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United States
Environmental Protection
Agency
EPA/600/R-09/050 June 2009 wvw.epa.gov/ord
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Seagrasses and Protective Criteria:
A Review and Assessment of
Research Status
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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
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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.
11
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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, eel grass Zoster a 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 Zoster a 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.
in
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IV
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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
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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
VI
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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(•), Rvalues (A), 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
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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
Vlll
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List of Tables
Table 1.1. Sections of the Clean Water Act which pertain to the protect!on 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 uM gdw If1 h"1 is
the maximum uptake rate, Ks is the half saturation concentration (uM) 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 eutrophi cation on Z. marina
communities 3.7
Table 3.4. Water column and sediment nutrient concentrations (//M) 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
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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
x
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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,
1.1
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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.
1.2
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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.
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Table 1.1. Sections of the Clean Water Act which pertain to the protection of seagrasses.
SEC. 304. (a)(l)
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
1.4
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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
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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.
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
Measure progress
toward targets and
goals. Reassess as
needed.
Define and implement
nutrient management
strategies to achieve
load reduction goals
4
Determine nutrient
loads necessary to
maintain chl a targets
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).
1.6
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Numeric Nutrients Criteria Critical Development Path
Define Water Quality
Protection Goals for
Specific Spatial Scale
Regional——~
11
Watershed ,„
Local _
Classify Systems Based on
Set of Similarities
e.g. A) Assessment end points (3AV)
EL) Hydrology
C) Ecoregion
D) Great Lake
etc.
Conceptual mod els
Select Variables Needed
for Assessing. Nutrient Status
A) Causal variables - N, P, andfor other
B) Response variables - light, Chi, OO
SAV, macroalgae, benthos andtor other
Data Collection & Analysis
Biological Effects-Based
Approaches (or cause/effect
relationships}
A) Empirical Predictive' Models
B)Mech anisic Predictive Models
Biological Condition
Gradients
Data Distribution Approaches
A) Reference Condition
B)Percentiles
Propose recommended Criteria
Weight of Evidence -
Based on Original
Water Quality Goals
RTAGs, Slates, EPA
may develop recommendations
States Adopt Numeric
Criteria and Other
Components of WQ
Standard through legislative
process
Numeric standards for
nutrients
Establish compliance standards
Technical Guidance for
Compliance Criteria
e.g. establishing exceerfances,
frequency, duration,
monitoring design
States Implement
Nutrient Standards
Establish Control
Strategies, as
Required
Figure 1.2. Diagram of the critical path from Goal Statement to Implementation by the States/Tribes of numeric standards for
nutrients.
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Nutrient Criteria Development
Scaling Issues
Criterion Effectiveness
(at a given site)
High
Intermediate
Low
Spatial Scale
Local (stream segment,
sub-estuary, salinity zone)
Watershed (complete
hyd ro log i c c a tcli m e nt)
Regional (State,, Ecoregion
Level 111 or IV)
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.
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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//, 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
1.9
-------
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
Light
Nutrients
Salinity
Current and Wave Exposure
Sediment Characteristics (Grain size,
dissolved oxygen, sulfide)
Epiphytes
Macroalgae
Temperature (Freezing, Heat Stress,
Desiccation)
Bioturbation
Diseases
Herbivory
Subject to Anthropogenic
Modification
Yes
Yes
Yes
Yes
Yes
Yes, through indirect
mechanisms
Yes, through indirect
mechanisms
Yes, through both direct and
indirect mechanisms
Yes, through indirect
mechanisms
Yes, through indirect
mechanisms
Yes, through indirect
mechanisms
Chapter
2
O
4
5
6
7
8
9
10
10
7 for micrograzers
1.10
-------
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, I.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 resource-
based 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, I.E. and P.M. Eldridge (eds.). 2006. Development of mechanistic models to guide
establishment of protective criteria for seagrasses. Report on the Zoster a 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-l 1 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.
1.11
-------
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\OBS-
84\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.
1.12
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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 meteorologically-
induced 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
2.1
-------
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 QC):
a QC) = aw (A,) + aDoM (A,) + aphto (A,) + aTss (A,) Equation 2.1
where aw, aooM, aphto, and axss 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 (lOPs) of water and
in combination contribute to the spectral, diffusive attenuation coefficient, K^ 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":
/* = /0V-^*Z) Equation 2.2
2.2
-------
where 7^" is the irradiance of wavelength, A,, Z meters below the surface, and
irradiance.
is the surface
0.2
0.15
* 0.1
SJj
o
0.05
-0.05
alp
/'
******
-*—»—»-
50
100
150
200
250
300
350
\imol photons m" sec"
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 post-
illumination respiration phenomenon (Heichel, 1979; Falkowski and Raven, 1997).
2.:
-------
1800
Winter
Hsat=8h
Hcomp=9h
3:00
6:00
9:00
12:00
15:00
18:00
21:00
0:00
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;
2.4
-------
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 marind), 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 (ug I"1) + 0.094 * TSS
(mg I"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. Long-
term 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
2.5
-------
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 lOPs 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 Zoster a marina resulting from light limitation has been observed both during field
observations (Dennison and Alberte 1985; Moore et a/. 1996; 1997; Cabello-Pasini etal. 2002)
and experimentation (Alcoverro et al. 1999; Cabello-Pasini et al. 2002; Thorn 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 a/.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 prescriptive.
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
2.6
-------
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 (Ih) (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 umol 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 umol 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).
2.7
-------
Table 2.1. Observations of the death of Zostera marina plants in the field and laboratory. NR = not reported.
Location
MA, Great Haibor
MA, Great Haibor
VA, York estuary
Mexico, Baja
California, Pacific
coast
CA, Monterey Bay
WA, Sequim Bay
Mexico, Baja
California, Pacific
coast
NC, North River
Month/year
6/82
8/82
4-5/87
2-3/98
11-12/93
11/95
11/00
2-3/03
Study type
field experiment
field experiment
field monitoring
field monitoring
lab experiment
lab experiment
lab experiment
lab experiment
Temperature
°C
15
21
21
17
12
9
16
10 to 16
Irradiance
mol m"2 day"1
~2
~2
1.5-2.5
0-4
-0.3
1-2
0
0
Hsat
hour
day1
~6
~6
~0 (Hcomp)
NR
2
NR
0
0
Time to
death
weeks
~4
~4
3-4
~4
~4
~4
3-4
5-7
Reference
Dennison and
Alberte 1985
Dennison and
Alberte 1985
Moore etal. 1997
Cabello-Pasini et al.
2002
Alcoverro et al.
1999
Thorn et al. 2002
Cabello-Pasini et al.
2002
Biber et al. 2005
2.8
-------
Table 2.2. P vs. I experimental results for Zostera marina L. from various locations and measurement conditions
Location
AK
NC
CA
MA shall/deep
VA
VA
MA
MA shall/deep
MA
MA shall/deep
CA-Monterey
CA-SF
VA
CA-Monterey
CA-Monterey
Mexico-Baja
Mexico-Baja
Temp
°C
5-20
22
1-28
15/25
20
20
0-35
20
10/20
15
25
15
15
16-22
12-27
umol photons
m 2 sec"1
Ik
5
(15 °C)
80-385
67
7-90
78/71
35
210
1-350
50-140
Ic
1
(15 °C)
37
15-25
18
1-17
13/14
12
Parameters in published units
alpha
xlOOO
4.4-8.9
11.4
4.4
2.7-3.7
4-9
4.2/2.4
3.8
10-60
5-10
^max
931
0.27-1.8
14.2
(15 °C)
0.4
1.13-2.8
0.42/0.62
0.22
0.09-0.21
0.055-0.45
0.33/0.17
0.74
0.84
0.25
0.2
0.6-2.45
0.3-0.6
Respiration
1.4
(15 °C)
0.08-0.13
0.080
0.06-0.09
0.0083-0.13
0.053/0.034
5/8
0.06
0.06
0.02
0.085
0.03-0.06
Published
units
mg C h" gdw"
mg C h" gdw"
Hg C h" cm"2
|amol O2 min mg chl a
mg C h" gdw"
|amol O2 min mg chl a
|amol O2 min mg chl a
|amol O2 min" mg chl a"
|amol O2 min mg chl a
and
|amol O2 min"1 cm"2
|amol O2 min mg chl a
|amol O2 h"1 gfw"1
|amol O2 min" mg chl a"
lamol O2 min"1 mg chl a"1
|amol O2 min" mg chl a"
lamol O2 min"1 gfw"1
|j.mol O2 min"1 gfw"1
l^mol O2 min"1 gfw"1
mg chl a
[dm2] or
gfw1
[1.8]/[2.2]
1.5
[2.5]
[2.7]/[4]
[2.2]
[2.9]/[4]
1.1/1.5
2.2
1.5
2
Reference
McRoy 1974
Penhale 1977
Drew 1979
Dennison and Alberte
1982
Wetzel and Penhale 1983
Evans etal. 1986
Dennison and Alberte
1985
Dennison and Alberte
1986
Marsh etal. 1986
Dennison et al. 1987
Zimmerman et al. 1989
Zimmerman et al. 1991
Goodman etal. 1995
Zimmerman et al. 1996
Alcoverro et al. 1999
Cabello-Pasini et al.
2002
Cabello-Pasini, et al.
2003
2.9
-------
Table 2.3. Sucrose content of Zostera marina L. tissues. NR = not reported
Sucrose content
juniol sucrose gfw"1
(leaves / roots /
rhizomes)
220/30/300
NR/ 10/250
1307 10 /NR
40/15/250
100/30/200
90 / 20 / 200
233/NR/NR
350/NR/NR
Loss rate
juniol sucrose gfw"1 day *
(leaves / roots / rhizomes)
no significant loss
(whole plants in dark over 12 hr at unknown
temperature)
2.2 /NR/ 3.4
(whole plants in -0.3 mol photons m"2 day"1
for 30 days at 12 °C)
15±1/NR/NR
(whole plants in dark at 16 °C over 21 days)
Reference
Zimmerman et al. 1989
Kraemer and Alberte
1993
Zimmerman et al. 1995
Kraemer and Alberte
1995
Zimmerman et al. 1996
Alcoverro et al. 1999
Touchette and
Burkholder2001
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 umol sucrose gfw"1
and an average, healthy plant, concentration of 166 + 40 umol sucrose gfw"1 (+ 1SE). The
carbon equivalents of this range are -180 and 1992 ± 480 umol 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:
2.10
-------
gross Pmax = (0.036 ± 0.0053 * deg C) + 0.11 ± 0.096 (to 30 °C, n = 7) Equation 2.3
respiration = (0.0085 ± 0.00147 * deg C) - 0.02 ± 0.031 (to 35 °C, n = 8 Equation 2.4
where the units are umol QI 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, umol C>2 umol 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 umol C>2 umol 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 ofZostera 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
2.11
-------
Table 2.4. Conversion factors from various sources.
Units
Mean
SD
n
Andersen and Johnson1
gfw dm"
2.168
0.538
gdw dm"
0.350
0.081
gfw gdw"
6.26
0.99
72
Evans et
al. (1986)2
gfw gdw"
4.4
0.4
100
Nelson3
gdw dm"
0.25
0.062
215
Chlorophyll a from
Table 2.2
mg dm"
2.79
0.82
8
mg gfw"1
1.6
0.40
6
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.
mean
SD
n
Production
Respiration
jimol O2 dm" min"
5°C
0.290
0.199
7
20 °C
0.828
0.157
7
5°C
-0.027
0.070
8
20 °C
-0.154
0.049
8
alpha
(imol O2
jimol photons"1
0.0285
0.0113
9
Sucrose
Content
jimol gfw"1
166
107
7
Loss rate
jimol gfw"1 day"1
-15.11
1.97
4
-2.22
0.80
4
1 Cabello-Pasini, et al. (2002)
2Alcoverro, et al. (1999)
respiration rate of 0.12 umol 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:
r. • +- -70 j, Q n
Respiration = -79 ± 8.0
liimo1 C -0.12 /wo/ 02 dm2 1440 min 1 /wo/ C
— - = - ^ - - * - * - * — — -
gfw day dm min 2. 17 gfw day 1 jumol O2
,-,.<.
Eg. 2.5
Sucrose Loss = -180 ±12
C -15 fjmol sucrose 12 /wo/ C
gfw day
gfw day
/wo/ sucrose
Equation 2.6
where the rates and conversion factors are from Tables 2.3 to 2.5.
2.12
-------
240
10 20
degrees Celsius
30
Figure 2.3. Gross maximum productivity, gross Pmax(«), Rvalues (A), and dark respiration rates
(•) from Marsh et al. 1986.
While the dark respiration and the sucrose loss rates are significantly different (ts 6.95,
= 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 (r^, below, is the
correlation coefficient between x andj).
2.13
-------
for v = x*y*z, —- = S = J| ^M + ^^- + ^-M Equation 2.1 a
x ) { y ) \. z )
for v = x + y, SDv = ^SD2x + SD2y + 2r^ * SDX * SDy Equation 2.7b
for v = x-y, SDv = SD2x + SD2y - 2^ * SDX * SDy Equation 2.7c
SDv = SEv * n Equation 2.1 d
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 jimole 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:
~^ , A n C -2.2 umol sucrose1" . 12 umol C „
-26 + 4.8 — - = - — - * - — - Equation 2.8
gfw day gfw day jumol sucrose
where A 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:
o/io-uoA /WIO/C f0.54^wo/02V dm2 120 min \fmolC
24.9 ± 2.6 - = - - - - * - * - * - Equation 2.9
gfw day ^ dm min j 2.17 gfw day 1.2 /umol O2
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:
2.14
-------
^^ -0.086//mo/02 dm2 .1440 mm 1 /umol C .
-57 ±6.6- = * * *— Equation 2.10
gfw day [^ dm min j 2.17 g/w Joy 1 fmol O2
The net production rate under this limited light exposure is -32 + 9.3 umol C gfw"May"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 umol C/gfw-day, is
not significantly different than the net production rate (ts 0.88, t.osp] = 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 umol C gfw"
May"1 while the net production with 2 hours of light (Equations 2.9 and 2.10) went up to -64
umol C gfw" May"1 which is significantly larger (ts 3.29, t.osp] =2.26) than the sucrose loss rate of
-26 umol C gfw"May"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.
2.15
-------
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 umol photons m^sec"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:
T
i^net
gross
- R i
Equation 2.11
Where, / is the average irradiance over time interval, /', and t is the in situ temperature.
Ifk is calculated by dividing gross Pmax at temperature ^by the quantum yield, alpha. This yields
I'k with the following units:
/Ltmol photons _ jumol O2 ^ /Ltmol photons .,.100 dm2 ^ min
m2 sec dm2 min 0.029 jumol O2 m2 60 sec
Equation 2.12
The units for ,/L are:
/jmol C
gfwi
1 /Limol O2 fjmol C
1 dm2 min 1.2 /nmol O2
O2 jumol C
dm2 min /urnol O2
dm2
mm
2.11 gfw i
E0.2.13
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 umol sucrose gfw"1 (Table 2.3) or 1992 umol
Cgfw'1:
Sucrosemax > ;Sucrose = (;. i Sucrose +iP«ef/ 12) < 15
Equation 2.14
2.16
-------
where 12 is the number of carbon atoms per sucrose molecule. When -fnet is positive, carbon is
added to the sucrose pool up to Sucrosemax.
For these computations, plant death occurs at ^Sucrose ~ 15 umol 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 ^Sucrose accounts for varying light and temperature conditions across
time and would indicate when light and temperature conditions persisted long enough to sustain
the growth oi Zoster a 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:
mol photons -umol C .1.2 umol O7 . umol photons .100 dm2 „ _ ,,
f = —^ 2-*-^ Ł. * .— Equation 2.15
m day dm day umol C 0.029 umol O2 m
where jimolC dm^day"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.
2.17
-------
_i
— 711
CS 7U "
•a
C Łn
"S =n
3 so -
•o
~ 4U
%
J-* in
•o
9ft
10
•1
:]
•i
5
[•
:
1
:
;
f
\
.
1
ft
.1
1
1
T
it
1
rf,
' '
••
2.5 5.0 7.5 10.0 12.5 15.0 17.5 20.0
degrees Celsius
22.5 25.0 27.5
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 2.5
7.5 10
12.5 15 17.5 20 22.5 25 27.5 30
degrees Celsius
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.
2.18
-------
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 ofZostera marina populations. The ^Sucrose 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 lOPs 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, peat-
derived 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
2.19
-------
area normalization is a means to directly couple in situ light fluxes and laboratory-determined
photosynthetic rates.
2.4 Literature Cited
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Batiuk, R. A., R. J. Orth, K. A. Moore, W. C. Dennison, J. C. Stevenson, L. W. Staver, V. Carter,
N. B. Rybicki, R. E. Hickman, S. Kollar, S. Bieber, and P. Heasly 1992. Chesapeake Bay
Submerged Aquatic Vegetation Habitat Requirements and Restoration Targets: A
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Beers, Y. 1957. Introduction to the Theory of Error, 2nd edition. Addison-Wesley Publishing
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seagrass stress to light limitation, pp. 167-176. Chapter 13. In: S.A. Bortone (ed.)
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marina L.). Marine Biology 90:111-116.
Bulthuis, D. A. 1987. Effects of temperature on photosynthesis and growth of seagrasses.
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photosynthesis in Zostera marina at its southern end of distribution in the North Pacific.
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Cabello-Pasini, A., C. Lara-Turrent, and R. C. Zimmerman. 2002. Effects of storms on
photosynthesis, carbohydrate content and survival of eelgrass populations from a coastal
lagoon and the adjacent open ocean. Aquatic Botany 74:149-164.
Cerco, C. F. and K. A. Moore. 2001. System-wide submerged aquatic vegetation model for
Chesapeake Bay. Estuaries 24:522-534.
Dennison, W. C. and R. S. Alberte. 1982. Photosynthetic responses of Zostera marina L.
(eelgrass) to in situ manipulations of light intensity. Oecologia 55:137-144.
Dennison, W. C. and R. S. Alberte. 1985. Role of daily light period in the depth distribution of
Zostera marina (eelgrass). Marine Ecology Progress Series 25:51-61.
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(eelgrass) transplants along a depth gradient. Journal of Experimental Marine Biology
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Dennison, W. C., R. C. Aller, and R. S. Alberte. 1987. Sediment ammonium availability and
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Bergstrom, and R. A. Batiuk. 1993. Assessing water quality with submerged aquatic
vegetation. BioScience 43:86-94.
2.20
-------
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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7:139-150.
Drew, E. A. 1983. Sugars, cyclitols and seagrass phytogeny. Aquatic Botany 15:387-408.
Duarte, C. M. 1991. Seagrass depth limits. Aquatic Botany 40:363-377.
Evans, A. S., K. L. Webb, and P. A. Penhale. 1986. Photosynthetic temperature acclimation in
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Falkowski, P. G. and J. A. Raven. 1997. Aquatic Photosynthesis. Blackwell Science, Maiden,
MA.
Gallegos, C. L. 1994. Refining habitat requirements of submersed aquatic vegetation: role of
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Gallegos, C. L. 2001. Calculating optical water quality targets to restore and protect submersed
aquatic vegetation: overcoming problems in partitioning the diffuse attenuation
coefficient for photosynthetically active radiation. Estuaries 24:381-397.
Gallegos, C. L. and W. J. Kenworthy. 1996. Seagrass depth limits in the Indian River lagoon
(Florida, U.S.A.): application of an optical quality model. Estuarine, Coastal and Shelf
Science 42:267-288.
Gallegos, C. L. and T. E. Jordan. 2002. Impact of the spring 2000 phytoplankton bloom in
Chesapeake Bay on optical properties and light penetration in the Rhode River,
Maryland. Estuaries 25:508-518.
Giesen, W. B. J. T., M. M. van Katijk, and C. den Hartog. 1990. Eelgrass condition and
turbidity in the Dutch Wadden Sea. Aquatic Botany 37:71-85.
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(Zostera marina L.) to light and sediment sulfide in a shallow barrier island lagoon.
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Heichel, G. H. 1970. Prior illumination and the respiration of maize leaves in the dark. Plant
Physiology 46:359-362.
Jerlov, N. G. 1976. Marine Optics. Elsevier, Amsterdam.
Kemp, W. M., R. Bartleson, and L. Murray. 2000. Epiphyte contributions to light attenuation at
the leaf surface, pp. 28-37. In R.A. Batiuk, et al. (eds.) Chesapeake Bay Submerged
Aquatic Vegetation Water Quality and Habitat-based Requirements and Restoration
Targets: A Second Technical Synthesis. U.S. Environmental Protection Agency,
Chesapeake Bay Program, Annapolis, Maryland.
Kemp, W. M. R. Batiuk, R. Bartleson, P. Bergstrom, V. Carter, C. L. Gallegos, W. Hunley, L.
Karrh, E. W. Koch, J. M. Landwehr, K. A. Moore, L. Murray, M. Naylor, N. B. Rybicki,
J. C. Stevenson, and D. J. Wilcox. 2004. Habitat requirements for submerged aquatic
vegetation in Chesapeake Bay: water quality, light regime, and physical-chemical factors.
Estuaries 27:363-377.
Kirk, J. T. O. 1994. Light and Photosynthesis in Aquatic Ecosystems, 2nd edition. Cambridge
University Press, Cambridge, U.K.
2.21
-------
Kraemer, G. B. and R. S. Alberte. 1993. Age-related patterns of metabolism and biomass in
subterranean tissues of Zoster a marina (eelgrass). Marine Ecology Progress Series
95:193-203.
Kraemer, G. B. and R. S. Alberte. 1995. Impact of daily photosynthetic period on protein
synthesis and carbohydrate stores in Zostera marina L. (eelgrass) roots: implications for
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
activity and abundance in root and leaf tissues of Zostera marina. Aquatic Botany
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.
Losee, R. J. and R. G. Wetzel. 1983. Selective light attenuation by the periphyton complex, pp.
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.
2.22
-------
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.
Thorn, 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 bottom-
up 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 transit!on 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.
2.23
-------
-------
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 seagrass-
nutrient 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).
3.1
-------
Table 3.1. Summary of nitrogen uptake rates for Z. marina. Vmax in units of uM gdw Ifl h"1 is the maximum uptake rate, Ks is the half
saturation concentration (uM). Units for uptake rates are denoted by superscripted letter.
Location
Leaf uptake
'max
Ks
Root uptake
'max
Ks
Reference
Zostera marina
Rhode Island
Netherlands
Netherlands
Netherlands
Japan
Alaska
0.03 -0.19a
0.01 -2.8b
0.35C
0.001d
1.05-3.2b
20.5
9.2
0.01 -3.9b
0.33C
0.011d
0.9 - 2.9b
211
104
Thursby & Harlin 1982
Pedersen & Borum 1993
Hemminga et al. 1994
Pedersen & Borum 1992
lizumi & Hattori 1982
Short & McRoy 1984
Thalassia hemprichii
Indonesia
32-37
21-60
Stapeletal. 1996.
Thalassia testudinum
Texas (NH4+)
Texas (NO3~)
8-16.5
3.4-6.5
5-19
2.2-38
8-73
34-765
Lee&Dunton 1999b
Lee&Dunton 1999b
a g N m'2 d'1
b uM N gdw'1 h'1
c mg N gdw'1 d'1
d g N (g plant N)'1 d'1
i.2
-------
Table 3.2. C:N:P ratios for Z. marina from the literature.
Seagrass
Multiple spp.
Location
global
%C
33.6
%N
1.92
%P
0.23
C:N
20
N:P
24
C:P
474
C:N:P
474:24:1
Reference
Duarte 1990
Zostera marina
various
California
Oregon
Oregon
Oregon
California
Rhode Island
Virginia
Virginia
Virginia
36
38.4
34
35
29
39
34
42
35-39
27-35
2.5
2.37
2.7
1.3
1.4
2-3.8
1-2
0.39
0.34
0.4
0.2
0.1
0.2-
0.3
0.06-
0.15
17
20
15
34
24
18
12
13
11
19
7
18
14
15
16
17
15
24
27
10
6
15
3
38
27
41
246
304
255
510
576
481
120
79
170
57
274
481
584
255:15:1
255:17:1
510:15:1
576:24:1
481:27:1
120:10:1
79:6:1
170:15:1
57:3:1
274:38:1
481:27:1
584:41:1
Duarte 1990
Fourqurean et al. 1997
2Kaldy 2006a
3Kaldy 2006a
4Kaldy 2006a
^erez-Llorens et al. 1991
^erez-Llorens et al. 1991
^erez-Llorens et al. 1991
^erez-Llorens et al. 1991
^erez-Llorens et al. 1991
Atkinson and Smith 1983
Atkinson and Smith 1983
Atkinson and Smith 1983
2Moore and Wetzel 2000
3Moore and Wetzel 2000
Zostera capricorni
Australia
Australia
Australia
20
58
33
17
8
9
349
465
302
349:17:1
465:8:1
302:9:1
2 Atkinson and Smith 1983
3 Atkinson and Smith 1983
4 Atkinson and Smith 1983
i.3
-------
Zostera noltii
Netherlands
12
21
16
9
6
7
104
127
114
104:9:1
127:6:1
114:7:1
2Perez-Llorens et al. 1991
3Perez-Llorens et al. 1991
4Perez-Llorens et al. 1991
Thalassia testudinum
Barbados
Barbados
Texas
Texas
Texas
Texas
Florida
Florida
33-37
35-37
34-36
36-37
34.6
36.9
1.7-
2.7
0.7-
0.9
1.7-2
0.7-
0.9
2.2
1.82
0.095
0.113
15-
25
45-
62
21-
24
46-
60
18.5
24.6
40.2
40.2
1070
937
445:32:1
601:20:1
2 Atkinson and Smith 1983
3 Atkinson and Smith 1983
2Kaldy and Dunton 2000
3 Kaldy and Dunton 2000
2 Lee and Dunton 1999a
3 Lee and Dunton 199a
2 Fourqurean et al. 1992
2 Fourqurean et al. 2002
Thalassia hemprichii
Indonesia
1.9
0.48
0.98
0.14
0.07
0.06
30
15
36
2 Erftemeijer & Middelburg
1993
3 Erftemeijer & Middelburg
1993
4 Erftemeijer & Middelburg
1993
1 Literature values from Table 1 for original citations see Perez-Llorens et al. 1991
2 Ratios for leaf material
3 Ratios for rhizome material
4 Ratios for root material
i.4
-------
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 uM 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 rubrd) 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.
3.5
-------
3.2 Nutrient Enrichment and Eutrophication
"Eutrophication" 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 "an 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
"cultural" 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).
3.6
-------
Table 3.3. Selected list of literature examining the effect of eutrophi cation 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
Community
Experimental
system
aquaria
aquaria
aquaria
aquaria
field
field
field
field
field
field
field
mesocosm
mesocosm
mesocosm
mesocosm
Lit. review
Lit. review
Response
+Epi, -Zm
+Zm, -Znf1
+ Epi
+Epi, -Zm1
+Zm, +Macro
+Zm
—
+Phyto,
+Macro, -Zm
- Epi, -Zm*
+Macro, -Zm4
+Macro, -Zm
+Epi, -SAV
-Zm1
+algae, -Zm
+Phyto, -Zm,
-Macro, -Epi
—
—
Location
Virginia
Netherlands
Washington
Virginia
Rhode Island
Netherlands
Maryland
Mass.
Washington
Finland
Mass.
Maryland
North
Carolina
New
Hampshire
Rhode Island
Reference
Necklesetal. 1993
van Katwijk et al. 1999
Williams & Ruckleshaus
1993
Moore & Wetzel 2000
Harlin & Thorne-Miller
1981
van Lent etal. 1995
Stevenson et al. 1993
Valielaetal. 1992
Williams & Ruckleshaus
1993
Bostrom et al. 2002
Hauxwell et al. 2003
Twilley etal. 1985
Burkholder et al. 1992,
1994
Short etal. 1995
Taylor etal. 1995
Worm et al. 2000
Nixon etal. 2001
* Suggested nutrient limitation of Z. marina
Suggested nitrate toxicity of Z. marina
2Concluded 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.
3.7
-------
J>* Nutrient
Nutrient
Nutrient
Phytoplankton
Epiphytes
Macroalgae
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.
3.8
-------
The responses of lakes to nutrient loading and eutrophication abatement have been
examined in a number of studies. Eutrophi cation 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.
3.9
-------
Table 3.4. Water column and sediment nutrient concentrations (//M) associated with selected studies ofZostera habitat.
Seagrass
Location
Water
NH4+
NO37NO2
P
Sediment
NH4+
N03
/N02=
P
Reference
Zostera marina
Alaska
Netherlands
Rhode Island
Denmark
Alaska
Mass.
Alaska
North Sea
Alaska
3-15
0.1-1.5
1-5.5
1-14
1-18
0.1-0.8
0.5-6
0-9
5-17
0-1
2-7
20-147
250-1100
10-1500
270-1550
50-130
0-2
1-15
4-28
5-75
Short 1983a
Herman etal. 1996
Harlin & Thorne-Miller
1981
Pedersen & Borum 1993
lizumi etal. 1982
Dennison et al. 1987
Me Roy etal. 1972
Hemminga et al. 1994
lizumi etal. 1980
Zoster a japonica
Oregon
Oregon
2.2-2.9
2-5
7-35
2-26
0.6-1.1
0.5-2
60-170
600-2400
0.8-2.0
0-30
3.3-5.6
7-40
Larned 2003
Kaldy 2006b.
3.10
-------
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 "new" or "old" nitrogen. New nitrogen
is typically inorganic (e.g. N2, NCV, 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 NCV (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 Thome-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
3.11
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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 2. 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 °C) 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 °C). 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
eutrophi cation and/or the potential for eutrophi cation 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
3.12
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types of habitat (e.g. marshes), as well as economically and recreationally important fisheries, is
poorly known.
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Havens, K. E., J. Hauxwell, A. C. Tyler, S. Thomas, K. J. McGlathery, J. Cebrian, I. Valiela, A.
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shallow marine and freshwater ecosystems: implications for community responses to
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supply to intertidal eelgrass (Zostera marina). Marine Biology 118:223-227.
Hemminga, M. A., N. Marba, and J. Stapel. 1999. Leaf nutrient resorption, leaf lifespan and the
retention of nutrients in seagrass systems. Aquatic Botany 65:141-158.
Hemminga, M. A. and C. M. Duarte. 2000. Seagrass Ecology. Cambridge University Press,
Cambridge, UK.
Herman, P. M. J., M. A. Hemminga, P. H. Nienhuis, J. M. Verschuure, and E. G. J. Wessel.
1996. Wax and wane of eelgrass Zostera marina and water column silicon levels.
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Howarth, R. W., D. M. Anderson, T. M. Chruch, H. Greening, C. S. Hopkinson, W. C. Huber, N.
Marcus, R. J. Naiman, K. Segerson, A. N. Sharpley, and W. J. Wiseman. 2000. Clean
coastal waters: understanding and reducing the effects of nutrient pollution. Ocean
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Studies Board and Water Science Technology Board, National Academy Press,
Washington, DC.
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eelgrass (Zostera marina L.) beds. Marine Biology 66:59-65.
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in temperate waters of the Pacific coast of Japan III. Kinetics of Nitrogen uptake.
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Jensen, H. S., K. J. McGlathery, R., Marino, and R. W. Howarth. 1998. Forms and availability
of sediment phosphorus in carbonate sand of Bermuda seagrass beds. Limnology and
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Kaldy, J. E. 2006a. Carbon, nitrogen, phosphorus and heavy metal budgets: How large is the
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Kraufvelin P, B. Sinisalo, E. Leppakoski, J. Mattila, and E. Bonsdorff. 2001. Changes in
zoobenthic community structure after pollution abatement from fish farms in the
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nutrient fluxes between the water column and benthos in a NW Pacific estuary. Marine
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testudinum: Development of a whole-plant nitrogen budget. Limnology and
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Florida, USA.
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sources. Limnology and Oceanography 43:577-585.
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enriched coastal waters. Journal of Phycology 37:453-456.
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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 Bjork 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).
4.1
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Table 4.1. Reported salinity tolerance ranges for eelgrass (Zostera marina) populations.
Location
World Wide
World Wide
Northern Hemisphere
Denmark
Denmark
Denmark (estuaries)
Baltic Sea
Netherlands (marine environments)
Netherlands (estuaries)
Netherlands (Lake Grevelingen)
France (Thau)
Italy (Venice)
Chesapeake Bay
Yaquina Bay, OR
Salinity
Range
0-35
0-42
5-35
13-31
12-31
9-23
6-12
-30
15-25
22-32
27-41
25-33
14-22
25-33
Citation
Thayeretal. 1984
Phillips 1984
denHartog 1970
Pinnerup 1980
Sand- Jensen and Borum 1983
Wium- Anderson and Borum 1984
Hellblom and Bjork 1999
van Katwijk et al. 1999
van Katwijk et al. 1999
Kamermans et al. 1999
Rigollet et al. 1998
Rigollet et al. 1998
Wetzel and Penhale 1983
Kentula and DeWitt 2003
Table 4.2. Reported salinity tolerance ranges for turtlegrass (Thalassia testudinum) populations.
Location
Florida
Florida
Florida
Texas
Florida
Dry Tortugas
Everglades National Park
Florida (west coast)
Salinity
Range
28-36
17-36
24-35
30-40
22-36
35-38.5
28-48
25-34
Citation
Phillips 1960
Zimmermann and Livingston 1976*
Zieman and Zieman 1989*
Zieman 1982*
Adairetal. 1994*
Doering and Chamberlain 2000
Phillips 1960
Phillips 1960
Phillips 1960
*As cited in Doering and Chamberlain 2000.
4.2
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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 (Thorn 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
4.3
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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).
4.4
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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.
4.5
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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. Bjork. 1999. Photosynthetic responses in Zostera marina to decreasing
salinity, inorganic carbon content and osmolality. Aquatic Botany 65:97-104.
Herrera-Silveira J. A., J. Ramirez-Ramirez, N.Gomez, 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.
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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, Syringodiumfiliforme, 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 Casablanca, 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/OBS-
84/02. 147 pp
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Thorn 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 eel grass (Zoster a
marina L.) in Pacific Northwest, USA, System (Chapter 5). PINCERS 2001 Annual
Report.
Thorn 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/OBS-
82/25, 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),
155pp.
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.
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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
5.1
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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
5.2
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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
5.3
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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 ofH. 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.,
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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 eel grass 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
Scoffm (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.
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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 uni-
directional 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
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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). Enriquez and Rodrigues-Roman (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 self-
shading increases with increasing current. Zimmerman (2003) proposed that the production-
enhancing 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.
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High
Medium
Low
Cross-Sectional View
Aerial View
At high flow, there is increased
mounding of the substratum.
(Fonseca etal.,1983)
There is a reduction in seagrass cover
under high flow conditions (Fonseca et
al., 1983) and when subjected to uni-
directional flow seagrass grow in rows
(Fonseca et al, 2007). Black and white
areas indicate seagrass and bare
regions, respectively
Individual
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)
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 ofHalodule
wrightii and T. testudimim 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.
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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
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current speed of 10 cm s"1) sites epiphyte coverage was negligible. Results of cross-
transplantation 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^and 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.
O
O
CD
Diffusional
Limitation
< 5 cm s"1(1)
< 5.4 cm s"1(2)
Deposition of
Sediment on
Leaves
< 8 cm s"1 (3)
Optimal
Increasing Productivity &
Growth with Flow
Flow range of observations
10 to 40 cms (4,5)
Oto 16 cms"1 (6)
2 to 34 cm s"1 (7)
1 to 35 cm s"1 (8)
Physical Damage
to Plants
70 cm s"1 (9)
Erosion of sediment
resulting in exposure of
roots & rhizomes:
50-150 cms"1 (9)
Threshold for damage:
33-85 cms"1 (10)
Severe damage:
85-110 cms"1 (10)
Self Shading
(11)
Current Speed, cm s"
(1) Koch 1994; (2) Enrfquez and Rodrfgues-Roman 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) Thorn et al. 1996; (11) Zimmerman 2003
Figure 5.2. Growth conditions of seagrass as a function of current speed.
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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
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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
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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.
Thorn et al. (1996) conducted flume experiments to assess the impact of propeller washes
on Z. marina. Flume experiments were conducted on intact patches of eel grass 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.
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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 2. 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//, 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).
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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,
5.15
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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.
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6.0 Interactions of Zostera marina and Thalassia testudinum with
Sediments
Peter M. Eldridge, MarkG. 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 water-
column 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
Zoster a 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
6.1
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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 2. marina growing in a range of sediment texture classes in Yaquina
Bay, OR. Although a majority of the 2. marina habitat was in the lower estuary associated with
sediments characterized by 75-100% sand (up to 25% silt/clay), 2. 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. Thorn et al. (2001) cultivated 2. 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 2. 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
2. 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 2. 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 2. 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
6.2
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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.
Light
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)
6.3
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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
Zostera marina
Zostera
japonica
Zostera
noltii
Zostera
muellerl
Zostera
novelandica
Thalassia
testudinum
Organic
Matter (%)
1.2
1.2
2.2-2.7
1.7
0.4- 1.0
6.4 - 16
0.4-3.5
0.4-1.4
0.4- 12
1.7-3.8
0.8-7.3
3.5-4.9
Size Class
very fine -
fine sand2
medium -
muddy sand
medium-fine
sand2
fine sand -
silty clay
medium -
fine sand2
fine sand2
medium sand
-fine
sand/silt
medium -
very fine
sand2
fine sand2
fine sand2
very fine
sand2
fine sand
Median Grain
Size (mm)
0.100 -0.1461
0.12-0.50
0.17-0.34
0.14-0.271
0.17-0.201
0.06-0.353
0.13-0.231'3
0.13-0.23
0.07
0.193
Silt+Clay
(%)
14
11-13
6-24
5-35
8-13
4-6
2-56
40-70
1-72
5
Source
Marshall and Lucas (1970)
Orth(1977)
Nienhuis and DeBree (1977)
Kenworthy el al. (1982)
Peterson etal. (1984)
Kenworthy and Fonseca (1992)
Short etal. (1983)
Lalumiere et al. (1994)
Townsend and Fonseca (1998)
Dan etal. (1998)
Webster etal. (1998)
Frost etal. (1999)
Koch (2001)
Thorn etal. (2001)
Lee et al. (2002)
Posey (1988)
van Lent etal. (1991)
Sprung (1994)
Edgar and Shaw (1995)
Heiss et al. (2000)
Wood etal. (1969)
6.4
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Thalassia
hemprichii
0.8
0.5-0.6
0.5-2.3
3.5- 10
medium
sand2
medium
sand2
fine sand2
medium -
coarse sand2
fine - coarse
sand2
medium
sand2
0.35-0.361
0.241'3
0.15-0.20
0.24 - 0.593
0.19-0.501
0.24 - 0.371'3
15
1-34
3
12-34
22
2-9
2- 17
5-12
23-35
Scoffin(1970)
Burrell and Schubel (1977)
Orth(1977)
Grady (1981)
Wanless (1981)
Hoskin(1983)
Lee and Dunton( 1996)
McGlathery et al. (1994)
Kuenen and Debrot (1995)
Kalbfleisch and Jones (1998)
Livingston et al. (1998)
Koch (2001)
Terrados et al. (1998)
Kaldy and Dunton (2000)
Paula etal. (2001)
1. Calculated from phi (e) values: (0) = -Iog2(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.
6.5
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(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. Zoster a 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
PW
PW
PW
W
PW
PW
PW
PW
PW
Marine/Estuarine
Plant
Zostera marina
Zostera marina
Zostera marina
Zostera marina
Thalassia testudinum
Thalassia testudinum
Thalassia testudinum
Thalassia testudinum
Thalassia testudinum
Effect
Cone.
(MM)
2300
400- 800
70
100-1000
~ 1,500
80
2000-6000
6000
5500
Duration
24 hr
21 day
2 mo
1-3 weeks
in water
4-6 weeks
6 mo
48 hr
14 day
38 day
Response Variable
root respiration not reduced
decreased photosynthesis
decreased leaf elongation rate;
shoot density unaffected
photosynthesis stopped
leaf elongation rate stopped
leaves/shoot decreased 37%
root non- structural carbohydrate decreased
81%
above-ground biomass (shoot/root )
decreased 55%
sulfide suggested as synergistic secondary
stressor
seedling mortality 100%
leaf elongation rate decreased 43%
root energy charge decreased 22%
root ATP production decreased
no visual signs of acute toxicity
high salinity (S), 50% mortality
high temperature (T), 33% mortality
high S and T, 100% mortality
net shoot loss 65% at 34-35°c
no new leaf emergence at 34-3 5 °c
Citation
Penhale and
Wetzel (1983)
Goodman et al.
(1995)
Terrados et al.
(1999)
Holmer and
Bondgaard (2001)
Carlson et al.
(1994)
Koch (1999)
Erskine and Koch
(2000)
Koch and Erskine
(2001)
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.
6.6
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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 rotundatd) 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 NFLf
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(n)) and manganese
(Mn2T) 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(in) compounds. An
advantage of this proposition would be competitive anaerobic mineralization of organic matter
by Fe(ni) 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 (Hoberg and Jensen 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 H^S 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 |j,mol. This study was done
at saturating irradiance (400-500 (imol photon m"2 s"1). Koch and Erskine (2001) determined the
response ofThalassia 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 (35°C). 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,
6.7
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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 as 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 CC>2 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).
6.8
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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 all984;
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
6.9
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and utilization of stored carbohydrates, the availability of NH4+ and N(V 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 //M saturate the growth response of Z. marina. In contrast, Van Katwijk et al. (1997)
reported that an NH4+ concentration of 125 //M was toxic to eelgrass, and suggested that the
toxicity threshold was as low as 25 //M. However, Hebert et al. (2007) reported up to 2500 fjM
concentration of NHt+ 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
6.10
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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 fine-
grained 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 MLt+
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.
6.11
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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 Zoster a 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 brackish-
water 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-
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Burrell, D. C. and J. R. Schubel. 1977. Seagrass ecosystem oceanography. Pp. 195-232. In: C.
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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 andBorum 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
7.1
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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 andNeckles 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.
7.2
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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 imiflora) 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,
7.3
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Increasing:
blade age
epiphyte % cover and
biomass
light attenuation
uv protection
epiphyte diversity
canopy shading
Decreasing:
transpiration
blade structural
integrity
blade chlorophyll
content
>'*
&
w
fc
%
c
t
i.
V
fc
Increasing leaf turnover rates,
or grazer abundance
Tip
Increasing environmental nutrients
Base
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 Mxon 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
7.4
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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 ofHalodule 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. angustifolid) 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.
7.5
-------
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
0resund, Denmark
0resund, Denmark
Roskilde, Denmark
Mesocosms, VA
Mesocosms, RI
Elkhorn Slough, CA
Suspension
Suspension
Suspension
Suspension
Scraped and unscraped
leaves
Scraped and unscraped
leaves
A
A
A
A
N
A
1.9(DW)
9.5 (DW)
.0039 (chl a)
20(DW)
99 (DW)
99.7
99.7
76
90
60
Borum & Wium-Andersen 1980
Same absorbance data as 1980; Borum et al. 1984
May underestimate total absorption, Borum 1987
Measured spectral absorption, Neckles (1993 and
unpubl. data), cited in Brush & Nixon 2002
Effect varied with type of epiphyte, Brush & Nixon
2002
Measured spectral absorption, underestimated
epiphyte biomass by use of phospholipid measure
only; Drake et al. 2003
Thalassia testudinum
Terminos Lagoon, Mexico
Tampa Bay, FL
Indian River Lagoon, FL
Bahamas
Suspension
Suspension
Scraped and unscraped
leaves
Scraped and unscraped
leaves
N
A
A
A
17(DW)
18(DW)
81(51)
85(61)
36
Kemp et al. (1988), cited in Brush & Nixon 2002
Excluded necrotic blade tips, underestimates total
absorbance; Dixon & Kirkpatrick 1995
Harden 1994
Measured spectral absorption, underestimated
epiphyte biomass by use of phospholipid measure
only; Drake et al. 2003
Halodule wrightii
Tampa Bay, FL
Indian River Lagoon, FL
Suspension
Scraped and unscraped
leaves
A
A
30 (DW)
99(48)
90 (60)
Excluded necrotic blade tips, underestimates total
absorbance; Dixon & Kirkpatrick 1995
Harden 1994
7.6
-------
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, Cebrian et al. 1999
Heterozostera tasmanica
Victoria, Australia
Scraped and unscraped
leaves
A
3.3 (DW)
75
Bulthuis & Woelkerling 1983
Artificial substrates
Cockburn Sound, Australia
Success Bank, Australia
Bane d'Arguin, Mauritania
Delaware Inland Bays
Virginia, Maryland Bays
Growth on glass slides
Growth on plastic slides
Growth on glass slides
Scraped and unscraped
mylar strips
Scraped and unscraped
plastic strips
A
A
N
N
A
0.004 (chl a)
4.2 (AFDW)
3(DW)
17(DW)
4(DW)
96(63)
96
Silberstein et al. 1986
Burt et al. (1995), cited in Brush & Nixon 2002
Hootsmans et al. (1993), cited in Brush & Nixon
2002
Glazer (1999), cited in Brush & Nixon 2002
Measured spectral absorption, Brandt & Koch 2003
7.7
-------
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
7.8
-------
corresponding to total N concentrations in winter of 38.8, 30.3, 21.6, and 12.8 jiM, 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
7.9
-------
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 jiM 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 1 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
7.10
-------
turnover (5% per day), while all tanks included a variety offish 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, 2. 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 1 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 1101 aquaria to test effects of elevated
nutrients and reduced light on Zoster a 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
7.11
-------
Table 7.2. Summary of laboratory microcosm and mesocosm nitrogen and phosphorus enrichment assays using seagrasses.
Treatment Abbreviations: C = control (ambient conditions), +N = NOs 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
meso-
cosm:
-2
m
Algae
Grazers
Experimental Results and
Comments
Zoster a marina
Netherlands:
van Katwijk et
al.1997
Washington:
Williams &
Ruckelshaus 1993
Rhode Island:
Lin et al. 1996
Temperature: 15, 20
°C
Sediment: Mud, Sand
Nutrients: +AN
(9:3;25:25;25:50;25:12
5;75:75;125:25 \iM)
Sediment: C, D, +A (5,
20 |^M)
Water: C, +A (10, 33
^M)
Nutrients:
C,+P(.7^iM),+N(7.6|i
M), +NP(7.6,.7|^M),
+A(7.6,.7|^M)
Continuous
feed from
stock
solutions
Continuous
feed from
stock
solutions
Once per
day,
dissolved
19.4 1 : 1.2 d"1
75 1: 61 d'1
4554 1: .05 d'1
40
26-62
518:
250 m'2
Epiphytes
not
apparently
measured
Epiphyte
biomass
measured
Epiphyte
biomass
measured
No
Amphipods
possibly
present
Added inverts
& fish, no
epiphyte
grazers
+A - toxic at 125 uM, probably
toxic as low as 25 [iM
+N - no effect
Toxicity twith t temperature.
Epiphyte biomass t with t water
column NH4. Eelgrass leaf
growth 1 with t epiphyte
biomass.
No t in epiphytes in nutrient
additions. Phytoplankton
blooms may have limited
epiphyte response.
7.12
-------
Rhode Island:
Taylor etal. 1999
Rhode Island:
Kopp 1999
Rhode Island:
Bintz et al. 2003
Virginia:
Neckles et al.
1993
Virginia:
Moore & Wetzel
2000
North Carolina:
Burkholder et al.
1992
Nutrients: C, +AP
(9 I'l 8 236 3'7 3
6|iM)
Nutrients: C, +N (50,
75, 100, 150 |iM)
Light: C, 55%
Nutrients: C, +NP
(6. 5,. 5 |^M)
Temperature: 3 levels
(9 yr mean, 4° above,
4° below)
Nutrients: C, +AN
(16.4-37.8 |iM)P (1.8-
3.3 [iM)
Grazers: C, +G
Light: C, 9,28,42%
Nutrients: C, +ANP
(7-16.3; 6.6-7.7; 1.4-
1.8 |^M)
Nutrients: C,+N
(3.5,7, 35 |iM)
Once per
day,
dissolved
Continuous
feed from
stock header
tanks
Continuous
feed from
drip bags
Continuous
feed from
stock
solutions
Continuous
feed from
stock
solutions
Pulse, daily
4554 1: .05 d'1
150 1: 1 d'1
45541:. Id'1
1101: 16 d'1
1101: 16 d'1
15701:
0.5-0. Id'1
over a 2 hr
period only
518:
250 m'2
50: 200
m"2
580:
365 m"2
92
(15/pot):
1500 m'2
92
(15/pot):
1500 m'2
-2000:
700 m'2
Epiphyte
biomass
measured:
added 4
macroalgae
Epiphyte
biomass
measured
after 1
month
Epiphyte
biomass
measureSd:
added 2
macroalgal
species
Epiphyte
biomass
measured
Epiphyte
biomass
measured
Epiphytes
assessed as
cell counts
only
Added inverts
& fish,
bottom
grazing snails
@100
No
Grass shrimp
@ 8 per tank
Snails,
isopods,
amphipods 9-
11.4xl03 m'2
Snails;
5,200 m'1
Amphipods +
isopods @
25-58 xlO3
m"2
No significant treatment effects
on seagrass, epiphytes or drift
algae.
No sig. effect of nitrate treatment
on epiphytes (although trend for
t for high light, high nutrient),
low light level 1 epiphyte load
significantly.
Epiphytes t in warm, nutrient
enriched versus cool treatments.
Seasonally varied grazer
densities and nutrients.
Epiphyte biomass twith
t nutrients.
Shading inhibited epiphytes and
eelgrass. Epiphytes t only with
high light and nutrients in
Spring.
No significant treatment effects
on epiphytes. Variable
macroalgal response. Grazer
density was high in spring
experiment.
7.13
-------
North Carolina:
Burkholder et al.
1994
New Hampshire:
Short etal. 1995
Nutrients:
Spring C,+N (5 \iM)
Fall C, +N (10 \iM)
A)Light: 11,2 1,4 1,61,
94% of ambient
B) Nutrients: C,+AP
@6x
C)Light:ll,41,94%x
Nutrients: C,+AP (7-
13 |^M) (1.4-2.4 \iM)
Pulse, daily
Continuous,
from bags of
slow release
fertilizer
15701:
0.5-0. Id'1
over a 2 hr
period only
8001,2 d'1
-2500:
800 m"2
200:
133 m'2
Epiphytes
assessed as
cell counts
only
Macroalgae
and epiphyte
biomass
assessed
Amphipods
@ -7-80 xlO3
m"2
Added
amphipod
predators and
mud snails
Epiphyte biomass was
"negligible". Grazer densities
very high in both spring and fall
experiments
Epiphytes t compared to control
in high light treatment. Variable
responses to nutrients at lower
light levels.
Thalassia testudinum
Florida:
Tomasko and
Lapointe 1991
Florida:
LaPointe et. al
1994
Light: C, -33%
Nutrients: C, +ANP
(10,1,10 |^M)
Nutrients: C, +A (10
\iM) , +P (1 |^M)),
+AP(10,l^iM)
Pulse daily,
from
solutions
Pulse daily,
from
solutions
120 1: (?50 d'1)
static for 14 hr
d-1
1201:50 d'1
static for 14 hr
d'1
?9
?
Epiphyte
biomass
measured
Epiphyte
biomass
measured
Added mixed
grazers,
density?
Added mixed
grazers,
density ?
+ANP t epiphytes and 1 rhizome
growth.
Shading 1 epiphyte effect.
Both +A and +P t epiphytes and
1 rhizome growth.
Halodule wrightii
Florida:
LaPointe et. al
1994
North Carolina:
Burkholder et al.
1994
Nutrients: C, +A, +P,
+AP
Nutrients:
Fall C, +N (10 \iM)
Pulse daily,
from
solutions
Pulse, daily
1201:50 d'1
static for 14 hr
d'1
15701:
0.5-0. Id'1
over a 2 hr
period only
?
-3768:
1200 m'2
Epiphytes
assessed
Epiphytes
assessed as
cell counts
only
Added mixed
grazers,
density?
Amphipods
@ -7-80 xlO3
m"2
Both +A and +P t epiphytes;
only +P 1 rhizome growth.
Epiphyte biomass was
"negligible". Grazer densities
very high in both spring and fall
experiments
7.14
-------
Ruppia maritima
Maryland:
Twilley et al.
1985
North Carolina:
Burkholder et al.
1994
Nutrients: C, +NAP @
3 levels
Nutrients:
Fall C, +N (10 \iM)
Pulse, 6
@10d
intervals,
fertilizer
Pulse, daily
340,000 1:
pulse turnover
@8-9d
15701:
0.5-0. Id'1
over a 2 hr
period only
?
50-80%
cover
-3768:
1200 rn2
Epiphytes
measured
Epiphytes
assessed as
cell counts
only
?not
specified but
likely
Amphipods
@ -7-80 xlO3
m"2
+N at >60 |^M caused seagrass
1.
Epiphytes t .
Epiphyte biomass was
"negligible". Grazer densities
very high in both spring and fall
experiments
Potamogeton perfoliatus
Maryland:
Murray et al. 2000
A) C, +NP as Pulse,
+NP as Continuous
(bothSS^iMN, 3.8
I^MP)
B) Nutrients: D(2-4
\iM), +N (20-24 \iM)
Water exchange:
1,3,6,12^
C) Nutrients: +N (<10
l^mol I"1 N, >30 [imol
r'N)
Grazers: C, +G
A) Pulse and
continuous
B, C)
Continuous
A) 101
B) 100 1: see
Treatment
Variables
C) 100 1: 1 d'1
?
Epiphyte
biomass
measured
A) No
B)No
C) 0.25 grn2,
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.
7.15
-------
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 testiidmum, 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
7.16
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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.
7.17
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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
7.18
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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.
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species in the Indian River Lagoon, Fl. M.S. thesis, Florida Institute of Technology,
Melbourne Fl. pp. 130.
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Handbook of Seagrass Biology: An Ecosystem Perspective. Garland STPM Press, New
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on Thalassia hemprichii (Ehrenb.)Aschers. from Papua New Guinea. Aquatic Botany
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Hootsmans, M. J. M. and J. E. Vermaat. 1985. The effect of periphyton-grazing by three
epifaunal species on the growth of Zostera marina L. under experimental conditions.
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epiphyte grazers. Aquatic Botany 24:287-302.
Humm, H. J. 1964. Epiphytes of the seagrass, Thalassia testudinum, in Florida. Bulletin of
Marine Science of the Gulf and Caribbean 14:306-341.
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Jernakoff, P., A. Brearley, and J. Nielsen. 1996. Factors affecting grazer-epiphyte interactions
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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 eutrophi cation renders the negative effect
of marine eutrophi cation 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).
8.1
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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
Chaetomorpha
macroalgae
macroalgae
Ulva
Ulva
macroalgae
macroalgae
macroalgae
macroalgae
Ulva
Cladophora
Ulva
macroalgae
Ulva,
Chaetomorpha
macroalgae
macroalgae
Cladophora,
Gracilaria
Ulva
Ulva
Cladophora,
Gracilaria
macroalgae
macroalgae
macroalgae
macroalgae
Cladophora,
Gracilaria
macroalgae
macroalgae
Cladophora,
Gracilaria
macroalgae
Ulva
macroalgae
Zostera suffocated
macroalgae replaced Zostera
Zostera declined
no effect
Zostera uprooted, buried
macroalgae replaced Zostera
macroalgae replaced Zostera
Zostera declined
Zostera declined
macroalgae replaced Zostera
Zostera declined
Zostera declined
Zostera bed suffocated
macroalgae replaced Zostera
Zostera declined
macroalgae replaced Zostera
macroalgae replaced Zostera
decreased meadow area
shading killed Zostera
macroalgae replaced Zostera
shading killed Zostera
macroalgae replaced Zostera
loss of seagrass
macroalgae replaced Zostera
macroalgae replaced Zostera
removal increased Zostera
Zostera declined
macroalgae replaced Zostera
macroalgae replaced Zostera
no displacement
no effect
macroalgae replaced Zostera
Netherlands Coast
North Sea
Coast of Poland
Coast of Ireland
Northeast Pacific
Venice Lagoon
Coast of Denmark
Coast of Poland
Coast of Poland
Northwest Atlantic
Northwest Atlantic
Coast of Denmark
SW Coast of England
Northwest Atlantic
North Sea
Coast of Portugal
Baltic Sea
Northwest Atlantic
Venice Lagoon
Coast of Portugal
Northwest Atlantic
Temperate Zone
Northwest Atlantic
Northwest Atlantic
Northwest Atlantic
Northwest Atlantic
Northwest Atlantic
Coast of Denmark
Northeast Atlantic
Northeast Pacific
Northeast Pacific
Adriatic Coast of Italy
den Hartog and Polderman (1975)
Nienhuis (1983)
Plinski and Florczyk (1984)
Whelan and Cullinane (1985)
Kentula and Mclntire (1986)
Sfrisoetal. (1989)
Funen County Council (1991)
Kruk-Dowgiallo (1991)
Ciszewskietal. (1992)
Valielaetal. (1992)
Short etal. (1993)
Thybo-Christesen et al. (1993)
den Hartog (1994)
Lyons etal. (1995)
Nienhuis (1996)
Oliveira and Cabecadas (1996)
Schramm (1996)
Short and Burdick (1996)
Coffaro and Bocci (1997)
Flindt etal. (1997)
Hauxwell etal. (1998)
Raffaelli etal. (1998)
Bricker etal. (1999)
Valiela et al. (2000a)
Bowen and Valiela (2001)
Deegan et al. (2002)
Hughes et al. (2002)
Nielsen et al. (2002)
Hauxwell et al. (2003)
Kentula and DeWitt (2003)
Thorn et al. (2003)
Curiel et al. (2004)
8.2
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Zoster a noltii
Ulva
Ulva
macroalgae
Ulva
Ulva,
Chaetomorpha
Ulva
Ulva
Ulva
Ulva
Ulva
Zostera suffocated
macroalgae replaced Zostera
little effect
macroalgae replaced Zostera
Zostera declined
macroalgae replaced Zostera
macroalgae replaced Zostera
macroalgae replaced Zostera
macroalgae replaced Zostera
macroalgae replaced Zostera
Thalassia testudinum
macroalgae
Chaetomorpha,
Acetabularia
Eucheuma
macroalgae
macroalgae
macroalgae
decreased productivity
uprooted; stopped
recolonization
possible competition
no effect
possible competitor
potential threat
Netherlands Coast
Venice Lagoon
Dutch Wadden Sea
Spanish Atlantic Coast
North Sea
Portuguese Atlantic
Northwest Atlantic
Portuguese Atlantic
Portuguese Atlantic
Portuguese Atlantic
Gulf of Mexico
Mexican Caribbean
Gulf of Mexico
Gulf of Mexico
Columbian Caribbean
Gulf of Mexico
den Hartog and Polderman (1975)
Sfrisoetal. (1989)
Philippart and Dijkema (1995)
Nielletal. (1996)
Nienhuis (1996)
Oliveira and Cabecadas (1996)
Flindtetal. (1997)
Pardal et al. (2000)
Cardoso et al. (2004)
Patricio et al. (2004)
Cowper(1978)
Merino etal. (1992)
Perez-Enriquez (1996)
Bell and Hall (1997)
Angel and Polania (2001)
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
Katwijketal. 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
8.3
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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,
8.4
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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 Mclntire (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, Thorn 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 Acetabularid) 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
8.5
-------
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
8.6
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grazers decreased, resulting in negligible top-down control of macroalgal biomass. Then, using
macroalgal enclosure or exclusion field experiments, Hauxwell et al. (2001) compared eel grass
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 PC>4" 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).
8.7
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25
20 -
C\l
o
O
_c
CO
15 -
10 -
T
1
5 -
T
Random Samples Controls Enclosure Controls Algae Removal Algae Addition
Figure 8.1. Effects of experimental benthic green macroalgal (Ulva spp.) manipulations on shoot
density ofZostera 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 (Laurenciapoiteaui) 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
-------
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
8.9
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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.
8.10
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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
9.1
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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 - 20 °C, 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 513 C values indicates that seagrasses tend to become more 13C
depleted from tropical (warmer) to temperate (cooler) regions (Hemminga and Mateo 1996).
9.2
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Variation (in order of relative importance) of source of carbon, irradiance and temperature on
CC>2 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 15°C, 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
9.3
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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 10-18°C 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 6-21 °C
in Rhode Island, 10-20° C in Halifax, N.S., while Puget Sound temperatures varied only from 8-
9.4
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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 eel grasses, cite the agreement of their data on east
coast 2. 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-turbidity-
caused 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
9.5
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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 thatZ. marina seedling survival peaked at 10° C and 10-20700, Z.
noltii survived best at 10° C and 1700, and both species showed maximal germination at 30° C
and l%o 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 Nino 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 Nino 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 Nino episodes. Thorn et al. (2003)
found that warmer winters and cooler summers associated with the transition from el Nino to la
Nina 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
9.6
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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, light-
saturated gross photosynthetic rates increased in direct proportion to temperature increase up to a
point between 30-35 °C, 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
9.7
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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.Ox, i.e., l.Ox = 31, 3.Ox = -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.
Perez-Llorens 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 10°C
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.
9.8
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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 above-
ground biomass and leaf area index (LAI) of Z. marina from a Mexican coastal lagoon were
associated with water temperature.
9.9
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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. (Wyllie-
Echeverria 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.
Marba 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.
9.10
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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 28-31 °C 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 longh et al. 1995; Stapel et al. 1997). In an extreme case the upper margin of
a Zostera noltii bed was described as "burned" 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 Mclntire 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
9.11
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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 2. 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
Mclntire 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).
9.12
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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 (Zm;n) 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
-------
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 "windows"
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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10.0 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 parti culate 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
10.1
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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 Wyllie-
Echeverria 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
10.2
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plume of ejected
sediment covering.
feeding pit excavating
seagrasses
burrow mounds from
infaunal bioturbator.
such as burrowing
mp
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
Burrowing shrimp
Seagrass Species
Zostera marina, Z. japonica
Z. marina, Z. japonica
Z. japonica
Z. japonica, Z. capricorni
Z. noltii
Z. capensis
Z. capensis
Posidonia oceanica
Halodule uninervis
Thalassia testudinum
T. hemprichii
Syringodium isoetifolium
Reference
Thompson and Pritchard 1969
Harrison 1987
Dumbauld and Wyllie-
Echeverria 2003
Berkenbusch et al. 2007
Pranovietal. 1996
Siebert and Branch 2005, 2006
Angel et al. 2006
Molenaar and Meinesz 1995
Duarte et al. 1997
Suchanek 1983
Duarte etal. 1997
Duarte et al. 1997
10.3
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Stingrays
Crabs
Polychaetes
Echinoderms
Cymodocea rotundata
C. serrulata
Z. marina
Z. marina
Z. marina, H. wrightii
H. wrightii
T. testudinum
T. testudinum
T. testudinum
Syringodium sp.
Z. capricorni (=novazelandica)
Z. marina
Z. marina
T. testudinum
Z. noltii
Z. marina
Z. marina
Zostera spp.
Z. marina
Duarteetal. 1997
Duarteetal. 1997
Orth 1975
Merkel 1990
Townsend and Fonseca 1998
Fonsecaetal. 1994
Ogden 1980
Zieman 1982
Valentine et al. 1994
Fonsecaetal. 1998
Woods and Schiel 1997
Davis and Short 1997
Davis etal. 1998
Valentine et al. 1994
Phillipart 1994
Luckenbach and Orth 1999
Davis and Short 1997
Hughes et al. 2000
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
10.4
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«* \ Ł Great Pond Bay - Productivity
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OCT ' NOV ' DEC I JAN ' FEB ' MAR
1980 1981
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.
10.5
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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 ofCallianassa 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 ofPosidonia
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, Upogebiapusilla; 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 co-
occurring 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
10.6
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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 Wyllie-
Echeverria (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 (shrimp-
dominated 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 Wyllie-
Echeverria (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 TV. californiensis declined. They suggested that a
combination of rapid growth by the eelgrass and relatively low bioturbation by N. californiensis
10.7
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(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. californiemis and Upogebiapugettensis. 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.
Thorn 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 TV. 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 organic-
rich 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
10.8
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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
10.9
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narinarf) were observed to damage turtlegrass (T. testudinum) habitat while feeding (Ogden
1980; Zieman 1982 in Valentine et all994).
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 andHalodule 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//, 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 ofHalodule 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
10.10
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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'?, 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
10.11
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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 (Limuluspolyphemus) 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
10.12
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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 bioturbation-
induced damage to roots may have been an experimental design artifact, and thus the role of
bioturbation in this worm-seagrass interaction is uncertain.
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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 TV. virens, although their natural history
observations presented a plausible mechanism.
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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 arenicold) 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 quinquiesperforatd) 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-
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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
Callicmassa 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
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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 eel grass, 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 (7! 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; paniculate 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
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water. Typically, overlying water has higher concentrations of dissolved oxygen and lower
concentrations of nutrients, hydrogen sulfide, CC>2 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 physiologically-
based stress-response model for Thalassia testudimim, 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
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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 disturbance-
tolerant 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 (BJ. 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 ofZostera 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 ofZostera 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
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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 californiemis), 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
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virginicd) 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,
Corbiculafluminea. 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. virginicd) and hard clams (Mercenaria mercenarid) 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 Upogebiapusilla (Dworschak 1981) and U. pugettensis (DeWitt et al.
2004; Griff en 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
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.
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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;
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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.
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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 bioirrigator-
seagrass 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.
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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
11.1
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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
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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 Zoster a 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.
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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 oi Zoster a 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
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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
Hauxwell, J., J. Cebrian, C. Furlong, and I. Valiela. 2001. Macroalgal canopies contribute to
eel grass (Zoster a marina) decline in temperate estuarine ecosystems. Ecology 82:1007-
1022.
Lee, K. S., F. T. Short, and D. M. Burdick. 2004. Development of a nutrient pollution indicator
using the seagrass, Zostera marina, along nutrient gradients in three New England
estuaries. Aquatic Botany, 78: 197-216.
Nixon, S., B. Buckley, S. Granger, and J. Bintz. 2001. Responses of very shallow marine
ecosystems to nutrient enrichment. Human and Ecological Risk Assessment 7:1457-
1481.
U.S. EPA. 1998. National Strategy for the Development of Regional Nutrient Criteria. U.S.
EPA, Office of Water. EPA 822-R-98-002.
U.S. EPA. 2000. Ambient Water Quality Criteria Recommendations. Information Supporting
the Development of State and Tribal Nutrient Criteria. Wetlands in Nutrient Ecoregion
XIII. U.S. EPA, Office of Water, Office of Science and Technology. EPA 822-B-OO-
023.
U.S. EPA. 2001. Ambient Water Quality Criteria Recommendations. Information Supporting
the Development of State and Tribal Nutrient Criteria. Lakes and Reservoirs in Nutrient
Region III. U.S. EPA, Office of Water, Office of Science and Technology. EPA 822-B-
01-008.
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.
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 190:155-165.
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