and ammonia (den Hartog 1994). However, 

Whelan and Cullinane (1985) found no significant interaction between macroalgae and eelgrass 

in southwest Ireland. 

Although there was no evidence of causality, Thybo-Christesen et al. (1993) noted a 

significant change in the vegetation in Danish coastal waters, with the appearance of filamentous 

algal mats being associated with a substantial decrease in the abundance of eelgrass. Working in 

Venice Lagoon, Sfriso et al. (1989) found that high nutrient loads in the lagoon in the 1970s 

were followed in the 1980s by a major shift in macrophyte composition. Z. marina and Z. noltii 

beds were replaced by “nitrophile species,” principally Ulva rigida. In some shallow areas of the 
 8.4 

lagoons, this macroalga occupied 100% of the bottom at an abundance exceeding 10 kg/m2 wet 

weight, leading to periodic occurrences of water column anoxia and extensive mortality of 

macrofauna. 

Flindt et al. (1997) compared two other European estuaries with Venice Lagoon. In the 

Mondego River estuary of Portugal, major reductions of Z. noltii beds were observed, apparently 

as the result of blooms of the green macroalgae Ulva spp. (Pardal et al. 2000). In Denmark=s 

Roskilde Fjord, which experienced increased nutrient loading from urbanization of the 

watershed, increased agricultural fertilization, and increased atmospheric deposition, eelgrass 

meadows also were substantially reduced. However, in this case the direct cause appears to have 

been increased biomass of phytoplankton and epiphytic algae. Coffaro and Bocci (1997) 

modeled the competition of resources by the green macroalgal species U. rigida and Z. marina in 

Venice Lagoon, concluding that both nitrogen availability and water velocity influenced the 

structure of the primary producer community, and that competition for light was a major factor in 

the interaction between the macroalgae and eelgrass. 

Losses of eelgrass and other seagrass in recent decades also have been reported and 

extensively studied in the U.S.A. Orth and Moore (1983) analyzed extensive data sets on 

submerged aquatic vegetation (SAV) from Chesapeake Bay. They found that, in Virginia, the 

abundances of seagrass beds dominated by Z. marina and Ruppia maritima decreased sharply in 

the early 1970s; similar decreases in SAV in Maryland were observed. These authors suggested 

that this decline may be related to factors affecting the quantity and quality of light reaching the 

plants. However, they did not discuss a specific mechanism (e.g., macroalgal increases) for such 

an effect. 

Valiela et al. (1992) conducted an extensive study of the sources and effects of nutrient 

enrichment in Waquoit Bay, Massachusetts. They concluded that increased development of 

watersheds there led to increased groundwater nutrient concentrations, which in turn led to 

increased abundances of macroalgae and major reductions in distributions of eelgrass. 

Continuing this research, Lyons et al. (1995) concluded that, in this bay, the abundance of 

macroalgae increased linearly with nitrogen loading while the abundance of eelgrass decreased 

exponentially. They also reported that, as salt marsh area increased, eelgrass biomass also 

increased, suggesting that salt marshes might serve as a buffer against watershed inputs of 

nutrients. Short and Burdick (1996) analyzed data from aerial photographs and ground surveys 

in the Waquoit Bay National Estuarine Research Reserve, relating the loss in eelgrass areal 

extent to housing development and groundwater loadings of nitrogen in the sub-basins of the 

watershed. The direct effect of these loadings varied from basin to basin, but included 

stimulation of phytoplankton, epiphytic growth on the eelgrass, and macroalgae. The authors 

concluded that in one area, “the main algal competitor causing a decline in eelgrass habitats was 

unattached macroalgae [sic] (Gracilaria sp., Cladophora sp.), which smothered and crowded out 

eelgrass plants.” 

Valiela et al. (1997) continued and expanded this work. The increase of nutrient loading 

in estuaries was found to increase macroalgal nitrogen uptake rate, tissue nitrogen content, 
 8.5 

photosynthetic rate, and growth rate. Although the effects of such loadings are mitigated by 

fringing salt marshes and higher rates of tidal exchange, high nitrogen loadings and resultant 

macroalgal blooms can significantly alter estuarine ecosystems. Effects include interception of 

nutrients released from sediments, change in the flux of carbon through the food web, alteration 

of the oxygenation of the water and sediments, and changes in the benthic fauna. Valiela et al. 

(2000a; 2000b) subsequently reported a successful mathematical model of the watershed 

nitrogen loadings of Waquoit Bay. Hauxwell et al. (2003) extended their studies in Waquoit 

Bay, showing an exponential decrease in eelgrass shoot density and bed area as nitrogen loads 

increased. They recommended that these variables be used for routine monitoring of eelgrass 

health. They also noted that the relationship between nitrogen loading and eelgrass health was 

indirect; the direct effect was increased growth and standing stock of algae (water column algae, 

epiphytes, and macroalgae), causing light limitation of eelgrass. A major effect, however, was 

severe light limitation to newly recruiting shoots by shading from macroalgal canopies < 15 cm 

in height. 

There are few reports of the interaction between macroalgae and eelgrass on the Pacific 

Coast of the U.S.A. Phillips (1984) commented that large masses of loose macroalgae such as 

Ulva spp. commonly occur in seagrass meadows, especially where tidal currents are sluggish. In 

Netarts Bay on the Oregon coast, Kentula and McIntire (1986) found a decrease in shoot net 

primary production of Z. marina in mid-summer concurrent with a decrease in insolation and a 

rapid increase in the biomass of U. prolifera. They observed that as this macroalga drifted 

through the eelgrass meadow it became entangled with the eelgrass canopy, uprooting the plants. 

Where the Ulva was attached, sediment deposition was increased which partially buried the 

aboveground biomass of nearby eelgrass plants. Working in Yaquina Bay, Oregon, Kentula and 

DeWitt (2003) did not find any evidence of the displacement of eelgrass by macroalgae, but 

noted the potential for a negative interaction based on the fact that the biomass measured for the 

eelgrass and macroalgae were comparable. Similarly, Thom et al. (2003) did not find clear 

evidence of a decline in eelgrass distribution or abundance with increased green macroalgae in 

Coos Bay, Oregon, but they did suggest that green macroalgae such as Ulva spp. may negatively 

affect the eelgrass in the future. 

8.2.1 Field Evidence Suggesting Macroalgal Impacts on Thalassia

There also have been several reports of an inverse relationship between the abundance of 

macroalgae and the distribution and abundance of Thalassia and other seagrass species (Table 

8.1). Cowper (1978) found that drift algae competed for light with seagrasses, including T. 

testudinum, in Redfish Bay, Texas, with the algae having substantially higher growth rates than 

the seagrasses at irradiances less that 45% of surface irradiance. Merino et al. (1992) studied the 

relationship between T. testudinum and macroalgal communities (dominated by the genera 

Chaetomorpha and Acetabularia) on the Mexican Caribbean coast. They reported that as the 

algae became more common, the seagrass communities declined. In particular, the algae formed 

mats covering the bottom communities; then, as oxygen bubbles formed within the mats causing 

them to float, the remaining seagrass sprouts were uprooted. Perez-Enriquez (1996) found a 

negative association between the red seaweed Eucheuma isiforme and the seagrasses T. 

testudinum and Syringodium filiforme off the Peninsula of Yucatan, Mexico. However, in this 
 8.6 

case the author suggested that these seagrasses might be affecting the distribution of the 

seaweed. In contrast, in Tampa Bay, Florida, Bell and Hall (1997) found no significant 

relationship between percent cover or biomass of drift algae and mean blade length, shoot 

density, or above-ground biomass of the dominant seagrass species T. testudinum and Halodule 

wrightii. Angel and Polania (2001) studied the distribution of T. testudinum and S. filiforme

around San Andres Island in the Columbian Caribbean. They reported significant damage to the 

seagrass meadows around the island, suggesting a number of possible anthropogenic causes as 

well as possible competition for space by macroalgae. Working in two estuaries in the western 

Gulf of Mexico, Kopecky and Dunton (2006) found very high abundances of drift macroalgae 

which they characterized as a potential threat to T. testudinum and other seagrasses in those 

systems. 

8.3 Evidence of Macroalgal Impacts on Zostera from Laboratory or Field Manipulation

Harlin and Thorne-Miller (1981) conducted field experiments in Rhode Island, adding 

ammonium, nitrate, or phosphate to the water column. They found that ammonium additions 

caused the appearance of dense mats of the free-floating green macroalgae Ulva and stimulated 

eelgrass growth. This growth response of Z. marina was greater in the area where current 

velocity reached 12 cm/sec, presumably because the boundary area around the eelgrass leaves in 

the higher velocity area was decreased, enhancing nutrient uptake. In contrast, the nitrate 

additions enhanced the growth of the green macroalgae but not the eelgrass, while phosphate 

additions stimulated growth of eelgrass but not green macroalgae. None of the nutrient 

supplements had a significant effect on epiphytic algae or phytoplankton in the test areas. The 

authors concluded that the rapid growth of green macroalgae in the nitrogen-rich waters probably 

limited the growth of adjacent seagrasses. In mesocosm experiments, Bintz et al. (2003) 

observed that the negative effect of elevated water temperature on eelgrass was significantly 

increased with inorganic nutrient additions, which enhanced the accumulation of macroalgae, 

especially at higher temperatures. 

 Short et al. (1995) enriched mesocosms with nitrogen and phosphorus via dissolution of 

a slow-release fertilizer and found that stimulation of three different algal forms (phytoplankton, 

epiphytes, and macroalgae) occurred in different replicate treatments. However, the enrichment 

effects on eelgrass shoot density, biomass, and leaf length were similar for all replicates. In each 

case, the negative effect of algae on eelgrass occurred primarily through shading, and eelgrass 

growth decreased linearly with reduced light. Taylor et al. (1995) also conducted mesocosm 

experiments on eelgrass in Rhode Island with various combinations of nutrients. They reported 

that in the controls phytoplankton levels remained low while macroalgae and epiphytes were 

abundant, but in the nutrient-enriched mesocosms phytoplankton blooms dominated. These 

mesocosm studies do not appear to have produced definitive results regarding the effect of 

macroalgae on eelgrass. This point was emphasized by Raffaelli et al. (1998), who commented 

on the absence of controlled manipulative field experiments to explore such effects. 

Subsequently, Hauxwell et al. (1998) observed that benthic algal growth rates and 

biomass increased with nitrogen load in three Massachusetts estuaries, while abundance of 
 8.7 

grazers decreased, resulting in negligible top-down control of macroalgal biomass. Then, using 

macroalgal enclosure or exclusion field experiments, Hauxwell et al. (2001) compared eelgrass 

productivity in two estuaries that had a six-fold difference in nitrogen loading rate. The authors 

concluded that macroalgal cover was the primary cause of eelgrass loss in the high-nitrogen 

estuary. Upon removal of macroalgae, shoot density, summer growth, and summer aboveground 

net production all increased significantly. They observed low sediment redox conditions and 

potentially toxic concentrations of ammonium in the porewater, and identified an approximate 9-

12 cm critical macroalgal canopy height at which eelgrass declined. Working in Washington 

State, Nelson and Lee (2001) conducted similar field manipulations, finding that removal of the 

dominant green macroalgae Ulvaria obscura from eelgrass beds significantly reduced the loss of 

shoots during the summer bloom. They concluded that natural blooms of this macroalgae reduce 

eelgrass shoot density in the area. 

Brun et al. (2003a; 2003b) conducted laboratory and field experiments in Cadiz, Spain on 

the effect of shading by U. rigida canopies on Z. noltii. They reported that productivity and 

elongation rates of the seagrass decreased when subjected to overlying mats of the green 

macroalgae. They also reported the mobilization of starch in both above- and below-ground 

tissues, accompanied by enhanced protein turnover and changes in metabolic pathways. 

Cummins et al. (2004) added U. intestinalis (at levels equivalent to a naturally occurring bloom) 

to a seagrass meadow composed of Z. capricorni and other genera in New South Wales, 

Australia. Three months later, considerable gaps resulted in the seagrass canopy. 

In 1999-2000, Sullivan (unpub.) manipulated green macroalgae densities in 2-m2

enclosures in Yaquina Bay estuary, Oregon to examine impacts on sediment and water column 

processes in Z. marina habitat. Macroalgae biomass was removed or added, with enclosure and 

non-enclosure controls. Porewater NH4

+

 and PO4

 - increased with algae density in all habitats, 

while porewater nitrite and nitrate differed but did not respond to algal density. Water column 

nutrient and oxygen concentrations within algal canopies differed from open water levels. 

However, only the water column oxygen concentrations responded to algal manipulations. 

Macroalgal density affected seagrass density, with algae additions reducing shoot density relative 

to non-enclosure controls (Figure 8.1). 
 8.8 

Random Samples Controls Enclosure Controls Algae Removal Algae Addition

Shoots m-2

0

5

10

15

20

25

Figure 8.1. Effects of experimental benthic green macroalgal (Ulva spp.) manipulations on shoot 

density of Zostera marina in experimental enclosures as compared to controls, for 

Yaquina Bay, Oregon (G. Sullivan, unpub. data, The Wetlands Institute, Chicago IL). 

Algae addition treatment is significantly different from all other treatments except the 

enclosure controls (one-way ANOVA, square root transformed data, Holm-Sidak 

multiple comparisons, p < 0.0001).

8.3.1 Evidence of Macroalgal Impacts on Thalassia from Laboratory or Field Manipulation

McGlathery (1995) manipulated nutrient and grazing levels at a eutrophic and a 

mesotrophic/oligotrophic site in a Bermuda lagoon. Nutrient enrichment caused an increase in 

percent cover of the filamentous, mat-forming macroalga Spyridea hypnoides and a decline in 

the percent cover and above-ground biomass of T. testudinum at the eutrophic site. Holmquist 

(1997) manipulated algal mats (Laurencia poiteaui) over study plots of T. testudinum in 

southwestern Florida Bay. The algal canopy heights were 40 cm versus 18 cm for the seagrass. 

After 6 months of algal cover, the density of T. testudinum fell to 12% of the original value, and 

after 18 months of recovery the density had increased to only about 25% of the initial density. 

Macia (2000) manipulated drift algae and sea urchins (Lytechinus variegatus) within cages 

containing T. testudinum in Biscayne Bay, Florida. Under normal grazing, the drift algae blooms 

that formed large mats covering the seagrass canopy in winter did not have a significant negative 

effect on the seagrass. However, with increased grazing pressure there was a synergistic effect 

of grazing and drift algae on seagrass shoot density. Davis and Fourqurean (2001) manipulated 

densities of the rhizophytic algae Halimeda incrassata in plots of T. testudinum in the upper 
 8.9 

Florida Keys. They reported that the addition of the macroalgae had no significant impact on 

seagrass growth, but that the removal of the macroalgae significantly lowered the leaf tissue C:N 

ratio. The authors concluded that competition for nitrogen was the mechanism of the interaction. 

In comparison, Irlandi et al. (2004) did not find any significant effects of manipulated drift algae 

cover on T. testudinum at two different sites and seasons in Biscayne Bay, Florida. Further, 

Armitage et al. (2005) found no replacement of T. testudinum by macroalgae under nutrient 

enriched conditions in Florida Bay. 

8.4 Positive Effects of Macroalgae on Eelgrass

There is little evidence in the literature of positive effects of macroalgae on eelgrass. A 

possible exception is for eelgrass in the estuarine intertidal zones of the Pacific Northwest and 

elsewhere. The accumulation of green macroalgae within and upslope of eelgrass meadows may 

help to retain water on the mudflats during periods of daylight low-tide intervals, thus reducing 

desiccation of the eelgrass plants. Boese et al. (2003) have documented desiccation damage in 

intertidal eelgrass and suggested that this may be a limiting factor for upper intertidal 

distribution. In Yaquina Bay, Oregon, acute desiccation stress is often observed in late spring 

and early summer when daylight spring-tides, sunny and/or windy weather combine. In contrast, 

in the late summer and fall, the presence of large amounts of macroalgae tends to cover exposed 

sheaths which may provide some protection. Annual eelgrass shoots may not be as susceptible to 

desiccation stress since their sheaths are often more flexible than perennial eelgrass shoots, and 

tend to lie flat on the sediment surface (van Katwijk et al. 2000; Boese, unpub. data). 

8.5 Research Gaps 

The reports cited above provide substantial evidence that, in numerous locations around 

the world, elevated concentrations of nutrients in near shore estuarine and marine waters have 

stimulated algal growth, including that of macroalgae, which interferes with the physiology 

(photosynthesis, respiration, reproduction, etc.) of eelgrass. The most frequently cited impact is 

shading. Quantitative relationships between macroalgal canopy height (or corresponding 

measures of abundance) and specific impacts on eelgrass plants (such as those provided by 

Hauxwell et al. 2001), for different water body characteristics (temperature, current velocity, 

turbidity, grazing pressure, etc.) are needed. Similarly, relationships between macroalgal 

abundance and the causative anthropogenic activity (e.g., normalized nitrogen load rate) are 

needed to recommend corrective actions. 

8.6 Conclusions 

The studies summarized here generally identify excessive accumulations of green 

macroalgae as a principal cause of the decline or disappearance of the eelgrass Z. marina. The 

primary mechanism of impact is shading of the eelgrass, thus reducing its photosynthesis. 

However, a secondary mechanism is that excessive macroalgal loading of the sediments may 

lead to elevated porewater concentrations of ammonium and/or dissolved sulfide that may 

contribute to eelgrass decline. There also are reports of drifting macroalgae becoming entangled 
 8.10 

with eelgrass and uprooting it, or causing its burial via increased sediment deposition. On 

occasion these processes are lumped into the general terms outcompete, crowd out, suffocate, 

etc. Eutrophication of the near shore marine waters by anthropogenic inputs of nutrients 

(principally nitrogen) is most commonly cited as the direct cause of the macroalgal blooms 

impacting the eelgrass. 

In summary, there appears to be a substantial body of evidence that increases in nitrogen 

loadings to estuaries are accompanied by increases in macroalgal abundances and decreases in 

seagrass distributions. This suggests that, conceptually, macroalgal abundance should be 

included in estuarine nutrient loading criteria intended to protect seagrasses in those systems. 

Hauxwell et al. (2001) provided quantitative guidance for at least part of such a strategy in 

Waquoit Bay, Massachusetts. They identified a macroalgal canopy height of 9-12 cm above 

which eelgrass declines. In one estuarine watershed of the Bay, a six-fold increase in the 

nitrogen loading rate over that in a “pristine” watershed corresponded to an increase in 

macroalgal canopy height from 2 cm to 9 cm, the threshold for a measurable decline in eelgrass 

distribution suggested by their study. It seems likely that such relationships between nutrient 

loading, macroalgal abundance, and seagrass distribution are site-specific. For example, systems 

with long residence times are likely to be much more vulnerable to increased loadings than are 

those that are rapidly flushed. The logical sequence appears to be first a determination of the 

relationship between macroalgal and seagrass abundances in a given system, and then a 

determination of the relationship between nutrient loading and macroalgal abundance for that 

system. 
 8.11 

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Bay, southwest Ireland. Aquatic Botany 23:41-51.

9.1

9.0 The Effects of Temperature and Desiccation on the Seagrasses 

Zostera marina and Thalassia testudinum

David T. Specht and Bruce L. Boese

9.1 Background

Temperatures elevated above the ambient range have long been recognized as having 

deleterious or fatal effects on marine organisms (e.g. Sachs 1864, working with Vallisneria and 

Ceratophyllum; Mayer 1914). Environmental temperature influences the rate of critical 

biological processes of organisms in general (Billings 1952; Vernberg 1978), including processes 

such as photosynthesis and respiration in seagrasses. This review examines and summarizes 

work characterizing the effects of natural and anthropogenically influenced variation in 

temperature and desiccation, cyclic and abnormal, on aspects of physiology, growth, 

reproduction and distribution of primarily the seagrasses Zostera marina and Thalassia

testudinum in North America. 

Larkum et al. (1989) summarized work involving temperature effects on seagrasses, 

emphasizing that the literature is rife with conflicting views on the relative importance of 

temperature and irradiance, attributable to the fact that water temperature is largely determined 

by the amount of incoming solar radiation, so that it is often difficult to separate the effects of 

each contribution. Anthropogenic alterations of the temperature environment through thermal 

effluent discharge has been shown to have major and non-predictable impacts on marine benthic 

communities (Schiel et al. 2004), and may interact with other stressors such as nutrient elevation 

through multiple mechanisms; these stresses can negatively impact seagrass populations. 

Thorhaug, Segar and Roessler (1973) documented the dramatic effect of heated power 

plant effluent dilution on the local distribution and health of Thalassia testudinum (turtlegrass) in 

a subtropical habitat. Power plant effluents ~5° C above the ambient temperature of ~30° C 

completely denuded a ~9 ha expanse of seagrass. Concentric to that zone of initial dilution, at 

temperatures 3-4° C above ambient, plants showed “severe” damage. In the zone at ~1-2° C 

above ambient temperature, there was elevated productivity (see also Thorhaug, Blake and 

Schroeder 1978). 

Aerial exposure and resulting desiccation stress is probably the most important factor 

limiting the upper intertidal distribution of seagrass species. Although this is not directly an 

anthropogenic stressor, alterations in estuarine bathymetry which increase elevations in intertidal 

areas would reduce seagrass populations by increasing the frequency and duration of aerial 

exposure. For example, high rates of sedimentation from logging in coastal watersheds with 

steep slopes such as found in the Pacific Northwest can reduce average depth within coastal 

embayments (Komar 1997; McManus et al. 1998; Styllas 2001). Reduced light due to 

eutrophication and turbidity in subtidal eelgrass beds would also serve to restrict potential 
9.2

distribution of eelgrass to shallower areas where desiccation stress may be important on a 

seasonal basis. 

9.2 Temperature-driven geographic, regional and spatial distribution

Setchell (1929) was the first to comprehensively examine the effects of temperature on 

characteristics of populations of Zostera marina L., including growth rate, onset of reproduction, 

senescence, and germination of seeds. Based on field observations, he proposed an optimal 

temperature window of 10 - 20EC, outside of which growth apparently ceased. He proposed 

vegetative growth occurred chiefly between 10 - 15° C, and reproductive activity between 15 - 

20° C. He noted that along the geographic range, reproduction began earlier in southern 

populations and progressed northward. He found marked differences in plants in subtidal versus 

intertidal populations, and attributed these to temperature differences. 

Although morphological differentiation may be attributed to complex influences (see 

below) subsequent work has confirmed many of these early observations (Phillips and Backman 

1983). Comparison of flowering events along a latitudinal gradient from North Carolina to 

Canada indicated that reproduction occurred earlier in the south and at successively later dates 

with increasing latitude (Silberhorn et al. 1983). Puget Sound populations were found to have 

temperature optima of between 7.5 to 12.5° C, but tolerated temperature extremes of 6.5 to 18° C 

(Phillips 1972). Zostera populations can persist under more extreme temperature regimes. For 

example, Bering Sea populations experience temperatures from -6 to ~30° C, and even to 35° C 

in exposed intertidal pools (Zieman and Wetzel 1980; Biebl and McRoy 1971). At the southern 

extreme of the population range, the species apparently shifts its reproductive strategy based on 

the temperature range experienced. Perennial forms of Z. marina are found in northern Baja 

California (11-27° C), while the annual form is characteristic of the southern peninsula (12-32°

C) (Ibarra-Obando et al. 1997; Meling-Lopez and Ibarra-Obando 1999), and in the Sea of Cortez 

(Phillips and Backman 1983). 

Glynn (1968) suggested that temperature probably limited the northern distribution of 

Thalassia testudinum in Florida, although there were regional differences. In the Gulf of 

Mexico, T. testudinum is apparently capable of enduring a warm temperate climate, whereas 

along Florida's east coast exposure to temperatures of 35 - 40° C will kill the leaves. 

Subtidal and tidepool populations of Zostera sp. in Alaska differ in a range of 

characteristics that have been attributed to differences in temperature regimes (McRoy 1970). 

Subtidal plants were aseasonal with high root/rhizome biomass, low shoot density, few or no 

flowering shoots, and long, wide leaves; tidepool beds had low root/rhizome biomass, high shoot 

density with a large number of flowering shoots and short, narrow leaves, and exhibited marked 

seasonal cycles of biomass. Similar contrasts were observed for other Pacific coast locations 

(Phillips et al. 1983b). 

Analysis of seagrass δ13 C values indicates that seagrasses tend to become more 13C 

depleted from tropical (warmer) to temperate (cooler) regions (Hemminga and Mateo 1996). 
9.3

Variation (in order of relative importance) of source of carbon, irradiance and temperature on 

CO2 availability in seawater may partly explain this latitudinal trend, but gradients in day length 

may also be involved. 

Evans (1983) examined the occurrence of two thermally disjunct populations of Z. 

marina in the Woods Hole, MA area (27° C vs. 20-22° C). When grown under identical 

conditions at 15EC, the two isolates maintained growth differences. Their Pmax data under 

experimental culture indicated that increased photosynthetic performance at elevated 

temperatures was not a good predictor of ecological success, e.g., when transplanted to a habitat 

with a different temperature regime. Phillips and Lewis (1983) reciprocally transplanted Z. 

marina over temporal, spatial and ecogenetic gradients in North America; they demonstrated that 

initial leaf widths of transplanting stock are not a good predictor of success. Puget Sound 

populations, coming from the mid-range of west-coast geographic and genetic distribution, and 

low temperature variance habitat, exhibited broad adaptive tolerance; Alaskan stock, originating 

in a highly variable thermal habitat, did not survive in Puget Sound. Olesen and Sand-Jensen 

(1993) concluded that Danish Z. marina acclimated to winter conditions by altering biomass 

allocation, i.e., increasing leaf surface area and reducing weight proportionally and respiration of 

non-photosynthetic tissues. Pollard and Greenway (1993) showed that Z. capricorni from warm, 

turbid Australian waters had higher than expected photosynthetic efficiencies than the same 

species from a relatively high light environment; plants accommodated to low-light turbid 

waters, at high water temperatures (29-33° C) reached Hsat with small incremental light 

increases. One quarter of gross production was expended as respiration. 

9.2.1 Implications of climate change

The IPCC (2001) summary of historical data and predictions for climate change yields 

the following: The global average sea-surface temperature has increased since 1861 (beginning 

of “reliable” measurements) about 0.6 ± 0.2° C, which includes a great deal of variability - most 

of the warming occurred in two periods, 1910-1945 and 1976-2000. Analyses of proxy data for 

the northern hemisphere indicate that the increase in temperature in the 20th century is likely to 

have been the largest of any century during the past 1,000 years. The increase in sea-surface 

temperature has been about one half that of the air temperature increase (~0.2° C per decade) 

between 1950 and 1993. Tide gauge data show that global average sea level rose between 0.1 

and 0.2 m during the 20th century, and that the global ocean heat content has increased since the 

late 1950s, the period for which adequate observations of sub-surface ocean temperatures have 

been available. According to NOAA=s 2006 Annual Climate Review (NOAA 2006), the 2006 

average annual temperature for the contiguous U.S. was the warmest on record, nearly identical 

to that of 1998, and ~1.2° C above the 20th Century mean of ~11.7° C. The past nine years 

(1997-2006) are among the 25 warmest years on record, a span unprecedented in the historical 

record. 

Short and Neckles (1999) suggest that a variety of factors such as increases in seawater 

temperature, resultant rise in sea level, changing water depth and tidal range, and increased 

salinity intrusion all will impact seagrasses as an effect of global warming. The direct effect of 

temperature increases will tend to alter the geographic distribution of seagrasses, and will result 
9.4

in changes in the patterns of sexual reproduction. Another proposed effect of global warming is 

an increase in UV. However, Z. capricorni in Australia has been shown to acclimate readily to 

elevated UV and PAR by production of UV-blocking agents and, despite the enhancement of 

inhibitory effects of high PAR and UV-B by temperature stress, the relatively small temperature 

changes associated with climate change are unlikely to force widespread damage to the majority 

of eelgrass populations (Beardall et al. 1998). 

9.3 Effects of temperature on phenology and reproduction

Each of the three reproductive phenophases (initiation of bud, anthesis, and appearance of 

fruit) of Z. marina vary significantly in their dates of occurrence along latitudinal gradients in 

North America (Phillips et al. 1983b). The time and temperature at which onset of reproductive 

phases occur differs between populations apparently representing different genotypes on the east 

and west coasts at the same latitude (see also Gambi 1988; Phillips and Lewis 1983; Buia and 

Mazzella 1991; Inglis and Lincoln-Smith 1998; Phillips et al. 1983b; Silberhorn et al. 1983; and 

Churchill and Riner 1978). 

 Flowering of Thalassia testudinum is largely controlled by water temperature, optimally 

in the range of ~20-26° C, while lower spring temperatures in the range of 10-18EC cause 

delayed flowering, with subsequent dehiscence of immature fruit (Phillips 1960; Moffler and 

Durako 1987; McMillan 1982). Thus there tends to be increasing failure of sexual reproduction 

as one approaches the northern limit of distribution (Witz and Dawes 1995). Zieman (1975) 

observed that seedling success was an exceptionally rare event, having documented the failure to 

survive of germinated seeds, and that most Thalassia growth and spreading is likely due to 

vegetative reproduction, suggesting that most expansion was clonal, citing McMillan and 

Moseley (1967). 

Phillips et al. (1983a) noted a higher incidence of sexual reproduction in response to 

temperature at the geographic extremes of Zostera distribution than in the temperate central 

range, where temperatures are relatively moderate. Dawes, Phillips and Morrison (2004) 

observed that beds of Thalassia testudinum in more tropical regions contain a greater number of 

distinct genets (i.e., genetic variants) than do beds at higher latitudes, which may reflect success 

of seed production (and, hence, successful sexual reproduction), due to seasonal temperature 

effects (see also Davis et al. 1999; Witz 1994; Witz and Dawes 1995; Kirsten et al. 1998). They 

cite the example of lower winter water temperatures (i.e., 10-18° C) in the Tampa Bay area (a 

relatively “northern” population of Thalassia) as a possible cause of later flowering, resulting in 

loss of immature fruits in response to rapidly rising late spring temperatures. They credit 

Gessner (1970) with the general concept that sexual reproduction was less likely to be successful 

when aquatic plants encounter less-than-optimal temperatures - genetic variance at the cost of 

higher reproductive failure. 

Time of the initial appearance of visible floral buds of Z. marina is quite variable within a 

location. Over a four year period, Phillips et al. (1983b) noted ranges of temperature of 6-21EC 

in Rhode Island, 10-20° C in Halifax, N.S., while Puget Sound temperatures varied only from 8-
9.5

10° C over two sites. Distinguishable primordial anthers and pistils have been found in 

Chesapeake Bay populations in February at 3° C (Silberhorn et al. 1983), while Churchill and 

Riner (1978) observed microscopically visible buds in January in New York populations at water 

temperatures of 0.5-3° C. From these and other accounts (see especially Nienhus 1983; and 

Zimmerman et al. 1995; below, and Section 9.4), we may infer that the appearance of initial 

activity with respect to temperature and date may reflect the degree of success of the previous 

growth season, i.e., available stored energy from the rhizome (from depleted to plentiful), and 

the date of onset of significant increase in solar-induced diurnal temperature cycling. Burke et 

al. (1996), discussing carbohydrate reserves in eelgrasses, cite the agreement of their data on east 

coast Z. marina non-structural carbohydrate (nsc) reserve drawdown as fall progresses to winter 

with San Francisco Bay population data (Zimmerman et al. 1995), conditioned by high-turbiditycaused low light levels. Note that many authors cited elsewhere in this review observe the 

initiation and onset of growth of flowering buds in Zostera at very low temperatures in the dead 

of winter, precisely when nsc reserves would be at their lowest in most habitats. Zimmerman et 

al. (1995) also notes that plants with the highest level of nsc reserves show the highest level of 

asexual reproduction (i.e., transplant success = vegetative growth) at winter/spring onset of 

growth; there is a significant correlation of high nsc and growth with low turbidity and 

consequent higher light availability. de Cock (1981) suggested that the development of 

reproductive shoots is not inhibited by temperatures below 15° C (see Setchell 1929), but that 

maturation of inflorescences and timing of anthesis may be suppressed at lower temperatures. 

Ramage and Schiel (1998) noted that in New Zealand populations of Z. novazelandica in 

intertidal platforms, flowering shoots were more numerous in the low intertidal zone than in 

upper zones, and two times more in patches bordering tidepools than in patches not bordering 

tidepools; Z. novazelandica cultured at 5° C had ~3 times the number of inflorescences than 

those at 15° C, while none was formed at 25° C (see also McRoy=s (1970) Alaskan observations 

with respect to submergent and emergent Z. marina). 

Anthesis and pollination occur as temperatures rise from winter minima, with 

considerable interannual variation observed over a four year period, following a latitudinal 

gradient northward (Phillips et al. 1983b). Silberhorn et al. (1983) showed first evidence of 

Chesapeake Bay Z. marina pollen release at 14.3° C, stigma loss at 16° C in late April, with 

pollination essentially complete by mid-May. 

Fruiting and seed dispersal in Z. marina are typically complete by late May to early June, 

consistently occurring in a temperature range from 20-25° C across a wide latitudinal gradient 

(de Cock 1980; Phillips et al. 1983b; Silberhorn et al. 1983); more northerly distributions may 

tend to fruit later in the season, and within lower temperature bands, e.g., Phillips (1983a; 1983b) 

for Puget Sound, and Harrison and Mann (1975) for Nova Scotia. 

Nienhuis (1983) and Verhagen and Nienhuis (1983), modeling changes in biomass and 

distribution of Z. marina in the Netherlands, concluded that those changes could be attributed to 

temperature-induced changes in seed production: low water temperatures during the growing 

season caused a reduction in biomass, numbers of generative shoots and number of seeds; high 

water temperatures from August to late autumn would stimulate the production of vegetative 
9.6

shoots. This vigorous growth tended to exhaust below-ground resources, leading to a reduction 

of biomass and numbers of generative shoots in the succeeding summer and thus diminished 

seed production. Zimmerman et al. (1995) determined that if carbohydrate reserves were low in 

below-ground organs, survival of transplants during winters dominated by high turbidity (and 

consequent low light) and low temperatures would be severely compromised. 

A number of authors addressed the phenomenon of germination dormancy (Orth and 

Moore 1983; McMillan 1983; Loques et al. 1990). Mid-to northerly populations of Z. marina 

tend to exhibit temperature-mediated dormancy, reflecting the considerable change in seasonal 

temperature. More southerly populations (Sonora, Mexico, French Mediterranean), with 

seasonal temperature variation, exhibited salinity-mediated dormancy, or showed no dormancy at 

all. 

Conclusions regarding temperature effects on germination have changed over time, 

reflecting broader geographic investigations. Phillips and Menez (1988) determined that water 

temperature, and not salinity, was the primary germination control. Hootsmans et al. (1987) 

experimentally determined that Z. marina seedling survival peaked at 10° C and 10-20o

/oo, Z. 

noltii survived best at 10° C and 1o

/oo, and both species showed maximal germination at 30° C 

and 1‰ salinity. In contrast, Moore et al. (1993) and Brenchley and Probert (1998) 

demonstrated that Z. marina and Z. capricorni germination was highest at low temperatures and 

under anaerobic conditions, and lowest under aerobic conditions. 

El Niño events may have local or regional effects on seagrasses. Seddon et al. (2000) 

reported that Z. muelleri and Z. mucronata (= Z. muelleri subsp. mucronata) were among a 

number of intertidal and shallow subtidal Australian seagrasses that were drastically affected by 

high water temperatures associated with the 1993 El Niño event. Nelson (1997) concluded that 

intertidal eelgrass plants would tend to decline as a result of increased photoinhibition and 

desiccation due to increased temperature and light during El Niño episodes. Thom et al. (2003) 

found that warmer winters and cooler summers associated with the transition from el Niño to la 

Niña ocean conditions corresponded with an increase in eelgrass abundance and flowering. 

Harrison (1982a) attributed a decline of Z. marina in the upper portion of a tidal flat 

drainage channel to impacts of warmed water receding with the ebbing tide, while flowering 

peaked earlier than the subtidal population. Phillips and Backman (1983) report that Z. marina

in the Sea of Cortez, Mexico, completed all reproductive activities before the putative lethal 

upper limit for the species is reached (30° C), which suggests that the ultimate response to high 

water temperature is to behave as a true annual. 

Burkholder et al. (1992) found adverse effects of water column nitrate enrichment in Z. 

marina in mesocosms were exacerbated by increasing or high temperatures. The meristematic 

portion of the shoot disintegrated after several weeks of exposure when water temperatures were 

held at 4° C above the 10 year (local) mean. Touchette and Burkholder (2002), using similar 

exposure scenarios, found that cellulose accumulation in below-ground structures was 

substantially below that for Z. marina grown at ambient temperatures; higher cellulose content is 
9.7

shown to allow significant increase in new shoot productivity. Touchette et al. (2003), further 

exploring this phenomenon, noted that Z. marina grown under such conditions, typically that of 

the southern latitudinal limits of distribution, exhibited morphological and physiological 

symptoms of decline. Significantly reduced shoot density, leaf and root production, and altered 

internal C and N content, support the premise that in more southerly distributions, growth is 

inhibited by high temperature stress. The response of Z. marina to these conditions may be a 

reliable predictor for the impact of warming trends in climate change scenarios. 

9.4 Effects of temperature on physiology

McMillan (1978) demonstrated that Z. marina collected from Alaska and Washington 

produced leaves of significantly different widths under three different temperature regimes in 

culture, indicating some characteristics of seagrass phenology can be modified by temperature. 

 Drew (1979) found that for a range of seagrass species, including Z. marina, lightsaturated gross photosynthetic rates increased in direct proportion to temperature increase up to a 

point between 30-35EC, above which thermal damage caused a rapid reduction. Noting that 

respiration rates were not so dramatically affected, he concluded that effects on gross and net 

photosynthetic rates differed only in slope. Bulthuis (1985, 1987) comprehensively reviewed 

temperature effects on photosynthesis and growth of seagrasses. He found general agreement 

that the photosynthetic capacity of seagrasses (most seagrass genera represented) is reduced at 

35-40° C. Within the limits of physiological tolerance (5-30° C), the rate of photosynthesis at 

light saturation, the dark respiration rate and the light compensation point more than double as 

temperature increases within the range experienced by most temperate zone seagrasses. The 

optimum temperature for photosynthesis decreases from 25-35° C at light saturation to as low as 

5° C as irradiance decreases. Marsh et al. (1986) demonstrated that ratios of maximum 

photosynthetic rates to respiration rates were highest at 5° C and declined markedly at higher and 

lower temperatures in Z. marina. Even short-term (15 min) leaf exposure to high temperatures 

(e.g., ≥30° C) reduced net photosynthesis, increased respiration and led to a reduction in P:R 

ratios. Burke et al. (1996), Thayer et al. (1975) and Evans et al. (1986) support the thesis that 

25° C may be an important threshold in that they note a negative carbon balance when water 

clarity conditions are low enough to reduce photosynthetic rates below that at light-saturation. 

Such conditions tend to compromise the ability of Z. marina to survive suboptimal weather 

scenarios (rainy, turbid springtime conditions, paralleling those of the early 1930’s severe 

declines), especially if carbon reserves had been exhausted the previous summer under high 

temperature/high light conditions. Zimmerman et al. (1995) note that daily Hsat requirements for 

Z. marina range from 2.5 to 4 hours; for subtidal habitats, biweekly tidal cycling when high 

turbidity exists could severely stress deeper plants by severely limiting available light. 

Sand-Jensen and Borum (1983) concluded that leaf productivity in Z. marina is limited 

by temperature, in that increasing temperatures increase maintenance costs by increasing dark 

respiration rates. However, Zimmerman et al. (1989) concluded from laboratory experiments 

that since Z. marina shows evidence of thermal acclimation, seasonal changes in ambient 

temperature may not significantly affect light-saturated photosynthesis (Hsat) requirements and 
9.8

whole-plant C balance. Rapid mortality at high temperatures during summer may result instead 

from thermal disruption of metabolism. 

The impact of temperature on photosynthesis and respiration may also be affected by 

both salinity variations and preadaptation. Biebl and McRoy (1971) showed that both subtidal 

and intertidal forms of Z. marina from Alaska maintained plasmatic resistance for short term (24 

hr.) exposures over a range of salinities (0-3.0x, i.e., 1.0x = 31, 3.0x = ~93) and between -6°C 

and 34°C. Within these limits, photosynthesis increased with temperature in the intertidal plants 

up to 35° C, but up to only 30° C in the subtidal form. McRoy (1970) also demonstrated that the 

plants from tidepools shifted their entire temperature-respiration relationship toward the 

consumption of more oxygen in summer for any given temperature. None of the experiments 

showed effects of enzyme denaturation at 30° C, failing to support the 30° C lethal limit 

proposed by Setchell (1929). 

In a comparison among species, Lutova and Feldman (1981) demonstrated that the 

thermostability of selected cell functions and enzymes in Z. noltii is approximately 4-5° C higher 

than that of Z. marina, consistent with a shallower distribution pattern. Physiological responses 

to temperature in Z. marina and Ruppia maritima in lower Chesapeake Bay are suggested as a 

partial basis for the difference in depth distribution of the species (Evans et al. 1986). Dennison 

(1987) showed that southern populations of Z. marina exhibit bimodal seasonal patterns of net 

photosynthesis, due to high respiration during summer months, while northern populations 

exhibit a more unimodal seasonal pattern, presumably due to temperature differences between 

southern and northern populations. 

Pérez-Lloréns and Niell (1993, 1994) reported on the ability of two morphotypes of Z. 

noltii (in southern Spain) to photosynthesize in air (in the intertidal); the photosynthetic rates of 

the narrow-leaved variety (higher in the intertidal) were higher in air than the wider-leaved 

variety (lower in the intertidal), suggesting local adaptation to elevated insolation and 

temperature, as well as resistance to desiccation (see also McMillan 1984). 

Investigations on Thalassia testudinum in Florida, the Gulf coast, and Caribbean have 

revealed that temperatures in excess of ~32° C will interfere with maintenance of ionic condition 

(Schroeder 1975), and that physiological changes such as chlorophyll a fluorescence values will 

occur prior to observable morphometric changes (Byron and Fourqurean 2004). Capone and 

Taylor (1980) found that C2H2 (acetylene) reduction rates were halved by a decrease of 10EC 

and that rates of N2 fixation varied ~20-fold, being maximal in late summer and minimal in 

January. Zieman (1970) determined that net leaf productivity was significantly temperature 

dependent, with optimal growth occurring from 23-31° C (see also Barber and Behrens 1985). 

Borum et al. (2005) and Rudnick et al. (2005) reviewed the mass die-off of Thalassia

beds in Florida Bay in the 1980s (Robblee et al. 1991 and others), concluding that toxicity was 

caused by sulfide invasion of the rhizome, which was preceded by hypoxia caused by accelerated 

respiration in response to temperatures elevated above the seasonal ambient level. 
9.9

Sand-Jensen and Borum (1983) (Z. marina in Danish waters) and Kerr and Strother 

(1985, 1989) (Z. muelleri in Australian waters) concluded that leaf productivity and 

photosynthetic ability respectively either were not affected by water temperature, or could be 

maintained under extreme conditions; secondary mechanisms such at those affecting dark 

respiration rates and rates of mineralization of the sediments were more likely to be affected. 

Vergeer et al. (1995) demonstrated that Z. marina under culture produced lower levels of 

phenolic compounds when grown at high temperature, while those subjected to high light 

intensity increased production of phenols. Phenolic compounds are good bactericides and 

fungicides, which might offer some protection against infection of Z. marina by Labyrinthula 

zosterae, the putative cause of eelgrass wasting disease. However, infection with Labyrinthula

itself proves also to have pronounced effects on the production of phenolic compounds (ibid.), 

confounding the protective effect of lower temperatures. 

Interactive effects of global warming with other anthropogenic factors may be 

particularly stressful to seagrass populations. Mesocosm experiments with eelgrass (Bintz et al. 

2003) which combined nutrient enrichment with sustained temperature elevation to 4° C above a 

9-year mean caused significant declines in number of leaves per shoot, shoot surface area and 

shoot growth rate. The authors concluded that widespread eelgrass declines in the Northeast 

U.S. may be due to the combination of nutrient enrichment of coastal waters and the increasing 

frequency of warmer than average summer water temperatures. Johnson et al. (2003) conclude 

that persistent replacement of Z. marina by widgeongrass (Ruppia maritima) in two bays in the 

San Diego area during and following the 1997-1998 ENSO event foretell the possibility of 

widespread long-range habitat conversions if average water temperatures increase 1.5 - 2.5° C 

due to global warming. 

Warmer than usual water temperatures have been proposed as a cause for the wide scale 

loss of eelgrass in the 1930's in Europe due to wasting disease. Giesen et al. (1990) propose a 

combination of high water temperatures, below-average sunshine and increased turbidity of 

coastal waters may have caused eelgrass to succumb to the saprophyte Labyrinthula macrocystis

(cf. L. zosterae). Elevated water temperatures may have provided a higher overwintering 

survival rate for the parasite (see also Vergeer et al. 1995). 

9.5 Effects of temperature on leaf growth, density and biomass

In Pacific coast populations, the relative proportion of vegetative compared to 

reproductive biomass is correlated to the length of growing season with temperatures between 15 

and 20° C, although irradiance level is an important co-factor (Felger and McRoy 1974). 

Poumian-Tapia and Ibarra-Obando (1999) demonstrated that seasonal changes in aboveground biomass and leaf area index (LAI) of Z. marina from a Mexican coastal lagoon were 

associated with water temperature. 
9.10

At the other end of the temperature spectrum, Harrison and Mann (1975) discounted 

temperature as an important factor in eelgrass growth for a subtidal population in Nova Scotia, as 

most of the vegetative growth occurred at temperatures less than 10° C, concluding that this 

population may be adapted to very different temperature regimes than the optimal range 

proposed by Setchell (1929). Wium-Andersen and Borum (1984) also concluded that Setchell’s 

growth regulation by temperature theory was not supported by their observations of Z. marina

populations in Denmark, observing that ~75% of seasonal variance in leaf growth rate could be 

attributed to variation in surface irradiance, while only ~6% could be attributed to variation in 

temperature. Concurrently, Borum (1980), showed experimentally that temperature-dependent 

light-saturated photosynthesis only occurred at mid-day in the uppermost part of the leaf canopy. 

A recent comprehensive review of eelgrass research in San Francisco Bay, CA. (WyllieEcheverria and Fonseca 2003) supported the original proposals of Setchell (1922, op. cit.) 

regarding the effects of temperature on the biotic responses of eelgrass (e.g., Phillips et al. 1983; 

Merkel and Associates 1999 (cited in Wyllie-Echeverria and Fonseca 2003); and Zimmerman et 

al. 1995). They conclude that transplantation and restoration efforts pay particular attention to 

season and temperature regimes in planning and scheduling, as many wintertime transplants 

were significantly inhibited by a combination of high turbidity and low temperatures. 

Kirkman et al. (1982) (New South Wales, Australia) demonstrated a closer relationship 

for Z. capricorni productivity to water temperature than to solar radiation. The partial 

correlation coefficients between growth and water temperature were 0.95 (holding instantaneous 

solar radiation fixed) and 0.77 (holding solar radiation lagged by 1 month fixed), indicating an 

association between growth and water temperature not accounted for by solar radiation or solar 

radiation lagged by 1 month. 

Marbà et al. (1996) found leaf and shoot growth in Mediterranean Z. marina populations 

to be associated primarily with average irradiance (R2

 = 0.43 leaf, 0.70 shoot), temperature 

variation was secondary (R2

 = 0.57 leaf, 0.10 shoot); Z. noltii, growing in the same area, but 

elevated in the intertidal with respect to Z. marina, was co-dependent on temperature (R2

 = 0.62 

leaf, 0.32 shoot) and light (R2

 = 0.64 leaf, 0.37 shoot), although the light relationship lagged by 

~1 month. 

Harrison (1982a), Phillips and Backman (1983), Orth and Moore (1986), and Evans et al. 

(1986) all provide evidence that temperatures in excess of 30° C lead to declines in condition, 

i.e., loss of leaf or meristematic tissue or defoliation, or suppression of growth. Growth often 

resumes as temperatures fall below the 30° C threshold (often into the early fall season), 

explaining bimodal biomass peaks observed especially at the southern end of the distribution of 

seagrass populations in the northern hemisphere. Within single estuarine systems, biomass 

measures may follow gradients of both salinity and temperature, illustrated by the Yaquina Bay 

(Oregon) distribution of Z. marina, where summer coastal upwelling and large tidal prism 

“pumping” provides relatively cool, nutrient-rich water to the lower estuary, while temperatures 

increase upstream, reflecting watershed runoff influence. Biomass measures (shoots per unit 

area, plant size) are greatest near the ocean, tapering off proceeding upstream into warmer water. 
9.11

During the summers of 1998-99, the overlying water column downstream was, on average, 

colder, more saline, less turbid, had more available light and higher dissolved inorganic nitrogen 

(DIN) than upstream. Temperature was negatively (but not significantly) correlated with Z. 

marina and green macroalgal cover R = - 0.52 and - 0.47, respectively, p > 0.1); summertime 

DIN and phosphate concentrations were very poorly correlated with either seagrass or 

macroalgal cover R < 0.18, p > 0.1) (Kentula and DeWitt 2003). 

Vegetative (leaf) growth in Thalassia testudinum is positively correlated with 

temperature, with the range 20-30° C being optimal for growth (Phillips 1960; Macauley, 

Clark, and Price 1988; Fletcher and Fletcher 1995; Irlandi et al. 2002; Tomasko and Hall 

1999). Leaf length and width both decrease when stressed by elevated temperature (Linton and 

Fisher 2004). Zieman (1975) reported that leaf growth rates for T. testudinum in Biscayne Bay, 

Florida (data collected in 1969-70), was highest when the temperature was between 28-31EC and 

salinities ~30, and at stressed habitats, the minimum growth rates were when salinity was low 

(13-15) and temperature was highest (34-35° C). 

9. 6 Desiccation Effects on Seagrasses 

9.6.1 Minimal depth limit

Seagrasses in general are not tolerant of exposure to aerial conditions (excepting Zostera 

japonica, cf. Z. americana, Z. nana, Z. noltii? - see den Hartog and Kuo, 2006), suggesting that 

the shallowest distribution should be at a depth below the MLW (Koch 2001). This relationship 

has clearly been shown in tropical seagrasses where aerial exposure associated with extreme 

tides results in seasonal losses of above-ground biomass (Vermaat et al. 1993; Erftemeijer and 

Herman 1994; De Iongh et al. 1995; Stapel et al. 1997). In an extreme case the upper margin of 

a Zostera noltii bed was described as Aburned@ following such an exposure (van Lent et al. 1991). 

For intertidal Z. capensis in South African estuaries (Adams and Bate 1994), “scorched” leaves 

did not recover and were sloughed off, but regrowth from the basal meristem quickly replaced 

the lost tissue. For the temperate zone eelgrass, Zostera marina, numerous authors have implied 

that desiccation is the probable cause for changes in seagrass abundance and morphology across 

the tidal gradient (Bayer 1979; Jacobs 1979; Kentula and McIntire 1986; Keddy 1987; Koch and 

Beer 1996). Thalassia testudinum starts to become exposed at about a 15-cm water depth, 

because the blades are somewhat rigid (Phillips 1960). This exposure results in desiccation, 

which is often more severe in winter than in summer in subtropical habitats, where winter 

“spring” tides, high insolation and low-humidity polar air combine to enhance desiccation of 

tidally-exposed plants (Strawn 1953, 1961; Phillips 1960). Desiccation could be further 

accelerated in winter due to increased transpiration resulting from higher wind speeds 

(Holmquist et al. 1989). 

Intertidal Z. marina is not found in the northern (where presumed ice scouring) and 

southern extremes of its range (with high summer temperatures) (Phillips et al. 1983b), 

respectively, preclude its growth. However, in the central portions of its range it is often present 

intertidally, where it can be found in greater density than subtidal populations. In these central 

ranges it is often found in three zones which have been defined by plant growth characteristics 
9.12

and related to the degree of tidal exposure. For example Bayer (1979) and Boese et al. (2003) 

found and defined these zones in Yaquina Bay (Newport, OR, semi-diurnal tide, range ~2.4 m) 

as: 1) a subtidal and lower intertidal perennial zone (below 0.25 m MLLW) which consists 

mainly of perennial shoots which grow vegetatively from below-ground rhizomes which persists 

throughout the year, 2) a transition zone between 0.25-0.75 MLLW consisting of annual shoots 

and perennial patches, and 3) an upper intertidal zone (0.75 to 1.5 m MLW) which is 

characterized by annual shoots which grow from seeds and are absent in winter. Roughly the 

same growth pattern is representative of eelgrass from the Wadden Sea (Netherlands), where the 

perennial bed was never observed above 0.20 m above MLW, followed by the transition zone 

(bare sediment) and a mid-intertidal zone of annual plants (van Katwijk et al. 2000). 

Occasionally Z. marina annuals and perennials are found above these zones, and are 

associated with micro-topographical features (e.g. tide pools, drainage channels) which would 

tend to retain water longer during ebb tides. Thus it appears that the upper limit for Z. marina is 

controlled not directly by tidal elevation but by the duration of water coverage (Kentula and 

McIntire 1986; Jacobs 1979). It is also possible that eelgrass biomass can structurally retard 

water loss by trapping water during ebb tides as has been reported for turtle grass in Florida Bay 

(Powell and Schaffner 1991). They observed that water was retained by dense turtle grass for up 

to eight hours during low tides and prevented desiccation. Intertidal macroalgae may play a 

somewhat similar role, as we have observed the low tide trapping of water during periods of high 

macroalgae accumulations in Yaquina Bay (Boese unpubl. data). It is also possible that the 

presence of epiphytes on the leaves of seagrass may reduce the desiccating effects of aerial 

exposure. Penhale and Smith (1977) noted that Z. marina plants, which had epiphytes removed, 

lost five times the amount of dissolved organic carbon following exposure to 1 h of laboratory 

desiccating conditions, when compared to plants with encrusting epiphytes. This suggested to 

them that the encrusted epiphytes trapped water interstitially during the receding tide, which 

diminished or prevented desiccation damage during aerial exposure. 

The zonal differences between Z. marina annual and perennial forms appear to be due to 

morphological differences. van Katwijk et al. (2000) noted in the intertidal areas of the Wadden 

sea that eelgrass annuals when exposed during low tide tend to lie flat on the moist sediment 

surface as opposed to perennials which had stiffer sheaths that could not lie flat on the sediment. 

These upright sheaths tended to desiccate rapidly when exposed. These observations were 

confirmed by Boese et al. (2003), who noted the same phenomena in Yaquina Bay eelgrass. 

Studies done on Z. noltii and Z. japonica have observed that these two eelgrass species 

are more tolerant of desiccation stress than Z. marina and can grow higher in the intertidal zone 

(Harrison 1982a; Harrison 1982b; Leuschner et al. 1998). Both of these species have smaller 

leaves and appear to sustain photosynthesis at lower leaf water content than Z. marina

(Leuschner et al. 1998). Similar morphological differences have been noted in tropical seagrass 

species with tolerance to high temperature and aerial exposure, correlating with smaller and 

narrower leaves (McMillan 1984). Although Z. marina is less tolerant of desiccation, exposed 

leaves are capable of photosynthesis as long as leaves are moist, and are also able to recover 

from mild desiccation when leaves are re-wetted (Leuschner and Rees 1993). 
9.13

9.6.2 Effect of tidal amplitude 

The interaction between tidal exposure and light availability are the principal forcing 

factors determining the depth range within which seagrass can survive. As waters become more 

turbid seagrasses will be limited to shallow waters where they increasingly are affected by aerial 

exposure (Koch 2001). Assuming a low tolerance of desiccation, the minimum depth (Zmin) was 

defined by Koch (2001) as half the tidal amplitude. Tidal amplitude in this case was defined as 

the difference between mean high and mean low water in areas with diurnal tides and as the 

difference between mean higher high water and mean lower low water in areas with semi-diurnal 

tides (Koch 2001). If the maximum depth that seagrass can grow (Zmax) is due to light limitation, 

then no seagrass can survive if Zmax≤Zmin. Thus as the tidal range increases, the depth range 

where seagrasses can occur is linearly decreased (Koch 2001). 

This theoretical relationship is supported by the results of Koch and Beer (1996). That 

study attempted to explain the disappearance of Z. marina from the western portions of Long 

Island Sound while it persisted in the eastern portions. Light attenuation had increased due to 

eutrophication and the greater tide range in the western portion of the sound appeared to narrow 

the depth range within which Z. marina could grow, making this shallow population vulnerable 

to storm events (Koch and Beer 1996). Koch and Beer (1996) concluded that regulatory 

management of seagrass should take into account tides as well as light. 

 

9.6.3 Effects of desiccation 

At the population level, aerial exposure and associated desiccation stress affects Z. 

marina in a variety of ways. With increasing intertidal exposure, shoot density is reduced, leaves 

become narrower, canopy height is reduced, annual plants become more prevalent and flowering 

increases; these effects are associated with perennial eelgrass meadows becoming progressively 

more patchy with increasing tide height (Bayer 1979; Jacobs 1979; Kentula and McIntire1986; 

Harrrison 1982a; Keddy 1987; Ruckleshaus 1994; Koch and Beer 1996; Boese and Robbins 

unpublished data). 

At the individual shoot level, desiccation reduces the plant=s photosynthetic ability 

(Leuschner and Rees 1993), and leaves become physically damaged (van Katwijk et al. 2000; 

Boese et al. 2003). The progression of damage is an initial drying of the leaf in and immediately 

adjacent to the exposed sheaths, followed by gradual loss of pigment over a 7- 14 d interval. 

These non-pigmented areas often occur in bands across individual leaves which probably affect 

translocation of photosynthetic products, and weaken the leaf making it more vulnerable to 

breakage. Desiccation-induced breakage can contribute to the reduced canopy height often 

observed in high intertidal plants (Boese et al. 2003). 

Desiccating events are episodic, and often associated with daylight spring tides during the 

summer, as illustrated in the seasonal losses of intertidal seagrass that occur in the tropics 

(Vermaat et al. 1993; Erftemeijer and Herman 1994; De Iongh et al. 1995; Stapel et al. 1997; van 

Lent et al. 1991). Similar episodic event have been observed in Yaquina Bay, Oregon during 
9.14

daylight low tides in the spring and early summer, although winter low tide exposures to freezing 

temperatures may result in similar damage (Boese et al. 2003). 

9. 7 Research Gaps

There is a general lack of research on intertidal seagrass populations and how desiccation 

stress directly affects individual shoots and populations. Development of plant and population 

level models of the effects of exposure on eelgrass would be useful. 

With respect to temperature effects, knowledge of site-specific temperature Awindows@

regarding vegetative growth, the initiation of reproduction (anthesis, pollination, seed set and 

dispersal), and upper thresholds for inhibition of growth, leaf loss and dark respiration seem to 

be the criteria that would be required for any restoration or preservation effort. Elevated levels 

of nitrogen, either as nitrate or ammonia, in combination with even slightly elevated above norm 

water temperatures could be held as a violation of acceptable habitat conditions. Additional 

multifactorial, mesocosm experiments with varying temperature and nitrogen concentrations 

could help define acceptable boundaries from a regulatory standpoint. 

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10.1 

10.0 The Effects of Bioturbation and Bioirrigation on Seagrasses 

Theodore H. DeWitt

10.1 Background

 Marine and estuarine fauna can diminish or enhance seagrass condition simply as a result 

of their burrowing, sediment re-working, feeding activities (e.g., bioturbation), or their 

ventilation of burrows and tubes (e.g., bioirrigation). Compared to other limiting or facilitating 

factors, little mention is made of bioturbation or bioirrigation in the seagrass literature prior to 

the early 1980’s (i.e., prior to Suchanek 1983). Since then, however, several studies (see below) 

and one review (Short and Wyllie-Echeverria 1996) present compelling evidence that these 

processes may profoundly affect the local distribution, abundance, and productivity of 

seagrasses. 

 Generally speaking, bioturbation negatively affects seagrasses (although it may do the 

reverse in some circumstances), whereas bioirrigation has positive or neutral effects. These 

processes are ecologically distinct from other important plant-animal interactions that can affect 

seagrasses, such as herbivory and epifaunal colonization of seagrass leaves. Bioturbation is the 

biogenic transport of particulate matter (inorganic and organic) within the sediment column, 

resulting from burrowing, excavation, and feeding activities of infaunal and epifaunal 

invertebrates, demersal fish and marine mammals, and birds. Bioturbation includes the vertical 

and horizontal mixing of particulate matter and associated porewater within the sediment 

column, the deposition of sediment from depth to the sediment surface (or sediment water 

interface), and the resuspension of bedded sediments into the water column. The negative 

impacts of bioturbation on seagrasses are mediated through burial, shading, erosion, or damage 

to roots. Bioirrigation is the biogenic pumping of water and solutes from waters overlying the 

seafloor into the sediment column due to organismal activities within burrows or tubes (i.e., 

ventilation for respiration, feeding, defecation, or excavation) and passive ventilation of burrows 

or tubes caused by the Bernoulli-effect of bottom currents flowing over burrow openings 

(Allanson et al. 1992). The positive effects of bioirrigation to seagrasses, while less studied than 

the negative effects of bioturbation, potentially include oxygenation of sediments, import of 

nitrogen-rich particulate matter into the sediment (via suspension feeding), stimulation of organic 

matter remineralization, and enhancement of oxidation or removal of toxic substances in 

sediments (i.e., hydrogen sulfide, ammonia) (see reviews by Aller 1988; Kristensen 1988; 

Pearson 2001). Bioturbation can have similar impacts on sediment geochemistry as bioirrigation 

(i.e., enhancement of organic matter remineralization, oxygenation of sediments, and burial of 

sediment organic matter) and, therefore can potentially benefit seagrasses under some 

circumstances. 

 Bioturbation and bioirrigation may affect the impact of other stressors, particularly 

nutrient enrichment, on seagrasses. Little research has been conducted to specifically examine 

interactions between bioturbation/bioirrigation, nutrient enrichment and seagrass condition, 

although hypotheses regarding consequences of such interactions can be proposed. Both 
10.2 

bioturbation and bioirrigation can reduce the adverse impacts of eutrophication by increasing the 

rates of organic matter burial, below-ground decomposition, and nutrient cycling; by 

oxygenating sediment porewater; and by increasing the flux of dissolved metabolites and 

nutrients from sediments into the water column. Bioturbation can potentially increase the 

adverse effects of nutrient enrichment by decreasing water column light levels (i.e., by increasing 

turbidity as a result of sediment resuspension) or by acting as an independent stressor on 

nutrient-stressed plants. These interactions, their relevance to seagrass management, and gaps in 

scientific knowledge are discussed in the final section of this chapter. 

 As with all limiting or facilitating factors, the energetic and spatial magnitude of 

bioturbation and bioirrigation determine the importance (and the direction of the effect) of these 

processes on seagrasses. At low intensity, neither process is likely to have measurable effects on 

seagrasses. The research reviewed here focuses only on those cases where significant 

correlations or experimental evidence suggest a cause and effect relationship. 

10.2 Role of Bioturbation in Limiting Seagrass Populations

 Bioturbation by burrowing shrimp, sting rays, crabs, polychaete worms, and echinoderms 

have been reported to adversely affect seagrass recruitment, growth, and survival by burial or 

uprooting of seeds, seedlings, shoots and patches of seagrass (Table 10.1 and Figure 10.1). 

Dugongs and manatees mix and resuspend sediments as they forage for seagrass leaves and 

rhizomes (Packard 1984; Preen 1995; Domning 2001), but the impact of their bioturbation on 

seagrasses has not been evaluated; thus, it would be premature to include sirenians as 

bioturbators in this review. In some cases, cause and effect were inferred from disjunct 

distributions of populations of bioturbators and seagrasses, and observations of burial of seagrass 

shoots by excavated sediments where populations overlapped. Stronger evidence was provided 

in several studies featuring field or laboratory experiments in which seagrasses were transplanted 

into sediments containing different densities of bioturbators, or cages were erected to exclude 

bioturbators from seagrasses. The effects of each group of bioturbators on seagrasses are 

reviewed first, followed by a summary of the mechanisms by which bioturbation disturbs 

seagrasses. 

10.2.1 Burrowing Shrimp Bioturbation 

 The most frequently reported bioturbator-seagrass interactions are those involving 

thalassinid burrowing shrimp (Arthropoda: Decapoda: Thalassinidae) as the bioturbators. 

Burrowing shrimp have been reported to variously affect seagrass recruitment (Dumbauld and 

Wyllie-Echeverria 2003), shoot growth, productivity and survival (Suchanek 1983; Harrison 

1987; Molenaar and Meinesz 1995; Siebert and Branch 2007), population distribution (Suchanek 

1983; Harrison 1987; Pranovi et al.1996; Siebert and Branch 2005; Dumbauld and WyllieEcheverria 2003), and community structure (Duarte et al. 1997). Post-larval stages of burrowing 

shrimp live in extensive burrow galleries excavated in estuarine and marine sediments, have 

prodigious rates of sediment turnover (reviewed by Rowden and Jones 1993), and are important 

ecosystem engineering species in many coastal systems because of their influence on benthic 
10.3 

Figure 10.1. Illustration of some of the adverse effects of bioturbation on seagrasses (adapted 

from Suchanek 1983, Figure 4). Arrows show paths of sediment subduction, advection 

and resuspension as result of sediment reworking by burrowing shrimp. 

Table 10.1 Summary of sources of significant bioturbation impacts for different seagrasses 

Bioturbation Source Seagrass Species Reference 

Burrowing shrimp Zostera marina, Z. japonica Thompson and Pritchard 1969 

Z. marina, Z. japonica Harrison 1987 

Z. japonica Dumbauld and WyllieEcheverria 2003 

Z. japonica, Z. capricorni Berkenbusch et al. 2007 

Z. noltii Pranovi et al. 1996 

Z. capensis Siebert and Branch 2005, 2006 

Z. capensis Angel et al. 2006 

Posidonia oceanica Molenaar and Meinesz 1995

Halodule uninervis Duarte et al. 1997 

Thalassia testudinum Suchanek 1983 

T. hemprichii Duarte et al. 1997 

Syringodium isoetifolium Duarte et al. 1997 

current

plume of ejected

sediment covering

seagrasses

burrow under-cutting

seagrasses

feeding pit excavating

seagrasses

burrow mounds from

infaunal bioturbator,

such as burrowing

shrimp
10.4 

Cymodocea rotundata Duarte et al. 1997 

C. serrulata Duarte et al. 1997 

Stingrays Z. marina Orth 1975 

Z. marina Merkel 1990 

Z. marina, H. wrightii Townsend and Fonseca 1998 

H. wrightii Fonseca et al. 1994 

T. testudinum Ogden 1980 

T. testudinum Zieman 1982 

T. testudinum Valentine et al. 1994 

Syringodium sp. Fonseca et al. 1998 

Crabs Z. capricorni (=novazelandica) Woods and Schiel 1997 

Z. marina Davis and Short 1997 

Z. marina Davis et al. 1998 

T. testudinum Valentine et al. 1994 

Polychaetes Z. noltii Phillipart 1994 

Z. marina Luckenbach and Orth 1999 

Z. marina Davis and Short 1997 

Zostera spp. Hughes et al. 2000 

Echinoderms Z. marina Backman 1984 

community structure and geochemical processes (Berkenbusch and Rowden 2003 and 2007; 

DeWitt et al. 2004; Tamaki 2004; Siebert and Branch 2006; Berkenbusch et al. 2007). 

Burrowing shrimp species occur along virtually all coasts world-wide (Dworschak 2000) and 

frequently occur in habitats that can sustain seagrass populations. 

 Suchanek (1983) conducted the first published study that connected burrowing shrimp 

bioturbation with a decrease in seagrass condition and abundance. He noted a negative 

correlation in the spatial distribution of subtidal ghost shrimp (Callianassa spp.) and turtlegrass 

(Thalassia testudinum) in two bays on St. Croix (US Virgin Islands). Subtidal populations of 

four ghost shrimp species live in those bays, and one species (C. rathbunae) turns over ~2.6 kg 

sediment m-2 d-1, producing large mounds (ca. 19 cm height) of ejected sediment, at densities of 

~6-7 mounds m-2. Turtlegrass productivity and percent cover decreased in proportion with 

increasing density of ghost shrimp mounds (Figure 10.2.A). Shoots and leaves of T. testudinum

transplanted into areas of high ghost shrimp density steadily decline in abundance over a five 

month period whereas turtlegrass transplanted into areas with low shrimp densities were “lush 
10.5 

Figure 10.2. Effects of bioturbating burrowing shrimp (Callianassa spp.) on turtle grass 

(Thalassia testudinum) on subtidal sand flats on St. Croix island (US Virgin Islands); 

from Suchanek (1983). A. Field data showing inverse relationships between burrowing 

shrimp abundance and seagrass percent cover at Tague Bay (solid line) and Great Pond 

Bay (dashed line), and with seagrass productivity (dotted line) at Great Pond Bay. B. 

Results of a field experiment showing temporal change in number of seagrass blades and 

shoots (mean +/- 1 SE) from turtle grass transplanted into areas with high densities 

(experimental; dashed line) or low densities (control; solid line) of burrowing shrimp. 

Reproduced with permission from Journal of Marine Research and T.H. Suchanek. 

10

0

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B
10.6 

and healthy” (Figure 10.2.B). Suchanek (1983) proposed that burial by excavated sediments, or 

shading due to either turbidity or deposition of fine-grained resuspended sediment onto leaves 

were the likely mechanisms causing decreased turtlegrass productivity and survival. He 

concluded that burrowing shrimp could severely limit T. testudinum distribution and negatively 

affect the habitat and energetic value of Caribbean seagrass beds. Whereas Suchanek (1983) 

characterized the direct effects of Callianassa spp. on turtlegrass, Wanless et al. (1988) reported 

an indirect effect of bioturbation by these shrimp on T. testudinum, positing that hurricane 

transported sediments from shrimp’s mounds (which had a characteristic sediment grain size 

distribution) contributed to the smothering of seagrasses. 

 In subtidal sandy habitats (specifically the French Mediterranean), Molenaar and Meinesz 

(1995) reported that populations of the shrimp C. tyrrhena reduced survival of Posidonia 

oceanica transplanted by ~40% relative to sands not inhabited by burrowing shrimp. They 

proposed that the transplants were buried by sediments excavated by C. tyrrhena, which is a 

prodigious bioturbator; at densities of 16 shrimp m-2, they deposit ~600 g m-2 d-1 of sand on the 

seafloor (Ott 1976 in Molenaar and Meinesz 1995). Pranovi et al. (1996, F. Pranovi, personal 

communication) observed a sharp decrease in the presence of the seagrass, Z. noltii, in Venice 

Lagoon (Italy) during the early 1990’s and a coincident increase in the abundance of the mud 

shrimp, Upogebia pusilla; this pattern was also found in spatial surveys conducted in the lagoon. 

Bioturbation by the shrimp was proposed as a mechanism to explain the distribution patterns, but 

no experiments were conducted to test the hypothesis (F. Pranovi, personal communication). 

 Callianassid burrowing shrimp can limit the intertidal distribution of seagrasses. In 

Langebaan Lagoon (west coast of South Africa), Branch et al. (2003) and Siebert and Branch 

(2005) observed that the seagrass, Z. capensis, was relegated to the upper intertidal when cooccurring on tide flats with the ghost shrimp C. kraussi. Using an elegant field experiment, 

Siebert and Branch (2006) demonstrated that patches of seagrass increased in area when 

transplanted into lower intertidal areas from which the shrimp had been extirpated. However, if 

high densities of ghost shrimp were present, the seagrass shoots died and the patches declined in 

size. Sediment resuspension and deposition were much greater in shrimp-dominated habitats 

than in the seagrass beds, and smothering and burial by shrimp-excavated sediments were 

suspected to be main mechanisms that limited the lower limit of Z. capensis distribution (Siebert 

and Branch, 2005, 2006). However, they also found that these callianassid shrimp did not 

colonize areas immediately down-slope of the seagrass bed, nor open patches within the seagrass 

meadow; processes causing these patterns are discussed later in this section. In Papanui Inlet 

(South Island, New Zealand), Berkenbusch et al. (2007) reciprocally transplanted ghost shrimp 

(C. filholi) or small patches of seagrass (Z. capricorni) into field enclosures containing the other 

species (i.e., seagrass or shrimp, respectively), and compared the abundances of manipulated 

shrimp and seagrass relative to untransplanted, enclosed patches of ghost shrimp and seagrass 

(i.e., control treatments) over 6 months. In the presence of ghost shrimp, Z. filholi biomass and 

shoot density decreased. Bioturbation was considerably higher in enclosures containing ghost 

shrimp; burial by excavated sediments was presumably the mechanism causing the demise of Z. 

filholi transplants. However, in seagrass-dominated enclosures, transplanted ghost shrimp 

abundances declined whereas the seagrass abundance was unchanged relative to the 
10.7 

unmanipulated control treatment. They inferred that the density of the root mat prevented the 

ghost shrimp from burrowing into the sediment, as reported or inferred in other studies (i.e., 

Brenchley 1982; Harrison 1987; Siebert and Branch 2006). 

 Interactions between intertidal burrowing shrimp and seagrasses in NE Pacific estuaries 

have been the subject of several studies, particularly focused on the Callianassid ghost shrimp 

(Neotrypaea [=Callianassa] californiensis) and the native eelgrass Z. marina and the invasive 

Japanese eelgrass Z. japonica. The outcomes of the interactions, however, were not 

unidirectional. In British Columbia (Canada), Harrison (1987) observed negative correlations in 

the abundances of Z. marina, Z. japonica and ghost shrimp, and proposed that ghost shrimp 

bioturbation could prevent the seagrasses from colonizing unvegetated sediments. Z. japonica

shoots transplanted into dense patches N. californiensis disappeared within a few weeks due to 

burial by sediments excavated by the shrimp. He proposed that interactions between ghost 

shrimp and Zostera spp. could limit the distribution of seagrasses, but that the seagrasses also 

could limit the shrimp’s distribution if the water column was sufficiently clear to allow early 

lateral growth of the seagrasses in early Spring (see section 10.4). Dumbauld and WyllieEcheverria (2003) reported that burrowing shrimp bioturbation was an important factor 

determining recruitment success for Z. japonica seedlings in intertidal sediments of Willapa Bay 

estuary (Washington, USA). They observed that, following the eradication of burrowing shrimp 

(predominantly the ghost shrimp, N. californiensis) by pesticide application, Z. japonica and Z. 

marina often colonized intertidal habitats, whereas seagrass seedlings were sparse or absent from 

ghost shrimp dominated areas. They measured no significant difference in vertical distribution, 

viability or germination success of Z. japonica seeds among experimental plots (shrimpdominated sediments treated, or not, with pesticide to kill the shrimp), but decreased growth and 

loss of Z. japonica seedlings in plots containing ghost shrimp. Burial or light limitation due to 

sediment resuspension were suggested as the mechanisms that reduced seedling growth and 

survival. They proposed that bioturbation could affect Zostera seed germination if seeds were 

buried to >12 cm depth under sediment in light of Bigley’s (1981) finding that the hypocotyls of 

sprouts fail to reach the sediment surface if germinated below that depth. Dumbauld and WyllieEcheverria (2003) concluded that “seedling survival is important for recruitment from the seed 

bank and while shrimp may cause some loss and decreased germination success, the effect of 

bioturbation on seedling survival, in part due to the coincident timing of shrimp activity and 

sprouting in early spring, is more important at the population level.” (pg. 37). 

 Thompson and Pritchard (1969) noted that the upper intertidal limit of Z. marina in 

Oregon matched the lower limit of dense ghost shrimp populations and that seasonal expansion 

of either species of eelgrass on sand flats was correlated with declines in the abundance of ghost 

shrimp. In contrast to Dumbauld and Wyllie-Echeverria (2003), Thompson and Pritchard (1969) 

found that ghost shrimp densities declined in experimentally created plots of Z. japonica. 

Berkenbusch et al. (2007) reported much the same result from a field study in which ghost 

shrimp or small patches of Japanese eelgrass were reciprocally transplanted into field enclosures 

containing the other species (i.e., Z. japonica or shrimp, respectively). In all cases, the patches 

of Z. japonica grew whereas the abundance of N. californiensis declined. They suggested that a 

combination of rapid growth by the eelgrass and relatively low bioturbation by N. californiensis
10.8 

(as compared to it’s New Zealand counterpart, C. filholi) allowed the invasive eelgrass to be the 

superior competitor in these experiments. In laboratory experiments, Brenchley (1982) found 

that the root and rhizome mat of Z. marina greatly inhibited burrowing by the burrowing shrimp 

N. californiensis and Upogebia pugettensis. For presumably the same reasons, Harrison (1987) 

observed that N. californiensis were largely unsuccessful at colonizing dense Z. marina patches 

on a British Columbia tide flat. The key difference between the studies of Dumbauld and 

Wyllie-Echeverria (2003) and those of Thompson and Pritchard (1969) and Berkenbusch et al. 

(2007) is the life history stages of the seagrasses that were brought into conflict with the ghost 

shrimp. Synthesizing the results of experimental studies on interactions between N. 

californiensis and Zostera spp. in NE Pacific estuaries, it appears that bioturbation by ghost 

shrimp can suppress the germination of eelgrass seeds and the survival of seedlings (at least for 

Z. japonica and probably for Z. marina), but that established patches of adult eelgrass are 

resistant to adverse impacts from ghost shrimp probably because the root and rhizome mat 

inhibits burrowing by the shrimp. 

 Thom et al. (2003) found burrowing shrimp co-existing with intertidal Z. marina in Coos 

Bay (Oregon, USA) and Willapa Bay (Washington, USA) estuaries, and found no evidence for 

negative interactions between these organisms. However, they did not report which species of 

burrowing shrimp they encountered in the eelgrass beds, and it is possible that they saw the 

burrows of the mud shrimp, Upogebia pugettensis, rather than N. californiensis. Upogebia 

pugettensis is common in Z. marina beds in estuaries of California, Oregon and Washington 

(USA), including Coos Bay and Willapa Bay, whereas N. californiensis is relatively uncommon 

in eelgrass beds (T. H. DeWitt, unpublished data). However, N. californiensis is common in 

high intertidal beds of Z. japonica (Dumbauld and Wyllie-Echeverria 2003; T. H. DeWitt, 

personal observation). No research has been published investigating why these two species of 

burrowing shrimp have different patterns of coexistence with Z. marina; possible reasons include 

different rates of sediment turnover (which are lower for U. pugettensis) and thus less 

disturbance to eelgrass by mud shrimp (T. H. DeWitt, personal observation), different 

capabilities to burrow through the eelgrass root and rhizome mat (Brenchley 1982), or increased 

susceptibility of N. californiensis to predation in lower intertidal and subtidal eelgrass habitats 

(Posey 1986). In South Africa, Siebert and Branch (2005, 2006) found high densities of the mud 

shrimp U. africana coexisting with Z. capensis in the upper intertidal where both species 

apparently found refuge from interactions with the ghost shrimp, C. kraussi). Upogebiid shrimp 

are less vigorous bioturbators than Callianassid burrowing shrimp largely because Upogebiids 

are primarily suspension feeders and thus do not need to constantly excavate and ingest organicrich sediments as do deposit-feeding Callianassids. 

 In contrast to the suppression of populations of seagrasses in single-species meadows, 

burrowing shrimp bioturbation can increase seagrass species diversity within mixed-species 

meadows. Cumulative disturbance by burrowing shrimp to several species of tropical seagrasses 

in the Silaqui and Santiago Islands (Philippines) was estimated to be greater than the disturbance 

caused by sediment transport associated with hurricanes and typhoons (Duarte et al. 1997). They 

conducted a field experiment to measure the effects of episodic deposition of sediment on the 

growth, survival, and demography of seven seagrass species (Thalassia hemprichii, Enhalus 
10.9 

acoroides, Cymodocea rotundata, C. serrulata, Halodule uninervis, Syringodium isoetifolium, 

and Halophilla ovalis, listed by relative abundance). Seagrass species responded variably to 

sediment deposition treatments, one (E. acoroides) was relatively unaffected by any of the 

deposition treatments; two (T. hemprichii, C. rotundata) exhibited sharp declines in response to 

the moderate and high sediment deposition treatments; three (H. uninervis, S. isoetifolium, C. 

serrulata) declined initially and then recovered in all deposition treatments; and one (H. ovalis) 

opportunistically increased in abundance in most treatments. Differential responses to sediment 

loading was proposed be an important mechanism for maintaining high seagrass species diversity 

in this meadow by having the greatest negative impact on abundant seagrass species (presumably 

competitive dominant species, although that was not tested directly), and opening space for 

colonization by competitively subordinate seagrass species. Duarte et al. (1997) noted that 

natural disturbances affecting sediment transport ranged in scale from sub-meter (burrowing 

shrimp) to 103

 m (hurricanes and typhoons), with the frequency of occurrence being 

approximately inverse of their size (i.e., monthly for burrowing shrimp mound duration vs. 

“rarely” for hurricanes and typhoons). Hence, the cumulative disturbance of small-scale but 

frequent sediment reworking by burrowing shrimp resulted in an estimated twice-yearly 

reworking of the seagrass meadow. The scale and frequency of disturbance by burrowing 

shrimp bioturbation was sufficient to cause major growth and population responses by the 

seagrasses in this tropical meadow, and sustain higher seagrass species diversity than under 

conditions of very low or very high disturbance (sensu the intermediate disturbance hypothesis; 

Sousa 1984). 

 In summary, disturbance from burrowing shrimp bioturbation (particularly Callianassid 

ghost shrimp) significantly disturbs several species of seagrasses including those with deep 

rhizomes (e.g., Thalassia testudinum), can affect all life stages of seagrasses (e.g., seeds, 

seedlings and shoots), can limit the distribution and abundance of seagrasses in some locations, 

and can modify the biodiversity of seagrasses in mixed species meadows. In some cases, the 

early life stages of seagrasses are more susceptible to burrowing shrimp bioturbation than are 

adult life stages. The interaction between seagrasses and burrowing shrimp is not, however, 

always tilted in favor of the bioturbator. Dense root mats have been shown or suggested to 

prevent Callianassid ghost shrimp from burrowing into sediments within established eelgrass 

beds, and Upogebiid mud shrimp coexist with Zostera in at least two regions. Finally, the 

outcome of shrimp-seagrass interactions can be tilted to favor the shrimp if water quality 

(particularly, turbidity) slows seagrass growth (see section 10.4). 

10.2.2 Stingray Bioturbation 

 Stingrays (Vertebrata: Chondrichthyes: Myliobatiformes) forage for benthic invertebrate 

prey by excavating sediments using jets of water blown from the mouth and plunger-like suction 

created with the pectoral disc (Martin 2003). In the process, they excavate broad, shallow pits of 

width approximately that of a ray’s disc width and depths of ~10-20 cm (Valentine et al. 1994), 

often removing seagrass shoots and rhizomes in the process. Schools of cownose rays 

(Rhinoptera bonasus) destroyed large areas of one Z. marina meadow (ca. 4 km2

 y

-1) and were 

suspected of causing similar damage at six other sites in Chesapeake Bay (Virginia, USA) (Orth 

1975). Similarly, southern stingrays (Dasyatis americanus) and spotted eagle rays (Aetobatis 
10.10 

narinari) were observed to damage turtlegrass (T. testudinum) habitat while feeding (Ogden 

1980; Zieman 1982 in Valentine et al.1994). 

 Valentine et al. (1994) documented the persistence of large unvegetated patches within T. 

testudinum meadows in St. Joseph Bay (Florida, USA) that appeared to be created by herbivory 

or bioturbation, possibly by stingrays. Three stingray species (R. bonasus, D. americanus, and 

D. sabina) released within large experimental cages (72 m-2) over turtlegrass beds were observed 

to dig feeding pits and dislodge T. testudinum shoots. However, only the largest individuals of 

the southern stingray (D. americanus; disc width >90 cm) dug pits sufficiently deep to damage 

seagrass rhizomes and cause a decrease in below-ground biomass. Based on the scarcity of large 

southern stingrays and paucity of large sting ray pits recorded in field surveys, Valentine et al. 

(1994) concluded that stingrays were unable to create the persistent bare patches in these T. 

testudinum meadows. 

 By contrast, Townsend and Fonseca (1998) determined that large bioturbation pits, 

possibly created by stingrays, were deeper (mean 4 cm, maximum 12 cm) than the rhizome depth 

of the predominant seagrasses (Z. marina and Halodule wrightii; 1 to 5 cm) in mixed-species 

seagrass meadows (North Carolina, USA). Many smaller pits, possibly created by crabs, were 

also deeper than the rhizomes. Therefore bioturbators could damage seagrass beds through 

disruption of rhizomes and roots, or dislodge seeds or seedlings. Pits occupied only ~1% of the 

area, but owing to the frequency of pit formation and persistence of pits, Townsend and Fonseca 

(1998) estimated that every square meter of seafloor was disturbed at a rate of 1.2 y-1. They 

suggest that disturbance by bioturbation is an important process creating patchiness in seagrass 

beds and disrupting linkages among seagrass patches, increasing the risk of erosion of the 

seagrass patches by currents and waves (Fonseca and Bell 1998). Thus, bioturbation may be 

important for generating and maintaining landscape-scale distribution patterns of seagrasses. 

 In addition to affecting extant populations of seagrasses, bioturbation by stingrays has 

been identified as an important source of disturbance for seagrass restoration. Bioturbation by 

foraging round stingrays (Urolophus halleri) was identified as a potential hindrance to 

restoration of eelgrass (Z. marina) in San Diego Bay (California, USA) (Merkel 1990). 

Transplanted eelgrass protected by stingray exclusion barriers (i.e., fences, stakes, erosion mats) 

had higher short-term (23 d) survival than unprotected planting units in soft sediments where ray 

feeding pits were abundant (Merkel 1990). Stingrays caused a loss of >50% of seagrass (H. 

wrightii) transplant units in Tampa Bay (Florida, USA), and cages constructed of 2.5 cm mesh 

galvanized chicken wire improved H. wrightii survival to 60% relative to <1% survival in 

uncaged controls (Fonseca et al.1994). Stingrays’ impacts to transplanted seagrass can occur 

very rapidly; Fonseca et al. (1998) reported 100% loss of Halodule sp. and Syringodium sp. 

transplants within 24 h of planting where stingray-exclusion cages were not used. 

10.2.3 Crab Bioturbation 

 A handful of studies have reported that bioturbation by crabs and lobsters (Arthropoda: 

Decapoda) can damage seagrass shoots, rhizomes and roots, eroding edges of seagrass beds or 

opening space within beds, and thereby affecting seagrass distribution and abundance. Valentine 
10.11 

et al. (1994) determined that bioturbation by stone crabs (Menippe spp.) was the most likely 

cause of open patches in T. testudinum meadows in St. Joseph Bay, although stingrays had at 

first appeared to be the culprits (see above). Stone crabs were most common along margins of 

seagrass beds, where they build large burrows at oblique angles just under the turtle grass 

rhizosphere. Burrowing undermines the edge of the seagrass bed, causing sections of the grass 

mat to dislodge from the sediments. This often results in the in-filling or collapse of the stone 

crab burrows, requiring continual burrow maintenance by the crabs (e.g. extension under the new 

edge of the seagrass bed) and thus enlarging the area of seagrass disturbed by the crabs. As 

Valentine et al. (1994) point out, stone crab burrowing appears to create a positive feedback loop 

for recession of the edge of T. testudinum beds. Stone crabs caused twice as much recession in 

the edges of turtlegrass patches, compared to seagrass margins without crabs. Damage to 

turtlegrass rhizomes due to bioturbation may have long-lasting consequences because of slow 

rhizome growth and patch recolonization (Williams 1990). 

 On rocky intertidal platforms (Kaikoura, New Zealand), burrowing crabs 

(Macrophthalmus hirtipes) are responsible for significant erosion of the edges of seagrass 

patches (Z. novazelandica) (Woods and Schiel 1997). M. hirtipes burrows, found predominantly 

at the edges of the seagrass patches, persist longer in seagrass patches than on open mudflats, 

possibly because the interconnecting rhizomes and roots reduce cave-ins of burrow walls. As 

with stone crabs, burrowing by M. hirtipes appears to disrupt the sediment-binding properties of 

Z. novazelandica’s rhizome-root mat, dislodging the seagrass, requiring the crabs to extend their 

burrows further into the seagrass patch, etc., setting up a positive feedback loop that perpetuates 

erosion of the bed margin. Once the seagrass patches began to decline at their edges, the 

interiors of crab burrows were exposed, creating a greater surface area for further erosion by 

waves and currents. The area immediately surrounding the burrow is rapidly undermined and 

torn away, ultimately accelerating the erosion of the seagrass patch margin. In addition to being 

disturbed through this crab’s bioturbation, Z. novazelandica leaves, roots and rhizomes are eaten 

by the omnivorous M. hirtipes (Woods and Schiel 1997). 

 

 These two studies illustrate two important points about the effects of bioturbator impacts 

on seagrasses. First, bioturbation impacts to the edges of seagrass patches may be more 

damaging than those in the middle of the patches because exposure of the edge may make the 

patch more vulnerable to subsequent erosion by water movement. Second, damage to seagrass 

patch edges may set up a positive feedback process that accelerates erosion of the seagrass patch, 

either by increased bioturbation or by hydrodynamic forces. 

 Davis and Short (1997) and Davis et al. (1998) reported that non-native green crabs 

(Carcinus maenas) disturbed eelgrass transplants (Z. marina) at restoration sites in Great Bay 

Estuary (New Hampshire, USA). Garbary and Miller (2006) reported that bioturbation by green 

crabs was responsible for the nearly total loss of eelgrass from Antigonish Harbour (Nova Scotia, 

Canada). Davis et al. (1998) observed that green crabs damaged naturally occurring and 

transplanted eelgrass shoots by tearing or cutting the sheath bundle during their burrowing 

activities and while foraging for infaunal prey. Mesocosm experiments demonstrated that 

disturbance by green crabs resulted in loss of 39% of transplanted Z. marina shoots within 1 wk 
10.12 

of exposure to the crabs (Davis et al.1998). Cages alone were not successful at keeping C. 

maenas away from transplants, but using crab traps in addition to exclusion cages improved the 

survival rate of transplanted eelgrass (Davis and Short 1997). 

 Bioturbation associated with horseshoe crabs (Limulus polyphemus) foraging activity in 

Great Bay Estuary (New Hampshire, USA) uprooted unprotected established plants and 

transplanted shoots of Z. marina (Davis and Short 1997); wire mesh cages successfully protected 

transplants from disturbance by horseshoe crabs. Other epibenthic crabs (such as Callinectes 

sapidus, Cancer magister), and lobsters (Homerus americanus) also burrow into sediments 

within seagrass patches, and may excavate open space in the form of small pits in seagrass beds 

(Townsend and Fonseca 1998; Valentine et al.1994; Short and Wyllie-Echeverria 1996; Short et 

al. 2001). However, the extent to which bioturbation by these crabs significantly affects 

seagrasses distribution or survival has not been reported. 

10.2.4 Polychaete Bioturbation 

 Four studies reported that bioturbation by polychaete worms (Annelida: Polychaeta) can 

adversely affect seagrass populations, two of which provided experimental evidence to support 

that hypothesis. Beneficial aspects of polychaete bioturbation have also been reported, and will 

be discussed in a later section. Phillipart (1994) reported that populations of Z. noltii and the 

lugworm, Arenicola marina, maintained non-overlapping distributions on tide flats of the Dutch 

Wadden Sea (Terschelling, The Netherlands). The species’ distributions met at an abrupt border, 

even though both had similar substrate and emersion limitations. Seagrass shoots transplanted 

into tide flat plots containing high lugworm density (ca. 68 worms m-2) soon had sediments 

deposited upon them from the worms’ fecal castings and material excavated from burrows. 

Shoots transplanted into plots from which lugworms had been excluded had substantially higher 

biomass and abundance (cover) than in plots where lugworms were absent. In the presence of 

lugworms, Z. noltii shoots completely disappeared within 6 wk. Phillipart (1994) suggests that 

increased bioturbation, associated with the population expansion of A. marina between 1970 and 

1990, may have been a major contributor to decline in Z. noltii on Wadden Sea tide flats. In 

contrast, van Katwijk and Hermus (2000) report that lugworm bioturbation did not affect the 

survival of transplanted Z. marina shoots on wave-exposed Dutch Wadden Sea tide flats. This 

result was inferred from shoot survival in a series of exclosures designed to reduce wave and 

current effects that also excluded lugworms and other large macrofauna. However, van Katwijk 

and Hermus’ (2000) study did not include treatments that manipulated bioturbation 

independently of hydrodynamics, and thus, any effects of bioturbators were confounded with 

those of water movement. 

 Sediment reworking by the head-down deposit-feeding polychaete, Clymenella torquata, 

had both detrimental and beneficial effects on the dispersal and survival of Z. marina seeds in 

Chesapeake Bay (Virginia, USA) (Luckenbach and Orth 1999). In laboratory flumes, lateral 

transport of eelgrass seeds was reduced in the presence of “medium” and “high” densities of the 

worms (192 and 288 worms m-2, respectively). Worm bioturbation caused the sediment surface 

to have an enhanced topographic relief, and seeds became passively trapped within small 

biogenic depressions or pits. Subsequently, many trapped seeds were buried by sediments 
10.13 

reworked (defecated) by C. torquata. Trapping reduced the dispersal of eelgrass seeds, which 

potentially slows the expansion of seagrass populations. However, seeds buried by C. torquata

bioturbation were not buried too deeply for germination and they were protected from herbivory 

(Luckenbach and Orth 1999). 

 Hughes et al. (2000) proposed that the polychaete, Nereis diversicolor, may have 

contributed to the loss of coastal seagrasses (Zostera spp.) and to difficulties in restoring seagrass 

beds in south-east England (United Kingdom). In field and laboratory experiments, shoots of Z. 

noltii that were protected from N. diversicolor had higher biomass, higher survival, and less 

damage to roots than unprotected transplants (Figure 10.3). Worms apparently damaged Z. noltii

roots by burrowing and leaves by herbivory. However, the authors noted that bioturbationinduced damage to roots may have been an experimental design artifact, and thus the role of 

bioturbation in this worm-seagrass interaction is uncertain. 

Figure 10.3. Results of two laboratory experiments showing decreased survivorship of seagrass 

(Zostera noltii) planted into sediments containing polychaetes (Nereis diversicolor; open 

circles and triangles, and dashed or dotted line) or devoid of worms (controls; filled 

circles and triangles, and solid lines) (modified from Figure 6 of Hughes et al. 2000). 

Davis and Short (1997) noted that bioturbation by an ecologically similar polychaete, Neanthes 

virens, may have been responsible for 99% loss of subtidal transplanted Z. marina shoots. The 

worms appeared to pull the distal ends of seagrass leaves into their burrows, forcing the rest of 

the leaves flat against the sediment surface, after which the leaves were buried by bioturbated 

sediments leading to shoot death. No experiments were conducted to test the hypothesis that 

eelgrass shoot death was clearly caused by activities of N. virens, although their natural history 

observations presented a plausible mechanism. 

1.5

Survivorship (log)

0

0.5

1

2

012 5 34 6789

Days

Polychaetes present (Nov. 1993 expt.)

Polychaetes present (Feb. 1994 expt.)

Control (Nov. 1993 expt.)

Control (Feb. 1994 expt.)

1.5

Survivorship (log)

0

0.5

1

2

012 5 34 6789

Days

Polychaetes present (Nov. 1993 expt.)

Polychaetes present (Feb. 1994 expt.)

Control (Nov. 1993 expt.)

Control (Feb. 1994 expt.)
10.14 

10.2.5 Echinoderm Bioturbation 

 Sea cucumbers and sand dollars (Echinodermata: Holothuroidea and Echinoidea, 

respectively) have been observed to disturb seagrasses during their burrowing and feeding 

activities. However, only two studies have experimentally examined whether bioturbation by 

echinoderms can affect seagrass growth, survival, or populations. Burrowing sea cucumbers 

(Holothuria arenicola) in the Bahamas were observed to smother turtlegrass (T. testudinum) with 

defecated sediments (Mosher 1980, in Valentine et al. 1994); however, whether this adversely 

affected the seagrass was not discussed. In field experiments, Backman (1984) demonstrated 

that dense populations of sand dollars (Dendraster excentricus) uprooted eelgrass (Z. marina) 

and inhibited colonization of unvegetated habitats in Puget Sound, Washington. 

 Sand dollars (Mellita quinquiesperforata) were frequently observed in unvegetated 

patches within subtropical turtlegrass meadows of St. Joseph Bay (Florida, USA) and elsewhere 

(Valentine et al. 1994, and references therein). To test the hypothesis that bioturbation from 

burrowing activities of M. quinquiesperforata were responsible for the long-term persistence of 

these open spaces by preventing recolonization by T. testudinum, Valentine et al. (1994) 

measured the rate of change in the size of the turtlegrass perimeter surrounding open patches 

from which all sand dollars were periodically removed or patches containing undisturbed 

populations of sand dollars (ca. 5-15 m-2). Over a two year period, perimeters of replicate open 

patches expanded or contracted by up to 20%, but independently of sand dollar abundance, and 

at the conclusion of the study, no significant difference in mean patch perimeter was detected 

between these treatments (accounting for initial patch size). Valentine et al. (1994) concluded 

that sand dollars were ineffective at controlling open-space recolonization by T. testudinum; they 

ultimately determined that bioturbation by burrowing stone crabs was responsible for erosion of 

the edges of turtlegrass beds and therefore the biogenic creation of open space in those seagrass 

meadows (see above). 

 Although there are few studies of echinoderm bioturbation impacts on seagrasses, 

Valentine et al. (1994) point out that sediment reworking by other burrowing echinoids (i.e., 

spatangoid urchins) can exceed that of sand dollars (Thayer 1983), that they can occur in high 

densities in the vicinity of seagrass meadows (Chester 1969, in Valentine et al. 1994), and may 

therefore be a source of disturbance for tropical seagrasses. Most echinoderms are found only in 

marine waters. Thus, bioturbation by echinoderms is more likely to occur along open coasts 

rather than in estuaries. 

10.2.6 Mechanisms of Bioturbation-Induced Disturbance to Seagrasses 

 Relatively little is known about the actual mechanism by which the seagrasses are 

harmed, compared to observations and experiments that demonstrate the net effects of 

bioturbators on seagrass shoot growth and survival. The most commonly reported mechanisms 

in the studies reviewed above were burial of seagrass shoots and seeds, uprooting of shoots and 

patches, undermining edges of seagrass patches, damaging roots or rhizomes, and shading by 

deposition of resuspended sediments onto leaves. Indirect mechanisms by which bioturbators 

were reported to damage seagrasses include reducing water column light availability because of 

increased turbidity from resuspended fine sediments (Suchanek 1983), burial by storm-
10.15 

transported sediments previously excavated by bioturbators (Wanless et al. 1988), and increased 

susceptibility to hydrodynamic erosion (Valentine et al. 1994; Woods and Schiel 1997). 

 Rarely were the magnitudes or effects of those disturbance processes confirmed 

experimentally. In very few studies were the magnitudes of the disturbing process measured 

(e.g., burial rate, burrowing rate, sediment deposition rate, light reduction, etc.). In all of the 

studies reviewed above, the mechanism of disturbance was inferred from short-term observations 

of bioturbator activity (i.e., burrowing, digging, sediment expulsion), the condition of seagrasses 

(i.e., partial burial, uprooted shoots, exposed roots and rhizomes, sediment on leaves), biogenic 

structures in the environment (i.e., burrows, mounds, and pits), or feeding habits of the 

bioturbators (i.e., absence of herbivory). Transplant experiments have provided the strongest 

evidence that organisms that are bioturbators can decrease seagrass growth or survival, but few 

studies independently tested that the proposed disturbance mechanism, presented at the 

magnitude affected by the bioturbators, could cause the growth and survival responses seen in 

the seagrasses (but see Duarte et al. 1997). Furthermore, few of the studies quantified the spatial 

scale over which the disturbance could be expected to occur. Ideally, this would require a 

comparison of the landscape and regional distributions of the bioturbator species, of the seagrass 

species, and of the habitat suitable for sustainable seagrass growth, and experimental 

demonstration of the responses of seagrasses to bioturbation at random locations throughout that 

landscape. No study has put it all together, measuring the natural rates and magnitudes of the 

causal process of disturbance (i.e., burial, etc), measuring the responses of appropriate seagrass 

life history stages to those levels of disturbance, and determining the spatial scales over which 

bioturbation could affect seagrass populations. That is perhaps a tall order, but not inappropriate 

given the inferred ecological importance of bioturbator impacts to seagrass populations and the 

functions of seagrass habitats, as stated in most of the papers (for example, “bioturbation by 

Callianassa has a direct effect on seagrass beds but may indirectly influence a multitude of 

faunal relationships both within grass beds and in nearby communities such as coral reefs, 

mangroves, and the deep sea”, Suchanek 1983, p. 296). 

10.3 Role of Bioturbation and Bioirrigation in Enhancing Seagrass Populations 

 Bioturbation and bioirrigation can benefit seagrass populations and communities in two 

ways: 1) modification of sediment geochemistry to create conditions favorable for seagrass 

growth, and 2) facilitation of seagrass recruitment and colonization. Relatively few studies have 

been conducted on processes that may facilitate seagrass population growth. However, as 

interest increases in expanding or restoring seagrass populations, study of the role of these 

natural facilitating processes in improving the condition of seagrass meadows can be expected to 

also increase. 

10.3.1 Modification of Sediment Geochemistry 

 Bioirrigation and bioturbation increase the exchange of solutes and organic matter 

between the water column and the sediment column, strongly affecting sedimentary microbial 

processes such as organic matter decomposition, ammonification, nitrification-denitrification, 

and sulfide oxidation-reduction (Aller 1988; Kristensen 1988; Hopkinson et al. 1999; Fukukawa 
10.16 

2001). To the extent that these sediment geochemical processes limit seagrass growth and 

survival (Chapter 6), bioirrigation and bioturbation can stimulate processes favorable for 

seagrass growth (e.g., fertilization), reduce conditions harmful to seagrass survival (i.e., 

eutrophication, hypoxia), and reduce toxic substances within the rhizosphere (i.e., hydrogen 

sulfide, ammonia). 

 The role of bioturbating and bioirrigating benthic fauna in elevating sediment nutrient 

concentrations is well documented (Aller 1988; Kristensen 1988; Schaffner et al.1992; 

Hopkinson et al. 1999; Felder 2001; Eyre and Ferguson 2002; Webb and Eyre 2004). 

Bioturbation and bioirrigation transfer organic matter from the sediment surface and water 

column into the sediment column, enhancing aerobic remineralization of sedimentary organic 

matter, thereby increasing sediment porewater nutrient concentrations, and enhance dissolved 

nutrient efflux back to the water column. Rates of bioturbation, deposit feeding, and suspension 

feeding typically increase in response to increasing organic matter concentration, at least up to 

some threshold where ingestion rate is maximized or organic matter loading causes toxicity (i.e., 

hypoxia or toxic contaminants associated with organic matter) (Jumars and Wheatcroft 1988; 

Wheatcroft and Martin 1996; Prins et al. 1997; DeWitt et al. 2004). However, little research has 

been conducted to investigate the occurrence of these processes in seagrass beds (particularly 

increasing nutrient availability) and their role in elevating seagrass productivity or growth. 

Thayer and Fonseca (1984) suggested that under conditions of nutrient limitation, “nitrogen 

availability [for eelgrass, Z. marina]...may be limited by the requirements of the heterotrophic 

community responsible for the decomposition of organic matter in sediments” (p. 31). Phillipart 

(1994) suggested that low densities of juvenile lugworms (Arenicola marina) might be beneficial 

to Z. noltii as a result of the worms’ bioturbating activities increasing nutrient fluxes, as reported 

by Huettel (1990). Reusch et al. (1994) and Reusch and Williams (1998) demonstrated that the 

mussels, Mytilus edulis and Muscalista senhousia, respectively, caused porewater concentrations 

of ammonium and phosphate to increase dramatically, and both papers suggested that these 

suspension feeders fertilized beds of Z. marina. Peterson and Heck (1999, 2001a, 2001b) 

demonstrated that the suspension feeding mussel (Modiolus americanus) enhanced productivity 

of turtlegrass (T. testudinum) through fertilization of sediments under nutrient-limited conditions 

in St. Joseph Bay (Florida, USA). The mechanism by which mussels fertilized seagrass beds 

was basically the same in all five papers: mussels ingested nitrogen-rich seston and transferred 

this material to the sediment surface via deposition of feces and pseudofeces; particulate organic 

nitrogen in the fecal matter was microbially remineralized into dissolved inorganic nitrogen 

through ammonification and nitrification; and the dissolved inorganic nitrogen was absorbed by 

seagrass roots and transformed into new tissue. Suspension feeding and deposit feeding benthic 

invertebrates (i.e., bioirrigators and small-scale bioturbators) are common in seagrass beds, 

sometimes in great abundance (Orth et al. 1984; Beal 1994; Bachelet et al. 2000; Bowden et al. 

2001), and it is highly likely that they also fertilize sediments and seagrasses (Peterson and Heck 

1999, 2001a, 2001b; Peterson et al. 2003). 

Bioirrigation also acts to enhance oxygenation of sediments and to advect toxic solutes 

back to the water column (Aller 1988, 1994; Fukukawa 2001; Meile et al. 2001). Feeding and 

respiratory currents generated by burrowing infauna exchanges burrow water with overlying 
10.17 

water. Typically, overlying water has higher concentrations of dissolved oxygen and lower 

concentrations of nutrients, hydrogen sulfide, CO2 and other porewater constituents than burrow 

water. Sediment porewater adjacent to the burrows diffusively exchanges with burrow waters 

(Aller 1988; Kristensen 1988), and thus ventilation of burrows results in increased porewater 

oxygen concentrations, and decreased concentrations of nutrients, hydrogen sulfide, etc. 

(Kristensen 1988; Aller 1994; Furukawa 2001). Elevated levels of ammonium and hydrogen 

sulfide have adverse effects on seagrasses as discussed in Chapters 3 and 6. Thus, the presence 

of bioirrigators has the potential to increase the survival and growth of seagrasses by reducing 

toxic concentrations of those compounds in sediment porewater. Little research has been 

conducted on this interaction between bioirrigators and seagrasses. Using a physiologicallybased stress-response model for Thalassia testudinum, Eldridge and Kaldy (2004) suggests that 

bioirrigation has a very large impact on turtlegrass patch growth due to the oxidation of 

sediments in the root zone. This effect remains to be investigated experimentally with this and 

other seagrass species. 

10.3.2 Enhancing Seagrass Recruitment and Colonization 

 While bioturbation can act to inhibit seagrass recruitment by burial of seeds (Luckenbach 

and Orth 1999; Dumbauld and Wylie-Echeverria 2003), burial of seeds by bioturbation can also 

increase seagrass recruitment by protecting those seeds from herbivory (Luckenbach and Orth 

1999), by trapping seeds in shallow-water sediments that might otherwise be transported by 

currents to deep waters (i.e., below the species’ photic zone) (Orth et al.1994), or by transporting 

seeds to geochemical microenvironments that stimulate germination (Moore et al.1993). 

 Bioturbation may also benefit seagrasses by modifying sedimentary habitats such that 

seagrasses could recolonize them. Extending processes described above, bioturbation can 

oxygenate and remediate sediments that are anoxic or hypoxic due to high organic matter loading 

(e.g., eutrophication) and are toxic to seagrasses (van Katwijk et al. 1997; Koch and Erskine 

2001), possibly resulting in geochemical conditions tolerable to seagrass colonization, 

particularly if remineralization and burial exceed organic matter loading (Aller 1988; Fukukawa 

2001). Bioturbation can also facilitate restoration of sediment texture altered by episodic 

sedimentation. Norkko et al. (2002) reported that bioturbation facilitated the recovery of 

macrofaunal communities following the experimental deposition of terrigenous clay onto tide 

flats, simulating massive sediment runoff associated with poor land management. Duarte et al. 

(1997) noted that sediment reworking by burrowing shrimp reduced the thickness of sediment 

layers experimentally deposited on seagrass patches; this may have benefited seagrasses in those 

patches to some extent given that several of the seagrass species were adversely affected by 

sediment deposition. Aside from the comment in Duarte et al. (1997), I know of no studies 

specifically demonstrating the role of bioturbation in modifying sediments and thereby 

facilitating seagrass colonization. Whether this is a reflection of the lack of investigation into 

this process, or because it rarely occurs, is unknown; however, I believe the process is plausible 

and potentially important for seagrass persistence in eutrophic systems. 

 Disturbance by bioturbation can create open space within areas monopolized by 

dominant space competitor(s). Those areas are then available for colonization by weaker 
10.18 

competitors. This is an example of the intermediate disturbance hypothesis, which predicts that 

in communities structured by competition, biodiversity is greatest under conditions of 

intermediate disturbance, and reduced under conditions of low disturbance (where the dominant 

competitor(s) dominate all resources) or high disturbance (where only the most disturbancetolerant species persist) (Sousa 1984; Dial and Roughgarden 1998). As discussed in section 

10.2.1, Duarte et al. (1997) observed that bioturbation by burrowing shrimp could be partially 

responsible for maintaining high seagrass diversity in Philippine seagrass meadows by creating 

open patches within dense Thalassia hemprichii stands that could be colonized by more 

opportunistic species of seagrass. A similar scenario may maintain diversity in mixed-species 

seagrass beds of south Florida (B.J. Peterson, personal communication). 

10.4 Effects of Interactions Between Bioturbation/Bioirrigation and Water Quality on 

Seagrasses

 Previous sections have described how bioturbation can act to limit or facilitate the growth 

of seagrass populations; a third seagrass response could be “no effect”. (To the best of my 

knowledge, bioirrigation only has neutral or facilitating effects on seagrasses). As with other 

environmental factors that can be limiting or facilitating to seagrasses (i.e., nutrients, 

temperature, light), the direction of the seagrass response will to some extent be determined by 

the magnitude and timing of the bioturbation, and on the presence of other limiting factors 

(“stressors”). Low intensity bioturbation (such as that caused by small species, young life stages 

of large species, species with low sediment reworking rates, or by low population densities of 

large bioturbators) would be likely to have neutral direct effects or could facilitate seagrass 

condition by stimulating nutrient cycling and organic matter remineralization. High intensity 

bioturbation (such as that caused by species with high sediment reworking rates, by large-bodied 

bioturbators, or by dense populations of bioturbators) would be likely to directly disturb one or 

more life history stages of seagrasses. Variation in seasonality of sediment reworking rates 

among bioturbators could have important consequences for seagrasses. Seagrasses whose 

growth rates were seasonally synchronous with the reworking rates of bioturbators would have a 

better chance of growing away (vertically or horizontally) from the disturbance. Seagrasses 

encountering bioturbators that had no seasonal variation in sediment reworking rates would be 

less able to grow away from the disturbance during low growth seasons (i.e., winter). 

 Other factors (such as water quality) that limit seagrass condition could affect the 

outcome of interactions between bioturbators, bioirrigators, and seagrasses. Little is known 

about the effects of interactions between bioturbation/bioirrigation and water quality on seagrass 

growth or survival. The issue has been addressed in only three studies. Phillipart (1994) 

suggested that, following the historical decline of Zostera spp. in the Wadden Sea due to disease 

and eutrophication, expansion of populations of bioturbating lugworms may have prevented 

recovery of seagrass populations. Similarly, Hughes et al. (2000) suggested that expansion of 

populations of the bioturbating polychaete Nereis diversicolor may have hindered natural and 

intentional restoration of Zostera spp. populations in south-east England following their decline 

due to wasting disease and eutrophication. In these two cases, the bioturbator-water quality 

interaction is temporally out of phase, with bioturbation impacts expressed after impacts of poor 
10.19 

water quality and wasting disease (which itself may be linked to poor water quality, Hawkins et 

al. 1999). One could argue that it is not valid to consider these cases as examples of the 

interaction; however, the net detrimental effect to seagrass populations (if true) would be likely a 

consequence of both sources of disturbance acting sequentially. 

 Over a 15 year period, Harrison (1987) observed an increase in the size of seagrass beds 

(Zostera marina and Z. japonica.) on a British Columbia (Canada) tide flat, and a decrease in the 

area dominated by burrowing shrimp (Neotrypaea californiensis), following the construction of a 

breakwater. The breakwater greatly reduced currents over the tide flat, and water clarity 

increased substantially. Transplant experiments demonstrated that survival and growth of Z. 

japonica was sharply reduced in presence of high densities of burrowing shrimp; however, the 

shrimp were not successful at burrowing through the root-rhizome mat of extant seagrass beds 

(see section 10.2.1). Burrowing shrimp bioturbation activity was seasonal, being greatest in the 

summer and early fall, and least in the winter and early spring, probably in response to 

temperature (Fritz 2002). Harrison (1987) proposed that the increase in water clarity allowed the 

seagrass patches to start growing laterally earlier in the spring than when the water had been 

turbid. Each year, the seagrass patches expanded gradually over shrimp beds, smothering 

seasonally-quiescent shrimp that could not burrow through the seagrass bed. Lateral expansion 

of the seagrass bed stopped as shrimp bioturbation increased, which reduced survival and growth 

of seagrass shoots at the shrimp-seagrass boundary. However, there was a net gain in area for 

the seagrass beds. The converse of this interaction hypothesis is that when water clarity was 

reduced, burrowing shrimp bioturbation limited growth of Zostera spp. patches. No experiments 

were conducted to test this interaction hypothesis (other than the previously discussed transplant 

experiments), and it is possible that the seagrass bed expansion occurred only because of the 

increase in water clarity. In any case, the hypothesis illustrates one way in which water quality 

and bioturbation may interact to limit seagrass populations. 

 Bioturbation and bioirrigation may help to offset adverse impacts of poor water quality 

(such as eutrophication and sedimentation) that could otherwise be detrimental to seagrasses. 

The capability of bioturbators and bioirrigators to enhance nutrient cycling and organic matter 

remineralization, and to oxygenate sediments, were discussed in Section 10.3.1. Similarly, the 

capability of bioturbators to facilitate restoration of benthic communities following episodic 

sedimentation events was discussed in Section 10.3.2. To the extent that some benthic 

invertebrates are suspension feeders as well as bioturbators or bioirrigators, feeding by 

populations of those organisms could potentially increase water clarity, increase light 

availability, and thus enhance seagrass growth. 

 Several studies have demonstrated the capacity of populations of benthic suspension 

feeders to daily filter large proportions (i.e., >33%) of the water column within embayments and 

substantially reduce the concentration of phytoplankton (for bivalves see Dame and Prins 1998; 

for burrowing shrimp see DeWitt et al. 2004 and Griffen et al. 2004). Other investigators have 

speculated that reduction of populations of suspension feeders could contribute to an increase in 

turbidity and, consequently, reduce populations of submerged aquatic vegetation (SAV). Newell 

(1988) and Ulanowicz and Tuttle (1992) suggested that over harvesting of oyster (Crassostrea 
10.20 

virginica) populations may have contributed to the decline of seagrasses in Chesapeake Bay. 

Phelps (1994) suggested that the rise and fall of SAV in the upper Potomac R. estuary 

(Washington, D.C., USA) during 1980-1991 was caused in part by changes in water clarity due 

to filter feeding by increasing and then declining populations of the invasive Asiatic bivalve, 

Corbicula fluminea. Only one study has explicitly examined whether benthic suspension feeders 

could remove sufficient particulate material to increase the light field enough to affect seagrass 

growth and survival. Newell and Koch (2004) developed a model to test whether the filtering 

capacity of oysters (C. virginica) and hard clams (Mercenaria mercenaria) at various population 

densities could affect Ruppia maritima growth in Chesapeake Bay. They determined that 

filtering by high densities of oysters, as occurred historically, could have reduced turbidity 

sufficiently to enhance seagrass growth. However, hard clams had much lower filtration rates 

than oysters, and in their model, high population densities of the clams were unable to reduce 

turbidity enough to benefit R. maritima. 

 Clearly not all benthic suspension feeders have the capability to substantially reduce 

turbidity, and the best example of an organism that does have this ability (i.e., oysters) is not a 

bioturbator/bioirrigator. Newell and Koch (2004) concluded that the potential for benthic 

suspension feeders to reduce turbidity and enhance seagrass growth is limited to those 

invertebrates that 1) are able to sustain high filtration rates under conditions of high turbidity 

(which can clog the filtering mechanisms of some species) and 2) occur at high population 

density within or adjacent to seagrass beds. One group of bioirrigating infauna that does have 

the potential to significantly reduce turbidity are suspension-feeding upogebiid burrowing 

shrimp. Populations of Upogebia pusilla (Dworschak 1981) and U. pugettensis (DeWitt et al. 

2004; Griffen et al. 2004) are estimated to daily remove 60-100% of phytoplankton from the 

water over their habitats. Whereas U. pugettensis have similar individual and population filter 

capacities to oysters (Crassostrea gigas; Griffen et al. 2004), then populations of Upogebid 

burrowing shrimp may be able to reduce turbidity sufficiently to benefit seagrass growth. 

 Note that the mechanism by which benthic suspension feeders improve water quality for 

seagrass growth is primarily the result of feeding rather than bioturbation or bioirrigation per se. 

But, suspension feeding is not uncommon among bioturbating and bioirrigating crustaceans 

(amphipods, thalassinids) and bivalves. Furthermore, particles entrained in water passing 

through U. pugettensis burrows (e.g., during bioirrigation) can become trapped on the burrow 

walls and thus removed from the water column (Griffen et al. 2004). Those authors estimate that 

~40% of particle removal by the suspension-feeding and bioirrigating burrowing shrimp, U. 

pugettensis, is caused by burrow-wall entrapment. While this process likely occurs for many 

other bioirrigating infauna, it has not been studied for other species. But, to the extent that U. 

pugettensis populations have the potential to significantly reduce turbidity, then burrow-wall 

entrapment of particles is an important mechanisms by which bioirrigating infauna (or 

Upogebiids at least) can reduce turbidity sufficiently to benefit seagrasses. 
10.21 

10.5 Implications for Seagrass Protection and Restoration

 Narrative or numerical criteria for protection of seagrasses may need to incorporate a 

safety margin to account for the adverse effects of bioturbation when large bioturbators co-occur 

with anthropogenic stressors that adversely affect seagrasses. As discussed in Section 10.4, 

bioturbators can potentially increase the adverse effects of nutrient enrichment by decreasing 

water column light levels (i.e., by increasing turbidity as a result of sediment resuspension) or by 

acting as an independent stressor on nutrient-stressed plants. However, these interactions have 

been specifically examined in only a very few studies and consequently are poorly understood. 

An initial step to account for adverse interactions between water quality and bioturbators would 

be development of guidelines to identify the magnitude of risk of seagrass loss caused by the 

presence of bioturbators. Those guidelines could simply be 1) determining whether large 

bioturbating organisms (similar to those listed in Table 10.1) are present within areas habitable 

by seagrasses, and 2) if present, categorizing the population density of bioturbators as lower, 

equal, or higher than the densities reported to damage seagrasses. Short et al. (2002) recommend 

using similar information as part of their Transplant Suitability Index to rank sites for their 

potential for successful seagrass (Z. marina) restoration. Determination of an appropriate safety 

margin to account for bioturbator-stressor interactions will require additional information on how 

those interactions magnify adverse effects of the stressor alone. 

 To complicate matters, bioirrigators and bioturbators can reduce some adverse effects of 

poor water quality, particularly those caused by nutrient enrichment (i.e., reducing turbidity or 

phytoplankton concentration; oxygenating sediments; flushing ammonia, H2S, or other toxins 

from sediments; see Sections 10.3 and 10.4). Thus, the presence of bioirrigators and bioturbators 

in or near seagrass beds may help explain seemingly anomalous situations wherein seagrasses 

are observed growing under conditions of poor water or sediment quality. Knowledge of the 

abundance of bioirrigators and bioturbators may thus be a useful covariate for interpreting 

outliers in field-based stress-response data sets, and thereby improve the scientific basis for 

establishing numerical protective criteria for seagrasses. 

 Bioturbation has been a significant detriment to seagrass bed restoration primarily due to 

uprooting of transplanted shoots (i.e., stingrays: Merkle 1990; Fonseca et al. 1994, 1998; crabs: 

Davis and Short 1997; Davis et al. 1998; polychaetes: Davis and Short 1997; Hughes et al. 2000; 

see discussions above). Additionally, manipulative experiments have demonstrated that other 

bioturbators taxa can kill transplanted shoots or seedlings (i.e., burrowing shrimp: Suchanek 

1983; Molenaar and Meinesz 1995; Dumbauld and Wyllie-Echeverria 2003; Siebert and Branch 

2006; Berkenbusch et al. 2007; echinoderms: Backman 1984) and therefore can potentially affect 

seagrass restoration. Physical barriers, such as cages or fences, have been used to exclude 

bioturbating stingrays and crabs from seagrass restoration sites, at least on a small scale (Merkel 

1990; Fonseca et al. 1998). Physical barriers may also be useful for excluding echinoderms, 

however they are unlikely to be useful for excluding burrowing shrimp or polychaetes because 

these organisms are relatively small bodied and live underground. Pesticide application to kill 

polychaetes or burrowing shrimp can enhance survival of seagrasses (de Deckere et al. 2001; 
10.22 

Dumbauld and Wylie-Echeverria 2003), although use of broad-spectrum pesticides to control 

indigenous infauna is controversial (Feldman et al. 2000). 

 Alternatively, the presence of bioirrigators or low-intensity bioturbators may enhance the 

success of seagrass restoration in sediments with high organic matter concentrations by virtue of 

their oxygenating sediments and increasing the efflux of toxic hydrogen sulfide or ammonia 

from sediment porewater, and thereby increase the probability of seagrass growth or survival. 

Also, suspension-feeding, bioirrigating benthic fauna can reduce turbidity sufficiently to enhance 

seagrass growth. Newell and Koch (2004) suggest that enhancement of population densities of 

suspension feeders with high filtration rate could enhance restoration of seagrass beds. While 

their example was for epibenthic oysters, the principle holds for bioirrigating or bioturbating 

infauna such as Upogebiid burrowing shrimp or other bivalves that also have high filtration rates 

(Gerritsen et al. 1994; Dame and Prins 1998; Griffen et al. 2004). 

10.6 Knowledge Gaps & Research Needs

 Compared to other abiotic processes, little is known about the importance of bioturbation 

and bioirrigation as either limiting factors or facilitators of seagrass population growth. The 

three critical knowledge gaps concerning the negative and positive effects of bioturbation and 

bioirrigation on seagrass populations are: 1) effects of bioirrigation on seagrasses, particularly 

with respect to oxygenating root zone sediments, 2) interactions between 

bioturbation/bioirrigation and abiotic limiting factors (especially anthropogenic stressors), and 3) 

scaling the effects of bioturbator/bioirrigator populations to seagrasses,. 

 Recent models of seagrass-sediment interactions (Eldridge and Morse 2000; Eldridge and 

Kaldy in press) suggest that bioirrigation can profoundly affect seagrass growth and survival 

through oxidation of the root zone. To the extent that seagrass growth or survival is limited by 

the concentration of toxic metabolites (such as ammonium or hydrogen sulfide) in porewater, 

and that oxidation of those metabolites is limited by the exchange of porewater with oxygenated 

overlying water, then bioirrigators have the potential to increase the growth and survival of 

seagrasses by enhancing the oxidation of those toxic metabolites. While these geochemical 

processes have been empirically demonstrated for unvegetated sediments, their impacts on 

seagrasses have not been investigated experimentally in the field or laboratory. 

 As summarized in Section 10.4, various authors (Harrison 1987; Phillipart 1994; Hughes 

et al. 2000) have suggested that co-occurrence of bioturbators and poor water quality might have 

an increased adverse effect on seagrass populations, compared to the effect of either stressor 

alone. Other research suggests that the populations of bioirrigators or bioturbators might reduce 

one or more adverse effects of nutrient enrichment to seagrass populations (see Section 10.3). 

Further investigation of these interactions, particularly for benthic invertebrate bioturbators and 

bioirrigators that are common in habitats that suitable for seagrasses. The potential for 

bioturbation or bioirrigation to reduce adverse effects of certain pollutants may be valuable for 

protection or restoration of seagrass populations. 
10.23 

 Little is known about scaling the effects of bioturbator or bioirrigator populations to 

seagrasses. While several investigators have discussed the potential for dense populations of 

specific bioturbators or bioirrigators to affect seagrasses, the density-dependence of those effects 

have not been studied. Trivially, no effect would be expected in the absence of 

bioturbator/bioirrigators, but at what population density of those animals would effects to 

seagrasses be expected to occur? Secondly, one might expect that the effects of bioturbation and 

bioirrigation might scale with the animal body size, and thus with the size- or age-structure of the 

animal populations. Finally, the spatial and temporal scale over which the effects of 

bioturbation/bioirrigation are exerted upon seagrass populations have not been examined. Direct 

effects of bioturbation might be largely local (i.e., limited to the area disturbed by the animals). 

However, nutrient efflux or suspension feeding by bioirrigators potentially affect water quality 

parameters (and hence, seagrass growth) over a larger area than that which the animals occupy. 

Related to this is the question of the geographic ubiquity of bioturbator-seagrass and bioirrigatorseagrass interactions. Whereas study sites investigating these interactions were specifically 

selected for all of the studies reviewed above, no one has surveyed randomly selected sites 

within seagrass beds to assess the frequency or magnitude of bioturbation and bioirrigation. 

Understanding how the effects of bioturbator or bioirrigator populations scale relative to seagrass 

populations will be critical for estimating the risks or benefits that seagrass populations obtain 

from the presence of those animals. 
10.24 

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11.1 

11.0 Research Gaps in Relation to Setting Protective Criteria

Walter G. Nelson

11.1 Overview 

The U.S. EPA sponsored a workshop in Baltimore, Maryland in 2003 entitled: Towards a 

National Basis for Protective Criteria for Submerged Aquatic Vegetation, in conjunction with the 

National Conference on Coastal and Estuarine Habitat Restoration. A group of more than 60 

scientists and resource managers from 21 coastal states discussed the scientific knowledge base 

and the challenges in developing a national framework for protection of submerged aquatic 

vegetation. Through a series of regional presentations on the status of SAV research, followed 

by break-out discussion groups, several key questions were addressed. The principal questions 

were: Does our current state of scientific knowledge allow drafting of a national guidance for 

setting criteria protective of SAV, or are there still critical research gaps? Does it appear 

possible to set criteria on a regional basis, or is it likely that criteria will have to be set on a 

system or even sub-system basis? I will address these two questions in reverse order. 

The US EPA produced a National Strategy for the Development of Regional Nutrient 

Criteria (US EPA, 1998), which proposed an operational approach for developing nutrient 

criteria guidance for the states at a regional level, rather than at a national level. The Nutrient 

Strategy proposed the use of the Level III Ecoregional classification system as the regional 

framework for development of nutrient criteria. A variety of guidance documents 

recommending ambient water quality criteria for different water body types (e.g. wetlands, US 

EPA 2000; lakes and reservoirs, US EPA, 2001) have been developed for various Level III or 

sublevel III ecoregions. The tradeoffs of developing protective criteria at larger, aggregated 

ecological scales versus individual water bodies is clearly one of accuracy and relevance of a 

criterion versus the data collection effort needed for establishing a criterion at an individual 

aquatic system level (Chapter 1, Figure 1.3). The tremendous backlog at the state level in the 

development of Total Maximum Daily Load (TMDL) values for individual stream segments is a 

graphic expression of the potential disadvantages of trying to develop criteria at the system or 

sub-system scale. On the other hand, if a criterion developed at regional scale is not applicable 

or protective of individual systems or sub-systems, then it is of little practical use. 

The consensus of workshop participants was that establishing adequately protective 

criteria for SAV even on a regional basis would be difficult, given the inherent variability of 

estuarine systems at such scales. Apart from the scientific and technical issues of regional scale 

criteria, participants from the coastal management community noted that most resource 

management of estuaries occurs at the local scale. Thus, criteria established at the local scale 

and with local input will tend to resonate better with the public. Establishment of effective 

criteria at a spatial scale which avoids a “Balkanization” of standards may be possible. The 

guidance document for Chesapeake Bay water quality criteria (US EPA 2003) illustrates one 

possible approach to establishing criteria for seagrasses, where criteria within this large and 
11.2 

complex system are determined based on four separate salinity regimes, but are applicable 

throughout the Bay. 

In contrast to the development of what was at least a general consensus on an appropriate 

scale for setting seagrass criteria (subregional, system, or sub-system), the workshop participants 

identified a wide range of issues relating to the knowledge base needed to establish protective 

criteria. There were questions raised concerning the utility of using classical water quality 

criteria based on ambient concentrations of nutrients, versus attempting to frame criteria in terms 

of loadings to the system. If a nutrient concentration or loading criterion is established, should it 

be set at levels that insure survival of seagrasses or should it be set at more stringent levels to 

insure that seagrasses thrive? These questions of general principle have not yet been definitely 

resolved. 

11.2 Summary of Research Gaps Identified 

11.2.1 Light 

The review of light information proposed that one present gap in current approaches to 

establishing light requirements for seagrass populations is the ability to adequately account for 

the effects of varying periods of diminished irradiance on seagrass survival. An alternative is to 

determine a benchmark level of leaf sucrose representing the ability of the plant to maintain a 

positive carbon balance. Additional studies are needed on the rates of translocation of sucrose 

from rhizomes to shoots in Zostera marina during times of negative carbon balance to assess the 

importance of this mechanism of mitigating low light conditions. Physiological studies of P vs. I, 

sucrose content et cetera, should always be accompanied with a range of normalizing measures. 

11.2.2 Nutrients 

 The question of whether either nitrate or ammonium is directly toxic to seagrasses needs 

resolution. Apparently contradictory results have been reported, but a potential interactive effect 

of temperature has been suggested, and should be examined for all seagrasses. There is little 

apparent relationship between nutrient inputs and the rate and type of dominant primary 

producers (Nixon et al. 2001), making it difficult to predict whether a criterion will be protective. 

(See also 11.2.7). 

11.2.3 Salinity 

 While nutrient additions may have greater impact on eelgrass populations in high salinity 

areas, (van Katwijk et al. 1999), the interactions of nutrient effects with salinity have not been 

widely studied. Additional information on the effect of increasing the variance in salinity on Z. 

marina would be helpful, especially in terms of variation of population response. 

11.2.4 Hydrodynamic Factors

 Hydrodynamic factors clearly influence the survival and distribution of seagrasses, and 

more data may be needed to define current velocity and wave tolerance ranges for various 

species and systems. Relative differences in sensitivity of various life stages (e.g. seedlings) to 

hydrodynamic factors may be needed. Because many aspects of hydrodynamic influences on 
11.3 

seagrasses are determined by natural factors (wind speed, depth), it is most likely that 

hydrodynamic influences would be considered in terms of whether sufficient safety margins are 

present in setting water quality criteria. Hydrodynamics considerations are clearly critical in 

planning seagrass restoration projects. 

11.2.5 Sediment Characteristics 

Grain size of the substratum influences the distribution and health of Zostera marina, 

although there remain uncertainties concerning the mechanisms. In general, we need a better 

understanding of how sediment pore water concentration of potentially toxic substances, such as 

ammonium and dissolved sulfides, interact with the effects of varying light levels in determining 

plant survival. Such information would be helpful in insuring that protective criteria based on 

water quality variables alone would be adequately protective. 

11.2.6 Epiphytes

 Epiphytes influence available light quality and quantity and may partly determine the 

ability of seagrasses to grow and survive. The role of epiphytic cover in affecting light 

availability may be an essential element to include in development of management criteria for 

protection of coastal seagrass beds. The technical guidance for ambient water quality criteria for 

Chesapeake Bay (U.S. EPA 2003) provides an explicit formulation for including epiphyte loads 

in estimating light available at the seagrass leaf surface. Model formulations need to be 

validated for other systems. Mesocosm research has shown that different algal components may 

dominate in the biomass response to nutrient enrichment in spite of similar initial conditions, and 

that temperature, nutrient exposure regime and other factors such as grazing levels may all 

influence the outcome of nutrient enrichment. Thus, models used for development of seagrass 

criteria must be able to account for impacts from multiple pathways, and must account for effects 

of trophic cascades. 

11.2.7 Macroalgal interactions

 Quantitative relationships between macroalgal canopy height (or corresponding measures 

of abundance) and specific impacts on eelgrass plants (such as those provided by Hauxwell et al. 

2001), for different water body characteristics (temperature, current velocity, turbidity, grazing 

pressure, etc.) are needed. Similarly, relationships between macroalgal abundance and the 

causative anthropogenic activity (e.g., normalized nitrogen load rate) are needed to recommend 

corrective actions. It seems likely that such relationships between nutrient loading, macroalgal 

abundance, and seagrass distribution are site-specific, but this supposition needs to be confirmed. 

11.2.8 Desiccation and Temperature Impacts

 For a region such as the west coast, where intertidal seagrass can be a large percentage of 

the total population, better knowledge of desiccation effects on individual shoots and populations 

is needed to support multi-stressor, seagrass stress-response models. Better definition of the 

interactive effects of temperature and nutrient effects is a critical need, both for site specific 

permitting of heated discharge waters, and to be able to set protective criteria under conditions of 

global warming. 
11.4 

11.2.9 Bioturbation 

 The effects of bioturbation and bioirrigation on seagrass populations and how they 

interact with anthropogenic stressors remain poorly known. Failure to account for both the 

positive and negative effects of organisms in the sediment, e.g. alteration of sediment oxygen 

levels and sediment resuspension rates, could result in criteria either overly conservative or 

insufficiently protective of seagrasses. 

11.3 Conclusions 

 The current review confirms that there is a great deal of scientific information currently 

available concerning the responses of Zostera marina and Thalassia testudinum to a wide range 

of environmental factors. A main theme that has emerged from the review is that the interactive 

effects among factors influencing seagrass survival remain relatively poorly known, especially 

across broader regional scales. This appears true even for such fundamental environmental 

characteristics as salinity and temperature and their interactions in the expression of nutrient or 

sediment impacts on SAV, although research is beginning to fill this gap. 

Clearly of concern is whether current modeling approaches, whether empirical or 

mechanistic, are adequate to predict the response of seagrasses to even single stressors. The 

independent reviews of nutrients (Chapter 3), epiphytes (Chapter 7) and macroalgal interactions 

(Chapter 8) all identified the concern that there is currently a high level of uncertainty in being 

able to predict the trophic pathway for expression of nutrient impacts on seagrasses. Thus water 

quality criteria based on nutrient concentrations may not be adequately protective of seagrass 

resources. Alternate standards based on water clarity or water column chlorophyll a criteria may 

not be adequately protective if the principle expression of nutrient impacts occur through the 

epiphyte or macroalgal pathways. These same concerns were also identified by scientists and 

managers at the Baltimore workshop. 

Current protective criteria for seagrasses (e.g. US EPA 2003) are water column based. 

The review of sediment influences (Chapter 6) suggests that there are important influences on 

plant survival through sediment associated mechanisms that may not be adequately captured by 

water quality criteria alone. The review of effects of light limitation (Chapter 2) suggested there 

may be advantages to looking for integrative, plant based metrics such as sucrose content that 

relate to the ability of seagrasses to survive within a temporally varying environment. The 

Baltimore workshop similarly proposed the possibility of using such integrative measures, 

specifically the Nutrient Pollution Indicator (NPI) (Lee et al. 2004). Such measures may be an 

appropriate method to integrate water column and sediment impacts into single protective 

criteria. 

In conclusion, there remain some fundamental research needs which would greatly 

improve the ability to set protective criteria for seagrasses. However, there is also a considerable 

scientific knowledge base that can be used in establishing criteria for protection of seagrasses. 

The concept of protective criteria is actually one that embodies principles of adaptive 

management. Initial criteria may be based on incomplete data or modeling approaches, but it is 
11.5 

expected that as better data and approaches become available, they will be applied to improve the 

protective criteria. Various localized efforts have made significant progress in establishing 

seagrass protective criteria, e.g. Tampa Bay, Chesapeake Bay. While there are still many issues 

in moving from local criteria to guidance for criteria across multiple systems, it would appear 

that creation of a first level guidance is certainly feasible with the knowledge currently available. 

11. 4 Literature Cited 

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