EPA/600/R-09/050

June 2009



## Seagrasses and Protec tive Criteria: A Review and Assessment of

Research Status

Edited by:

Walter G. Nelson

Western Ecology Division

National Health and Environmenta l Effects Research Laboratory U.S. Environmental Protection Agency Newport OR 97365

Contributors:

Cheryl A. Brown, Bruce L. Boese, Theodore H. De Witt, Peter M. Eldridge (deceased), Mark G. Johnson, James E. Kaldy III, Walter G. Nelson, Da vid 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 97365ii 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 pub lication. 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 doc ument for their many perceptive and helpful comments. Dr. Timothy Nelson provided review co mments for all chapters of the report. Dr. Robert Virnstein and Ms. Lori Morris kindly pr ovided 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 data base of bibliographic entries for seagrass literature. Expert secretar ial assistance was provided by Ms. Karen Ebert and Ms. Jimmie Cheney.

This report is dedicated to the memory of Dr. Pe ter 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 Prot ective Criteria: A Review and Assessment of

Research Status. Office of Research and Deve lopment, National Health and Environmental

Effects Research Labor atory, EPA/600/R-09/050.iii Abstract

This report is a review and summary of the current status of scientific information relevant to the establishment of protective criteria for the most widely distributed seagrass species of the United States, eelgrass Zostera marina , and turtlegrass Thalassia testudinum . The report focuses on scientific information related to major limiting factors for seagrass survival, and assesses the degree to which environmenta l factors may need to be included in the development of adequately protective criteria. The review confirmed that th ere is a great deal of scientific information currently av ailable concerning the responses of Zostera marina and

Thalassia testudinum to a wide range of environmental f actors. However, interactive effects among factors influencing seagrass survival re main relatively poorly known, especially across broader regional scales. This appears true even for such fundamental environmental characteristics as salinity and temperature and thei r interactions in the ex pression of nutrient or sediment impacts on SAV, although research is be ginning to fill this gap. The question remains as to whether current modeling approaches, whet her empirical or mechan istic, are adequate to predict the response of seagrasses to even single stressors. A ke y concern is that there is a high level of uncertainty in being able to predic t the trophic pathway for expression of nutrient impacts on seagrasses. Thus, wa ter quality criteria based on nutri ent concentrations may not be adequately protective of seagrass resources. Alte rnate standards based on water clarity or water column chlorophyll a criteria may not be adequately prot ective if the principle expression of nutrient impacts occur through the epiphyte or macroalgal pathways . There are also important influences on seagrass survival through sedime nt associated mechanisms that may not be adequately captured by water qua lity criteria alone. There may be advantages to looking for integrative, plant based seagrass condition indicators, such as sucros e content, that relate to the ability of seagrasses to survive within a tempora lly varying environment. Such measures may be an appropriate method to integrate water column and sediment impacts into single protective

criteria.ivv 

## 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 Nutr ients as Limits to Growth of Zostera marina and

Thalassia testudinum ..................................................................................................................3.1 James E. Kaldy III

Chapter 4 Salinity as a Limiting Fa ctor 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. Boesevi Chapter 10

The Effects of Bioturbation a nd Bioirrigation on Seagrasses ............................................10.1

Theodore H. DeWitt

Chapter 11 Research Gaps in Relation to Setting Protective Criteria ...................................................11.1

Walter G. Nelsonvii 

## List of Figures

Figure 1.1. Example of pathway for establis hing and implementing seagrass protection and restoration targets. .......................................................................................................... ..1.6

Figure 1.2. Diagram of the critical path fr om 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 wh ich the criterion is set. .........1.8

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

Figure 2.2. An idealized change in incident light intensity th roughout a summer and winter day at 45 degrees N. .............................................................................................................. ..2.4

Figure 2.3. Gross maximu m productivity, gross P

max ( ■ ), I k values ( ▲ ), and dark respiration

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

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

Figure 2.5. Photon flux necessary to replace ca rbon lost to respira tion as a function of temperature (assuming equal daylight and nocturnal respiration rates). .......................2.18

Figure 3.1. Schematic diagram of the eff ects of increased nut rient 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 seagra ss as a function of current speed. ...............................5.10

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

Figure 7.1. Schematic diagram of seagrass plant illu strating typical patterns of distribution of epiphytes within and among blades. ................................................................................7.3

Figure 7.2. Horizontal arrows indicate external factors in fluencing abundance of seagrass epiphytes. .................................................................................................................... .....7.4

Figure 8.1. Effects of experiment al 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. ...........10.3viii Figure 10.2. Effects of biot urbating burrowing shrimp ( Callianassa spp.) on turtle grass ( Thalassia testudinum ) on subtidal sand flat s on St. Croix island. ................................10.5

Figure 10.3. Results of two laboratory experime nts showing decreased survivorship of

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

## List of Tables

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

Table 1.2. List of factors poten tially 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 factor s 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 . V

max in units of μ M gdw lf -1 h -1 is the maximum uptake rate, K

s is the half saturation concentration ( μ M). .......................3.2

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

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

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

Table 3.4. Water column and sedi ment nutrient concentrations ( F M) associated with selected

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

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

Table 4.2. Reported salinity tole rance ranges for turtlegrass ( Thalassia testudinum ) populations. .................................................................................................................. ....4.2

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

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

Table 7.1. Studies of light attenuation by epiphyt es in coastal marine systems, revised and expanded from Table 1, Br ush and Nixon (2002). ..........................................................7.6x Table 7.2. Summary of laboratory microcos m 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 impact s for different seagrasses 10.31.1 

## 1.0 Conceptual Framework for a Review of Research Needs for Development of National Water Quality Criteria Protective of Seagrasses

Walter G. Nelson

1.1 Background of EPA Submerged Aqua tic Vegetation Research Program

U.S. coastal waters provide tremendous na tural, economic, and public health benefits. Excess nutrients, together with excess carbon, cau se 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 ecol ogical effects such as loss of submerged aquatic vegetation (SAV), degradation of benthos, and ar eas of hypoxia/anoxia. Some of these changes affect use of the Nation’s aquatic resources, and pose risks to human heal th 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 c onditions due to eutrophication (Bricker et al. 1999). Evidence suggests that nutrient inputs from point and nonpoint source s 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, a nd predict the risks that excessive nutrients pose to the coastal marine environment. While nutrients can be a stre ssor resulting from human activities, they can also have beneficial effect s. Therefore, understa nding 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 E nvironmental Effects Research Laboratory (NHEERL) has undertaken an exte nsive research program to impr ove 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 nutrien t 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 polyha line portions of estuaries, are critical to maintaining estuarine “h ealth” and ecosystem function, and are widely regarded as a cornerstone of estuarine productivity. As su ch there is an increased awaren ess 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 lo ading can result in an accumulation of epiphytic, macroalgal and phytoplankton biomass that shades s eagrass and results in loss of areal coverage. Losses may result from direct impacts such as physical removal from the action of dense algal mats, or via indirect effects that result from sulfide toxicity and light reduction. Because estuaries are complex ecosystems, it is important to understand both the stressors (e.g., light,1.2 temperature, plankton blooms, t oxicity, algal mats, etc.) and e nd-point response variables (i.e. seagrass growth, biomass, C:N, carbohydrate reserves, etc.). Th e 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 effect s research on SAV has focused on developing nutrient load-response relationships for seagrass habitats. Th is research effort is designed to provide a scientific basis for developing nutrient cr iteria that would help protect seagrass habitats from degradation or loss, and to provide manage ment 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-respon se models, the ultimate goal is to develop and couple the stress-response models, which ar e plant scale and physio logically based, to population scale models. The Western Ecol ogy Division (WED) has developed a seagrass stress-response model for Z. marina (Kaldy and Eldridge 2006), th at has been parameterized with data from Yaquina Bay, Oregon, as representa tive of conditions in the Pacific Northwest. In parallel with this effort, Gulf Ecology Division (GED), Gulf Breeze, Florida, has validated a similar stress-response mode l 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 re sponse to nutrient stre ss over the range of distribution of the targ et seagrass species.

The specific objective of the eelgrass stress-r esponse 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 enhancemen t in estuaries. The initial ee lgrass stress-response model is one which predicts response at the individual seag rass 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 accuratel y predict seagrass nutrient responses.

While it would be desirable to link the patch scale models to populat ion 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 condi tions, water temperature) at sta tions located across the span of the estuarine system of interest, the model can be used to estimate seag rass plant responses to

increased loads of nutrients and/or sediment s at these measurement points. Under the assumption that stations are at least locally re presentative, model predictions at these multiple locations can be used to examine whether propos ed nutrient standards are adequately protective across the entire estuarine system, or whether they may need to be varied across an estuarine gradient.1.3 1.2 Regulatory Background

The U.S. Environmental Protection Agency is concerned with the protection of seagrasses under two sections of the Clean Water Act (33 U.S.C. 1252 et seq.), Section 304(a) and Section 404(c). Under section 304(a) (1) the agency is charged with development of water quality criteria reflecting the latest scientific kno wledge on effects of pollu tants on aquatic biota, including “plant life”. The ag ency is further charged (Sectio n 304(a) (2)) with providing timely scientific information on factors necessary to maintain the chemical, physical and biological integrity of the nations waters. Under sect ion 404 which regulates dredging and dredged material disposal under the lead of the Army Co rps of Engineers, EPA is authorized to deny issuance of dredged material disposal permits where such activity will have “an unacceptable adverse effect on municipal wate r 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 w ithin 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 exte nsive overview of the issues of nutrient over enrichment in marine waters. It provides gui dance on steps for developm ent of nutrient criteria (see section 1.4 below), and provi des 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. Ultimate ly, it is the states an d tribes that adopt water quality criteria and standa rds, and they would be respon sible for adopting any criteria specifically protective of seagrasses.

1.3 Pathways to Water Quality Criteria Setting

Kenworthy (1992) provides a detailed histor y of the early evolu tion of water quality criteria relevant to submerged aquatic vegetati on, tracing the original guidelines for developing state standards back to the Federal Water Qua lity Act of 1965. His review similarly addressed the history and development of the State of Florid a water transparency and turbidity standards.

His assessment at that time was that both the fe deral 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 th e past, the process of setting water quality criteria might propa gate uncertainties contained in guidance documents in such a way as to lead to flawed criteria. As a specifi c example, the state standard at the time included a numerical definition of a light compensati on point derived from phytoplankton which was inappropriate for the survival of seagrasses.1.4 Table 1.1. Sections of the Clean Water Act wh ich pertain to the protection of seagrasses.

S EC . 304. (a)(1) The Administrator, after consultation with appr opriate Federal and Stat e 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) criteri a 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 recr eation 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.

S EC . 304. (a)(2) The Administrator, after consultation with appr opriate Federal and Stat e agencies and other interested persons, shall develop and publish, within one year after the date of enactment of this title (and from time to ti me 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 ac tivities 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 fo r maximum daily load m easurement correlated with the achievement of water quality objectives.

S EC . 404. (c) The Administrator is authorized to prohibit the specification (including the withdrawal of specification) of any defined area as a disposal s ite, and he is authorized to deny or restrict the use of any defined area for specification (inc luding 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 r ecreational 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 th e 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 - R ecreation, Propagation and Maintenance of a Healthy, Well Ba lanced 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 derive d directly from the language of the Clean Water Act (e.g. Sec. 304. (a)(2), Table 1.1). While en compassing seagrasses, in doe s 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 im portant strategy in the Te xas plan was to propose that “Seagrass Habitat” be added to the list of Designated Uses for the state of Texas. With this1.5 basis, water quality criteria specifically protectiv e 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 desi gnated use to address seagrass protection comes from Chesapeake Bay (EPA 2003). In or der 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 de fined “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 onl y 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 rest oration 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 fo r extent of seagrass rest oration are first set. Setting of quantitative seagrass rest oration targets may be based on a variety of lines of evidence. Seagrass conservation targets developed for the Indian River Lagoon of Fl orida (Virnstein and Morris 2000, Steward et al. 2005) in volves three approaches, deve lopment of a potential target

based on distribution in the best available habitat (a reference condition approach), the use of a historically based distri bution target, and a “critical minimu m” approach which establishes the

lowest threshold below which s eagrass 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 (Figur e 1.2) to a critical process path (e.g. U.S. EPA 2001; Batelle 2008). With in the U.S. EPA Regional Offices, there are Nutrient Coordinators who establish Regional Te chnical Assistance Groups to assure that the best available current information is brought to the criterion development process, and inappropriate guidance is weeded out. As sugge sted 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 gui dance for water quality criteria protective of the seagrass resource on a nationa l basis is presaged by Kenwort hy (1992) in his consideration of development of a general state standard for Fl orida. He suggests that because of the broad variation in physical systems, a nd in the sources and causes of water quality impairments, “a general standard will fail to adequately prot ect 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 crit erion is set is summarized in Fi gure 1.3. Variation in physical1.6 systems becomes greatly magnified as one move s from state to national scales. The highly specific designated use boundary delineation de fined 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 limiti ng the ability to achi eve broadly protective water quality criteria for seagrasses are the ec onomic limitations and constraints imposed by the necessity to develop water-body specif ic criteria. In contrast to st ates 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 pr ocess of considering how best to develop water quality criteria protective of seagrass by first evaluating current knowledge on seagrass ecology in the context of developing protective standards for seagrasses.

Figure 1.1. Example of pathway for establis hing and implementing seagrass protection and restoration targets. The approach illustr ated is for the Tampa Bay National Estuary Program (modified from Johansson and Greening 2000). Adopt quantitative seagrass restoration goals Determine light requirements needed

to meet seagrass restoration goals Determine water quality necessary to maintain light requirements

Determine chl a concentration targets necessary to maintain water clarity requirements

Determine nutrient loads necessary to

maintain chl a targets Define and implement nutrient management strategies to achieve load reduction goals Measure progress toward targets and goals. Reassess as

needed.1.7Figure 1.2. Diagram of the critical path from Goal Statement to Implementation by the States/Tribes of numeric standards fornutrients.1.8Figure 1.3. Diagram of the presumed relationship of effectiveness of anutrient criterion in protecting resources at the sitescale versusthe scale at which the criterion is set.1.9 1.4 Goals and Limitations: Reviewing the En vironmental Requirements of Major U.S. Seagrass Species

During the 1980's, a series of summary docum ents on seagrass communities of the U.S. were produced by the U.S, Fish and Wildlife Serv ice (Phillips 1984; Thayer et al. 1984; Zieman 1982; Zieman and Zieman 1989). These broad re views encompassed asp ects 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 prepari ng for the present review, Western Ecology Division

scientists established a Micros oft Access bibliographic database of greater than 10,000 entries related to seagrass biology.

The aim of the present repor t 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 re levant to the process of developing protective criteria of major estuarine SAV resources of the continental U.S. As originally conceived, the review effort was to cover the three most widely distributed species of seagrasses in continental

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

effort was to include authors from other EPA divisions who could provi de regional knowledge on seagrass ecology, together with specialized knowledge on aspects of seagrass ecology. Unfortunately, this concept was not realized as divi sional research priorities shifted and scientists were unable to fully participate in the project. Thus, individual chap ter 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 de gree. A second limitation to the report is that publication has been repeatedly de layed, again due to shifting prio rities, and therefore the most recent seagrass literature may not be included. Desp ite this, the report synthesizes a great deal of information that it highly relevant to the manage ment and protection of se agrasses as a critical aquatic resource.

The principle goal of the review is to highlig ht 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-sp ecific requirements for the range of principal factors potentially limiting seagrass growth and survival within th e limits described above (Table 1.2). The review is structured in chapters wh ich describe individual lim iting 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 gr owth, survival, reproduction and distribution of seagrasses. The review of resear ch has a narrow focus on work that is directly relevant to the1.10 limiting factor and seagrass conditio n. For example, while a comp rehensive evaluation of what is known about taxonomic composition of epiphytic cover may be of significant academic interest, it would not be a focu s of the review unless it provide s some essential insight into seagrass survival. The focus is instead on whethe r there is adequate evidence to indicate whether for example, epiphytes can limit seagrass gr owth, 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 sec tion 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 potentia l target?”

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

this review.

Limiting Factor Subject to Anthropogenic

Modification Chapter

Light Yes 2

Nutrients Yes 3 Salinity Yes 4 Current and Wave Exposure Yes 5 Sediment Characteristics (Grain size, dissolved oxygen, sulfide) Yes 6

Epiphytes Yes, through indirect

mechanisms 7

Macroalgae Yes, through indirect

mechanisms 8

Temperature (Freezing, Heat Stress, Desiccation) Yes, through both direct and

indirect mechanisms 9

Bioturbation Yes, through indirect

mechanisms 10

Diseases Yes, through indirect

mechanisms 10

Herbivory Yes, through indirect

mechanisms 7 for micrograzers1.11 1.5 Literature Cited

Batelle. 2008. Conceptual Plan for Nutrient Criteria Development in Maine Coastal Waters. Report prepared for EPA Region 1, Maine Dept . of Environmental Protection, and U.S.

EPA Oceans and Coastal Protection Division. Batelle, Brunswick ME. EPA Contract No. 68-C-03-041. pp. 31. http://www.maine.gov/dep/blwq/report/ 2008/nutrient_crite ria_report_2008.pdf Bricker, S. B., C. G. Clement, D. E. Pirhalla , S. P. Orlando, and D. R. G. Farrow. 1999. National Estuarine Eutrophication Assessmen t: Effects of Nutrient En richment 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. Bo ese, T.H. DeWitt, P.M. Eldridge, J.E. Kaldy, H. Lee II, J.H. Power, and D.R. Young. 2007. An Approach to Developing Nutrient Criteria for Pacific Northwest Estuaries: A Case Study of Yaquina Estuary, Oregon. USEPA Office of Research and Development, National Hea lth and Environmenta l Effects Laboratory, Western Ecology Divi sion. EPA/600/R-07/046. Johansson, J. O. R. and H. S. Greening. 2000. Seagrass restoration in Tampa Bay: A resourcebased approach to estuarine management. pp. 279-294. In: S. A. Bortone (ed.). Seagrasses: Monitoring, Ecology, Physiol ogy, and Management. CRC Press. Boca Raton. pp. 318. Kaldy, J.E. and P.M. Eldridge (eds.). 2006. De velopment of mechanistic models to guide establishment of protec tive criteria for seagrasses. Report on the Zostera marina and

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

National Health and Environmental Effects Research Laboratory, EPA Internal Report. Kenworthy, W. J. 1992. Do federal water qualit y criteria and Florida state water quality standards protect seagrasses? A feasibility evaluation of altern ative criteria and standards to monitor, evaluate and regulate water tran sparency. Appendix II, pp. A-11 to A-29, In: A. Hart (ed.) Proceedings and Conclusions for the Submerged Aquatic Vegetation Initiative. Indian River Lagoon, National Estu ary 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 Vegeta tion and Photosynthetically Active Radiation. Special Publication SJ93- SP13. St. Johns River Water Management District, Palatka, Florida.

National Research Council. 2000. Clean coastal waters: understan ding and reducing the effects of nutrient pollution. National Acad emy Press, Washington, DC. pp. 405. Phillips, R. C. 1984. The ecology of seagrass meadows in the Pacific No rthwest: 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 fo r the Indian River Lagoon system , Florida. Estuaries 28:923-

935.1.12 Texas Parks and Wildlife. 1999. Seagrass conservation plan for Texas. Texas Parks and Wildlife Department, Resource Protec tion Division. Austin, Texas. pp. 79. Thayer, G. W., W. J. Kenworthy, and M. S. F onseca. 1984. The ecology of eelgrass meadows of the Atlantic coast: a community profile. U.S. Fish and Wildlife Service FWS\OBS84\02. pp. 147.

U.S. EPA. 2001. Nutrient cr iteria 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 a nd Implementation Plan for Effects Research. US EPA, Office of Research and Developm ent, National Health and Environmental Effects Laboratory, pp. 179. U.S. EPA. 2003. Ambient water quality crit eria for dissolved oxygen, water clarity and chlorophyll a for the Chesapeake Bay and its tidal tri butaries. 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. Ecologi cal Applications 7:737-750. Virnstein, R. W. and L. J. Morris. 2000. Se tting seagrass targets for the Indian River Lagoon, Florida. pp. 211-218, In S. A. Bortone, ed. S eagrasses: 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 Bi ol. Rep. 85(7.25). pp.

155.2.1 

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

marina Growth and Distribution

Robert J. Ozretich

2.1 Background

Light is a fundamental requirement for s eagrasses. The energy derived from photons is used to reduce carbon dioxide and fuel the biosyn thesis of carbohydrates that make up the bulk of these plants, amino acids a nd lipids. Without light consis ting of a sufficient quantity of photons of wavelengths overlapping the abso rption 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 av ailable to seagrasses

can be direct and indirect. Uncontrolled constr uction 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 potential ly reach seagrasses. Similarly, addition of

nutrients to a watershed from ex cess run off of nutrients from ag ricultural 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 indivi dual plants and from consid eration of the maximum depth distributions of a given species . Individual plant requirement s have focused on the plant’s response to light through the e xperimental determination of P vs. I (photosynthesis versus irradiance) curves (Figure 2.1). By comparing ra tes of photosynthesis with respiration, the daily number of hours of light saturating irradiance (H sat , Figure 2.2) necessary to balance respiratory

demands has been estimated. Estimates of H sat 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 acc ount for the astronomica lly and meteorologicallyinduced variation in time and la titude of the incident light field (Figure 2.2) and the in situ changes in water composition to which it is expose d. 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 fa vorable conditions.

To reach the leaf surface of seagrasses, s unlight 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 understa nding of the spectral and intensity changes in light prior to reaching the leaf surface has resulted in bio-optic al-physical models of varying complexities. Some models have progressed to the point where the spectral dyna mics of light through a stand of seagrass has been coupled with primary production models to predict in situ growth and2.2 distribution. Modeling of intensiv ely and extensively studied syst ems 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 li ght with dissolved and pa rticulate material in water as well as the in-water consequences of meteorological and astronomical changes (Mobley 1994). This understanding is expr essed 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 availa ble to the public. Other solutions of the RTEs have been published that produce comparable results in certai n 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 be cause the scattering in teraction of matter and light is a function of matter’s dimensions compared to light’s wave lengths, and because interactions with matter in which energy is absorbed are a quantum pheno mena 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 wave length) absorption coefficient, a ( λ ):

a ( λ ) = a

w ( λ ) + a DOM ( λ ) + a phto ( λ ) + a TSS ( λ ) Equation 2.1

where a

w , a DOM , a phto , and a TSS are the contributions to absorp tion by water molecules, dissolved organic matter (DOM), phytoplankton, and tota l suspended solids, respectively. Each is wavelength-dependent and the last thr ee are also concentration dependent.

Absorption and scattering are considered inhe rent optical properties (IOPs) of water and
in combination contribute to the spectr al, diffusive attenuation coefficient, λ

D K which is
considered an apparent optical property (AOP) b ecause it varies with the in-air geometry and intensity of the light source.

The consequence of absorption and scattering of light in water is expressed through the following, familiar expression of “Beer’s Law”:

)*(

0 ZK

Z D eII λ

λλ

−

= Equation 2.22.3 where λ

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

0 I is the surface

irradiance.

-0.05 0 0.05

0.1 0.15

0.2

0 50 100 150 200 250 300 350



### μ mol photons m -2 

### sec -1μmol O2mg chl a-1min-1 P

max

I k I

c R alpha

Figure 2.1. Photosynthesis vs. Irra diance curve where alpha is the initial slope of the curve, I k is the light intensity at the onset of light-saturated photosynthesis, P

max ; I
c , the compensation light intensity, is the light intensit y at which dark respir ation, R, is equal to the rate of photosynthesis. The dark respir ation determined graphically from P vs. I measurement represents dark respiration in th e light but may not be representative of a lower nocturnal dark respiration because of the plant’s response to the enhanced postillumination respiration phenomenon (Hei chel, 1979; Falkowski and Raven, 1997).2.4 0 200 400 600 800 1000 1200 1400 1600 1800

0:00 3:00 6:00 9:00 12:00 15:00 18:00 21:00 0:00μmol photons m-2sec-1

Summer

H sat = 13 h

H comp = 14.5 h

I k = 100

I

c = 25 Winter

H sat = 8 h

H comp = 9 h

Figure 2.2. An idealized change in incident light intensity th roughout a summer and winter day at 45 degrees N latitude where I k and I

c are from the P vs. I curve in Figure 2.1. H sat are the hours during which I k is exceeded and H comp are the hours during which production exceeds respiration and the light intensity exceeds I

c .

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 avai lable and conversion to K D (PAR) for modern mode ling applications has been found to be site-specific, especially for estuarine settin gs (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 sens ors (LiCor Environmental, Lincoln, NE; Biospherical Instruments, San Diego, CA). The use of photoelectric inst ruments in permanent moorings or deployments from vessels have s upplanted the Secchi depth measurement as their outputs are directly useful in calculating PAR attenuation coeffi cients 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 availabl e radiation (PAR). Measured at the deepest sites wh ere seagrasses are found, PAR has been used to assess long-term light requirements as a percen tage of PAR incident on the water’s surface (Duarte 1991;2.5 Dennison et al.1993). Multi-varia nt regression models using TSS (or turbidity ), chlorophyll a , and DOM concentrations as combined predictors of K D (PAR) were found to be insufficiently accurate in comparison to RTE-derived estimates of K D (PAR) for the purposes of establishing management options for seagrasses in Chesap eake Bay (Gallegos 2001). The consequences of different mitigation strategies could be asse ssed in achieving water quality goals when the contributions of these water column components to light attenuation have been determined.

Using parameterized absorption and scatte ring for two different seagrass habitats (including Zostera marina ), Zimmerman (2003) used the Hydrolight model in combination with his own model to investigate the light environm ent 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 i rradiances 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, c oncentration-based models would be universally applicable. One such model parameterization, K D = 0.32 + 0.016 * chl a ( μ g l -1 ) + 0.094 * TSS (mg l -1 ), has been done for Chesapeake Bay (Kemp et al. 2004) where, 0.32 is the contribution to attenuation by water molecules. Unfortunately, the spectral characteristics of the light absorbing and scattering components of natural waters appe ar to be at least system dependent, if not temporally dependent. For example, terrestrially derived material (abi otic particles and DOM) from different locales (Gallego s and Kenworthy 1996) can have di fferent 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 optic ally important components of natu ral waters will therefore be required for each studied system (Gallegos 2001; Zimmerman 2003).

While solution of the RTEs computes PAR reach ing the depth of a plant’s leaves, it is not often the PAR that actually reaches the chlorophy ll-containing surface of the leaves. This is

because of the colonization of th ese surfaces by a variety of organi sms 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 K D (PAR) through the epiphytic la yer as a function of both growing season nitrogen or phosphorous concentrati ons and TSS. These algorithms were used to reduce the percentage of water-in cident irradiance supporting seag rass populations that had been

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

2.2.3 Instrumentation Used in Support of RTE Parameterizations

Validation of modeli ng results requires in situ measurement of the light fields. Longterm and shipboard cast deployments of broad bandwidth, PAR sensors (LiCor Environmental, Lincoln, NE; Biospherical Instruments, San Diego, CA) have historically contributed greatly to understanding the temporal and spatial variability of this critical measurement. Field-deployable2.6 full spectrum scanning radiometers have also been developed such as LiCor’s LI-1800 and the much smaller HydroRad (HOBI Labs Inc., Tucson, AZ). In recent years, instruments have been developed that can simultaneously measure a ttenuation and absorption/scattering of multiple narrow bandwidths. Examples of these types of instruments are the ac-9 (WetLabs, Philomath, OR) and HOBI Lab’s HydroScat. The ac-9 has been used extensively in estimating water

column IOPs in situ (Zimmerman 2003; Gallegos and Jordan 2002). However, these field measurements require labor inte nsive complimentary laboratory determinations that involve filtering of samples to spectrally characterize the dissolved matter using standard scanning spectrophotometers and those w ith integrating sphere attach ments (Drake et al. 2003) for particulate matter.

2.2.4 Failure to Survive in Op tically Sufficient Systems Although the correlations betw een the distribution of persistent seagrass meadows and seasonally-averaged K D (PAR), optically important wate r 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. Su ch circumstances are not captured in annualized

data. Loss of Zostera marina resulting from light limitation ha s been observed both during field observations (Dennison and Alberte 1985; Moore et al .1996; 1997; Cabello-Pasini et al . 2002) and experimentation (Alcoverro et al. 1999; Cabello-Pasini et al . 2002; Thom et al . 2002; Biber

et al . 2005). The field and experimental condit ions, 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 requir ements 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 dimini shed light exposure. For the plan ts in the dark (Cabello-Pasini

et al . 2002; Biber et al .2005, Table 2.1), respiration was the do minant process. While the aim of these studies differed, each i nvoked either, the concept of H sat (the daily period of light-saturated photosynthesis, Figure 2.2), or H comp (Dennison and Alberte 1982), the daily period of light

above I

c , 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 H sat per day to meet the demands of growth and respiration (Dennison and Alberte 1982; Marsh et al.1986; Zimmerman et al.1989). Give n 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 re commended wide range in the number of hours of

H sat or H comp is insufficiently proscriptive.

2.2.5 Laboratory Studies of Zostera marina Photosynthesis Laboratory experiments to determine seagra ss production in response to light have been performed with leaf segments (scraped of epiphytes), or whole plants. These tissues are2.7 incubated in varying PAR intensity while mon itoring the change of dissolved oxygen (Figure 2.1) or other constituents. The reported units of P in the P vs. I curve ar e varied (Table 2.2) and are often not easily converted to common units becau se 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, dm 2 ). In addition, the variability of the P vs. I components is usually not reported and is often high because of the few da ta points and the steep in itial slope (alpha) of the relationship (Zimmerman et al. 1991). The e ffect of temperature on the P vs. I components have also been investigated (Drew 1979; Mars h et al. 1986; Bulthuis 1987; Zimmerman et al. 1989) (see Chapter 9). Instruments and techni ques have been deve loped to monitor the fluorescence of various pigments involved in pho tosynthesis. These include the photosynthetic efficiency analyzer and pulse amplified modulated (PAM) fluorometry which have been used to assess various acute stresses (Major and D unton 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 primar y carbon reserve during periods of light limitation (Smith et al. 1988; Smith 1989; Kraemer and Al berte 1995; Zimmerman et al. 1995; Zimmerman and Alberte 1996). Sucrose is the carbon source for dark resp iration during respiratory or fermentative processes. During periods of light , sucrose and oxygen produ ced 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 s hortly afterwards (1h) (Zimmerman and Alberte 1996), at which time the roots become fully a noxic (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. Transloc ation of inorganic phosphate from the roots to shoots is strongly light depende nt (Brix and Lingby 1985), whereas a sucrose gradient between rhizomes and shoots may offset the reduced dow nward 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 μ mol 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 su rvive (Alcoverro et al. 1999) while the rhizome sucrose content was much less reduced. Although the fresh weight concen tration of sucrose in shoots, roots, and rhizomes is on average 166, 20, and 240 μ mol 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).2.8Table 2.1. Observations of the death ofZostera marinaplants in the field andlaboratory. NR = not reported.LocationMonth/yearStudy typeTemperature°CIrradiancemol m-2day-1Hsathourday-1Time todeathweeksReferenceMA, Great Harbor 6/82 field experiment 15~2~6 ~4 Dennison andAlberte 1985MA, Great Harbor 8/82 field experiment 21~2~6 ~4 Dennison andAlberte 1985VA, York estuary 4-5/87 field monitoring 211.5-2.5 ~0 (Hcomp) 3-4 Moore et al. 1997Mexico, BajaCalifornia, Pacificcoast2-3/98 field monitoring 17 0-4 NR ~4 Cabello-Pasini et al.2002CA, Monterey Bay 11-12/93 lab experiment 12~0.32~4 Alcoverro et al.1999WA, Sequim Bay 11/95 lab experiment 91-2NR ~4 Thom et al. 2002Mexico, BajaCalifornia, Pacificcoast11/00 lab experiment 1600 3-4 Cabello-Pasini et al.2002NC, North River 2-3/03 lab experiment 10 to 1600 5-7 Biber et al. 20052.9Table 2.2. P vs. I experimental results forZostera marinaL. from various locations and measurement conditionsμmol photonsm-2sec-1Parameters in published unitsLocationTemp°CIkIcalphax1000PmaxRespirationPublishedunitsmg chl a[dm-2] orgfw-1ReferenceAK931mg C h-1gdw-1McRoy 1974NC 0.27-1.8mg C h-1gdw-1Penhale 1977CA 5-20 5(15 °C)1(15 °C)14.2(15 °C)1.4(15 °C)μg C h-1cm-2Drew 1979MA shall/deep 220.4 0.08 -0.13μmol O2min-1mg chl a-1[1.8]/[2.2] Dennison and Alberte1982VA 1-28 80-385 4.4-8.9 1.13-2.8mg C h-1gdw-1Wetzel and Penhale 1983VA 15/25 37 11.4 0.42/0.62μmol O2min-1mg chl a-11.5 Evans et al. 1986MA20 15-25 4.4 0.22 0.080μmol O2min-1mg chl a-1[2.5] Dennison and Alberte1985MA shall/deep 20 67 18 2.7-3.7 0.09-0.21 0.06-0.09μmol O2min-1mg chl a-1[2.7]/[4] Dennison and Alberte1986MA 0-35 7-90 1-17 4-9 0.055-0.45 0.0083-0.13μmol O2min-1mg chl a-1andμmol O2min-1cm-2[2.2] Marsh et al. 1986MA shall/deep 20 78/71 13/14 4.2/2.4 0.33/0.17 0.053/0.034μmol O2min-1mg chl a-1[2.9]/[4] Dennison et al. 1987CA-Monterey 10/20 5/8μmol O2h-1gfw-11.1/1.5 Zimmerman et al. 1989CA-SF 15 35 0.74 0.06μmol O2min-1mg chl a-1Zimmerman et al. 1991VA 25 210 12 3.8 0.84 0.06μmol O2min-1mg chl a-1Goodman et al. 1995CA-Monterey 15 0.25 0.02μmol O2min-1mg chl a-12.2 Zimmerman et al. 1996CA-Monterey 150.2 0.085μmol O2min-1gfw-11.5 Alcoverro et al. 1999Mexico-Baja 16-22 1-350 10-60 0.6-2.45μmol O2min-1gfw-12 Cabello-Pasini et al.2002Mexico-Baja 12-27 50-140 5-10 0.3-0.6 0.03-0.06μmol O2min-1gfw-1Cabello-Pasini, et al.20032.10 Table 2.3. Sucrose content of Zostera marina L. tissues. NR = not reported

Sucrose content μmol sucrose gfw -1

(leaves / roots / rhizomes) Loss rate μmol sucrose gfw -1 day -1 (leaves / roots / rhizomes) Reference

220 / 30 / 300 Zimmerman et al. 1989

NR/ 10 / 250 Kraemer and Alberte

1993

130 / 10 / NR no significant loss (whole plants in dark over 12 hr at unknown temperature) Zimmerman et al. 1995

40 / 15 / 250 Kraemer and Alberte

1995

100 / 30 / 200 Zimmerman et al. 1996

90 / 20 / 200 2.2 / NR / 3.4 (whole plants in ~0.3 mol photons m -2 day -1
for 30 days at 12 oC) Alcoverro et al. 1999

233 /NR /NR Touchette and

Burkholder 2001

350 / NR / NR 15±1 / NR/ NR (whole plants in dark at 16 oC over 21 days) Cabello-Pasini et al.

2002

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

Zostera marina

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

an initial reservoir of a relative ly uniform concentration of resp irable compounds. Sucrose is the likely compound that would sustain Zostera marina plants in the absence of light and would be replenished when irradiance exceeded I comp for a long enough time. When its concentration was reduced to near zero, the plants could not recover a nd 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 c ontent that is between a low of ~15 μ mol sucrose gfw -1

and an average, healthy plant, concentration of 166 ± 40 μ mol sucrose gfw -1 ( ± 1SE). The carbon equivalents of this range are ~180 and 1992 ± 480 μ mol C gfw -1

.

Both P

max 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

P

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

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

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

where the units are μ mol O 2 mg dm -2 min -1 and ± is 1 SE of the regression coefficients and intercepts, respectively.

The similarities between P

max values (on a leaf segment ar ea basis) from photoacclimated plants with different chlorophyll contents (De nnison and Alberte 1982 and 1985) suggest that the temperature-dependent dark bioche mical reactions control the optim um photosynthetic rate. It is

for this reason that the area-norm alized 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 Ra ven 1997). Mean alpha values from Table 2.2 were converted to quantum yields, μ mol O 2 μ mol 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 μ mol O 2 μ mol photons -1 (n = 9). Quantum yields are used to compute the photosynthesis-saturati ng irradiance, I k , from gross P

max (equation 2.12). Re latively invariant alphas, including some from likely photoacclimat ed plants, reflect the consequences of the tightly coupled, species dependent, chlorophyll a mediated light reactio ns (Falkowski and Raven 1997). A consequence of an invarian t alpha, but temperature-dependent P

max , is that I k , is also a function of temperature (Figure 2.3).

Sucrose constitutes approximately 90% of the soluble carbohydrate of Zostera marina (Drew 1983; Smith 1989; Alcoverro et al. 1999). U nder 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 struct ural carbohydrates, it can be readily respired e ither aerobically or thr ough 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 condi tions. Cabello-Pasini et al. m onitored sucrose in leaves from a lagoon and offshore, and leaves that were expe rimentally 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 tr ansplanted to the laborat ory 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 leaf2.12 Table 2.4. Conversion factors from various sources.

Andersen and Johnson 1 Evans et al. (1986) 2 Nelson 3 Chlorophyll a from

Table 2.2 Units gfw dm -2 gdw dm -2 gfw gdw -1 gfw gdw -1 gdw dm -2 mg dm -2 mg gfw -1

Mean 2.168 0.350 6.26 4.4 0.25 2.79 1.6

SD 0.538 0.081 0.99 0.4 0.062 0.82 0.40

n 72 100 215 8 6

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

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

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

Table 2.5. Statistics of variab les used for calculations.

Production Respiration Sucrose μmol O 2 dm -2 min -1 alpha Content Loss rate

5 oC 20 oC 5 oC 20 oC μmol O 2 μmol photons -1 μmol gfw -1 μmol gfw -1 day -1

mean 0.290 0.828 -0.027 -0.154 0.0285 166 -15.1 1 -2.2 2

SD 0.199 0.157 0.070 0.049 0.0113 107 1.97 0.80

n 7 7 8 8 9 7 4 4

1 Cabello-Pasini, et al. (2002)

2 Alcoverro, et al. (1999)

respiration rate of 0.12 μ mol O 2 dm -2 min -1 (from Equation 2.4). The following substitutions with the experimentally-determined fresh weight to leaf area conversions allow comparisons of the respiratory carbon and sucrose carbon losses:

6.2 12
* 15 12180 Loss Sucrose 5.2. 1 1
* min1440

* 17.2 * min 12.0 0.879n Respiratio

2 2

2 2

Equation sucrosemol Cmol

daygfw

sucrosemol

daygfw

Cmol Eq Omol Cmol

daygfw dm

dm Omol

daygfw

Cmol

μ μ μ μ μ μ μ μ

−

= ±−= −

= ±−=

where the rates and conversion factor s are from Tables 2.3 to 2.5.2.13 Ik

R 2 = 0.90 gross P

max
R 2 = 0.90

respiration

R 2 = 0.85

0 0.2 0.4 0.6 0.8

1 1.2

0 1 0 2 0 3 0



### degrees Celsiusμmol O2dm-2min-1

0 40 80 120 160 200 240μmol photons m-2sec-1

Figure 2.3. Gross maximu m productivity, gross P

max ( ■ ), I k values ( ▲ ), and dark respiration rates ( ● ) from Marsh et al. 1986.

While the dark respiration and the sucrose loss rates are significantly different (t

s 6.95,

t .05[9] = 2.262), the higher sucrose loss rate appe ars 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 (B eers 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 S

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

v ) is the square root of the sum of the sq uared 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 xy , below, is the

correlation coefficient between x and y ).2.14 d Equation nSESD c Equation SDSDrSDSDSDyxvfor b Equation SDSDrSDSDSDyxvfor a Equation

z SD

y SD

x SD S

v SD zyxvfor

vv yx xyy x v yx xyy x v z y x

v v

7.2 7.2 2

, 7.2 2

, 7.2

,

2 2 2 2 2 2 2

∗= ∗∗−+= −= ∗∗++= += ⎟ ⎠ ⎞ ⎜ ⎝ ⎛

+ ⎟ ⎟ ⎠ ⎞ ⎜ ⎜ ⎝ ⎛

+ ⎟ ⎠ ⎞ ⎜ ⎝ ⎛

==∗∗=

To conservatively compute the standard error of the calculated values from SD

v , n from the variable with the smallest sample size was used (Tables 2.4 and 2.5).

The lower sucrose loss rate (2.2 vs. 15 μmole sucrose gfw -1 day -1 ) in the Alcoverro et al. (1999) experiments (Table 2.5) reflects the cont ributions 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 experi ment at 12 °C is as follows:

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

8.2 12
* 2.2 8.426 Equation sucrosemol Cmol

daygfw sucrosemol

daygfw

Cmol

μ μ μ μ Δ

−

= ±−

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

Carbon balance from computed carbon fixation and respiration:

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

9.2 2.1

1

* min120

* 17.2 * min 54.0 6.29.24

2 2

2 2 Equation Omol Cmol

daygfw dm

dm Omol

daygfw

Cmol

μ μ μ μ ⎟ ⎟ ⎠ ⎞ ⎜ ⎜ ⎝ ⎛

= ±

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

24 hours of respiration in leaf tissue at 12 °C from Equation 2.4 yields:2.15 10.2 1 1

* min1440

* 17.2 * min 086.0 6.657

2 2

2 2 Equation Omol Cmol

day gfw dm

dm Omol

daygfw

Cmol

μ μ μ μ ⎟ ⎟ ⎠ ⎞ ⎜ ⎜ ⎝ ⎛

−

= ±−

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

, where the SE was computed by propagating th e uncertainty from the gross production and respiration using Equation 2.7a into their differences while ac counting for the correlation of production and respiration (Equation 2. 7b). They are strongly correlated (r pr =0.847) because gross production includes respiration. The carbon loss as sucrose, -26 ± 4.7 μ mol C/gfw-day, is not significantly different than the net production rate (t

s 0.88, t .05[9] = 2.262).

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

conversion factors that were used. It was assu med 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 demons tration or quantificati on 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, le af 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 da rk respiration and sucrose utilization rates (Equations 2.5 and 2.6) become closer an d not significantly di fferent (157 and 180 μ mol C gfw 1 day -1 while the net production with 2 hours of light (Equations 2.9 and 2.10) went up to -64

μ mol C gfw -1 day -1 which is significantly larger (t

s 3.29, t .05[9] =2.26) than the sucrose loss rate of

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

the various conversion factors needed for these calculations, the net produ ction 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 betw een 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 pl ant. Attaining and maintaining this maximum level is an indication of a healthy meadow. Co ncentrations much lower would be suggestive of plants under light stress.2.16 Periodic sampling and analysis of leaf sucr ose content could be us ed 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 P

max 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 P

max and respiration rates at the experimental temperatures to compute net production. In the field, produ ction will not always be maximal (for the temperature) or zero (darkness), but will range from zero to gross P

max because photons can

contribute to carbon-fixation at fluxes down to ~0.3 μ mol photons m -2 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, I k . Therefore, net production computed from the in situ temperature and time-interval averaged irradiances is the following:

11.2 *

max Equation

RPgross I I P t t

t
k i

neti − =

Where, I

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

I

k t is calculated by dividing gross P

max at temperature t by the quantum yield, alpha. This yields

I

k t with the following units:

12.2 sec60

min

* 100
* 029.0 * min sec 2 2

2 2 2

2 Equation

m

dm

Omol photons mol

dm

Omol

m photons mol

μ μ μ μ

=

The units for
i net P are:

13.2. min

* 17.2 * * min 2.1 * min * 1 1 2

2 2 2

2 2 2 Eq igfw dm

Omol Cmol

dm

Omol

Omol Cmol

dm

Omol

igfw Cmol ⎥ ⎦ ⎤ ⎢ ⎣ ⎡ ⎟ ⎟ ⎠ ⎞ ⎜ ⎜ ⎝ ⎛ ⎟ ⎟ ⎠ ⎞ ⎜ ⎜ ⎝ ⎛

− ⎥ ⎦ ⎤ ⎢ ⎣ ⎡

= μ μ μ

μ μ μ μ

To meet the expectations of the hypothesis, that sucrose mainta ined within a range of leaf concentrations is indicative of healthy Zostera marina plants, the following must occur when the optimal sucrose concentration (Sucrose

max ) is 166 μ mol sucrose gfw -1 (Table 2.3) or 1992 μ mol C gfw -1 :

Sucrose

max ≥ i Sucrose = ( i-1 Sucrose + i P net /12) ≤ 15 Equation 2.142.17 where 12 is the number of carbon atoms per sucrose molecule. When i P net is positive, carbon is added to the sucrose pool up to Sucrose

max .

For these computations, plant death occurs at i Sucrose

∼ 15 μ mol 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 i Sucrose accounts for varying light and temperature conditions across time and would indicate when light and temperat ure conditions persisted long enough to sustain the growth of Zostera marina . A specific time under a given light flux, e.g. H sat , 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 th e sucrose pool in complete darkness from its maximum level can be estimated from the respir ation rate alone as a function of temperature (Figure 2.4).

The minimum photon flux necessary to sust ain a plant (no net production, Sucrose

max maintained) was calculated in the following:

15.2 100

* 029.0 * 2.1

*

2 2

2 2

2 2 Equation

m

dm

Omol photons mol

Cmol

Omol

daydm Cmol

daym

photons mol

μ μ

μ

μ μ

−

=

where μmolC dm -2 day -1 is the temperature dependent daily re spiration rate that is converted to photon flux through the quantum yield and normaliz ing 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 sa me logic as is invoked by the concept of a

minimum number of H sat , or H comp hours. However, by themselves, neither is especially useful in regard to contributing to testable hypotheses about minimum light requirements because of the dynamic nature of in situ light fluxes.2.18 0 10 20 30 40 50 60 70 80 90 100

2.5 5.0 7.5 10.0 12.5 15.0 17.5 20.0 22.5 25.0 27.5

degrees Celsiusdays to death in dark476 days103 days

Figure 2.4. Days to plant death in dark assu ming the sole respirator y carbon source is leaf sucrose. Error bars ( ± 1 SE) are from the computations incorporating the uncertainties of the underlying measurements (Tables 2.4 and 2.5).

0.0 0.5 1.0 1.5 2.0

0 2.5 5 7.5 10 12.5 15 17.5 20 22.5 25 27.5 30 degrees Celsiusmol photons m-2day-1to meet respiratory demands

Figure 2.5. Photon flux necessary to replace carbon lost to respira tion as a function of temperature (assuming equal daylight and noctu rnal respiration rates). Error bars ( ± 1 SE) are from the computations incorpor ating the uncertainties of the underlying

measurements.2.19 2.3 Research Gaps in Relation to Setting Protective Criteria

The correlative approach utilizing grow ing season measurements for setting system habitat requirements is incapable of accounting for the varying periods of diminished irradiance

that have led to loss of Zostera marina populations. The i Sucrose model (above), utilizing irradiance and temperature values from continuous data sets collected fr om strategically placed locations within a system, would provide the mean s 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 photosynthesi s (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 u tility of leaf sucrose concentrat ions in assessing the health of this plant depends on the isolation of its l eaf sucrose pools from other tissues additional simultaneous studies of these sucrose pools unde r low light conditions are warranted. Such studies would assess the importance of this mechan ism 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 us ing the same laboratory techniques, same water samples, and to the extent possible, should in corporate the use of multi-spectral instruments,

such as the WetLab ac-9. To reduce costs of such surveys, the resulting IOPs could be processed by the analytically robust, and freely-available RTE model of Ga llegos (1994). This model is currently included as a Fortran module (C. Gall egos, 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 s cattering spectra (e.g., abiotic-mineral-derived turbidity, peatderived DOM) and ranges of spectral responses fo r, say, mono-specific and mixed-species algal blooms. In addition, such surveys would demonstr ate 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 sens ors as the conversion to K D (PAR) from Secchi readings have introduced unnecessary uncertainty in this critical measurement.

Physiological studies of P vs. I, sucrose c ontent etc., should always include a range of normalizing measures. The minimum, most cri tical measurements are chlorophyll per fresh weight and surface area per fr esh weight of blades, and fr esh weight for below ground components, if studied. Chlorophyll normalization has been shown to reduce the variability of several physiological parameters while normaliz ation to fresh weight may minimize confounding photoacclimation issues resulting from varying chlorophyll content. Th e use of blade surface2.20 area normalization is a means to directly couple in situ light fluxes and laboratory-determined photosynthetic rates.

2.4 Literature Cited

Alcoverro, T., R. C. Zimmerman, D. G. Kohrs, a nd R. S. Alberte. 1999. Resource allocation and

sucrose mobilization in light-limited eelgrass Zostera marina . Marine Ecology Progress

Series 187:121-131. Batiuk, R. A., R. J. Orth, K. A. Moore, W. C. De nnison, 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 Re quirements and Restoration Targets: A Technical Synthesis. CBP/TRS 83/92, Chesapeake Bay Program, Annapolis, MD. Beers, Y. 1957. Introduction to the Theory of Error, 2 nd edition. Addison- Wesley Publishing Company, Reading, MA. Biber, P. D., H. W. Paerl, C. L. Gallegos, and W. J. Kenworthy. 2005. Evaluating indicators of seagrass stress to light limitation. pp. 167-176. Chapter 13. In: S.A. Bortone (ed.) Estuarine Indicators. CRC Press, Boca Raton, FL. pp. 318. Bjork, M., J. Uku, A. Weil, and S. Beer. 1999. Photosynthetic tolera nces to desiccation of tropical intertidal seagra sses. Marine Ecology Progress Series 191:121-126. Brix, H. and J. W. Lingby. 1985. Uptake a nd translocation of phos phorous in eelgrass ( Zostera

marina L.). Marine Biology 90:111-116. Bulthuis, D. A. 1987. Effect s of temperature on photosynthesis and growth of seagrasses. Aquatic Botany 27:27-40. Cabello-Pasini, A., R. Muñiz-Salazar, and D. H. Ward. 2003. Annual variations of biomass and photosynthesis in Zostera marina at its southern end of distri bution in the North Pacific. Aquatic Botany 76:31-47. Cabello-Pasini, A., C. Lara-Turrent, and R. C. Zimmerman. 2002. Effects of storms on photosynthesis, carbohydrate cont ent 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. Syst em-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 Ecol ogy Progress Series 25:51-61. Dennison, W. C. and R. S. Alberte. 1986. Photoadaptation and growth of Zostera marina L. (eelgrass) transplants along a depth gradient. Journal of Experimental Marine Biology and Ecology 98:265-282. Dennison, W. C., R. C. Aller, and R. S. Albert e. 1987. Sediment ammonium availability and eelgrass ( Zostera marina ) growth. Marine Biology 94:469-477. Dennnison, W. C., R. J. Orth, K. A. Moore, J. C. Stevenson, V. Cart er, S. Kollar, P. W. Bergstrom, and R. A. Batiuk. 1993. As sessing water quality with submerged aquatic vegetation. BioScience 43:86-94.2.21 Drake, L. A., F. C. Dobbs, and R. C. Zimmerm an. 2003. Effects of epiphyte load on optical properties and photosynthetic pot ential of the seagrasses Thalassia testudinum Banks ex König and Zostera marina L. Limnology and Oceanography 48:456-463. Drew, E. A. 1979. Physiological aspects of primary production in seagrasses. Aquatic Botany

7:139-150. Drew, E. A. 1983. Sugars, cyclitols and seagrass phylogeny. 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 two coexisting seagrasses, Zostera marina L. and Ruppia maritima L. Aquatic Botany

24:185-197. Falkowski, P. G. and J. A. Raven. 1997. Aqua tic Photosynthesis. Blackwell Science, Malden,

MA. Gallegos, C. L. 1994. Refining habitat requiremen ts of submersed aquatic vegetation: role of optical models. Estuaries 17:187-199. 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. S eagrass 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 lig ht 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. Goodman, J. L., K. A. Moore, and W. C. Denni son. 1995. Photosynthetic responses of eelgrass ( Zostera marina L.) to light and sediment sulfide in a shallow barrier island lagoon. Aquatic Botany 50:37-47. Heichel, G. H. 1970. Prior illumination and th e respiration of maize l eaves 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 Hab itat-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. Bergst rom, V. Carter, C. L. Gallegos, W. Hunley, L. Karrh, E. W. Koch, J. M. Landwehr, K. A. M oore, 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 Phot osynthesis in Aquatic Ecosystems, 2 nd edition. Cambridge University Press, Cambridge, U.K.2.22 Kraemer, G. B. and R. S. Alberte. 1993. Ag e-related patterns of metabolism and biomass in

subterranean tissues of Zostera 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) root s: implications for survival in light-limited environments. J ournal of Experimental Marine Biology and Ecology 185:191-202. Kraemer, G., P. D. Kohrs, and R. S. Alberte. 1998. Developmental cha nges of sucrose synthase activity and abundance in r oot and leaf tissues of Zostera marina . Aquatic Botany

62:189-198. Lee, K. S. and K. H. Dunton. 1996. Produc tion and carbon reserve dyna mics 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. Select ive light attenuation by th e 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 Variatio ns 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. Albe rte. 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 experi ments in eelgrass Zostera

marina and its epiphytes. Aquaculture 4:131-137. Mobley, C. D. 1994. Light and Water Radiative Tr ansfer in Natural Waters. Academic Press,

New York. Moore, K. A., R. L. Wetzel, and R. J. Orth. 1997. Seasonal pulses of turb idity and their relations to eelgrass ( Zostera marina L.) survival. Journal of E xperimental 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 turb idity in the lower Chesapeake Bay. Marine Ecology Progress Series 142:247-259. Penhale, P. A. 1977. Macrophyte-epiphyte bi omass and productivity in an eelgrass ( Zostera

marina L.) community. Journal of Experime ntal 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 seagra ss photosynthesis in root aerobic processes. Plan t Physiology 74:1055-1058. Smith, R. D. 1989. Anaerobic metabolism in the roots of the seagrass Zostera marina L. PhD dissertation, The University of Chicago.2.23 Smith, R. D., A. M. Pregnall, and R. S. Albe rte. 1988. Effects of anaerobiosis on root

metabolism of Zostera marina (eelgrass): implications for survival in reducing sediments. Marine Biology 98:131-141. Thom, R., A. Borde, S. Rumrill, D. L. Woodruff, G. D. Williams, J. Southard, and S. L. Blanton. 2002. Factors influencing the sp atial and annual vari ability in eelgrass (Zostera marina L.) meadows in Pacific Northwest, USA, Sy stems. Chapter 5, In: J.K Parrish and K. Litle, eds. Pacific Northwest Coastal Ec osystems Regional Study, 2001 Annual Report. K. Litle, Seattle, WA. Touchette, B.W.and J. Burkholde r. 2001. Nitrate reductase ac tivity in a submersed marine angiosperm: controlling influences of envi ronmental and physiological factors. Plant Physiology and Biochemistry 39:583-593. Twilley, R. R., W. M. Kemp, K. W. Staver, J. C. Stevenson, a nd W. R. Boynton. 1985. Nutrient enrichment of estuarine submersed vascular plant communities: I. Algal growth and effects on production of plants and associat ed 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. Al berte. 1996. Top-down impact through a bottomup mechanism: the effect of limpet grazing on growth, productivity and carbon allocation

of Zostera marina L. (eelgrass). Oecologia 107:560-567. Zimmerman, R. C. and R. S. Alberte. 1996. Effect of light/dark transition on carbon translocation in eelgrass Zostera marina seedlings. Marine Ec ology 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-Echever ria, 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. Al berte. 1989. Thermal acclimation and whole-plant

carbon balance in Zostera marina L. (eelgrass). Journal of Experimental Marine Biology and Ecology 130:93-109.3.1 

## 3.0 Water Column and Sediment Nutrients as Limits to Growth of

Zostera marina and Thalassia testudinum

James E. Kaldy

3.1 Background

Seagrasses are vascular plants that have returned to the aquatic habit. In contrast to the algae (e.g. seaweeds), seagrasses have highly differ entiated 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 th e 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 envir onmental conditions and

nutrient concentrations. This review is intended to give the reader a current basic understanding of seagrass-nutrient interacti ons and dynamics. Additional co mprehensive detail on seagrassnutrient interactions are availa ble (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 fo r 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 19 80; Brix and Lyngby 1985; Perez-Llorens et al. 1993). Touchette and Burkholder (2000a,b) provid e excellent reviews of carbon, nitrogen and phosphorus metabolism in seagrasses addressing many of the ecophysiologi cal and biochemical aspects in detail. Unfortunately, there has been little attempt within th e literature to report uptake rates with consistent units (Table 3.1). The units for phosphorus uptake are especially variable. Consequently, comparison of rates be tween systems is difficult without first making site specific conversions base d 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 nutrien ts is an important adaptation that has permitted seagrasses to persist and to out-compete alg ae (seaweeds) under oligotrophic conditions. Additionally, recent work has show n that internal resorption of N and P from senescing leaves can meet part of the plant = s nutrient requirements (H emminga et al. 1999).3.2Table 3.1. Summary of nitrogen uptake rates forZ. marina. Vmaxin units ofμM gdw lf-1h-1is the maximum uptake rate, Ksis the halfsaturation concentration (μM). Units for uptake rates aredenoted by superscripted letter.LocationLeaf uptakeRoot uptakeReferenceVmaxKsVmaxKsZostera marinaRhode Island 20.5 9.2211 104 Thursby & Harlin 1982Netherlands0.03B0.19aPedersen & Borum 1993Netherlands0.01B2.8b0.01B3.9bHemminga et al. 1994Netherlands 0.35c0.33cPedersen & Borum 1992Japan0.001d0.011dIizumi & Hattori 1982Alaska1.05B3.2b0.9B2.9bShort & McRoy 1984Thalassia hemprichiiIndonesia32-37 21-60Stapel et al. 1996.Thalassia testudinumTexas (NH4+) 8-16.5 5-198-73 34-765 Lee & Dunton 1999bTexas (NO3-) 3.4-6.5 2.2-38Lee & Dunton 1999bag N m-2d-1bμM N gdw-1h-1cmg N gdw-1d-1dg N (g plant N)-1d-13.3Table 3.2. C:N:P ratios forZ. marinafrom the literature.SeagrassLocation%C%N%PC:NN:PC:PC:N:PReferenceMultiple spp.global33.61.920.232024474474:24:1Duarte 1990Zostera marinavarious362.50.391715246255:15:1Duarte 1990California38.42.370.342016304Fourqurean et al. 1997Oregon342.70.41517255255:17:12Kaldy 2006aOregon351.30.23415510510:15:13Kaldy 2006aOregon291.40.12424576576:24:14Kaldy 2006a1827481481:27:11Pérez-Lloréns et al. 19911210120120:10:11Pérez-Lloréns et al. 19911367979:6:11Pérez-Lloréns et al. 19911115170170:15:11Pérez-Lloréns et al. 19911935757:3:11Pérez-Lloréns et al. 1991California39738274274:38:1Atkinson and Smith 1983Rhode Island341827481481:27:1Atkinson and Smith 1983Virginia421441584584:41:1Atkinson and Smith 1983Virginia35-392-3.80.2-0.32Moore and Wetzel 2000Virginia27-351-20.06-0.153Moore and Wetzel 2000Zostera capricorniAustralia2017349349:17:12Atkinson and Smith 1983Australia588465465:8:13Atkinson and Smith 1983Australia339302302:9:14Atkinson and Smith 19833.4Zostera noltiiNetherlands129104104:9:12Pérez-Lloréns et al. 1991216127127:6:13Pérez-Lloréns et al. 1991167114114:7:14Pérez-Lloréns et al. 1991Thalassia testudinumBarbados445:32:12Atkinson and Smith 1983Barbados601:20:13Atkinson and Smith 1983Texas33-371.7-2.715-252Kaldy and Dunton 2000Texas35-370.7-0.945-623Kaldy and Dunton 2000Texas34-361.7-221-242Lee and Dunton 1999aTexas36-370.7-0.946-603Lee and Dunton 199aFlorida34.62.20.09518.540.210702Fourqurean et al. 1992Florida36.91.820.11324.640.29372Fourqurean et al. 2002Thalassia hemprichiiIndonesia1.90.14302Erftemeijer & Middelburg19930.480.07153Erftemeijer & Middelburg19930.980.06364Erftemeijer & Middelburg19931Literature values from Table 1 for original citationssee Pérez-Lloréns et al. 19912Ratios for leaf material3Ratios for rhizome material4Ratios for root material3.5 In temperate systems with silastic mud sediments the general paradigm is that nitrogen is the limiting nutrient (Orth 1977; van Lent et al. 1 995). However, there is some debate about the degree of nutrient limitation in temperate systems. Zimmerman et al. (1987) concluded from a modeling study that n itrogen limitation of Z. marina is probably very rare. In tropical systems with carbonate sediments, there is substantial evidence for phosphor us 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 carbona te 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 sy stems given that coasta l upwelling occurs during summer bringing nutrient rich o cean water to the surface (10-25 μ M nitrate; C. Brown unpubl. data). During winter, terres trial run-off through anthropogenically impacted watersheds also results in large N loading. Terre strial run-off through red alder ( Alnus rubra ) rich secondary growth forest can provide substa ntial N inputs because of the nitr ogen 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 relativ ely fixed ratio of elements. Deviations from these ratios provides prelimin ary 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 th e system may be nitrogen limited (excess phosphorus), N:P > 16 system may be phosphorus lim ited (excess nitrogen). The Redfield ratio may be reflective of unicellular organisms (phyto- and bacterioplankton); however because of the structural components associated with m acrophytes (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 ra te limiting steps. Literature reviews indicate that the median leaf seagrass C:N:P is about 400: 20:1 with considerable va riability (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 materi al 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 a nd sediment nutrient pools as well as internal transport tissues (i.e. xylem and phloem). Conseque ntly, nutrient ratios alone should not be used to infer limitation (Touchette and Burkholder 20 00a). Manipulative experiments should be conducted to determine limiting factors. Additio nally, other nutrients can play a role in controlling seagrass production, for example Herman et al (1996) present a case study where they suggest that decreases in dissolved silica may have been a factor in seagrass decline.3.6 3.2 Nutrient Enrichment and Eutrophication

A Eutrophication @ is frequently used to describe the increased input of nutrients, primarily nitrogen and phosphorus to receiving waters. Re cent work has suggested that eutrophication be

redefined as A an increase in the rate of supply of organic carbon to an ecosystem @ (Nixon 1995). Sources of organic carbon supply are from e ither allochthonous or autochthonous primary production. Nixon (1995) goes on to suggest a trophic classifica tion scheme based on organic carbon supply. Eutrophication is often incorrectly used to incl ude not only the process of

increased nutrient status but also the effects (e.g. hypoxia, algal bl ooms, etc.) of this enrichment (Richardson and Jorgensen 1996).

Eutrophication sensu Nixon is caused by both natural a nd anthropogenic alterations of nutrient supply (Jorgensen and Richardson 1996). Coastal upwelling, typi cally associated with western continental margins, is a wind-driven phenomenon that results in increased primary production and often leads to enhanced fisher ies production (Thurman 19 88). Anthropogenic or

A cultural @ eutrophication alters the availability of nutrient elem ents within receiving waters which has primary, secondary and tertiary le vel impacts on biogeochemistry, primary and secondary production (Jorgensen and Richardson 1996; Vollenweider et al. 1992; Howarth et al. 2000; Livingston 2001).

Numerous studies have i nvestigated the effects of eutrophication on seagrass communities (Table 3.3). Additiona lly, the physiological response of Z. marina to light (Chapter 2) and nutrients (this chapter) have been intensiv ely studied. Most eutr ophication studies have examined the community level response in experimental systems ranging from aquaria to

mesocosms to the natural envi ronment (Table 3.3). For Z. marina much of this work has been conducted along the East Coast of North America a nd has resulted in a general theory of seagrass response. Specifically, that enhanced nut rient 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). Ho wever, there does not appear to be a relationship between nutrient input and the algal type (epiphyte vs. macroalgae vs. phytoplankton) supplying primar y production or between nutrien t input and the amount of primary production (Nixon et al. 2001).3.7 Table 3.3. Selected list of literature examining the effect of eutrophication on Z. marina communities. Abbreviations are as follows: Zm = Zostera marina, SAV = Submerged Aquatic Vegetation, Epi = Epiphytes, Phyto = Phytoplan kton, Macro= Macroalgae, Algae = epiphytes + phytoplankton + macroalgae. Positive (+) or negative (-) response in biomass is denoted.

Organizational

level Experimental
system Response Location Reference

Community aquaria +Epi, - Zm Virginia Neckles et al. 1993
aquaria + Zm , - Zm 3 Netherlands van Katwijk et al. 1999
aquaria + Epi Washington Williams & Ruckleshaus

1993
aquaria +Epi, - Zm 2 Virginia Moore & Wetzel 2000

field + Zm , +Macro Rhode Island Harlin & Thorne-Miller

1981

field + Zm Netherlands van Lent et al. 1995

field

-- Maryland Stevenson et al. 1993

field +Phyto,
+Macro, - Zm Mass. Valiela et al. 1992

field - Epi, - Zm * Washington Williams & Ruckleshaus

1993

field +Macro, - Zm 4 Finland Bostrom et al. 2002

field +Macro, - Zm Mass. Hauxwell et al. 2003

mesocosm +Epi, -SAV Maryland Twilley et al. 1985

mesocosm

- Zm 1 North

Carolina Burkholder et al. 1992,

1994

mesocosm +algae, - Zm New Hampshire Short et al. 1995

mesocosm +Phyto, - Zm

, -Macro, -Epi Rhode Island Taylor et al. 1995

Lit. review

-- Worm et al. 2000

Lit. review

-- Nixon et al. 2001
*Suggested nutrient limitation of Z. marina

1 Suggested nitrate toxicity of Z. marina

2 Concluded light dominant factor, only +E pi and -Zm at highest light level.

3 Positive and negative effects were depe ndent on source of seagrass and salinity.

4 Conclusion based on inference.3.8 Figure 3.1. Schematic diagram of the effects of increased nutrient loading, which decreases seagrass density and biomass as well as promoting a shift toward phytoplankton, epiphyte or macroalgal community dominance. Adapted from Short et al. 1991.

This paradigm is based on several characteri stics of East Coast systems and consequently may not be applicable to Pacific Northwest estuar ies. Specifically, this paradigm applies to systems with relatively long residence times (mon ths to years) and to places where nutrient inputs are dominated by atmosphere and freshwat er. Most PNW estuaries exhibit very short residence times (1-30 days) as a result of large freshwater inputs and tida l 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 nut rient sources shift seasonally between river input during winter and ocean i nput during summer. The application of the nutrient loading/algal re sponse paradigm to the PNW requires further scientific investigation.3.9 The responses of lakes to nutrient loadi ng and eutrophication abatement have been examined in a number of studies. Eutrophication ab atement has also been an important research topic for the aquaculture industr y (e.g. Kraufvelin et al. 2001 ; Porrello et al. 2003a, b) since

effluent from mariculture activitie s is often high in nutrients. However, relatively few studies have examined the response of temperate estuar ies to reduced nutrient loading. Although not directly linked to seagrass, O = Shea and Brosnan (2000) conclude that despite reductions in municipal and industrial waste wa ter discharges to Long Island Sound, water quality, especially since bottom water hypoxia continues to be a probl em. Thus nutrient load ing and environmental problems may not exhibit linear responses, depending on the t ype 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 th e leaves and the roots and appears to be partially concentration depende nt (Thursby and Harlin 1982). Nitrogen uptake

rates from the literature are summarized in Ta ble 3.1. Ammonium is the preferred N source, since it is the reduced form (Touchette and Bu rkholder 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; Stap el and Hemminga 1997; Hemmi nga 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 ( 32 P, 14 C) and stable ( 15 N) 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 excret ion of phosphate is of minor importance (Brix and Lyngby 1985; Perez-Llorens et al. 1993). Am bient phosphorus concentrations in the environment are reported in Table 3.4.

Experimental manipulations of nutrient supply suggested that seagra sses can be nutrient

limited. Some of the first field enrichment e xperiments added fertilizer to the sediments and showed increased leaf length, biomass and s hoot 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 nutrien t conditions need to be consid ered 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 lim itation in tropical car bonate sediments.3.10Table 3.4. Water column and sediment nutrient concentrations (FM) associated with selected studies ofZosterahabitat.SeagrassLocationWaterSedimentReferenceNH4+NO3-/NO2=PNH4+NO3-/NO2=PZostera marinaAlaska20-1474-28Short 1983aNetherlands3-151-185-17Herman et al. 1996Rhode Island0.1-1.50.1-0.80-1Harlin & Thorne-Miller1981Denmark1-5.50.5-6250-1100Pedersen & Borum 1993Alaska10-1500Iizumi et al. 1982Mass.270-1550Dennison et al. 1987Alaska2-75-75Mc Roy et al. 1972North Sea1-140-950-1300-2Hemminga et al. 1994Alaska1-15Iizumi et al. 1980Zostera japonicaOregon2.2-2.97-350.6-1.160-1700.8-2.03.3-5.6Larned 2003Oregon2-52-260.5-2600-24000-307-40Kaldy 2006b.3.11 Research has clearly shown that seagrasses ar e affected by nutrients. Various processes, both natural and anthropogenic influe nce nutrient loading to estuaries. It is also apparent that nutrient loading proces ses are regionally variable and that this variability needs to be accounted for in the development of protect ive 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. ag riculture). 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 linka ge 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 nutri ent criteria may need to be tailored to account for regionally impo rtant processes (e.g. ocean currents).

Nitrogen supply is frequently described as either A new @ or A old @ nitrogen. New nitrogen is typically inorganic (e.g. N 2 , NO 3 - , NO 2 = , NO

x ) and is generally made available through processes such as upwelling and nitrogen fixati on. Old nitrogen is N made available through biological excretion and recycling and is typical ly composed of more reduced compounds such as amines, NH 4 + , urea, etc. Seagrasses prefer NH 4 + since it requires less energy for incorporation, although th ey can also utilize NO 3 - (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 indir ect effects. The shift in nutri ent 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 a nd 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 re ported in the lite rature (Burkholder et al. 1992, 1994; van Katwijk et al. 1997). For example, Har lin and Thorne-Miller (1981) performed field nutrient addition experiments in Rhode Island, adding ammonium, nitrate, or phosphate to the water column rather than to the sediment. The nitrate supplements, while not causing a noticeable change in gr owth of the above-ground plant, did inhibit the root-rhizome fraction of the seagrass. They sugge sted that this might indicate to xicity of the nitrate addition to the test plants. Consistent w ith this hypothesis was the obser vation 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 th at nitrate enrichment of the water column caused declines of eelgrass, especially at hi gher temperatures, and that this was a direct physiological effect independent of shading by macroa lgae. They attributed the effect to internal

imbalances in nutrient ratios from sustained nitr ate uptake through the leaf tissue. However, the evidence to support direct toxic e ffects is still considered somewh at tenuous and is a topic of debate in the literature (Moore a nd Wetzel, 2000). For example, Zostera marina thrives in at3.12 least one Oregon estuary (Yaquina Bay) wher e both ambient water column and sediment nitrogen (nitrate and ammonium) concentrations ar e 3 to 10 times higher than the levels used in any of the experiments that exhi bit toxic effects (Bro wn and Kaldy, US EPA, unpublished data). Touchette and Burkholder (2002) outline a physio logical mechanism to support the nitrate toxicity hypothesis. However, it is important to note that temperature stress may be a confounding factor leading to seag rass decline. Bintz et al. (2003) observed that the negative effect of elevated water temp erature on eelgrass was significan tly increased when inorganic

nutrient concentrations also were increased.

Experimental and observational evidence indi cate that increased organic loading (i.e. eutrophication) in estuarine ecosystems results in a shift toward algal dominated production and degradation of seagrass habitat, thus altering the organization of trophic levels and the flow of energy through the system (See Chapter 8).

3.4 Gaps in Knowledge

While much of the physiology of nutrient uptake in Z. marina has been examined in a crude manner, the details of allocation within the plant and the dominant sources under given conditions remain somewhat ambiguous. The issu e 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 tim es higher than reported toxic levels for long periods (days to months) as a re sult of coastal upwelling and rive rine loading with no apparent negative impact to Z. marina . Relatively cold water te mperatures (annual average 10 E C) may help to ameliorate stress. In comparison, areas th at 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 E C). As a result temperature and nut rient effects may be confounded and work synergistically. Detailed physiological work utilizing newly available technology (e.g. compound specific nitrogen isotope analyses, micr oelectrodes/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; howev er, few studies (if a ny) 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 domin ant primary producers (Nixon et al. 2001).

The absolute and relative contributions of different primary producers to net ecosystem primary production may provide an important metric for determining the degree of eutrophication and/or the potential for eutrophication related impairment of resources. In the PNW, the production ecology of benthic marine macrophytes (seagrasses and macroalgae) is poorly understood. The production ecology of microa lgae 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, sinc e 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 other3.13 types of habitat (e.g. marshes), as well as econom ically and recreationally important fisheries, is poorly known.

3.5 Literature Cited

Atkinson, M. J. and S. V. Smith. 1983. C:N:P ratios of benthic marine plants. Limnology and Oceanography 28:568-574. Bintz, J. C., S. W. Nixon, B. B. Buckley, and S. L. Granger. 2003. Impacts of temperature and nutrients on coastal lagoon plant co mmunities. Estuaries 26:765-776. 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 Shelf Science 55:795-804. Brix, H. and J. E. Lyngby 1985. Uptake a nd translocation of phosphorus in eelgrass ( Zostera

marina ). Marine Biology 90:111-116. Brown, C. A. and H. Lee II. 2006. Chapter 3: Environmental characte ristics of the target estuaries. In: Lee, II., H., C. A. Brown, B. L. Boese and D. A. Young (eds.). Proposed classification scheme for coastal receiving wa ters based on SAV and food web sensitivity

to nutrients. Volume 2. Nutrient driv ers, seagrass distri bution and regional

classifications of Pacific Northwest estuarie s. US Environmental Protection Agency. Internal Report. pp. 279. Brown, C. A and J. E. Kaldy, US EPA, unpublished data. Burdige, D. J. and R. C. Zimmerman. 2002. Im pact of seagrass density on carbonate dissolution in Bahamian sediments. Limnology and Oceanography 47:1751-1763. Burkholder, J. M., H. B. Glasgow, and J. E. C ooke. 1994. Comparative effects of water-column nitrate enrichment on eelgrass Zostera marina , shoal grass Halodule wrightii and widgeongrass Ruppia maritima . Marine Ecology Progr ess Series 105:121-138. Burkholder, J. M., K. M. Mason, and H. B. Glasgow. 1992. Water-column nitrate enrichment promotes decline of eelgrass Zostera marina : evidence from seasonal mesocosm experiments. Marine Ecology Progress Series 81:163-178. Butler, A. and P. Jernakoff. 1999. Seagrass in Australia: Strategic review and development of an R&D plan. CSIRO publishing, Collingwood, Australia. pp 210. Compton, J. E., M. R. Church, S. T. Larned, a nd W. E. Hogsett. 2003. Nitrogen export from forested watersheds in the Oregon Coast Range: The role of N 2 -fixing red alder. Ecosystems 6:773-785. Correll, D. L., T. E. Jordan, and D. E. Weller. 1992. Nutrient flux in a landscape: effects of coastal land used and terrestrial community mosaic on nutrient transport to coastal

waters. Estuaries 15:431-442. Dennison, W. C., R. C. Aller, and R. S. Albert e. 1987. Sediment ammonium availability and eelgrass ( Zostera marina ) growth. Marine Biology 94:469-477. Driscoll, C. T., D. Whitall, J. Aber, E. Boyer, M .Castro, C. Cronan, C. L. Goodale, P. Groffman, C. Hopkinson, K. Lambert, G .Lawrence, and S. Ollinger. 2003. Nitrogen pollution in the northeastern United States: Sources, e ffects and management options. BioScience

53:357-374.3.14 Duarte, C. M., M. Merino, and M. Gallegos. 1 995. Evidence for iron deficiency in seagrasses growing above carbonate sediments. Limnology and Oceanography 40:1153-1158. Duarte, C. M. 1990. Seagrass Nutrient Content. Marine Ecology Progress Series 67:201-207 Eftemeijer, P. L. A. and J. J. Middelburg. 1993. Sediment nutrient interactions in tropical seagrass beds: a comparison between a terrigenous and carbonate sedimentary environment in South Sulawesi (Indonesia) . Marine Ecology Progr ess Series 102: 187-

198. Fenn, M., L. Geiser, J. Peterson, E. Waddell, and E. Porter. 2004. Why federal land managers in the Northwest are concerned about Nitrogen emissions. http://nature.nps.gov/air/Pubs/pdf/NOxPaper2004.pdf Fourqurean, J. W., T. O. Moore, B. Fry, and J. T. Hollibaugh. 1997. Spatial and temporal variation in C:N:P ratios, δ 15 N and δ 13 C of eelgrass Zostera marina as indicators of ecosystem processes, Tomales Bay, Califor nia, USA. Marine Ecology Progress Series

157:147-157. Fourqurean, J. W. and Y. Cai. 2001. Arsenic an d phosphorus in seagrass leaves from the Gulf of Mexico. Aquatic Botany 71:247-258. Fourqurean, J. W., J. C. Zieman, and G. V. N. Powell. 1992. Phosphorus limitation of primary production in Florida Bay: Evidence from C:N:P ratios of the dominant seagrass

Thalassia testudinum . Limnology and Oceanography 37: 162-171. Fourqurean, J. W. and J. C. Zieman. 2002. Nutrient content of the seagrass Thalassia

testudinum reveals regional patterns of relative availability of nitrogen and phosphorus in the Florida Keys, USA. Biogeochemistry 61:229-245. Harlin, M. M. and B. Thorne-Miller. 1981. Nu trient enrichment of s eagrass beds in a Rhode Island coastal lagoon. Marine Biology 65:221-229. Hauxwell, J., J. Cebrian, and I. Valiela. 2003. Eelgrass Zostera marina loss in temperate estuaries: relationship to land-derived nitr ogen loads and effect of light limitation imposed by algae. Marine Ec ology Progress Series 247:59-73. Havens, K. E., J. Hauxwell, A. C. Tyler, S. Thom as, K. J. McGlathery, J. Cebrian, I. Valiela, A. D. Steinman, and S.-J. Hwang. 2001. Comp lex interactions between autotrophs in shallow marine and freshwater ecosystems: implications for community responses to

nutrient stress. Environmental Pollution 113:95-107. Hemminga, M. A., B. P. Koutstaal, J. van Soel en, and A. G. A. Merks. 1994. The nitrogen supply to intertidal eelgrass ( Zostera marina ). Marine Biology 118:223-227. Hemminga, M. A., N. Marba, and J. Stapel. 1999. Leaf nutrient resorpti on, leaf lifespan and the retention of nutrients in seagrass systems. Aquatic Botany 65:141-158. Hemminga, M. A. and C. M. Duarte. 2000. Seag rass Ecology. Cambridge University Press, Cambridge, UK. Herman, P. M. J., M. A. Hemminga, P. H. Nienhui s, J. M. Verschuure, and E. G. J. Wessel. 1996. Wax and wane of eelgrass Zostera marina and water column silicon levels. Marine Ecology Progress Series 144:303-307. 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 a nd reducing the effects of nutrient pollution. Ocean3.15 Studies Board and Water Science Tec hnology Board, National Academy Press, Washington, DC. Iizumi, H., A. Hattori and C. P. McRoy. 1980. Nitrate and nitrite in in terstitial waters of eelgrass beds in relation to the rhizosphere. Journal Expe rimental Mari ne Biology and Ecology 47:191-201. Iizumi, H., A. Hattori and C. P. McRoy. 1982. Ammonium regeneration and assimilation in eelgrass ( Zostera marina L.) beds. Marine Biology 66:59-65. Iizumi, H. and A. Hattori. 1982. Grow th and organic production of eelgrass ( Zostera marina L.) in temperate waters of the Pacific coast of Japan III. Kinetics of Nitrogen uptake. Aquatic Botany 12:245-256. 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 Oceanography 43:799-810. Jorgensen, B. B. and K. Richardson (eds). 1996. Eutrophication in Coastal Marine Ecosystems.

Coastal and Estuarine Studies Volume 52. American Geophysical Union, Washington,

DC. Kaldy, J. E. and K. H. Dunton. 2000. Above- and below-ground production, biomass and reproductive ecology of Thalassia testudinum (turtle grass) in a subtropical coastal lagoon. Marine Ecology Progress Series 193: 271-283. Kaldy, J. E. 2006a. Carbon, nitrogen, phosphorus and heavy metal budge ts: How large is the eelgrass (Zostera marina L.) sink in a temperate estuary? Marine Pollution Bulletin 52:

332-356. Kaldy, J. E. 2006b. Production ecology of th e non-indigenous seagrass, dwarf eelgrass ( Zostera japonica Ascher. & Graeb.), in a Pacific Northw est Estuary, USA. Hydrobiologia 553:

201-217. Kraufvelin P, B. Sinisalo, E. Leppakoski, J. Mattila, and E. Bonsdorff. 2001. Changes in zoobenthic community structure after pollu tion abatement from fish farms in the Archipelago Sea (N. Baltic Sea). Mari ne Environmental Research 51:229-245 Larkum, A. W. D., R. J. Orth, and C. M. Du arte. 2006. Seagrasses : Biology, Ecology and Conservation. Springer, Netherlands. pp. 649. Larned, S. T. 2003. Effects of the invasive, nonindigenous seagrass Zostera japonica on nutrient fluxes between the water column and benthos in a NW Pacific estuary. Marine Ecology Progress Series 254:69-80. Lee, K. S. and K. H. Dunton. 1999a. Influence of sediment nitrogen av ailability on carbon and nitrogen dynamics in the seagrass Thalassia testudinum . Marine Biology 134: 217-226. Lee, K. S. and K. H. Dunton. 1999b. Inor ganic nitrogen acquisi tion in the seagrass Thalassia

testudinum : Development of a whole-plant nitrogen budget. Limnology and Oceanography 44: 1204-1215. Livingston, R. J. 2001. Eutrophication processe s in coastal system. CRC Press, Boca Raton, Florida, USA. McClelland, J. W. and I. Valiela. 1998. Linking nitrogen in estuarine producers to land-derived sources. Limnology and Oceanography 43:577-585. McGlathery, K. J. 2001. Macroalgal blooms c ontribute to the decline of seagrass in nutrient

enriched coastal waters. J ournal of Phycology 37:453-456.3.16 McRoy, C. P. and R. J. Barsdate. 1970. Phosphate absorption in eelgrass. Limnology and Oceanography 15:6-13. McRoy, C. P., R. J. Barsdate, and M. Nebert . 1972. Phosphorus cycling in an eelgrass ( Zostera

marina L.) ecosystem. Limnology and Oceanography 17:58-67. Moore, K. A. and R. L. Wetzel. 200 0. Seasonal varia tions in eelgrass ( Zostera marina L.) responses to nutrient enrichment and redu ced light availability in experimental ecosystems. Journal Experimental Marine Biology Ecology 244:1-28. Neckles, H. A., R. L. Wetzel, and R. J. Orth. 1993. Relative effects of nutrient enrichment and grazing on epiphyte-macrophyte ( Zostera marina L.) dynamics. Oecologia 93:285-295. Nelson, T.A., A. Lee. 2001. A manipulative experiment demonstrates that blooms of the macroalga Ulvaria obscura can reduce eelgrass shoot de nsity. Aquatic Botany 79:149-

154. Nixon, S. W. 1995. Coastal marine eutrophicat ion: a definition, social causes and future concerns. Ophelia 41:199-220. Nixon, S., B. Buckley, S. Granger, and J. Bint z. 2001. Responses of very shallow marine ecosystems to nutrient enrichment. Human and Ecological Risk Assessment 7:1457-

1481.

O = Shea, M. L. and T. M. Brosnan. 2000. Tre nds in indicators of eu trophication in Western Long Island Sound and the Hudson-Rarita n Estuary. Estuaries 23:877-901. Orth, R. J. 1977. Effect of nutrient enrichment on growth of the eelgrass, Zostera marina in the Chesapeake Bay, Virginia, USA. Marine Biology 44:187-194. Pedersen, M. F. and J. Borum. 1992. Nitrogen dynamics of eelgrass Zostera marina during a late summer period of high growth and lo w nutrient availability. Marine Ecology Progress series 80:65-73. Pedersen, M. F. and J. Borum. 1993. An annual nitrogen budget for a seagrass Zostera marina population. Marine Ecology Progress Series 101:169-177. Penhale, P. A. and G. W. Thayer 1980. Uptake and transfer of carbon and phosphorus by eelgrass ( Zostera marina L.) and it epiphytes. Journal E xperimental Marine Biology and Ecology 42:113-123. Perez-Llorens, J. L., P. de Visscher, P. H. Nie nhuis, and F. X. Niell. 1993. Light dependent uptake, translocation and foliar release of phosphorus by the intertidal seagrass Zostera

noltii Hornem. Journal of Experiment al Marine Biology Ecology 166:165-174. Perez-Llorens, J. L, M. Muchtar, F. X. Niell, and P. H. Nienhuis. 1991. Particulate organic carbon, nitrogen and phosphorus content in roots, rhizomes and differen tly aged leaves of

Zostera noltii Hornem. in Oosterschelde Estu ary (Southwestern Netherlands). Botanica

Marina 34:319-322. Porrello, S., M. Lenzi, E. Persia, P. Tomass etti, and M. G. Finoia. 2003a. Reduction of aquaculture wastewater eutroph ication by phytotreatment ponds system I. Dissolved and particulate nitrogen and phosphor us. Aquaculture 219:515-529. Porrello, S., M. Lenzi, E. Persia, P. Tomasset ti, M. G. Finoia, and I. Mercatali. 2003b. Reduction of aquaculture wastewater eutroph ication by phytotreatment ponds system II. Nitrogen and phosphorus content in macroa lgae and sediment. Aquaculture 219:531-

544.3.17 Richardson, K. and B. B. Jorgensen. 1996. Eutr ophication: Definition, History and Effects. In: BB Jorgensen and K Richardson (eds). Eutroph ication in Coastal Marine Ecosystems.

Coastal and Estuarine Studies Volume 52. American Geophysical Union, Washington,

DC. Short, F. T. 1983a. The response of interstitial ammoni um in eelgrass ( Zostera marina L.) beds to environmental perturbations. Journa l of Experimental Marine Biology Ecology

68:195-208. Short, F. T. 1983b. The seagrass Zostera marina L.: Plant morphology and bed structure in

relation to sediment ammonium in Izembe k Lagoon, Alaska. Aquatic Botany 16:149-

161. Short, F. T. 1987. Effects of sediment nutrien ts on seagrasses: literature review and mesocosm experiment. Aquatic Botany 27:41-57. Short, F. T. and D. M. Burdick. 1996. Quan tifying eelgrass habitat loss in relation to housing development and nitrogen loading in Waquo it Bay, Massachusetts. Estuaries 19:730-

739. Short, F. T., D. M. Burdick, and J. E. Kaldy. 1995. Mesocosm experiments quantify the effects of eutrophication on eelgrass, Zostera marina . Limnology and Oceanography 40:740-

749. Short, F. T. and R. G. Coles. 2001. Global S eagrass Research Methods. Elsevier, Amsterdam. Short, F. T. and C. P. McRoy. 1984. Nitrogen uptake by leaves and roots of the seagrass

Zostera marina L. Botanica Marina 27:547-555. Short, F. T., W. C. Dennison, and D. G. Capone. 1990. Phosphorus-limited growth of the tropical seagrass Syringodium filiforme in carbonate sediments. Marine Ecology Progress Series 62:169-174. Short, F. T., G. E. Jones, and D. M. Burdick. 1991. Seagrass decline: Problems and solutions.

Coastal Zone = 91 Conference

B ASCE. Long Beach, CA/July 1991. Pp 439-453. Stapel, J., T. L. Aarts, B. H. M. van Duynhoven, J. D. deGroot, H. H. W. van den Hoogen, and M. A. Hemminga. 1996. Nu trient uptake by leaves and roots of the seagrass Thalassia hemprichii in the Spermonde Archipelago, Indonesia. Marine Ecology Progress Series

134:195-206. Stapel, J. and M. A. Hemminga. 1997. Nutrie nt resorption from seagrass leaves. Marine Biology 128:197-206. Stevenson, J. C., L. W. Staver, and K. W. Staver. 1993. Water qua lity associated with survival of submersed aquatic vegetation along an estuarine gradient. Estuaries 16:346-361. Taylor, D. I., S. W. Nixon, S. L. Granger, B. A. Buckley, J. P. McMahon, and H.-J. Lin. 1995.Responses of coastal lagoon plant communities to different forms of nutrient enrichmenta mesocosm experiment. Aquatic Botany 52:19-34. Thurman, H. V. 1988. Introductory Oceanograp hy. Merrill Publishing Company, Columbus,

Ohio. Thursby, G. B. and M. M. Harlin. 1982. Leaf -Root interaction in the uptake of ammonium by

Zostera marina . Marine Biology 72:109-112. Touchette, B. W. and J. M. Burkholder. 2000a. Review of nitrogen and phosphorus metabolism in seagrasses. Journal Experiment al Marine Biology Ecology 250:133-167.3.18 Touchette, B. W. and J. M. Burkholder. 2000b. Overview of the physiological ecology of carbon metabolism in seagrasses. Journal Experimental Marine Biology Ecology 250:

169-205. Touchette, B. W. and J. M. Burkholder. 2002. Seasonal variations in carbon and nitrogen constituents in eelgrass ( Zostera marina L.) as influenced by increased temperature and

water-column nitrate. Botanica Marina 45:23-34. 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. 1. Algal growth and effects of production of plan ts and associated communities. Marine Ecology Progress

Series 23:179-191. Valiela, I. and J. L .Bowen. 2002. Nitrogen sour ces to watersheds and es tuaries: role of land

cover mosaics and losses within waters heds. Environmental Pollution 118:239-248. Valiela, I., K. Foreman, M. LaMontagne, D. Hersh, J. Costa, P. Peckol, B. DeMeo-Anderson, C.

D = Avanzo, M. Babione, C.-H. Sham, J. Brawley, and K. Lajtha. 1992. Couplings of

watersheds and coastal waters: sources and consequences of nutri ent enrichments in Waquoit Bay, Massachusetts. Estuaries 15:443-457. van Katwijk, M. M., L. H. T. Vergeer, G. H. W. Schmitz, and J. G. M. Roelofs. 1997. Ammonium toxicity in eelgrass Zostera marina . Marine Ecology Progress Series

157:159-173. van Katwijk, M. M., G. H. W. Schmitz, A. P. Gasse ling, 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. van Lent, F., J. M. Vershuure, and M. L. J. van Veghel. 1995. Comparative study on populations of Zostera marina L. (eelgrass): in situ nitrogen enrichment and light manipulation. Journal Experimental Marine Biology Ecology 185:55-76. Vollenweider, R. A., R. Marchett i, and R. Viviani. 1992. Marine Coastal Eutrophication. Elsevier, Amsterdam. Williams, S. L. and M. H. Ruckelshaus. 1993. Effects of nitrogen availability and herbivory on eelgrass ( Zostera marina ) and epiphytes. Ecology 74:904-918. Worm, B., T. B. H. Reusch, and H. K. Lotze. 2000. In situ nutrient enrichment: methods for marine benthic ecology. International Review of Hydrobiologia 85:359-375. Zimmerman, R. C., R. D. Smith, and R. S. Albe rte. 1987. Is growth of eelgrass nitrogen

limited? A numerical simulation of the eff ects of light and nitr ogen on the growth dynamics of Zostera marina . Marine Ecology Progress Series 41:167-176.4.1 

## 4.0 Salinity as a Limiting Factor for the Seagrasses Zostera marina

and Thalassia testudinum

Bruce L. Boese

4.1 Background

Eelgrass, Zostera marina is a euryhaline species (Phill ips 1984) found world wide in coastal waters that vary from m ildly hypersaline to nearly fresh (T able 4.1). However, it appears to grow best in estuarine waters with salinities in the range from approxi mately 5 to just below

that of normal seawater. In cont rast, 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 tolera te extreme salinity fluxes nor pr olonged exposures to fresh water (Doering and Chamberlain 2000; Montague and Ley 1993).

Alterations in estuarine hydrology due to fr esh 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 ar eas 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 implicat ed in the historic wasting

disease die-offs that occurred in eastern and we stern Atlantic eelgrass pop ulations in the early 1930’s (den Hartog 1987; Giesen et al . 1990). Geographically isolated Z. marina populations may have genetically narrower salinity optima th an 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 complic ate 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 Diggele n et al. 1987). Although Z. marina is able to survive for a time in fresh water, net leaf photosynthesi s decreases in waters below 5 and totally ce ases in completely fresh water (Hellbom and Björk 1999; Biebl an d 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 report ed as 25 to 38.5 (Phillips 1960), and 17-36 Zimmermann and Livingston 1976).4.2 Table 4.1. Reported salinity tolerance ranges for eelgrass ( Zostera marina ) populations.

Location Salinity Range Citation

World Wide 0-35 Thayer et al. 1984

World Wide 0-42 Phillips 1984 Northern Hemisphere 5-35 den Hartog 1970

Denmark 13-31 Pinnerup 1980

Denmark 12-31 Sand-Jensen and Borum 1983 Denmark (estuaries) 9-23 Wium-Anderson and Borum 1984

Baltic Sea 6-12 Hellblom and Björk 1999 Netherlands (marine environments) ~30 van Katwijk et al. 1999 Netherlands (estuaries) 15-25 van Katwijk et al. 1999 Netherlands (Lake Grevelinge n) 22-32 Kamermans et al. 1999 France (Thau) 27-41 Rigollet et al. 1998 Italy (Venice) 25-33 Rigollet et al. 1998 Chesapeake Bay 14-22 Wetzel and Penhale 1983 Yaquina Bay, OR 25-33 Kentula and DeWitt 2003

Table 4.2. Reported salinity to lerance ranges for turtlegrass ( Thalassia testudinum ) populations.

Location Salinity Range Citation

Florida 28-36 Phillips 1960

Florida 17-36 Zimmermann and Livingston 1976*

Florida 24-35 Zieman and Zieman 1989*

Zieman 1982*

Texas 30-40 Adair et al. 1994*

Florida 22-36 Doering and Chamberlain 2000 Dry Tortugas 35-38.5 Phillips 1960 Everglades National Park 28-48 Phillips 1960 Florida (west coast) 25-34 Phillips 1960

*As cited in Doering and Chamberlain 2000.4.3 4.3 Growth/Shoot/Competition Effects

While Z. marina is tolerant of some exposure to fresh and hypersaline waters, optimal growth rates vary with populations. Zostera marina population increases and declines were correlated with relatively slight alteration in th e salinity structure of Lake Grevelingen in The Netherlands, within the range of 22 to 32 (Kam ermans et al. 1999). Laboratory experiments supported this interpretation show ing optimal growth rates were higher at 22 than 32. However

, Z. marina collected from an area of relatively great er salinity (Roscoff and Bay of Archchon, France) did not exhibit differences in growth ra tes when subjected to th e same salinity regimes (Kamermans et al. 1999). Optimal sa linities for Pacifi c Northwest (PNW) Z. marina populations have not been determined, but Z. marina appears to grow best at 20-32 in Puget Sound (Philips 1984). Laboratory experiments on PNW Z. marina suggested that higher growth rates occurred at 30 than 10 and that the highest densities of Z. marina in the field were in areas of high salinities and lowest temperature (Thom et al. 2001, 2003). This apparent relationship between plant densities (as estimated from percent c over) and high salinities combined with low temperature was also evident in the Yaquina es tuary, 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 le ss 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) exposur e 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 la boratory. 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 (s outhwest 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 an d moderate hypersalinity events in Laguna Madre (southeast Texas). This resulted in a decrease in H. wrightii with its replacement by S. filiforme and T. testud inum (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) a nd field studies which exam ined the responses of a

T. testudinum meadow to brine discharges from a desa lination plant show no apparent adverse effects (Tomasko et al. 2000). The most likely cause of this die off was an interaction between4.4 high salinities, high temperatur es (Florida Bay Science Plan 1994). High temperatures and salinities also may affect photos ynthetic oxygen production, resulti ng 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 (He rrera-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 result ing 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 Mediterra nean lagoon, began to degrad e after the completion of a hydroelectric dam in 1996 which erratically di verts 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 co mmunity (Stora and Arnoux 1983). Montegue and Ley (1993) noted that overall s eagrass biomass in Florida Bay decreased with increasing variation in salinity, and that these de creases were greatest wher e salinity was lowest. Their work suggests that even when mean salinit y changes of water diversion projects are small, a small increase in salinity vari ance may have drastic effects on seagrass populations. Similarly

fresh water diversions which re duce salinities below the optimal range may adversely affect T.

testudinum populations, especially recruitment from seed lings. In mesocosm experiments, Kahn and Durako (2006) found that while turtle grass seed lings tolerated a reduc tion 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 detrimenta l to recruit survival (Kahn and Duranko 2006).

In Z. marina there appears to be a significant in teraction between tolerance to higher salinity and nutrients. Laboratory experiment s 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 adversel y 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 fr om which these populations originated, suggesting genotypic differences in salinity tolerances (K am ermans et al. 1999; va n Katwijk et al. 1999, 2000). Genotypic differences in eel grass populations have also been implicated in the apparent geographic differences in l eaf widths (McMillan 1978).4.5 4.4 Wasting Disease

The slime mold ( Labyrinthula sp.) is a secondary decompos er 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 (Mar tin 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 ar e high (Blakesly et al. 2002). However, it is likely not the major cause of the seagrass die- off noted in Florida Ba y (Boesch et al. 1993; Zieman et al. 1999; Blakesly et al. 2002). However, as Labyrinthula prefers more saline waters, fresh water diversions and po ssible droughts resulting from globa l climate change will have a positive influence on its growth thus making it mo re likely to adversely affect seagrasses.

4.5 Summary/Research Gaps

Generally both turtle and eelgrass have wide sa linity tolerances with turtle grass being the more stenohaline species. In ar eas of suboptimal or highly vari able salinities both species are often precluded by complex interactions with other stressors such as high temperatures, suboptimal lighting conditions, wasting disease, a nd 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 regulator y 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 seagrass es 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 th at address these research gaps should be conducted on a variety of seagrass genets to assess effects on localized populations.4.6 4.6 Literature Cited

Adair, S. E., J. L. Moore, and C. P. Onuf. 1994. Distribution and status of submerged vegetation in estuaries of the Upper Te xas Coast. Wetlands 14:110-121. Biebl, R. and C. P. McRoy. 1971. Plasma tic resistance and rate of respiration and photosynthesis in the seagrass Zostera marina at different saliniti es 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 Ecos ystem: An evaluation of the scientific evidence. Report to the Interagency Work ing Group on Florida Bay. Sept. 15, 1993. pp.30. denHartog, C. 1970. The Seagrasses of the Wo rld. North-Holland Publication Amsterdam. pp.

275. denHartog, C. 1987. 'Wasting disease' and ot her dynamic phenomena in Zostera beds. Aquatic Botany 27:3-14. Doering P. H. and R. H. Chamberlain. 2000. Experimental studies on th e salinity tolerance of turtle grass, Thalassia testudinum . pp. 81-98 In: S. A. Bortone (ed)., Seagrasses: Monitoring, Ecology, Physiology and Manage ment. CRC Press, Boca Raton. pp. 318. Estevez, E. D. 2000. Matching Salinity Metric s to Estuarine Seagrasses for Freshwater Inflow Management. pp. 295-307. In: S. A. Bort one (ed)., Seagrasses: Monitoring, Ecology, Physiology and Management. CRC Press, Boca Raton. pp. 318. Florida Bay Science Plan. 1994. Flor ida 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 relati on to salinity fluctuations in Charlotte Harbor, Florida. Florida Scientist 69: 24-35. Hellblom, F. and M. Björk. 1999. Photosynthetic responses in Zostera marina to decreasing salinity, inorganic carbon content and osmolality. Aquatic Botany 65:97-104. Herrera-Silveira J. A., J. Ramír ez-Ramírez, N.Gómez, and A. Zaldivar-Jimenez. 2000. Seagrass bed recovery after hydrologica l restoration in a coasta l lagoon with groundwater discharges in the north of Yu catan (southeastern Mexico). pp 219-229. In: S. A. Bortone (ed)., Seagrasses: Monitoring, Ecology, Physiology and Management. CRC Press, Boca Raton. pp. 318 Jagels, R. 1983. Further evidence for osmoregulation in epidermal leaf cells of seagrasses. American Journal of Botany 70:327-333.4.7 Kahn, A. E. and M. J. Durako. 2006. Thalassia testudinum seedling response to changes in salinity and nitrogen levels. Journal of Experimental Ma rine 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-temper ature gradient in Ya quina Bay, Oregon,

USA. Estuaries 48:1130-1141. Koch, M. S. and J. M. Erskine. 2001. Su lfide as a phytotoxin in the tropical seagrass Thalassia

testudinum : interations with light, salinity and te mperature. Journal of Experimental Marine Biology and Ecology 266:81-95. Koch, M. S., S. A. Schopmeyer, C. Kyhn-Hansen, C. J. Madden, a nd J.S. Peters. 2007. Tropical seagrass species tolerance to hypersal inity stress. Aquatic Botany 86:14-24.. Lirman, D. and W. P. Cropper Jr. 2003. Th e influence of salinity on seagrass growth, survivorship, and distribution within Biscayne Bay, Florid a: Field, experimental, and modeling studies. Estuaries 26:131-141. Martin, A. C. 1954. A clue to the eelgrass myst ery. Transactions of the 19th North American Wildlife Conference, Washington, D.C 441-449. McMillan, C. 1978. Morphogeograhic variation unde r controlled conditions in five seagrasses, Thalassia testudinum, Halodule wrightii, Syringodium filiforme, Halophila engelmannii

, and Zostera marina . Aquatic Botany 4:169-189. Montagure, C. L. and J. A. Ley. 1993. A possi ble effect of salinity fl uctuation on abundance of benthic vegetation and associated fauna in nor theastern Florida Bay. Estuaries 16:703-

717. Phillips R. C. 1960. Observations on the ecolo gy and distribution of the Florida seagrasses. Professional papers series, No 2, Cont ribution No. 44, Florida Sate Board of

Conservation Marine Laborator y, 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. La guan Madre: Seagrass changes continue decades after salinity reduction. Estuaries 2:302-310. Rigollet, V., T. Laugier, M. L. De Casabianca, A. Sfriso, and A. Marcomini. 1998. Seasonal biomass and nutrient dynamics of Zostera marina L. in two Medite rranean 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 la rge freshwater diversi ons 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 meadowsof the Atlantic coast: A community profile. U.S. Fish and Wildlife Service. FWS/OBS84/02. 147 pp4.8 Thom R. M., A. B. Borde, S. Rumrill, D. L. Wo odruff, G. D. Williams, J. Southard, and S. L. Blanton. 2001. Factors influencing the spatial and annual vari ability in eelgrass ( Zostera

marina L.) in Pacific Northwest, USA, Sy stem (Chapter 5). PINCERS 2001 Annual Report. Thom R. M., A. B. Borde, S. Rumrill, D. L. Wood ruff, G. D. Williams, J. A. Southard, and S. L. Sargeant. 2003. Factors influencing sp atial 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, Physiol ogy and Management. CRC Press, Boca Raton. pp. 318. van Diggelen, J., J. Rozema, and R. Broekman . 1987. Mineral compos ition 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. Gasse ling, 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 Wa dden Sea for restoration of Zostera marina beds. Helgoland

Marine Research 54:117-128. Vergeer L. H. T., T. L. Aarts, and J. D. de Gr oot. 1995. The ‘wasting dis ease’ and the effect of abiotic factors (light intensity, temper ature, 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. Biom ass variations and auto trophic production of an epiphyte-macrophyte community in a coas tal 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 co mmunity profile. U.S. Fish and Wildlife Services, Office of Biol ogical Services, Washington, D.C., FWS/OBS82/25, pp. 158. Zieman, J. C., J. W. Fourqurean, and T. A. Fra nkovich. 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 a nd Wildlife Service Biological Report 85(7.25), 155 pp. Zimmerman, M. S. and R. J. Livingston. 1976. Seasonality and physico-chemical ranges of benthic macrophytes from a North Florida estu ary (Apalachee Bay). Contributions in

Marine Science 20:33-45.5.1 

## 5.0 The Effects of Hydrodyn amic 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 presen ce 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 dur ing low tides (Powell and Schaffner 1991). Water motions have the potential to affect the growth, survival, and di stribution of seagrasses through both direct and indirect mechanisms.

This chapter reviews the effect of hydrodyna mic 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-i nduced flows are oscillatory. Turbulence is temporally and spatially irregular water motions that are superimposed over the larger flow pattern, such as unidirectional cu rrent or oscillatory wave actio n (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. H ydrodynamic stressors can influence seagrass through numerous mechanisms, including direct damage or uprooting of plants due to the hydrodynamic forces, erosion of sediment surrounding the pl ant, and by limiting productivity through either shading or diffusional boundary layer limitation. In addition, water motions influence seagrass populations through dispersal of pollen (Ack erman 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 ecosyst ems, such as freshwater diversions, channelization, and damming, have been identifie d as one of the major environmental issues of coastal regions (National Research Council 1994). Hydrological and hydrodynamic alterations influence salinity patterns, tidal dynamics, circula tion 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 pr oduces 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 th e interaction of hydrodynamics and seagrass focuses on the effect of seagrass on the hydrodyna mics with relatively few studi es of the impact of the5.2 hydrodynamics on seagrass productivity and distribution. In addition, most of the studies on the effects on hydrodynamic stressors on seagrass are correlative rather than causative.

5.2 Effect of Seagrass on Currents, Waves, and Turbulence

The effect of seagrass on currents, waves and turbulence is complicated and depends on the hydrodynamic regime as well as specifics of the seagrass habitat (e.g., seagrass type and morphology, orientation of leaves and canopy, location within the seagrass bed and depth, shoot density, presence of epiphytes and macroalgae). Th ere have been field, laboratory, and modeling studies on the interaction of seag rasses and currents, waves and tu rbulent 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 occurri ng 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 re sult 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 availa bility for seagrass phot osynthesis (Zimmerman 2003) as well as the transport of particulate a nd dissolved materials between the water column and the canopy.

The water flow inside seagrass beds is ofte n reduced by a factor of 2 to 10 compared to adjacent bare areas (e.g., Scoffin 1970; Wor cester 1995; Komatsu 1996; Koch 1996; Koch et al. 2002; Peterson et al. 2004). Numerous field a nd 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; Fons eca and Fisher 1986; Gambi et al. 1990; Ackerman and Okubo 1993; Komatsu 1996; Nepf a nd Vivoni 2000; van Keulen and Borowitzka 2000; Peterson et al. 2004); howev er, 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 Borow itzka 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 rela tive 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 wa ter in seagrass beds. Dense seagrass beds in shallow regions have been shown to retain a th in layer (< 20 cm) of water during falling tides and this water trapping ability appears to be related to the morphol ogy of the seagrass and seagrass density (Powell and Schaffner 1991). Th is water trapping ability may be important in5.3 preventing desiccation (see Chapter 10) of the seagrass and other organisms that live in the

habitat.

Generally, water flow is reduced with increasing seagrass density and a recent study found that the vertically integrat ed 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 fr om 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 upo n 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 re ducing 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 coe fficient decreases. Ther e is often accelerated flow above the seagrass canopy (Gambi et al. 1990; Worcester 1995) which, when combined with the reduced flow within the canopy, result s 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 th e sheath region of the canopy (Fonseca and Kenworthy 1987; Ackerman and Okubo 1993; Koch 1996, Nepf and Vivoni 2000). This near bottom increase in fl ow speed may result in elevated sediment resuspension (Koch 1999b). Some have suggest ed that there is acceleration around seagrass beds due to flow being deflected above and arou nd the seagrass bed (Gambi et al.1990). Granata et al. (2001) found that the inte raction of hydrodynamics and seagra ss distribution (i.e., spatial variation in seagrass density, and presence of edges and gaps) results in three-dimensional circulation patterns, including incr ease in current above meadows, upward flow at the edge of the canopy, and recirculation patterns at gaps.

The interaction of current flow and the eel grass blades can result in large-amplitude synchronous waving of the blades (Fonseca a nd Kenworthy 1987; Acke rman and Okubo 1993; Grizzle et al. 1996), which has been termed “m onami” 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 curr ent velocities of about 30 cm s -1 (Grizzle et al. 1996). The

turbulent vertical transf er of momentum is enhanced during monami resulting in more vertical exchange between overlying water column and in terior of the canopy (G hisalberti and Nepf 2002). These coherent eddies ha ve implications for scalar fl uxes 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; Ghisalbe rti and Nepf 2002). Nepf and Koch (1999) demonstrated that submerged plant-like arrays exposed to gradients in longitudinal velocity in the laboratory produced ve rtical pressure gradients that drove vertical5.4 secondary flows. These vertical secondary fl ows can reach up to 15% of local longitudinal velocity and may affect the exchange of nutri ents between the sediment and water column. However, these vertical velo cities 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 reducti on of wave energy (which is proportional to wave height squared) in seagrass habitat (F onseca 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). Ge nerally, near bed orbital wave velocities are lower in seagrass beds and decrease with increasing plant density (Granata et al. 2001). Fonseca and Cahalan (1992) used a wave tank to examine the effect of H. wrightii, S. filiforme, T. testudinum, and Z. marina on wave energy under variou s combinations of shoot density and water depth to leaf le ngth 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 (rel ative to blade length), wave attenuation was reduced. Negligible wave energy reduction occurr ed 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 thr ee 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 va ried with tidal stage with ma ximum observed attenuations of 50% observed during low tides. Wave attenua tion was only observed at shoot density > 1000

shoots m -2 , and highest wave attenuation was observe d when the plants were reproductive, occupying the entire water column.

The interaction of seagrass with hydrodynami cs can have various effects on turbulence and turbulent mixing between the water column a nd 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 th e top of the seagrass can opy (Gambi et al. 1990; Nepf and Vivoni 2000), which is gene rated in part by the large shear stress in this region. There is a vertical reduction in turbulence with dept h inside the canopy (Koch and Gust 1999). Some

studies have documented a reduction of tur bulence (Koch 1996; Granata et al. 2001) and turbulent mixing within seagrass canopies (A ckerman 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 turb ulent flow with mean turbulent intensity and amount of the water column influenced by the presence of the canopy in creasing 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 a nd the configuration of the seagrass bed (e.g.,5.5 continuous versus patchy distribution, locati on within the meadow and canopy, presence of epiphytes, vegetation density). Worcester (1995 ) found that at low seagrass density and low current flow, the presence of eelgrass had no e ffect on turbulent mixing compared to adjacent bare areas, while at sites with continuous eelgr ass cover there was an increase in turbulent mixing above the seagrass canopy relative to adjacen t bare areas. Granat a 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 ener gy and generating high frequency energy (Koch 1996). Presence of epiphytes has been found to re sult 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 blad es producing a skimming flow over the seagrass canopy. This closure of the canopy results in re duction of turbulence inside the canopy and reduced mixing between th e overlying water column and the seagrass blades. In contrast, at wave-dominated sites os cillatory wave action causes seagrass blades to flap back and forth, and the canopy is repeat edly 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 dist ribution of a constituent in the water column determines whether the canopy will enhance or re duce the transport of the constituent. For example, a constituent with a vertical profile w ith maximum concentrations at the surface would have transport enhanced by 20%, whereas a cons tituent with maximum concentrations at the

bottom would have trans port reduced by 30%.

5.3 Effects of Water Velocity on Seagrass Growth and Distribution

The reduction of current velocities by seag rasses has positive and negative effects on their growth. Advantages of reduced flow include reduced se lf-shading, reduced sediment resuspension, increased settlement of organic an d inorganic particles, and high water residence time increasing potential for nutri ent uptake (Koch 2001). Detrimen tal effects of reduced water velocity include increased phytotoxin concentrati ons in the sediment and an increase in the thickness of the diffusional boundary layer, which may limit photosynthesis (Koch 2001; Koch et al. 2002).

5.3.1. Direct Damage to Seagrass resulting from Currents Scoffin (1970) conducted flume experiments to examine the effect of unidirectional

currents on erosion of T. testudinum . At current velocities of about 70 cm s -1 flapping of seagrass blades occurred, dislodging attached epiphytes and sometimes causing breakage of the

blades. Sediment removal around the base of the shoots was dependent upon current speed and density of the seagrass bed. Extensive sediment removal around the rhizomes and roots occurred at current speed of 50 cm s -1 in a sparse grass bed, 100 cm s -1 in a medium density bed, and at

150 cm s -1 in a dense grass bed (current velocity meas ured just above the blades). Fonseca et al. (1983) proposed that the maxi mum current velocity that Z. marina can tolerate is 120 to 150 cm

s -1

.5.6 To determine the susceptibility of seagra sses to hydrodynamic stressors, one can measure the biomechanical properties of seagrass, such as the breaking stress, breaking strain, elastic modulus, and toughness. Comparison of thes e properties to the hydrodynamic forces

encountered in different environm ents allows prediction of the pr obability 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 natura l populations of Z. marina there are always a few st rong reproductive shoots that would be resistant to extreme hydrodynamic even ts (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 unidi rectional 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 obs erved storm conditions, and that damage was more likely when the seagrass blades were subjected to oscillatory versus unidi rectional flow.

Previous studies have demons trated that the morphology of eelgrass leaves is dependent upon the nutrient availabili ty (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 we eks. 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 Morphol ogy 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 rela ted to current speeds (Figure 5.1). There was increased mounding of the substratum with increasing current velocity. In regions of unidirectional flow, seagrass is ofte n observed growing in rows perpe ndicular 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 e nvironment. 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 Pe ralta 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 cr oss-sections of the rhizomes and leaves increased, and the ratio of above to be low ground biomass decreased. Similar changes in morphometry (decreased leaf width and length) were observed in natural populations of Z. noltii5.7 exposed to more energetic environment associated with the opening of a new tidal inlet (Peralta et al., 2005).

5.3.3 Effects of Current on Photosynthesis and Growth Some studies have found that seagrass photos ynthesis and growth is related to current speed. Conover (1964 and 1968) found that oxyge n evolution from i ndividual shoots of Z.

marina increased as current velocity increase d 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 maximu m 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 in creasing current veloc ity (up to 34 cm s -1 ). Peralta et al. (2006) found a similar in crease 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 in creasing 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 re sulted in temporal and spatial variability in the boundary layer thickness, th ereby reducing the poten tial for boundary layer

limitation. Field measurements of blade fric tion 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 re latively quiescent conditions, the boundary layer thickness oscillates between non-limiting and limiting at high frequency, suggesting that diffusional boundary layer limitation is a transi ent 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 thic kness, then water will flow over the epiphyte cover rather than through it, incr easing the thickness of the boundary layer and result in further limiting conditions (Koch 1994). Enríquez a nd Rodrígues-Román (2006) found that the photosynthetic electron transport rates of T. testudinum was reduced under low flow conditions (< 5.4 cm s -1 ); however, this seagrass also appears to be able acclimate to flow conditions reducing their sensitivity to low flow by about 64%.

Zimmerman (2003) developed a bio-optical model of irradiance distribution and photosynthesis in Z. marina and T. testudinum canopies. This model incorporated the effects of canopy architecture on light availability fo r 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 an gles greater than 20 ° limit photosynthesis becaus e a larger fraction of light is absorbed by the upper la yers of the seagrass canopy where photosynthesis is already light saturated. As discussed previous ly, bending angle is a function of current velocity therefore selfshading increases with increas ing current. Zimmerman (2003) proposed that the productionenhancing aspects of flow (e.g., reduction in diffusive boundary layer thickness) may be offset by increased self-shading as leaves bend in response to the flow.5.8 At high flow, there is increased mounding of the substratum. (Fonseca et al.,1983) Low Medium High

There is a reduction in seagrass cover under high flow conditions (Fonseca etal., 1983) and when subjected to unidirectional flow seagrass grow in rows (Fonseca et al, 2007). Black and white areas indicate seagrass and bare regions, respectively Morphometry of the plant is dependent upon flow conditions. Under high flow
conditions, the leaves are

shorter and wider and the

rhizomes increased. (Peralta et al., 2006) Cross-Sectional View Aerial View Individual

Figure 5.1. Relationship between configurati on of seagrass beds and shoot morphology and current regime. At higher current velocities, there is increased m ounding 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 incr ease (Peralta et al., 2006).

5.3.4 Effect of Currents on Nutrient Uptake Thomas et al. (2000) conducted flume experiments in natural communities of Halodule wrightii and T. testudinum over a range of uni-directional cu rrent 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 velo cities 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 be tween the epiphytes and flow than with seagrass leaves and flow. Seagrass leaves contributed less than epiphytes to the total uptake of the seagrass community.

5.3.5 Effect of Currents on Sediment Geochemistry and Seedling Survival Koch (1999a) found that the interaction of currents and porewater chemistry may influence the growth and development of T. testudinum seedlings. In this study, seedling mortality was higher for stagnant conditions compared to medium and high current velocities.5.9 Seedling morphology was influenced by current ve locity with those s eedlings exposed to stagnant conditions having reduced biomass with shorter blades and roots, while those exposed to medium velocities had the longe st blades, largest blade area, a nd greatest biomass and number of roots. In addition, nitroge n, phosphorous and sulfide concentra tions in the porewater were the highest under stagnant conditions. This study suggests that an in termediate level of porewater flux may be beneficial to seagrass growth, while stagnant and hi gh 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 permeab le 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 re gion 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 sediment s 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 re moved from leaves at current velocities greater than 8 cm s -1 . Field experiments indicated that suspended solids were deposited on transplanted eelgra ss 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 infl uence the expansion of seagrasses to unvegetated regions through disper sal 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. Th e 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 fr om 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 (average5.10 current speed of 10 cm s -1 ) sites epiphyte coverage was ne gligible. Results of crosstransplantation and enclosure experime nts 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 maximu m current velocity for Z. marina growth and occurrence based on physiological and mechanical limits are a minimum flow of 3 cm s -1 and a maximum current

of 50-180 cm s -1 (Koch 2001). The growth conditions of s eagrass as a function of current speed reviewed in this chapter are summarized in Figure 5.2.

Deposition of

Sediment on

Leaves

< 8 cm s -1 (3)Growth Conditions

Current Speed, cm s -1 Optimal

Diffusional

Limitation

< 5 cm s -1 (1)

< 5.4 cm s -1 (2) Physical Damage

to Plants

70 cm s -1 (9)

Erosion of sediment resulting in exposure of

roots & rhizomes:

50-150 cm s -1 (9) Threshold for damage:

33-85 cm s -1 (10) Severe damage:

85-110 cm s -1 (10)

Self Shading (11) Increasing Productivity &

Growth with Flow

Flow range of observations

10 to 40 cm s -1 (4,5)

0 to 16 cm s -1 (6)

2 to 34 cm s -1 (7)

1 to 35 cm s -1 (8)

(1) Koch 1994; (2) Enríquez and Rodrígues-Román 2006; (3) Tamaki et al. 2002; (4) Conover 1964; (5) Conover 1968; (6) Nixon & Oviatt 1972; (7) Fonseca & Kenworthy 1987; (8) Peralta et al. 2006; (9) Scoffin 1970; (10) Thom et al. 1996; (11) Zimmerman 2003

Figure 5.2. Growth conditions of seag rass as a function of current speed.5.11 5.4 Effects of Wave Exposure on Se agrass Growth and Distribution

Wave exposure can influence seagrass growth and distribution in se veral ways, including plant breakage, erosion of edges of seagrass beds, influencing plant morphology, modifying water exchange between canopy a nd overlying water column, incr easing flux between porewater and the overlying water column, and causing sedime nt resuspension which can result in reduced light availability. In a recent review, Koch (2001) suggest ed that seagrass distribution and growth appear to be limited by hi gh 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 unve getated due to continuous sand movement (Koch et al. 2006a). In the deeper regions, seagrass typically establish be low the maximum wave penetration depth.

Waves can impact seagrasses by influencing the concentration of su spended particulates, which may alter the light environment and availa bility 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 suspe nded particulates is usually reduced in seagrass beds due to reduced resu spension and enhanced se ttling resulting from

reduced current velocities in the canopy, reduced turbulence en ergy within the canopy and at the sediment surface, attenuation of wave-induced curr ents, as well as binding of sediments by the roots and rhizomes. Sediment resuspension ha s 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, exposur e to waves and relative water de pths 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 speed s, 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 pe rcent silt-clay declined with increasing REI and current speed. There were species-specific differe nces 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 area s, 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 a nd 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 exposur e (REI) were the main factors influencing eelgrass cover in Danish coastal waters. Eelgrass cover was inversel y related to REI in water depths of 0 to 4 m. However, there was a high occurrence (~50% of obs ervations) of absence of eelgrass in areas5.12 that the regression model predicted as suitable ha bitat. Krause-Jensen et al. (2003) suggested that these discrepancies may be associated with stochastic phenomena, such as extreme storms, that are not adequately descri bed by average conditions. Recove ry 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 fluctuat ions in spatial coverage over interannual time scales occurred at exposed sites. Shallow ee lgrass populations form ch aracteristic 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 th e 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 eelgr ass populations (i.e., variability in shoot density and biomass) is highe st in shallow regions where they are subjected to frequent disturbance, due to hydrodynamics a nd other environmental fact ors (Middleboe et al. 2003; Krause-Jensen et al. 2000; Kr ause-Jensen et al. 2003). In addition, the wave environment may influence the plant morphology. Krause-J ensen et al. (2000) found that the shoot to rhizome ratio increases with depth (more alloca tion to photosynthetic tiss ue) 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 condition s. 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) a nd sediment resuspension and movement. At a high energy transplant site, wh ere maximum orbital velocity at the sediment surface frequently exceeded 60 cm s -1 , none of the transplanted plan ts 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 dept hs 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 pr esence of waves substantially influences the porewater flux between the sediment and overlyi ng water column (Pr echt and Huettel 2003; Precht and Huettel 2004), which has the potenti al to influence seag rass productivity. In laboratory flume experiments, Precht and Huet tel (2003) found that shallow water waves can increase the fluid exchange between sandy sedi ments (no vegetation) and the overlying water

column as much as 50-fold relative to mo lecular diffusion rates. They identified two mechanisms which increase the flux of porewat er, hydrostatic pressure induced wave pumping and topography-related filtering. Oldham and Lave ry (1999) found an increase in water column5.13 ammonium that they attributed to increased porewater fluxes associated with interactions between hydrodynamics (increased currents, waves and turbulence) and the sediment surface.

Thomas and Cornelisen (2003) conducte d flume experiments in the field in T. testudinum

habitat to examine the effect of unidirectional and os cillatory flow on ammonium uptake rates. The uptake of ammonium from the water colu mn by the seagrass community was 1.5 times higher in oscillatory flow than in unidirectiona l flow. Uptake rates were positively dependent on water velocity and turbulent energy in the water column.

5.5 Effects of Turbulence on Se agrass Growth and Distribution

Increased turbulence can be bot h beneficial and detrimental to seagrass growth. Benefits

associated with enhanced turbulence include faster removal of undesirable substances, and enhanced transport of nutrients and carbon th rough the blade boundary la yers (Koch 1996). On the negative side, increased turbulence can also le ad to more sediment resuspension, resulting in reduced light availability (K och 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 dist ribution 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 di rect impacts, including propeller scarring and vessel groundings. However, seagrasses may also be impacted by hydrodynamic stressors associated with boating activity. Boating activ ities can modify the hydrodynamics through the generation of wakes and currents. Potential impacts associated with boat wakes include increased sediment resuspension, release of se diment 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 poten tial benefit of increased wave activity was the dislodgement of epiphytes and particulate matter on leaves.

Thom et al. (1996) conducted flume experiment s to assess the impact of propeller washes

on Z. marina . Flume experiments were conducted on in tact patches of eelgrass subjected to current velocities ranging from 0 to 3.25 m s -1 . The lower threshold for plant damage, including loss of plants and exposure of rhizome and roots, occurred at current velocities between 33 and

85 cm s -1 . Severe damage occurred when the cu rrent speeds were between 85 and 110 cm s -1

.5.14 Eleuterius (1987) proposed th at seagrass damaged by motor vessel impacts, such as propeller scars and vessel landi ngs, are more susceptible to fu rther 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 pass age of a class 2 hurrican e, 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 th at 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 projec ts that alter estuarine hydrodyn amics have the potential to affect salinity patterns wh ich in turn can influence the distribu tion of seagrasses. Nienhuis et al. (1996) suggested that alterations in salinity pa tterns associated with the emplacement of dikes may be responsible for some of the observed changes in Z. marina distribution in the Netherlands. Eleuterius (1987) pr oposed that the alteration of fr eshwater discharges associated with the Mississippi River caused the decline of S. filiforme and elimination of H. engelmannii in Mississippi Sound. Quammen and Onuf (1993) proposed that a species shift of the seagrasses in Laguna Madre, Texas resulted from moderation of the sa linity associated with dredging of the Gulf Intracoastal Waterway. For a review of salin ity 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 di sease has been due to hydrodynamic stressors (De Jonge and De Jong 1992). Hydrologic modificati ons in the region resulted in an increase in tidal range of 15 to 30 cm and increased current ve locities 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 dept h) is not responsible for the changes in the underwater light regime. More recently, hydrodynamics has been proposed as a contributing factor to th e 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 infilli ng 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 larg e in shallow water where the populations are disturbed by wave action and other physical pa rameters (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 sh ift in seagrass distri butions to shallower environments due to light limitation, there may be an increase in the occurrence of hydrodynamic stressors influencing the distributi on of seagrass since hydrodynamic stressors are elevated in shallow environments (Patters on et al. 2001; Middleboe et al. 2003).5.15 5.8 Natural Hydrodynamic Stressor Events

Extreme events, such as storms, hurricanes, and floods, have been reported to damage seagrass beds (e.g., Eleuterius and Miller 1976; Preen et al. 1995; Aioi and Komatsu 1996). The effects of hurricanes on seagrasses vary wide ly 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 sm othering 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 km 2 of seagrass was lost from Hervey Bay, Australia, following two major floods and a cyclone within a three week inte rval. The effect of hurricanes on seagrass can be localized with so me regions showing no eff ect, while other regions are damaged (Van Tussenbroek 1994). In additio n to the physical stre sses 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 a nd Moore 1983). After a major storm disturbance, eelgrass populations may exhib it extensive grow th and increased survival of new shoots (Krause-Jensen et al . 2000; Aioi and Komatsu 1996). Fonseca et al. (2000) proposed that wave-exposed seagrass hab itat may be particularly vulnera ble 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 in fluencing 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 li fe stage of the plant needs to be taken into account when setting protective criteria. S eagrasses may be more susceptible to hydrodynamic stressors when they ar e 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. Alte rations to the hydrodynamics often results in changes in the environment (e.g., sa linity, turbidity, nutrien ts, 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 criter ia, we need more research on the effects of hydrodynamics on the productivity, survival and dist ribution of seagrass. One difficulty in setting protective criteria for hydrodynamic stre ssors 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 turbul ence). In order to further our understanding of the effect of hydrodynamic stressors, we need to be sure to collect sufficien t ancillary data for interpretation of study results. When colle cting hydrodynamic data in seagrass habitat,5.16 information also needs to be collected on water de pth, location of measurem ents (both within the bed and elevation), canopy height, percentage of water column occupied by seagrass canopy, seagrass density, size and patchiness of seagra ss 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., o ccurrence of monami, sk imming flow, and blade flapping) (Koch and Verduin 2001).

Koch et al. (2006b) suggeste d that numerical models of the interaction of waves and seagrasses are useful for developing testable hypotheses, explaini ng observations, designing

observational studies or restorati on efforts, and interpolating sp arse data in space and time. Hydrodynamic and sediment resuspension and tr ansport 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 transpor t modeling in shallow, vege tated 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 fi eld measurements are needed to improve model formulations for sedime nt resuspension within seagrass beds.

Although more research is needed on the effect of hydrodynamic stressors on seagrasses,

several studies include the eff ect of hydrodynamic stressors in models/indices to predict the success of seagrass restoration projects and prov ide 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 se lection tool for transplantation sites that included hydrodynamic stressors in the selection pro cedure. Selection of su itable transplant sites were based on sediment composition, emersion time, current velocity and wave action. Kelly et al. (2001) take into account hydr odynamics, in determining whic h regions you would expect to have high probability of restoration success. Short et al. (2002) presented a site-selection tool for transplantation of Z. marina which included wave exposure.5.17 5.10 Literature Cited

Abdelrhman, M. A. 2003. Effect of eelgrass Zostera marina canopies on flow and transport. Marine Ecology Progress Series 248:67-83. Ackerman, J. D. 2002. Diffusivity in a marine macrophyte canopy: Implications for submarine pollination and dispersal. American Journal of Botany 89:1119-1127. Ackerman, J. D. and A. Okubo. 1993. Redu ced mixing in a marine macrophyte canopy. Functional Ecology 7:305-309. Aioi, K. and T. Komatsu. 1996. A preliminary report on the r ecovery process of Zostera marina in Otsuchi Bay, Northern Japan after distur bance by large atmospheri c depressions. pp. 319-322. In: Kuo, J., Phillips, R.C., Walker, D.I., and Kirkman H. (eds.) Seagrass Biology: Proceedings of an International Wo rkshop.Faculty of Sciences, The University of Western Australia, Nedl ands, Western Australia. Conover, J. T. 1964. The importance of natural diffusion gradients and transport of substances related to benthic marine plant metabolism. Environmental relationships of benthos in salt ponds (Plant Relationships), vol 1, Technical Report 3, Graduate School of Oceanography, University of Rhode Island, Kingston, Rhode Island. Conover, J. T. 1968. The importance of natural diffusion gradients and transport of substances related to benthic marine plant me tabolism. Botanica Marina 11:1-9. Cornelisen, C. D. and F. I. M. Thomas. 2002. Ammonium uptake by seagrass epiphytes:

isolation of the effects of water velocity using an is otope label. Limnology and Oceanography 47:1223-1229. De Jonge, V. N., D. J. de Jong, and M. M. van Katwijk. 2000. Policy plans and management measures to restore eelgrass ( Zostera marina L.) in the Dutch Wadden Sea. Helgoland

Marine Research 54:151-158. De Jonge, V. N. and D. J. de Jong. 1992. Role of tide, light and fish eries in the decline of

Zostera marina L. in the Dutch Wadden Sea. Netherlands Institute for Sea Research

20:161-176. Eleuterius, L. N. 1987. Seagrass ecology al ong the coasts of Alab ama, Louisiana, and Mississippi. pp. 11-24. In: Durako, M. J., Phil lips, R.C., Lewis, R.R. (eds). Proceedings of the symposium on subtropical -tropical seagrasses of the southeastern United States. Florida Marine Research Publication No. 42, Florida Department of Natural Resources, St. Petersburg, FL. Eleuterius, L. N. and G. J. Miller. 1976. Observations on seagrasses and seaweeds in Mississippi Sound since Hurricane Camille. J ournal of the Mississippi Academy of

Science 21:58-63. Enríquez, S. and A. Rodríguez-Román. 2006. Effect of water flow on the photosynthesis of three marine macrophytes from a fringing-re ef lagoon. Marine Ecology Progress Series

323:119-132. Fonseca, M.S., M. A. R. Koehl, and B. S. Kopp. 2007. Biomechanical factors contributing to self-organization in s eagrass landscapes. Journal of Experimental Marine Biology and Ecology. 340:227-246. Fonseca, M. S., W. J. Kenworthy, and P. E. Whitfield. 2000. Temporal dynamics of seagrass landscapes: a preliminary comparison of chroni c and extreme disturbance events. p. 373-5.18 376. In: Pergent G, Pergent-Martini C, Buia MC, Gambi MC (eds). Proceedings 4th Int. Seagrass Biology Workshop, Sept 25-Oct 2, 2000, Corsica, France. Fonseca, M. S. and S. S. Bell. 1998. Influen ce of physical setting on seagrass landscapes near Beaufort, North Carolina, USA. Marine Ecology Progress Series 171:109-121. Fonseca, M. S. and J. A. Cahalan. 1992. A preliminary evaluation of wave attenuation by four species of seagrass. Es tuarine and Coastal Shelf Science 35:565-576.

Fonseca M. S. and J. S. Fisher. 1986. A co mparison of canopy friction and sediment movement between four species of seagrass with refere nce to their ecology and restoration. Marine Ecology Progress Series 29:15-22. Fonseca, M. S., J. S. Fisher, J. C. Zieman, and G. W. Thayer. 1982. Influence of the seagrass,

Zostera marina L., on current flow. Estuarine an d Coastal Shelf Science 15:351-364. Fonseca, M. S. and W. J. Kenworthy. 1987. Effects of current on photosynthesis and distribution of seagrasses. Aquatic Botany 27:59-78. Fonseca, M. S., J. C. Zieman, G. W. Thayer, and J. S. Fisher. 1983. The role of current velocity in structuring eelgrass ( Zostera marina ) meadows. Estuarine and Coastal Shelf Science

17:367-380. Fourqurean, J. W. and M. B. Robblee. 1999. Florida Bay: a histor y of recent ecological changes. Estuaries 22(2B):345-357 Frederiksen M., D. Krause-Jensen, M. Holmer, a nd J. S. Laursen. 2004. Spatial and temporal variation in eelgrass ( Zostera marina ) landscapes: influence of physical setting. Aquatic Botany 78:147-165. Gambi, M. C., A. R. M. Nowell, and P. A. Ju mars. 1990. Flume observations of flow dynamics

in Zostera marina (eelgrass) beds. Marine Ec ology Progress Series 61:159-169. Ghisalberti M. and H. M. Nepf . 2002. Mixing layers and c oherent structures in vegetated aquatic flows. Journal of Ge ophysical Research 107(C2):1-11. Granata, T. C., T. Serra, J. Colomer, X. Casamitj ana, C. M. Duarte, and E. Gacia. 2001. Flow and particle distributions in a nearshore seagrass meadow before and after a storm. Marine Ecology Progress Series 218:95-106. Grizzle, R. E., F. T. Short, C. R. Newell, H. Hoven, and L. Kindblom. 1996. Hydrodynamically induced synchronous wavi ng of seagrasses: ‘monami’ and its possible effects on larval mussel settlement. Journal of Experiment al Marine Biology and Ecology 206:165-177. Harwell, M. C. and R. J. Orth. 2002. Long-di stance dispersal potential in a marine macrophyte. Ecology 83:3319-3330. Kelly, N. M., M. Fonseca, and P. Whitfield. 2001. Predictive mapping for management and conservation of seagrass beds in North Ca rolina. Aquatic Conservation: Marine Freshwater Ecosystems 11:437-451. Koch, E. W. 2002. Impact of boat-generated wa ves on a seagrass habitat. Journal of Coastal

Research 37:66-74. Koch, E. W. 2001 Beyond light: Physical, geolog ical and geochemical parameters as possible submersed aquatic vegetation habitat re quirements. Estuaries 24:1-17. Koch, E. W. 1999a. Preliminary evidence on the interdependent effect of currents and porewater geochemistry on Thalassia testudinum Banks ex König seedlings. Aquatic Botany 63:95-102.5.19 Koch, E. W. 1999b. Sediment resuspension in a shallow Thalassia testudinum Banks ex König bed. Aquatic Botany 65:269-280. Koch, E. W. 1996. Hydrodynamics of a shallow Thalassia testudinum bed in Florida, USA. In: Seagrass Biology: Proceedings of an international workshop, 105-110. Koch, E.W., J. D. Ackerman, J. Verduin, a nd M. van Keulen. 2006a. Fluid dynamics in seagrass ecology-from molecules to ecosystems . pp. 193-225. In: A.W.D. Larkum, R.J. Orth and C. M. Duarte (eds.), Seagrasses: Biology, Ecology and C onservation, Springer,

Netherlands. Koch, E. W., S. Ailstock, and J. C. Stevens on. 2002. Chapter VI, Beyond Light: Physical, geological and chemical habitat requirem ents. pp. 71-94. In: Chesapeake Bay Submerged Aquatic Vegetation Water Qual ity and Habitat-Based Requirements and Restoration Targets: A Sec ond Technical Synthesis. U. S. Environmental Protection Agency. Koch, E. W. and G. Gust. 1999. Water flow in tide- and wave-dominated beds of the seagrass

Thalassia testudinum . Marine Ecology Progress Series 184:63-72. Koch, E. W. 1994. Hydrodynamics, diffusi on-boundary layers and photosynthesis of the

seagrasses Thalassia testudinum and Cymodocea nodosa . Marine Biology 118:767-776. Koch, E. W. and M. Huettel. 2000. The imp act of single seagrass shoots on solute fluxes between the water column and permeable se diments. Biologia Marina Mediterranea

7:235-239. Koch, E.W., L. P. Sanford, S.-N. Chen, D. J. Shafer, and J. McKee Smith. 2006b. Waves in seagrass systems: Review and techni cal recommendations. Report ERDC-TR-06-15, Engineer Research and Development Cent er, U.S. Army Corps of Engineers, Washington, DC. Koch, E. W. and J. J. Verduin. 2001. M easurements of physical parameters in seagrass habitats. pp. 325-344. In: F. T. Short and R. G. Coles (eds.). Global Seagrass Research Methods. Elsevier, Amsterdam. Komatsu, T. 1996. Influence of a Zostera bed on the spatial distribut ion of water flow over a broad geographic area. pp. 111-116. In: Kuo, J., Phillips, R.C., Walker, D.I., and Kirkman H. (eds.) Seagrass Biology: Proceedings of an international workshop. Kopp, B. S. 1999. Effects of nitrate fe rtilization and shading on physiological and biomechanical properties of eelgrass ( Zostera marina L.) Ph.D. dissertation, University of Rhode Island. pp. 187. Krause-Jensen, D., A. L. Middleb oe, K. Sand-Jensen, and P. B. Christensen. 2000. Eelgrass

Zostera marina , growth along depth gradients: uppe r boundaries of the variation as a powerful predictive tool. Oikos 91:233-244. Krause-Jensen, D., M. F. Pedersen, and C. Jensen. 2003. Regulation of eelgrass ( Zostera

marina ) cover along depth gradients in Danish coastal waters. Estuaries 26:866-877. Middleboe, A. L., K. Sand-Jensen, and D. Krau se-Jensen. 2003. Spatial and interannual variations with depth in eelgrass populations . Journal of Experimental Marine Biology and Ecology 291:1-15. Morin, J. and J. W. Morse. 1999. Ammoni um release from resuspended sediments in the Laguna Madre estuary. Marine Chemistry 65: 97-110.5.20 National Research Council. 1994. Priorities fo r coastal ecosystem science. National Academy Press, Washington, D.C. Nepf, H. M. and E. W. Koch. 1999. Vertical s econdary flows in submersed plant-like arrays. Limnology and Oceanography 44:1072-1080. Nepf, H. M. and E. R. Vivoni. 2000. Flow struct ure in depth-limited, vegetated flow. Journal of Geophysical Research 105:28547-28557. Newell, R. I. E. and E. W. Koch. 2004. M odeling seagrass density and distribution in response to changes in turbidity stemming from bi valve filtration and seagrass sediment

stabilization. Estuaries 27:793-806. Nienhuis, P.H., B. H. H. De Bree, P. M. J. Herm an, A. M. B. Holland, J. M. Verschuure, and E.

G. J

. Wessel. 1996. Twenty five years of ch anges in the distribution and biomass of eelgrass, Zostera Marina , in Greverlingen Lagoon, the Ne therlands. Netherlands Journal of Aquatic Ecology 20:107-117. Nixon, S. W. and C. A. Oviatt. 1972. Pre liminary measurements of midsummer metabolism in beds of eelgrass, Zostera marina . Ecology 53:150-153. Oldham, C. E. and P. S. Lavery. 1999. Pore water nutrient fluxes in a shallow fetch-limited estuary. Marine Ecology Progress Series 183:39-47. Oppenheimer, C. H. 1963. Effects of hurrican e Carla on the ecology of Redfish Bay, Texas.

Bulletin of Marine Science of the Gulf and the Caribbean 13:59-72. Orth, R. J., M. Luckenbach, and K. A. Moore. 1994. Seed dispersal in a marine macrophyte: implications for colonization and re storation. Ecology 75:1927-1939. Orth, R. J. and K. A. Moore. 1983. Chesap eake Bay: An unprecedented decline in submerged aquatic vegetation. Science 222:51-53. Paling, E. I., M. van Keulen, K. D. Wheeler, J. Phillips, and R. Dyhrberg. 2003. Influence of spacing on mechanically transplanted seagrass survival in a high wave energy regime. Restoration Ecology 11:56-61. Patterson, M. R., M. C. Harwell, L. M. Orth, an d R. J. Orth. 2001. Biomechanical properties of the reproductive shoots of eelgr ass. Aquatic Botany 69:27-40. Peralta, G., F. G. Brun, J. L. Pérez-Lloréns, an d T. J. Bouma. 2006. Direct effects of current velocity on the growth, morphometry, and ar chitecture of seagrasses: a case study on

Zostera noltii . Marine Ecology Progress Series 327:135-142. Peralta, G., F. G. Brun, I. Hernández, J. J. Vergara, and J. L. Pérez-Lloréns. 2005. Morphometric variations as acclimation mechanisms in Zostera noltii bed. Estuarine,

Coastal and Shelf Science 64:347-356. Peterson, C. H., R. A. Luettich, Jr., F. Micheli, and G. A. Skilleter. 2004. Attenuation of water flow inside seagrass canopies of differing structure. Marine Ecology Progress Series

268:81-92. Powell, G. V. N. and F. C. Schaffner. 1991. Water trapping by seagrasses occupying bank habitats in Florida Bay. Estuarine and Coastal Shelf Science 32:43-60. Precht, E. and M. Huettel. 2003. Advective porew ater exchange driven by surface gravity waves and its ecological implications. Limnology and Oceanography 48:1674-1684. Precht, E. and M. Huettel. 2004. Rapid wave -driven advective pore water exchange in a permeable coastal sediment. Jour nal of Sea Research 51:93-107.5.21 Preen, A. R., W. J. Lee Long, and R. G. Coles. 1995. Flood and cyclone re lated loss, and partial recovery, of more than 1000 km 2 of seagrass in Hervey Ba y, Queensland, Australia. Aquatic Botany 52:3-17. Quammen, M. L. and C. P. Onuf. 1993. La guna Madre: Seagrass changes continue decades after salinity reduction. Estuaries 16:302-310. Schanz, A. and H. Asmus. 2003. Impact of hydrodynamics on development and morphology of intertidal seagrasses in th e Wadden Sea. Marine Ecol ogy Progress Series 261:123-134. Scoffin, T. P. 1970. The trapping and binding of subtidal carbonate sediments by marine vegetation in Bimini Lagoon, Bahamas. J ournal of Sedimentary Petrology 40:249-273. Short, F. T. 1983. The seagrass, Zostera marina L.: Plant morphology and bed structure in

relation to sediment ammonium in Izembe k Lagoon, Alaska. Aquatic Botany 16:149-161. Short, F. T., R. C. Davis, B. S. Kopp, C. A. Short, and D. M. Burdick. 2002. Site-selection model for optimal transplantation of eelgrass Zostera marina in the northeastern US. Marine Ecology Progress Series 227:253-267. Short, F. T. and H. A. Neckles. 1999. The effects of global climate change on seagrasses. Aquatic Botany 63:169-196. Tamaki, H., M. Tokuoka, W. Nishijima, T. Terawa ki, and M. Okada. 2002. Deterioration of eelgrass, Zostera marina L., meadows by water pollution in Seto Inland Sea, Japan.

Marine Pollution Bulletin 44:1253-1258. Teeter, A. M., B. H. Johnson, C. Berger, G. Ste lling, N. W. Scheffner, M .H. Garcia, and T. M. Parchure. 2001. Hydrodynamic and sedime nt transport modeling with emphasis on shallow-water, vegetated areas (lakes, reserv oirs, estuaries, and la goons). Hydrobiologia

444:1-23. Thom, R. M., A. B. Borde, P. J. Farley, M. C. Horn, and A. Ogston. 1996. Passenger-only ferry propellor wash study: Threshold velocity determinations and field study, Vashon

Terminal. Battelle Marine Scienc es Laboratory, Sequim, WA. Thomas, F. I. M. and C. D. Cornelisen. 2003. Ammonium uptake by seagrass communities: effects of oscillatory versus unidirectional flow. Marine Ecology Progress Series 247:51-

57. Thomas, F. I. M., C. D. Cornelisen, and J. M. Zande. 2000. Effects of water velocity and canopy morphology on ammonium uptake by seagrass communities. Ecology 81:2704-

2713. Thomas, L. P., D. R. Moore, and R. C. Work . 1961. Effects of hurricane Donna on the Turtle Grass beds of Biscayne Bay, Florida. Bu lletin of Marine Science of the Gulf and

Caribbean 11:191-197. Tilmant, J. T., R. W. Curry, R. Jones, A. Szma nt, J. C. Zieman, M. Flora, M. B. Robblee, D. Smith, R. W. Snow, and H. Wanless. 1994. Hurricane Andrew’s effects on marine

resources. Bios cience 44:230-237. van Katwijk, M. M. and D. C. R. He rmus. 2000. Effects of water dynamics on Zostera marina : Transplantation experiments in the inter tidal Dutch Wadden Sea. Marine Ecology Progress Series 208:107-118. van Keulen, M. and M. A. Borowitzka. 2000. Comparison of water velocity profiles through morphologically dissimilar seagrasses. Bi ologia Marina Medite rranea 7:143-146.5.22 van Keulen, M., E. I. Paling, and C. J. Walker. 2003. Effect of planting unit size and sediment stabilization on seagrass transplants in We stern Australia. Restoration Ecology 11:50-55. van Tussenbroek, B. I. 1994. The impact of Hu rricane Gilbert on the vegetative development of

Thalassia testudinum in Puerto Morelos coral reef la goon, Mexico: a retrospective study.

Botanica Marina 37:421-428. Ward, L. G. W., M. Kemp, and W. R. Boynton. 1984. The influence of waves and seagrass communities on suspended particulates in an estuarine embayment. Marine Geology

59:85-103. Whitfield, P. E., W. J. Kenworthy, K. K. Hammerstrom, and M.S. Fonseca. 2002. The role of storms in the expansion of di sturbances initiated by motor vessels on subtropical seagrass

–coral banks. Journal of Co astal Research 37:86-99. Worcester, S. E. 1995. Effects of eelgrass beds on advection and turbulent mixing in low current and low shoot density environments. Marine Ecology Progress Series 126:223-

232. Zimmerman, R. C. 2003. A biooptical model of irradiance distributi on and photosynthesis in seagrass canopies. Limnology and Oceanography 48:568-585.6.1 

## 6.0 Interactions of Zostera marina and Thalassia testudinum with

Sediments

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

6.1 Background

As rooted aquatic plants, seagrasses are influenced by the sedimentary environment in which they grow. Sediment characteristics such as grain size, mineral composition and organic matter content may influence the overall biogeoche mical environment of the root zone. Human

actions can also alter sediment grain size distribution (i.e., dr edging activities, soil erosion, production of excessive fine sediments through jett y and dike construction, prop wash and wakes from boats) which in turn may affect the sedi mentary geochemical environment and the rooting environment. Further, human actions can affect sediment organic matter content through watercolumn nutrient elevations and the subsequent development of phytoplankton, epiphytes and macroalgal blooms. The resulti ng sedimentation of organic matter (OM) to the benthos leads to changes in sediment mineralization rates and re dox conditions. Of pa rticular concern are generation of phytotoxic substances in anaerobic sediments (e.g., elevated sediment ammonia, sulfide, metal ions, and other reduced chemical species) (Figure 6.1). Here we review and

consider the effects of sediment characteristics on the establishment, survival or growth of

Zostera marina and Thalassia testudinum

.

Seagrasses interact physically, biological ly and geochemically with both the water column and the sediment in which they are root ed. 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 geoche mistry, 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 fo r particulate matter (Nepf and Koch 1999) (Figure 6.1). Because of mineralizati on of organic matter input to the root zone through burial, DOM release from seagrass (Kaldy et al. 2006) and par ticle 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 re dox processes to non-toxic forms, phytotoxic

substances can concentrate around th e 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 Sedi ment 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 typically6.2 found growing in substrata of fine or muddy sand, although this genus is also found growing over a wide range of sediment size classes. In the Pacific Northwest, Boese and Young (unpublished data) found Z. marina growing in a range of sedime nt texture classes in Yaquina Bay, OR. Although a majority of the Z. marina habitat was in the lower estuary associated with sediments characterized by 75-100% sand (up to 25% silt/clay), Z. marina meadows also occurred up estuary in sediments containing up to 75% silt/clay. Unfortunately, it is not possible to determine if there is an optimum grain size composition for Zostera growth and survival with these data. In an experimental study of eelgrass using mesocosms, Short (1987) showed that leaf biomass, weight, and shoot height were signific antly greater in plants growing on fine grained mud than on coarse grained sand although the differ ences between treatments in this case may

have been due to nutrient limitation. Thom et al. (2001) cultivated Z. marina for 13 weeks in sediment types typically occupied by this specie s in the Pacific Northwest, as well as coarse, organic-poor sand and gravel, whic h typically are not inhabited by Z. marina (Phillips 1984). The greatest growth was observed in the finer gr ained sediments containi ng organic matter, and lowest growth was measured in the gravel subs tratum. Coarse-grained sand and a sand/gravel mixture produced intermediate growth rates. Thes e experimental results are consistent with the

Z. marina distribution data summarized in Table 6.1, but as in previous experiments, may be confounded by differences in nutrient availability or mineralogy.

6.2.2 Interactions of Sedi ment Grain Size and Thalassia

Sediments found in Thalassia beds typically fall in the “m edium sand” size class, and range from fine to coarse sand (T able 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, rangi ng from 1 to 34%, and the overall median value (taking the mid-point of the range for a given stu dy) 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 le af 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 c onditions (Brenchley and Probert 1998; Moore et al. 1993). Moore et al. (1993) reported that Z. marina seed germination was often triggered by anoxic conditions (in either sediment or water). Such conditions typically occur seasonally in finer, organic-rich sediments where the excha nge between overlying oxyg enated water and the sediment is restricted (Koch 2001). Seeds bur ied to 5 mm in the sediments showed lower germination success in the autumn than seed burie d at 15 to 25 mm, but in the winter there was no apparent effect of burial de pth in the sediment. In ge neral, germination of buried Z. marina seeds began when water temperature dropped belo w 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. Furthermor e, 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 seeds6.3 characterized by an intact, hard seed coat of da rk color, and a fully developed embryo were used in the Moore et al. (1993) experiment. Harw ell and Orth (1999) found high (41-56%) seedling survival in burlap bag treatments but only 5-15% in bare sediment treatment due to predation, burial, or lateral transport.

Current Light

Chlorophyll a TSS

DIN

DIP

Epiphytes SOM TSS Waves

SO 4 - O 2 Root O 2 Flux MeO

x

HS

AVS Me x NO 3 - Oxidized

NH 4 + Reduced

Figure 6.1. Both the water column and sediment environments influence seagrasses. While the physical and light attenuating water-column stressor may be the most important to the survival of the plant, the sediment geochemi cal processes, stimulated by sedimentation of reactive organic material from the water co lumn, 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 (Me x ) 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 e nvironment (i.e., the rh izosphere) even under conditions of relatively high organic matter input to the se diments. (DIN – dissolved inorganic nitrogen, DIP – dissolved inor ganic phosphorus, SOM – settling organic matter)6.4 The typical range of salinity (~15-30) generally found in estuaries also has minimal effects on seed germination (Brenchley a nd 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. Anaer obic conditions seem to promote greater germination success and shorter germinati on periods (Brenchley and Probert 1998).

Table 6.1. Distribution of Zostera species in relation to sediment texture or grain size. Texture was available in some publications while grain size was reported in others.

Seagrass Organic Matter (%) Size Class Median Grain Size (mm) Silt+Clay (%) Source

Zostera marina 1.2 14 Marshall and Lucas (1970)

1.2 very fine -

fine sand2 0.100 - 0.1461 Orth (1977)

medium - muddy sand 0.12 - 0.50 Nienhuis and DeBree (1977)

2.2 - 2.7 11 - 13 Kenworthy el al. (1982)

1.7 6 - 24 Peterson et al. (1984)

0.4 - 1.0 medium-fine

sand2 0.17 - 0.34 Kenworthy and Fonseca (1992)

6.4 - 16 Short et al. (1983)

fine sand - silty clay Lalumiere et al. (1994)

0.4 - 3.5 5 - 35 Townsend and Fonseca (1998)

0.4 – 1.4 Dan et al. (1998)

medium -

fine sand2 0.14 - 0.271 8 - 13 Webster et al. (1998)

fine sand2 0.17 - 0.201 4 - 6 Frost et al. (1999)

0.4- 12 2 - 56 Koch (2001)

medium sand

- fine

sand/silt Thom et al. (2001)

medium - very fine

sand2 0.06 - 0.353 Lee et al. (2002)

Zostera japonica 1.7 - 3.8 fine sand2 0.13 - 0.231 , 3 Posey (1988)

Zostera

noltii fine sand2 0.13 - 0.23 40 - 70 van Lent et al. (1991)

very fine

sand2 0.07 Sprung (1994)

Zostera

muelleri 0.8 - 7.3 1 - 72 Edgar and Shaw (1995)

Zostera

novelandica fine sand 0.193 Heiss et al. (2000)

Thalassia

testudinum 3.5 – 4.9 5 Wood et al. (1969)6.5 15 Scoffin (1970)

1 – 34 Burrell and Schubel (1977)

medium

sand2 0.35 - 0.361 Orth (1977)

0.8 medium

sand2 0.241 , 3 3 Grady (1981)

fine sand2 0.15 - 0.20 Wanless (1981)
12 – 34 Hoskin (1983)

22 Lee and Dunton (1996)

0.5 - 0.6 McGlathery et al. (1994)

medium -

coarse sand2 0.24 - 0.593 Kuenen and Debrot (1995)

fine - coarse

sand2 0.19 - 0.501 Kalbfleisch and Jones (1998)

2 - 9 Livingston et al. (1998)

0.5 – 2.3 2 - 17 Koch (2001)
5 - 12 Terrados et al. (1998)
23 - 35 Kaldy and Dunton (2000)

Thalassia hemprichii 3.5 - 10 medium

sand2 0.24 - 0.371 , 3 Paula et al. (2001)

1. Calculated from phi (ø) values: (ø) = -log 2 (mm) (Krumbein and Pettijohn 1938) 2. Wentworth size class obtained from grain si ze (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 (H 2 S and HS - ) may inhibit seagrass photosynthesis, growth, and survival through sulfide toxicity (Goodman et al. 1995; Terrados et al. 1999). The high conc entration 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 abund ant 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 se dimentation (Eldridge et al. 2004).

To counteract sulfide accumulation in the rhizosphere, seagrasses transport photosynthetically produced O 2 through lacunae to the roots (Smith et al. 1988; Caffrey and Kemp 1991; Kraemer and Alberte 1993). This tran sport mechanism is probably an adaptation to support aerobic root resp iration, but excess O 2 diffusing from the roots into the rhizosphere has the added benefit of oxidizing sulfide to non-toxi c sulfate (Caffrey and Kemp 1991). In addition to the photosynthetically produced O 2 , dissolved O 2 in the water column can diffuse through Z.

marina and T. testudinum lacunal systems, thereby reducing th e exposure of seagrass to sulfides at night when photosynthetic processes are not active (Koch and Erskine 2001). Binzer et al.6.6 (2005) showed that in the dark the degree of O 2 saturation in water, and flow velocity were the primary determinates of in ternal oxygen conditions in Cymodocea nodosa. Zostera marina may also exhibit some anatomical plasticity in the de velopment 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 ma rine/estuarine plants.

Pore

Water (PW) or Water (W) Exposure Marine/Estuarine

Plant Effect

Conc. (μM) Duration Response Variable Citation

PW Zostera marina 2300 24 hr root respiration not reduced Penhale and Wetzel (1983)

PW Zostera marina 400 - 800 21 day decreased photosynthesis Goodman et al. (1995)

PW Zostera marina 70 2 mo decreased leaf elongation rate; shoot density unaffected Terrados et al. (1999)

W Zostera marina 100-1000 1-3 weeks

in water photosynthesis stopped leaf elongation rate stopped

leaves/shoot decreased 37%

root non-structural ca rbohydrate decreased

81% above-ground biomass (shoot/root )

decreased 55% Holmer and Bondgaard (2001)

PW Thalassia testudinum ~ 1,500 4-6 weeks sulfide suggested as synergistic secondary

stressor Carlson et al. (1994)

PW Thalassia testudinum 80 6 mo seedling mortality 100% Koch (1999)

PW Thalassia testudinum 2000-6000 48 hr leaf elongation rate decreased 43% root energy charge decreased 22% root ATP production decreased no visual signs of acute toxicity Erskine and Koch (2000)

PW Thalassia testudinum 6000 14 day high salinity (S), 50% mortality high temperature (T), 33% mortality high S and T, 100% mortality Koch and Erskine (2001)

PW Thalassia testudinum 5500 38 day net shoot loss 65% at 34-35°c no new leaf emergence at 34-35°c Koch et al. (2007)

In marine environments, there is a constant composition of major ions relative to salinity so that in high salinity waters there is, as discussed above, an abundance of sulfate to support anaerobic diagenesis (Stumm and Morgan 1981). Owing to the potential for sulfide formation, plants in euryhaline and mesohaline waters ma y require sediments which are more oxygenated (i.e., coarser sediments) and ha ve higher rates of pore water exchange (Koch 2001) than is needed in brackish waters to reduce the effect of sulfide toxicity.6.7 Although sulfate competes poorly as an oxida nt, its abundance in seawater ensures that sulfate reduction is a dominant geochemical pr ocess in near shore se diments (Berner 1980). However, the extrapolation of the general trends in near shore sediment diagenesis to seagrass

sediments is not direct. While Holmer and B ongaard (2001) found that th e depth distribution of sulfate reduction in a tropical seagrass ( Cymodocea rotundata ) was positively correlated with below-ground biomass, they also found that su lfate reduction was not the major diagenetic process leading to nutrient re-min eralization, possibly because of the ability of seagrasses to inject O 2 and other oxidants into the root zone.

The ability of seagrasses to promote minera lization of organic material to release NH 4 +
while regulating sulfate reduction in the root zo ne may be an important mechanism to insure seagrass survival (Eldridge et al. 2004). Th e regulation of sulfate reducing bacteria by seagrasses may have other benefi ts. Sulfate reducing bacteria of ten fix nitrogen (Welsh et al. 2001), which may lead to increased nitrogen availab ility and which can also be beneficial to the growth of seagrasses. By re gulating sulfate reduction the pl ant can maximize the benefit of sulfate reduction while minimizing its toxic e ffects. Other complicat ing factors that may mitigate the effects of sulfide concentration include the presence of iron (Fe(II)) and manganese (Mn 2+ ) in the rhizosphere which may also reduce sulfide toxicity by the formation of insoluble metal sulfides (Erskine and Koch 2000). This might support the hypothesis that accumulation of sulfides could be mitigated by fertilization of seagrass habitats with Fe(III) compounds. An advantage of this proposition would be competiti ve anaerobic mineraliza tion of organic matter by Fe(III) instead of sulfate. The products of iron reduction are not thought to be toxic to Z.

marina

.

In terrestrial systems, root exudates (i.e ., plant-derived organic compounds added to the rhizosphere) can bind with toxic metal ions or other compounds to provide protection under chemically adverse soil conditions (Höberg and Jensén 1994). It is possible that seagrasses have similar mechanisms to protect against toxic ch emicals in sediments; however, little is known about root exudation, root growth, and root longevity and morta lity 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 2 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 (20 01) showed that photosy nthesis stopped after 6 days at water column sulfide concentrations between 100 and 1000 μ mol. This study was done at saturating irradiance (400-500 μ mol photon m -2 s -1 ). Koch and Erskine (2001) determined the response of Thalassia testudinum to sulfides under varying c onditions of light, salinity, and

temperature. T. testudinum showed no response to sulfid es until the dose reached 6 mM concentration, and a response occurred only in the high temperature treatment (35oC). In contrast, longer-term studies by Carlson et al. (1994) showed T. testudinum died after being exposed to ~1.5 mM sulfide concentrations

. However, Carlson et al. (1994) note that the high sulfides at their research sites are probably just one of several factors th at contribute to die-off episodes rather than the primary cause of d eath. Other factors su ch as hyperthermia,6.8 hypersalinity and microbial pathogens may act in s ynergy with high sulfides to induce mortality in seagrass (Carlson et al. 1994).

There are several internal pl ant 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 mitochondr ial enzyme cytochrome a 3 oxidase, blocking the terminal step in the electron transp ort system. Sulfides can also bind Zn(II) in carbonic anhydrase replacing the bound hydroxyl n ecessary for the inter-conversion of CO 2 and water to bicarbonate (Beauchamp et al. 1984). Currently, there is litt le information on how various plant organelles i ndividually respond to sulfides. In particular, more information is

needed to determine the effect of sulfides on s eagrass plant meristems since this is the site of most anabolic processes. Further we have f ound no studies that addre ss 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 sp ecies 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 possibl e 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., Bo strom et al. 2002; Mattila et al. 1999; Webster et al. 1998). Fewer studies, however, have demonstrat ed 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 disc ussion 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 th an 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 (Eld ridge 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 buf fers the effects of current velocity and turbulence within the canopy (Koc h 2001) (Figure 6.1) thereby re ducing sediment resuspension, total suspended solids concentrations and increasing water clarity. In addition to the settling of particles due to decreased turbul ence and water flow (Koch 2001), particles physically adhere to seagrass leaf surfaces or are trapped by protoz oa 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 hi gh rates of organic carbo n input into seagrass sediments (Agawin and Duarte 2002).6.9 Field studies have shown that surficial sedi ments 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) se diments. For example, Marshall and Lukas (1970) found that surficial (0-1 cm) sediments fr om 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 unvegeta ted sediments were 1.89 + 0.34% and 0.88 + 0.17%. Orth (1977) reported that the median sediment diameter fo r 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, Peters on 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 sedime nt texture generally reflect di fferences in physical processes related to waves and currents (Burrell and Sc hubel 1977). Seagrasses of ten impede currents, reduce the flow velocity (Gin sburg 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; Ke nworthy et al. 1982; Peterson et al.1984;

Gambi et al.1990; Ackerm an 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 veloc ity of about 25 cm/sec, while this variance was much reduced at higher current speeds. They s uggested that the initiatio n 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 const itutes a disturbance thre shold 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 im portant to both the flushing of organic matter out of the seagrass beds (Hemminga and Duar te 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 seagra ss by preventing the dete rioration of sediment condition (Hemminga and Duarte 2000) or by providing organic material which, upon mineralization, provides sufficient NH 4 + to avoid a nitrogen limitation (Zimmerman et al. 1987).

6.2.7 Sediment Nutrient Effects Seagrasses assimilate nutrien ts 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 th e rhizosphere is aerobic . However, during the dark period when roots and rhizomes are ofte n subject to fermentati on, 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 photosynthesis6.10 and utilization of stored carbohydr ates, the availability of NH 4 + and NO 3 - in the water-column, and the rate of organic sediment s 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). Minera lization rates may be regulated to some degree by the seagrass plant. In sedime nts with low organic matter content (less than about 2.5%), seagrasses receivi ng high irradiance at the canopy can regulate the sediment redox

environment with lacunal O 2 diffusion from the roots and rhizomes. Maintaining healthy redox conditions in the sediments allows the plant to maintain normal physiologi cal nutrient uptake and photosynthetic capacity (Koch 2001 and references within; Eldridge et al. 2004). Sulfide concentrations (Hebert and Mors e 2003) and sulfate reduction ra tes (Blaabjerg et al. 1998) showed diel cycles with greater H 2 S concentrations during the ni ght and, surprisingly, higher sulfide production rates during th e day. High sulfate reduction ra tes during periods of maximum photosynthesis suggests that exces s carbon produced by primary production is released as DOM from the roots and stimulates sulfate reduction. Simultaneously, lacunal O 2 oxidizes this

released sulfide. Radioa ctive and stable isotope data from se veral studies suggests that there is a strong linkage between seagrass producti on and DOM release (Holmer and Laursen 2002; Kaldy et al. 2006). Using tracer experi ments, Kaldy et al. (2006) showed a direct link from seagrass to DOM, and then to sediment bacteria, and Ho lmer and Laursen (2002) showed a positive relationship between seagrass photosynthesis and sulf ate reduction. A conclusion we reach from these result, is that diel vari ations in seagrass photosynthe sis produce pulses of DOM exudates from the seagrass roots and rhizomes that simulate daily cycles in sediment sulfate reduction.

The interaction between seagrass production a nd organic matter in the sediment is highly variable, and there are studies showing that healt hy seagrass can occur in highly enriched organic sediments (Koch 2001) (Table 6.1). We assume th at either the organic ma tter in these sediments is relatively unreactive or that infaunal i rrigation (as discussed earlier) and lacunal O 2 release aerates the sediments in the vicinity of the seagrass. We note, however, that both infaunal irrigation and lacunal O 2 release also involves ex cretion 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). De nnison et al. (1987) established an upper limit for this uptake and utilization, finding in mesocosm experiments that interstitial water concentrations of NH 4 +

above 100 F M saturate the growth response of Z. marina. In contrast, Van Katwijk et al. (1997) reported that an NH 4 + concentration of 125 F M was toxic to eelgrass, and suggested that the toxicity threshold was as low as 25 F M. However, Hebert et al. (2007) reported up to 2500 F M

concentration of NH 4 + 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-g rained sediments into the near shore waters from sandy beaches in Massac husetts. The nitrate-rich groundwater induced6.11 nitrate reductase activity in eelgr ass plants. This finding suggests the importance of high-porosity sandy sediments as a conduit for nutrient-rich gr ound water to near shore aquatic vegetation, including seagrass and macroalgae. The interlayering of permeable sandy sediments with finegrained sediments in a Georgia estuary provided conduits for advective transport of pore water constituents out of the sediments (Jahnke et al . 2003). They concluded that such fluxes are concentrated into small layers, and as such may constitute a significant i nput of nutrients to the estuary even if permeable, sandy layers compri se a very small proporti on of the total bottom area. Again, these findings indicate a relationshi p between sediment grai n 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

NH 4 + in interstitial water and onto sediment partic les. He found that in highly-reduced eelgrass sediments, NH 4 + is lost from the interstitial pool by diffusion into the thin, oxidized sediment surface layer, by adsorption onto sediment partic les, and by uptake into bacterial cells and eelgrass roots. However, part of this pore water NH 4 + is termed “exchangeable” because it is easily released from the particle surfaces by i on exchange. Short noted that the amount of exchangeable NH 4 + 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 m uddy sand containing substantial organic matter. This type of sediment can cont ain elevated pore water concentra tions of substances such as NH 4 +

or dissolved sulfides. Whether or not these cons tituents 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 distribu tion 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 he nce to the sulfide concentrations in the sediment. The ability of seagrass to protect itself from high levels of su lfide in the root zone wi ll be directly dependent on availability of light to driv e 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 a nd 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 evalua tion factor in seagrass pr otective criteria. The reported sediment pore wate r concentrations of NH 4 + tolerated by Z. marina range over two orders of magnitude. Again a better understanding of the relationship between dissolved inorganic nutrients and Z. marina physiology is needed.6.12 6.4 Literature Cited

Abal, E. G.,N. R. Loneragan, P. Bowen, C. J. Perry,, J. W. Udy, and W. C. Dennison. 1994. Physiological and morphological responses of the seagrass Zostera capricorni Ashers. to light intensity. Journal of Experiment al Marine Biology and Ecology 178: 113-129. Ackerman, J. D. and A. Okubo. 1993. Redu ced mixing in a marine macrophyte canopy. Functional Ecology 7:305-309. Agawin, N. S. R and C. M. Duarte. 2002. Evid ence 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 hydroge n sulfide toxicity. CRC Critical Reviews in Toxicology 13:25-97. Berner, R. A. 1980. Early Diagenesis: A Theore tical Approach. Princeto n University Press. Princeton, N.J. Binzer, T., J. Borum, and O. Pedersen. 2005. Flow velocity affects internal oxygen conditions in the seagrass Cymodocea nodosa. Aquatic Botany, 83: 239-247. Blaabjerg, V., K. N. Mouritsen, and K. Finster. 1998. Diel cycles of sulphate reduction rates in

sediments of a Zostera marina bed (Denmark). Aquatic Microbial Ecology 15: 97-102.Bostrom C., E. Bonsdorff, P. Kangas, and A. Norkko. 2002. Long-term changes of a brackishwater eelgrass ( Zostera marina L.) community indicate effects of coastal eutrophication.

Estuarine Coastal and Shelf Science 55:795-804. Brenchley, J. L. and R. J. Probert. 1998. S eed germination responses to some environmental factors in the seagrass Zostera capricorni from eastern Australia. Aquatic Botany 63:177-

188. Burrell, D. C. and J. R. Schubel. 1977. Seagrass ecosystem oceanography. Pp. 195-232. In: C. P. McRoy and C. Helfferich (eds.), Seag rass Ecosystems. Marcel Dekker, New York,

NY. Caffrey, J. M. and W. M. Kemp. 1991. Seas onal and spatial patterns of oxygen production, respiration and root-r hizome release in Potamogeton perfoliatus L. and Zostera marina L. Aquatic Botany 40:109-128. Carlson, P. R., Jr., L. A. Yarbro, and T. R. Ba rber. 1994. Relationship of sediment sulfide to mortality of Thalassia testudinum in Florida Bay. Bulleti n of Marine Science 54:733-

746. Dan, A, A. Moriguchi, K. Mitsuhashi, a nd T. Terawaki. 1998. Relationship between Zostera

marina and bottom sediments, wave action offshor e in Naruto, Southern Japan. Fisheries Engineering 34: 229-204. Dennison, W. C., R. C. Aller, and R. S. Albert e. 1987. Sediment ammonium availability and eelgrass ( Zostera marina ) growth. Marine Biology 94:469-477. Edgar, G. .J. and C. Shaw. 1995. The production and trophic ecology of shallow-water fish assemblages in southern Australia. III. General relationships between sediments, seagrasses, invertebrates and fishes. J ournal of Experimental Marine Biology and Ecology 194:107-131.6.13 Eldridge, P. M., J. E. Kaldy, and A. Burd. 2004. Stress response model for the tropical seagrass

Thalassia testudinum : The interactions of light, te mperature, sedimentation, and geochemistry. Estuaries, 6: 923-937. Eldridge, P. M. and J. W. Morse. 2000. A di agenetic model for sediment-seagrass interactions, Marine Chemistry 70:89-103. Erskine, J. M. and M. S. Koch. 2000. Sulfide effects on Thalassia testudinum carbon balance and adenylate energy charge. Aquatic Botany 67:275-285. Fonseca, M. S. and S. S. Bell. 1998. Influen ce of physical setting on seagrass landscapes near Beaufort, North Carolina, USA. Marine Ecology Progress Series 171:109-121. Fonseca, M. S., J. C. Zieman, G. W. Thayer, and J. S. Fisher. 1983. The role of current velocity in structuring eelgrass ( Zostera marina L.) meadows. Estuarine, Coastal and Shelf

Science 17:367-380. Fonseca, M. S. 1981. The interaction of a seagrass, Zostera marina L. with current flow. M.Sc. Thesis, University of Virginia, Charlottesville, VA, pp. 35. Frost, M. T., A. A. Rowden, and M. J. Attrill. 1999. Effect of habitat fragmentation on the

macroinvertebrate infaunal communiti es associated with the seagrass Zostera marina L. Aquatic Conservation: Marine and Freshwater Ecosystems 9:255-263. Gambi, M. C., A. R. M. Nowell, and P. A. Ju mars. 1990. Flume observations on flow dynamics

in Zostera marina (eelgrass) beds. Marine Ec ology Progress Series 61:159-169. Ginsburg, R. N. and H. A. Lowenstam. 1958. The influence of marine bottom communities on the depositional environment of sedime nts. Journal of Geology 66:310-318. Goodman, J. L., K. A. Moore, and W. C. Denni son. 1995. Photosynthetic responses of eelgrass ( Zostera marina L.) to light and sediment sulfide in a shallow barrier island lagoon, Aquatic Botany 50:37-47. Grady, J. R. 1981 Properties of sea grass and sand flat sediments from the intertidal zone of St. Andrew Bay, Florida. Estuaries 4:335-344. Harwell, M. C. and R. J. Orth. 1999. Eelgrass ( Zostera marina L.) seed protection for field experiments and implications for large-sc ale restoration. Aquatic Botany, 64:51-61. Hebert, A. B., and J. W. Morse. 200 3. Microscale effects of light on H 2 S and Fe 2+ in vegetated ( Zostera marina ) sediments. Marine Chemistry 81:1-9. Hebert, A. B., J. W. Morse, and P. M. Eldr idge. 2007. Small-scal e heterogeneity in the geochemistry of seagrass vegetated and non- vegetated estuarine sediments: causes and consequences. Aquatic Geochemistry, 13:19-39. Heiss, W. M., A. M. Smith, and P. K. Probert. 2000. Influence of the small intertidal seagrass

Zostera novazelandica on linear water flow and sediment texture. New Zealand Journal

of Marine and Freshwat er Research 34:689-694. Hemminga, M. A. and C. M. Duarte. 2000. S eagrass Ecology. Cambridge University Press, Cambridge, UK. pp. 298. Höberg, P. and P. Jensén. 1994. Aluminum and uptak e of base cations by tr ee roots: a critique of the model proposed by Sverdrup et al. Wa ter Air and Soil Pollution 75:121-125. Holmer, M. and L. Laursen. 2002. Effect of shading of Zostera marina (eelgrass) on sulfur cycling in sediments with c ontrasting organic matter and sulfide pools. Journal of Experimental Marine Bi ology and Ecology 270:25-37.6.14 Holmer, M. and E. J. Bondgaard. 2001. Photos ynthetic and growth resp onse of eelgrass to low oxygen and high sulfide concentrations duri ng hypoxic events. A quatic Botany 70:29-

38. Hoskin, C.M. 1983. Sediment in seagrasses near Link Port, Indian River, Florida. Florida

Scientist. 46:153-161. Jahnke, R. A., C. R. Alexander, and J. E. Kostka . 2003. Advective pore wa ter input of nutrients to the Satilla River estuary, Georgia, USA. Estuarine, Coastal and Shelf Science 56:641-

653. Jones,W.B., L. A. Cifuentes, and J. E. Kaldy, 2003. Stable carbon isotope evidence for coupling between sedimentary bacteria and seagrasses in a sub-tropical lagoon. Marine Ecology Progress Series 255:15-25. Kalbfleisch, W. B. C. and B. Jones. 1998. Sedimentology of shallow, hurricane-affected lagoons: Grand Cayman, British West Indies. Journal of Coastal Research 14:140-160. Kaldy, J. E., P. M. Eldridge, L. A. Cifuentes, and W. B. Jones. 2006. Utilization of DOC from Seagrass Rhizomes by Sediment Bacteria: 13 C Tracer Experiments and Modeling. Marine Ecology Progress Series 317:41-55. Kaldy, J. E. and K. H. Dunston. 2000. Above- and below-ground production, biomass and reproductive ecology of Thalassia testudinum (turtle grass) in a subtropical coastal lagoon. Marine Ecology Progress Series 193:271-283. Kenworthy, W. J. and M. S. Fonseca. 1992. The use of fertilizer to enhance growth of transplanted seagrasses Zostera marina L. and Halodule wrightii Aschers. Journal of Experimental Biology and Ecology 163:141-161. Kenworthy, W. J., J. C. Zieman, and G. W. Thayer. 1982. Evidence for the influence of seagrasses on the benthic nitrogen cycle in a coastal plain estuary near Beaufort, North Carolina (USA). O ecologia 54:152-158. Koch, E. W. 2001. Beyond light: physical, geolog ical, and geochemical parameters as possible submersed aquatic vegetation habita t requirements. Estuaries 24:1-17. Koch, E. W. 1999. Preliminary evidence on the in terdependent effect of currents and porewater geochemistry on Thalassia testudinum Banks ex Konig seedlings. Aquatic Botany 63: 95-

102. Koch, M. S. and J. M. Erskine. 2001. Su lfide as a phytotoxin to the tropical seagrass Thalassia

testudinum : interactions with light, salinity and temperature. Journal of Experimental Marine Biology and Ecology 266:81-95. Koch, M. S., S. Schopmeyer, C. Kyhn-Hansen, a nd C. J. Madden. 2007. Synergistic effects of high temperature and sulfide on tropical seag rass. Journal of E xperimental Marine Biology and Ecology 241: 91-101. Kraemer, G. P. and R. S. Alberte. 1993. Ag e related patterns of metabolism and biomass in

subterranean tissues of Zostera marina . Marine Ecology Prog ress Series 95:193-203. Krumbein, W. C. and F. J. Pettijohn. 1938. Principl es and methods of size analysis, in Manual of Sedimentary Petrography, K.F. Mather Ed., A ppleton-Century-Crofts, New York. pp. 91. Kuenen, M. M. C. E. and A. O. Debrot. 1995 A quantitative study of the seagrass and algal meadows of the Spaanse Water, Curaçao, The Netherlands Antilles. Aquatic Botany

51:291-310.6.15 Lalumière, R., D. Messieor. J.-J. Fournier, and C. P. McRoy. 1994. Eelgrass meadows in a low arctic environment, the northeast coast of James Bay, Quèbec. Aquatic Botany 47:303-

315. Lee, K. S. and K. H. Dunton. 1996. Producti on and carbon reserve dynami cs of the seagrass Thalassia testudinum in Corpus Christi Ba y, Texas, USA. Marine Ecology Progress

Series 143:201-210. Lee, S. Y., C. J. Kwon, K. S. Lee, and C. Choi. 2002. Distribution of eelgrass, Zostera marina L.on coasts of the Korean Peninsula: Preliminary study for eelgrass restoration. Ocean

and Polar Research 24:55-61. Livingston, R. L., S. E. McGlynn, and X. Niu. 1998. Factors controlling seagrass growth in a gulf coastal system: Water and sediment qual ity and light. Aquatic Botany 60:135-159. Maier, C. M. and A. M. Pregnall. 1990. In creased macrophyte nitrate reductase activity as a consequence of groundwater input of nitr ate through sandy beaches. Marine Biology

107:263-271. Marshall, N. and K. Lukas. 1970. Prelim inary observations on the properties of bottom sediments with and without eelgrass, Zostera marina , cover. Proceedings of the National

Shellfish Association. 60:107-111. Mattila, J., G. Chaplin, M. R. Eilers , K. L. Heck, J. P. O’Neal, and J. F. Valentine. 1999. Spatial

and diurnal distribution of inve rtebrate and fish fauna of a Zostera marina bed and nearby unvegetated sediments in Damariscotta River, Maine (USA). Journa l of Sea Research

41:321-332. McGlathery, K. J., R. Marino, a nd R. W. Howarth. 1994 Variab le rates of phosphate uptake by

shallow marine carbonate sediments: M echanisms and ecological significance. Biogeochemistry 25:127-146. Moore, K. A., R. J. Orth, and J. F. Norwak. 1993. Environmental regulation of seed germination in Zostera marina L. (eelgrass) in Chesapeake Bay: effects of light, oxygen and sediment burial. Aquatic Botany 45:79-91. Nepf, H. M. and E. W. Koch. 1999. Vertical s econdary flows in submersed plant-like arrays. Limnology and Oceanography 44:1072-1080. Nienhuis, H. and B. H. H. DeBree. 1977. Production and ecology of eelgrass ( Zostera marina L.) in the Grevelingen estuary, The Netherla nds, before and after the closure. 1977. Hydrobiologia 52:55-66. Orth, R. J. 1977. The importance of sediment stability in seagrass communities. pp. 281-300. In: B. C. Coull (ed.), Ecology of Marine Be nthos. University of South Carolina Press, Columbia, SC. Paula, J., P. Fidalgo e Costa, A. Martins, and D. Gove. 2001. Patterns of abundance of seagrasses and associated infaunal communities at Inhaca Island, Mozambique.

Estuarine Coastal and Shelf Science 53:307-318. Penhale, P. A. and R. G. Wetzel. 1983. St ructural and functional adaptations of eelgrass ( Zostera marina L.) to the anaerobic sediment envir onment. Canadian Journal of Botany

61:1421-1428. Pedersen, M. F., C. M. Duarte, and J. Cebrian. 1997. Rates of changes in organic matter and nutrient stocks during seagrass colonizati on and stand development. Marine Ecology Progress Series 159:29-36.6.16 Pedersen, O., T. Binzer, and J. Borum. 2004. Sulphide intrusion in eelgrass ( Zostera marina L. ). Plant, Cell and Environment 27:595-602. Percival, J. B. and P. J. Lindsay. 1997. Measur ement of physical properties of sediments. Pp. 7- 44. In: A. Mudroch, J.M. Azcue, and P. Mudroch Eds. Manual of Physical-Chemical Analysis of Aquatic Sediments. Lewis Publishers. New York. Perez, M., J. Romero, C. M. Duarte, and K. Sand-Jensen. 1991. Phosphorus limitation of Cymodocea nodosa growth. Marine Biology 109:129-133. Perez, M., C. M. Duarte, J. Romero, K. Sand-Je nsen, and T. Alcoverro. 1994. Growth plasticity

in Cymodocea nodosa stands: The importance of nutri ent supply. Aquatic Botany 47:

249-264. Peterson, C. H., H. C. Summerson, and P. B. Du ncan. 1984. The influence of seagrass cover on population structure and indivi dual growth rate of a suspension-feeding bivalve,

Mercenaria mercenaria . Journal of Marine Research 42:123-138. Peterson, B. J. and K. L. Heck, Jr. 2001. Positive interactions between suspension-feeding bivalves and seagrass - a facultative mu tualism. Marine Ecology Progress Series.

213:143-155. Phillips, R. C. 1984. The ecology of eelgrass meadows in the Pacific Northwest: a community profile. U.S. Fish and Wildlife Se rvice, FWS/OBS-84/24 (repr.). pp. 8585. Posey, M. H. 1988. Community changes associat ed with the spread of an introduced seagrass, Zostera japonica. Ecology 69:974-983. Scoffin, T. P. 1970 The trapping and binding of subtidal carbonate sediments by marine vegetation in Bimini Lagoon, Bahamas. J ournal of Sedimentary Petrology 40:249-273. Short, F. T. 1983. The response of interstitial ammoni um in eelgrass ( Zostera marina L.) beds to environmental perturbations. Journal of Experimental Mari ne Biology and Ecology

68:195-208. Short, F. T. 1987. Effects of sediment nutrients on seagrasses: Literature review and mesocosm experiment. Aquatic Botany 27:41-57. Smith, R. D., A. M. Pregnall, and R. S. Albe rte. 1988. Effects of anaerobiosis on root

metabolism of Zostera marina (eelgrass); implications for survival in reducing sediments. Marine Biology 98:131-141. Sprung, M. 1994. Macrobenthic secondary producti on in the intertidal zo ne of the Ria Formosa – a lagoon in Southern Portugal. 1994. Es tuarine Coastal and Shelf Science 38:539-558. Stumm, W. and J. J. Morgan. 1981. Aquatic Chemistry: an introduction emphasizing chemical equilibria in natural waters, 2nd Ed ition, John Wiley & Sons, New York. Terrados, J., C. M. Duarte, M. D. Fortes, J. Boru m, N. S. R. Agawin, S. Bach, U. Thampanya, L. Kamp-Nielsen, W. J. Kenworthy, O. Geertz-Hansen, and J. Vermaar. 1998. Changes in community structure and biomass of seagrass communities along gradients of siltation in

SE Asia. Estuarine Coastal and Shelf Science 46:757-768. Terrados, J., C. M. Duarte, L. Kamp-Nielsen, N. S. R. Agawin, E. Gacia, D. Lacap, M. D. Fortes, J. Borum, M. Lubanski, and T. Greve. 1999. Are seagrass growth and survival constrained by the reducing conditions of the sediment? Aquatic Botany 65:175-197. Thayer, G. W., W. J. Kenworthy, and M. S. F onseca. 1984. The ecology of eelgrass meadows of the Atlantic coast: A community profile. US Fish and Wildlife Service, FWS/OBS84/02. pp. 147.6.17 Thom, R., A. Borde, S. Rumrill, D. L. Woodruff, G. D. Williams, J. Southard, and S. L. Blanton. 2001. Factors influencing the spatial and annual variability in eelgrass ( Zostera marina L.) meadows in Pacific Northwest, USA, System s. pp. 38-51. In: J. K. Parrish and K. Litle (eds.), PNCERS 2001 Annual Repor t. Coastal Oceans Program, National Oceanographic and Atmospheric Administration. Townsend, E. C. and M. S. Fonseca. 1998. Bi oturbation as a potential mechanism influencing spatial heterogeneity of North Carolina seag rass beds. Marine Ecology Progress Series

169:123-132. van Katwijk, M. M., L. H. T. Vergeer, G. H. W. Schmitz, and J. G. M. Roelofis. 1997. Ammonium toxicity in eelgrass Zostera marina . Marine Ecology Progress Series

157:159-173. van Lent, F., P. H. Neinhuis, and J. M. Ve rschuure. 1991. Production and biomass of the

seagrasses Zostera noltii Hornem, and Cymodocea nodosa (Ucria) Aschers. at the Banc d’Arguin (Mauritania, NW Africa): a prelim inary approach. Aquatic Botany 41:353-367. Wanless, H. R. 1981 Fining-upwards sediment ary sequences generated in seagrass beds. Journal of Sedimentary Petrology 51:445-454. Webster, P. J., A. A. Rowden, and M. J. Attril l. 1998. Effect of shoot density on the infaunal macro-invertebrate community within a Zostera marina seagrass bed. Estuarine, Coastal

and Shelf Science 47:351-357 Welsh, D. T., G. Castadelli, M. Bartoli, D. Poli, M. Careri, R. de Wit, and P. Viaroli. 2001. Denitrification in an intertidal seagrass meadow, a comparison of 15 N-isotope and acetylene-block techniques: dissimilatory n itrate reduction to ammonia as a source of

N 2 O? Marine Biology 139:1029-1036. Wentworth, C. K. 1922. A scale of grade and cl ass terms for clastic sediments. Journal of Geology, 30: 277 Wood, E. J. F., W. E. Odum, a nd J. C. Zieman. 1969. Influence of seagrasses on the productivity of coastal lagoons. pp. 495-502. In: Ayala Castan ares, A. and F.B. Phelger (eds.). Coastal lagoons. Universidad Nacional Autonoma de Mexico, Ciudad Universitaria, Mexico,

D.F. Zimmerman, R. C., R. D. Smith, and R. S. Albe rte. 1987. Is growth of eelgrass nitrogen

limited? A numerical simulation of the eff ects of light and nitr ogen on the growth dynamics of Zostera marina . Marine Ecology Progress Series 41:167-176.

Sources of Unpublished Data Boese, B. and D. Young. in prep. Pacific Coastal Ecology Branch, We stern Ecology Division,

U.S. EPA.7.1 

## 7.0 The Interaction of Epiphytes wi th Seagrasses under Nutrient

Enrichment

Walter G. Nelson

7.1 Background

The surfaces of seagrass blades are often c overed by a variable layer of epiphytes, which may respond in complex ways to variation in environmental nutrient co ncentrations and light availability. This chapter reviews the obser vational and experimental literature through approximately 2004 on the effects of nutrient enri chment on seagrass epiphytes. The goal was to

determine whether there is an adequate understanding of the e ffects 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 a nd Humm 1975; Hall and Eiseman 1981; Hall 1988), microalgae (e.g. Sand-Jensen 1977; Sand-Jensen and Borum 1984), and a variety of sessile animal groups including polychaetes, bryozoans, hydroids and tunicates (Kita and Harada 1962; Nagle 1968; Lewis and Hollingworth 1982). The term "epiphytes" will be used to refer to the combined microalgal, macroalgal, animal and inorganic components covering seagrass blades. Major reviews of seagrass epiphytes include Ha rlin (1980) and Borowitzka and Lethbridge (1989), while van Montfrans et al . (1984), Orth and van Montfran s (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 bene ficial 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 avai lable for photosynthesis, 2) a reduction in the rate of diffusion of materials such as CO 2 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 w ith heavy epiphyte cover may beco me more brittle and break off (Borowitzka and Lethbridge 1989), although quantitative data supporti ng 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 pl ants in the upper intertidal zone (Penhale and Smith 1977, and see Chapter 11).

Epiphytes are patchily distri buted on seagrass blades (Figur e 7.1), and are typically more abundant on the distal portions of all blades (F igure 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 Lethbrid ge 1989). Rates of blade turnover for7.2 seagrass can vary widely across seasons, and for Z. marina may range from 50-70 days in summer to a maximum leaf age of 200 days in wint er (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 Kir kpatrick (1995) observed that li ght attenuation by epiphytes was highest during winter months in Sarasota Bay wh en leaf turnover rates were reduced. However, at higher latitudes where winter light may be limiting, winter ma y 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 epi phytes, Borowitzka and Lethbridge (1989) concluded that there is little evidence that seagrasses have any means of inhibiting epiphyte colo nization and growth.

Borowitzka and Lethbridge (1989) reviewed the evidence that transl ocation 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 nut rient release reported from seagrasses are too low to support levels of epiphyte growth genera lly observed, so the major source of nitrogen and phosphorus appears to be from the water colu mn (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 cove r appears to be decr eased 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-Jens en and Revsbach 1987). Epiphytic shading has been suggested to be particularly important at lower ambient light levels (Morgan and Kitting 1984; Twilley et al. 1985; Wetzel and Neckles 1986).

Cambridge et al. (1986), working with seag rasses from Cockburn Sound, Australia, provided one of the first suggestions that high epiphytic loads resulting from eutrophic conditions directly causes loss of seagra sses. Transplantation experiments with Posidonia

sinuosa seedlings resulted in leaf area ~60% lo wer for plants growing in Cockburn Sound after 48 d. Seedlings were heavily covered with macr oepiphytes. 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 st udy noted but did not quant ify the presence of macroalgal blankets up to 1 m thick in Co ckburn Sound, and the relative role of these macroalgae versus epiphytes in seagrass loss is unclear.7.3 Figure 7.1. Schematic diagram of seagrass plan t illustrating typical patterns of distribution of epiphytes within and among blades. Epiphyte co ver increases from base to tip and from youngest to oldest blades.

For the freshwater macrophyte ( Littorella uniflora ) in Danish lakes, Sand-Jensen (1990) found a relation of epiphytic load, nutrient lo ad, and depth distribution. Maximum depth of macrophyte distribution was inversel y related to epiphytic load, a nd 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 acc ounted 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 th at 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 he terogeneity of epiphyte distribution on blades of different ages, average light reduction to a seag rass plant is more typically 50-60% (Harden 1994). Annual averages of light reduction, resul ting from seasonal variation in epiphytic load,7.4 Figure 7.2. Horizontal arrows indicate external factors in fluencing abundance of seagrass epiphytes. The vertical arrow indicates the variation in impor tant 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 199 5; Dixon 2000). Monthly averages of light reduction ranged from 17-59%, while shallower site s where light was not limiting tended to have greater light attenuation by epiphytes (Dixon 2000).

A review of seagrass light attenuation studies (Brush and Nixon 2002), suggests that

studies which have estimated li ght attenuation using epiphyte su spensions (see Table 7.1) may overestimate the attenuation of light for high ep iphytic loads. They observed that highest epiphytic biomass tended to be ge nerated by larger arborescent al gae which tended to float away from the seagrass blade when submersed, thus a llowing more light to reach the blade than would be predicted by the suspension approach.

There are few direct measurements of the e ffect of epiphytes on photosynthesis rates. Sand-Jensen (1977) demonstrated that diatom epiphytes reduced the pho tosynthetic 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 proportionally7.5 greater (58% vs. 31%) at lower light intensities, again suggesting the impact of epiphytes may be most severe when ambient light is low.

Using in situ incubations of clipped blades of Halodule wrightii , Morgan and Kitting (1984) examined the relative productivity of the s eagrass and its epiphytes over a range of light intensities. At light levels ≥ 50% of surface illumination, the ep iphytic 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, al though 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 e ffect at high light levels appears tenuous.

Extrapolation of results from productivity m easurements 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 shadi ng on seagrass survival is likely to be determined by complex interactions. For example, Neverauskas (1988) experimentally reduced light to Posidonia spp. ( P. sinuosa and P. angustifolia ) by 50% using shade cloth over small experimental quadrats (0.0625 m 2 ) with rhizome connections severed. Epiphyte biomass

decreased from an initial va lue 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 leve ls had greatly reduced epiphytic loads.

In spite of multiple uncertainties, the guida nce document for management of seagrasses for Chesapeake Bay (U.S. EPA 2003) incorporat es the concept of a percent of surface

illumination available at the leaf surface to a ccount 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 persis tence of seagrass in th e face of combined light attenuation in the water column, and that due to epiphytes. P LL values are calculated by the

formula:

PLL = 100 [e – (K d )(Z) ][e – (K

e ) (B

e ) ]

where K d is the light extinction coefficient in the water column, Z is the depth, K

e is the epiphyte biomass specific PAR attenuation coefficient, and B

e is the epiphyte biomass per unit seagrass biomass. Specific equations for determination of K

e and B

e 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), indica ted that the additiona l reduction in light intensity at the leaf surface due to epiphytes was 20-60 % in th e 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 calc ulation 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.7.6Table 7.1. Studies of light attenuation by epiphytes in coastalmarine 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 recordedinlight absorbance measurements, and may differfrom that reported by Brush and Nixon (2002). A: artificiallight; N: natural light.MacrophyteStudy area (Reference)Method LightsourceMax epiphytedensity(mg cm-2leaf)Max % lightabsorbance(mean)Comments, ReferenceZostera marinaØresund, Denmark SuspensionA 1.9 (DW)99.7 Borum & Wium-Andersen 1980Øresund, Denmark SuspensionA 9.5 (DW)99.7 Same absorbance data as 1980; Borum et al. 1984Roskilde, Denmark SuspensionA .0039 (chl a)76 May underestimate total absorption, Borum 1987Mesocosms, VA SuspensionA 20(DW)Measured spectral absorption, Neckles (1993 andunpubl. data), cited in Brush & Nixon 2002Mesocosms, RIScraped and unscrapedleavesN99 (DW)90 Effect varied with type of epiphyte, Brush & Nixon2002Elkhorn Slough, CA Scraped and unscrapedleavesA 60 Measured spectral absorption, underestimatedepiphyte biomass by use of phospholipid measureonly; Drake et al. 2003Thalassia testudinumTerminos Lagoon, Mexico SuspensionN17 (DW)Kemp et al. (1988), cited in Brush & Nixon 2002Tampa Bay, FLSuspensionA81(51) Excluded necrotic blade tips, underestimates totalabsorbance; Dixon & Kirkpatrick 1995Indian River Lagoon, FL Scraped and unscrapedleavesA 18 (DW) 85 (61) Harden 1994BahamasScraped and unscrapedleavesA36 Measured spectral absorption, underestimatedepiphyte biomass by use of phospholipid measureonly; Drake et al. 2003Halodule wrightiiTampa Bay, FLSuspensionA99(48) Excluded necrotic blade tips, underestimates totalabsorbance; Dixon & Kirkpatrick 1995Indian River Lagoon, FL Scraped and unscrapedleavesA 30 (DW) 90 (60) Harden 19947.7Syringodium filiformeIndian River Lagoon, FL Scraped and unscrapedleavesA 22 (DW) 97 (61) Harden 1994Posidonia oceanicaMediterranean Sea, Spain Collection on GF/F filters A71 (DW)95(~70) Measuredonlypigmenteffect, underestimated totallight absorbance, Cebrián et al. 1999Heterozostera tasmanicaVictoria, AustraliaScraped and unscrapedleavesA 3.3 (DW) 75 Bulthuis & Woelkerling 1983Artificial substratesCockburn Sound, Australia Growth on glass slidesA 0.004 (chl a) 96(63) Silberstein et al. 1986Success Bank, Australia Growth on plastic slides A 4.2 (AFDW) Burt etal. (1995), cited in Brush & Nixon 2002Banc d’Arguin, Mauritania Growth on glass slides N3(DW)Hootsmans et al. (1993), cited in Brush & Nixon2002Delaware Inland Bays Scraped and unscrapedmylar stripsN 17 (DW) Glazer (1999), cited in Brush & Nixon 2002Virginia, Maryland Bays Scraped and unscrapedplastic stripsA 4 (DW) 96 Measured spectral absorption, Brandt & Koch 20037.8 7.3 Role of Nutrient Loads in Determining Epiphytic Load

7.3.1. Field Observations

There are a number of correlative studies of field distribution of epiphyte load versus nutrient concentrations which sugge st nutrient enrichment increases epiphyte loading. Orth and Van Montfrans (1984) reviewed a number of studies from freshw ater systems that indicated a

relation between nutrient enrichme nt, epiphyte increases, and d ecrease or loss of macrophytes. Borum (1985) observed that th e biomass of epiphytes on Zostera marina increased exponentially with increasing total N concentration in the wa ter 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 initia ted, while partially affected areas at greater

distances from the outfalls showed high epiphytic loads. Experime nts placing artificial substrata along a distance gradient from the outfall showed highest epiphytic r ecruitment at sites closest to

outfall.

Examination of accumulation rates of epiphytes on Heterozostera tasmanica in Victoria, Australia showed that highest ep iphyte 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 dimi nish light below the compensation point within 36 days, about half the mean life span of the s eagrass leaves. However, the expression of impacts on the seagrass tended to be site specific.

Silberstein et al. ( 1986) compared characteri stics of the seagrass Posidonia australis and its epiphytes between two sites, on e of which was near a sewage di spersal line. Epiphytic loads, measured as chlorophyll per unit l eaf area, were higher at the sewa ge 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 sewa ge, 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 Baha mas, Jensen and Gibson (1986) found that the highest epiphytic biomass were at sites with the highest concentra tions of phosphorus and silicate. Both Tomasko and Lapointe (1991) and Lapointe et al. (1994) al so 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 po pulated island with septic ta nks and a bird rookery island, and were associated with highe r 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 hypereutrophi c, eutrophic, mesotrophic, or oligotrophic7.9 corresponding to total N concentrations in winter of 38.8, 30.3, 21.6, and 12.8 μM, respectively. Seagrasses from the oligotrophic zone typically had lowest epiphyte levels, while those in the hypereutrophic and eutrophic zones had high levels of epiphytes and mat-forming macroalgae, and low shoot densities and productivity.

Epiphyte loads on Thalassia testudinum were also measured along nutrient gradients of differing scales in Florida Bay by Frankovich and Fourquean (1997). At a fine scale, they found significantly higher ep iphyte loads nearest a bird rookery island, with the enhancement effect decreasing at between 15 and 30 m from the isla nd. 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 nutri ent enrichment. They s uggested that epiphytes played no role in generating the Florida Bay seag rass die off, but were instead stimulated by nutrients released by dying seagrass. They conc luded 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 th e expression of epiphyte standing stock in response to nutrients loads, Tomasko et al . (1996) found that alt hough the greates t epiphyte biomass at sites within Sarasota Bay, Florida wa s found on one date at the site with highest total nitrogen loading, the general pattern of epiphyte biomass showed little rela tion 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 importan ce of spatial variability and also emphasize the complexity of interpretations possible from fi eld data (Hanisak 2001). In spite of spatial differences in nutrient concentrations at study site s, 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 si ngle mean per site, above-ground seagrass biomass

at a site decreased as a functi on 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 s eagrass biomass increase s the amount of grazers which leads to reduced epiphytes.

The number of in situ experimental nutrie nt additions in seagrass beds was relatively limited until recently. Harlin and Thorne-Mil ler (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) amm onium enrichment experime nts of both water and

sediments in Z. marina beds in Puget Sound, WA. There wa s no significant increase in epiphyte7.10 biomass, and in fact epiphyte biomass was significantly reduced by water column ammonium

enrichment.

7.3.2 Mesocosm nutrient addition experiments Conditions of various mesocosm experime nts 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 me socosm size, flow rates, type of nutrient addition (pulse versus continuous addition), a nd presence of absence of grazers within experimental systems make generalizations difficult.

Large mesocosm ponds were used by Twilley et al. (1985) to expose Potamogeton perfoliatus and Ruppia maritima to pulse additions of nitrogen and phosphorus (Table 7.2). Concentrations of N greater than 60 μM caused declines in the va scular plants in the ponds in 6- 10 weeks. At the highest nutrient addition levels, phytoplankton blooms in the ponds were evident. Dense epiphytes devel oped on plants in all nutrient addition treatments. Epiphytic

cover was estimated to decrease >80 % of li ght 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, grow th rates were reduced by 50%.

Using mesocosm tanks of 800 l capacity with high water turnover ra tes (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 limitati on. With full light levels, large increases of phytoplankton, macroalgae and epiphyt es occurred relative to contro ls. 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 re duced light treatment significan tly 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 us ed pulse additions of nutrients to examine responses of Zostera marina . The experiment reported by Lin et al. (1996) had limited water7.11 turnover (5% per day), while all ta nks included a variety of fish and invertebrates, but without significant numbers of grazers. There was no incr ease in epiphyte biomass on eelgrass leaves in any treatment. The lack of epiphyte response wa s 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, wh ich is consistent with other mesocosm study results (Short et al., 1995; Moore and Wetzel, 200 0). The authors concluded that elevated nutrient levels are not necessarily pr edictive of an epiphyte increase.

Taylor et al. (1999) used th e 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 we re observed on epiphytes, Z. marina , or drift algae in the mesocosms. In these experiments, a brown tide bloom occurre d 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 contin uous addition of nutrien ts, 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) us ing mesocosms of 1570 l with pulse additions

of nutrients and limited water tur nover showed no effect of nutr ient additions on epiphytic loads

on Z. marina , measured as cell counts rath er than biomass. Grazer densities in both nutrient additions and controls tended to be quite high in these experiments, and may explain the lack of epiphyte response.

Moore and Wetzel (2000) used replicated 110 l aquaria to test effects of elevated nutrients and reduced light on Zostera marina . The experiment used 16 turnovers per day with

continuous flow nutrient additions , and tanks included moderate densities of gastropod grazers. Epiphyte responses were highly depended on treatme nt, 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 respon se to an epiphyte load of ~ 16 g g -1 of eelgrass. The experimental design did not examine the res ponse 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 Lapoi nte 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 dail y 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 rhizom e growth rates with nutrient additions in both7.12Table 7.2. Summary of laboratory microcosm and mesocosm nitrogen and phosphorusenrichment assays using seagrasses.Treatment Abbreviations: C = control (ambient conditions), +N = NO3addition, +P = PO4addition, +NP = NO3 and PO4addition, +A= NH4addition, +AN = NH4and NO3addition, +AP = NH4and PO4addition, +NAP = NO3and NH4and PO4addition, +G = grazeraddition, D = depleted NH4, ? = unable to determine from information in reference. All ratios presented in Treatment Variables arerelative nutrient concentrations.Species:LocationTreatmentVariablesNutrientAdditionMethodMesocosmVolume:TurnoverrateShootspermesocosm:m-2Algae GrazersExperimental Results andCommentsZostera marinaNetherlands:van Katwijk etal.1997Temperature: 15, 20ECSediment: Mud, SandNutrients: +AN(9:3;25:25;25:50;25:125;75:75;125:25:M)Continuousfeed fromstocksolutions19.4 l : 1.2 d-140 EpiphytesnotapparentlymeasuredNo+A - toxic at 125 μM, probablytoxic as low as 25:M+N - no effectToxicity8with8temperature.Washington:Williams &Ruckelshaus 1993Sediment: C, D, +A (5,20:M)Water: C, +A (10, 33:M)Continuousfeed fromstocksolutions75 l: 61 d-126-62 EpiphytebiomassmeasuredAmphipodspossiblypresentEpiphyte biomass8with8watercolumn NH4. Eelgrass leafgrowth9with8epiphytebiomass.Rhode Island:Lin et al. 1996Nutrients:C,+P(.7:M),+N(7.6:M), +NP(7.6,.7:M),+A(7.6,.7:M)Once perday,dissolved4554 l: .05 d-1518:250 m-2EpiphytebiomassmeasuredAdded inverts& fish, noepiphytegrazersNo8in epiphytes in nutrientadditions. Phytoplanktonblooms may have limitedepiphyte response.7.13Rhode Island:Taylor et al. 1999Nutrients: C, +AP(.9,.1;1.8,.2,3.6,.3;7.3,.6:M)Once perday,dissolved4554 l: .05 d-1518:250 m-2Epiphytebiomassmeasured:added 4macroalgaeAdded inverts& fish,bottomgrazing snails@100No significant treatment effectson seagrass, epiphytes or driftalgae.Rhode Island:Kopp 1999Nutrients: C, +N (50,75, 100, 150:M)Light: C, 55%Continuousfeed fromstock headertanks150 l: 1 d-150: 200m-2Epiphytebiomassmeasuredafter 1monthNoNo sig. effectof nitrate treatmenton epiphytes (although trend for8for high light, high nutrient),low light level9epiphyte loadsignificantly.Rhode Island:Bintz et al. 2003Nutrients: C, +NP(6.5,.5:M)Temperature: 3 levels(9 yr mean, 4Eabove,4Ebelow)Continuousfeed fromdrip bags4554 l: .1 d-1580:365 m-2Epiphytebiomassmeasure5d:added 2macroalgalspeciesGrass shrimp@ 8 per tankEpiphytes8in warm, nutrientenriched versus cool treatments.Virginia:Neckles et al.1993Nutrients: C, +AN(16.4-37.8:M) P (1.8-3.3:M)Grazers: C, +GContinuousfeed fromstocksolutions110 l: 16 d-192(15/pot):1500 m-2EpiphytebiomassmeasuredSnails,isopods,amphipods 9-11.4 x103m-2Seasonally varied grazerdensities and nutrients.Epiphyte biomass8with8nutrients.Virginia:Moore & Wetzel2000Light: C, 9,28,42%Nutrients: C, +ANP(7-16.3; 6.6-7.7; 1.4-1.8:M)Continuousfeed fromstocksolutions110 l: 16 d-192(15/pot):1500 m-2EpiphytebiomassmeasuredSnails;5,200 m-1Shading inhibited epiphytes andeelgrass. Epiphytes8only withhigh light and nutrients inSpring.North Carolina:Burkholder et al.1992Nutrients: C,+N(3.5,7, 35:M)Pulse, daily 1570 l:0.5-0.1 d-1over a 2 hrperiod only~2000:700 m-2Epiphytesassessed ascell countsonlyAmphipods +isopods @25-58 x103m-2No significant treatment effectson epiphytes. Variablemacroalgal response. Grazerdensity was high in springexperiment.7.14North Carolina:Burkholder et al.1994Nutrients:Spring C,+N (5:M)Fall C, +N (10:M)Pulse, daily 1570 l:0.5-0.1 d-1over a 2 hrperiod only~2500:800 m-2Epiphytesassessed ascell countsonlyAmphipods@ ~7-80 x103m-2Epiphyte biomass was“negligible”. Grazer densitiesvery high in both spring and fallexperimentsNew Hampshire:Short et al. 1995A)Light:11,21,41,61,94% of ambientB) Nutrients: C,+AP@ 6xC) Light:11,41,94% xNutrients: C,+AP (7-13:M) (1.4-2.4:M)Continuous,from bags ofslow releasefertilizer800 l, 2 d-1200:133 m-2Macroalgaeand epiphytebiomassassessedAddedamphipodpredators andmud snailsEpiphytes8compared to controlin high light treatment. Variableresponses to nutrients at lowerlight levels.Thalassia testudinumFlorida:Tomasko andLapointe 1991Light: C, -33%Nutrients: C, +ANP(10,1,10:M)Pulse daily,fromsolutions120 l: (?50 d-1)static for 14 hrd-1? 9 EpiphytebiomassmeasuredAdded mixedgrazers,density?+ANP8epiphytes and9rhizomegrowth.Shading9epiphyte effect.Florida:LaPointe et. al1994Nutrients: C, +A (10:M) , +P (1:M)),+AP (10,1:M)Pulse daily,fromsolutions120 l: 50 d-1static for 14 hrd-1? EpiphytebiomassmeasuredAdded mixedgrazers,density ?Both +A and +P8epiphytes and9rhizome growth.Halodule wrightiiFlorida:LaPointe et. al1994Nutrients: C, +A, +P,+APPulse daily,fromsolutions120 l: 50 d-1static for 14 hrd-1? EpiphytesassessedAdded mixedgrazers,density?Both +A and +P8epiphytes;only +P9rhizome growth.North Carolina:Burkholder et al.1994Nutrients:Fall C, +N (10:M)Pulse, daily 1570 l:0.5-0.1 d-1over a 2 hrperiod only~3768:1200 m-2Epiphytesassessed ascell countsonlyAmphipods@ ~7-80 x103m-2Epiphyte biomass was“negligible”. Grazer densitiesvery high in both spring and fallexperiments7.15Ruppia maritimaMaryland:Twilley et al.1985Nutrients: C, +NAP @3 levelsPulse, 6@10 dintervals,fertilizer340,000 l:pulse turnover@8-9 d?50-80%coverEpiphytesmeasured? notspecified butlikely+N at >60:M caused seagrass9.Epiphytes8.North Carolina:Burkholder et al.1994Nutrients:Fall C, +N (10:M)Pulse, daily 1570 l:0.5-0.1 d-1over a 2 hrperiod only~3768:1200 m-2Epiphytesassessed ascell countsonlyAmphipods@ ~7-80 x103m-2Epiphyte biomass was“negligible”. Grazer densitiesvery high in both spring and fallexperimentsPotamogeton perfoliatusMaryland:Murray et al. 2000A) C, +NP as Pulse,+NP as Continuous(both 38:M N, 3.8:M P)B) Nutrients: D(2-4:M), +N (20-24:M)Water exchange:1,3,6,12 d-1C) Nutrients: +N (<10:mol l-1N, >30:moll-1N)Grazers: C, +GA) Pulse andcontinuousB, C)ContinuousA) 10 lB) 100 l: seeTreatmentVariablesC) 100 l: 1 d-1? EpiphytebiomassmeasuredA) NoB) NoC) 0.25 gm-2,amphipods,density ?B, C) Used sequential runs fordifferent treatments, thereforeinterpretation of results isdifficult because of uncontrolledtime effect.See text for effects.7.16 Halodule and Thalassia . When light reduction was combin ed 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 emergi ng that complex interactions among trophic components in seagrass systems may determine the ul timate 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 de creased epiphytic load caused by grazer removal tends to improve seagra ss growth (e.g. Hootsman and Vermaat 1985; Howard and Short 1986; Phillipart 1995) or that there was inhibition of seagrass growth caused by increased epiphytic loads. These impacts were observed for a wide variety of seagrass species including Zostera marina and Thalassia testudinum , and Halodule wrightii . An important issue that remains is whether there is evidence that grazi ng can control epiphyte growth under high nutrient loading conditions, and whether there is any specific nutrient elevation level where grazer controls will like ly 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 Ruckel shaus (1993) did not conduct simultaneous experiments with both factors. Meso cosm treatments without the isopod grazer Idotea resecata had epiphyte biomass almost 300% higher than in ta nks with grazers at fiel d densities. However, the growth response of Z. marina after about two months was not significantly different, presumably because total epiphyte load in th e ungrazed treatment was still below the critical threshold for seagrass growth impacts.

In a series of mesocosm experiments, Neckle s et al. (1993) varied nu trient concentrations and epiphyte grazer densities to examine the re lative 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 relativ e composition all varying with season to better reflect natural conditions. Experimental outcomes also varied with season. Epiphyte biomass increased with nutrient enrichment in all expe riments, although margin ally so in the fall experiment. Grazer impacts were greatest in the two summer treatments, which also corresponded to the treatments with the highest de nsity 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, sugges ting 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 co mposition all varied simultaneously among the7.17 experiments makes it impossible to identify the ca uses of the differentia l seasonal responses of epiphytes to grazers that were observed.

In addition to the factors identified by N eckles 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 experi ments with the brackish water macrophyte Potamogeton perfoliatus to examine the relative imp acts of the frequency and timing of nutrient additions, the reside nce time of water within experimental systems, and relative trophic complexity of food chains in the me socosms. By scaling macrophyte and epiphyte responses relative to the contro ls 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 nutri ents, changes from pulse d to continuous nutrient addition and from high to low water exchange rate s, 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 bioma ss 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 th e 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 macr ophyte response, which was -88% for macrophyte growth in both grazed and ungrazed treatments relative to controls.

Any factor that influences eith er 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) obs erved that biomass of epiphytes on Z. marina was highest in sites exposed to wate r movement, and that there was little epiphyte coverage on seagrass in sheltered areas where a bundance of the grazing snail Hydrobia

ulvae was extremely high (151 x 10 3 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 inhib it feeding of micrograzers, thus releasing epiphytes from grazing pressure. However, Caine (1980) found an oppos ite pattern, with epiphyte biomass and abundance of the grazing amphipods Caprella laeviscula both being higher in quiet water sites

versus sites with active wave acti on. On individual seagrass bl ades, grazer biomass and epiphyte biomass were also positively correlated. Th ese differences suggest that species specific differences in the dominant grazers in diffe ring 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 Jerna koff et al. 1996, Table 3), grazer community composition will clearly be critical to determ ining the ultimate level of effect on seagrass epiphytes.7.18 7.5 Research Gaps

Epiphyte biomass appears to be a major res ponse variable in determining the ability of seagrasses to grow and survive under in situ conditions. In early st udies, determination of photosynthesis/ irradiance re lationships for seagrasses were typi cally 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 al tered. Effects of epiphyt es 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 epiphy tic algae all potential ly occurring. High loadings of epiphytes clearly can substantially reduce available i llumination to seagrass plants. In some systems, light reduction to seagrasses fr om epiphyte load may reach 60 to 80%, at least seasonally (Harden 1994; Dixon 2000). Epiphyte graz ing studies generally show that removal of epiphytes enhances seagrass growth. These obser vations are strongly sugg estive that persistent heavy epiphyte cover will lead to seagrass loss, bu t 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 li ght availability, and hence seagrass distribution, may be an essential element to include in develo pment of management crit eria for protection of coastal seagrass beds. Tomasko (pers. comm.) has found that, for Thalassia testudinum in several southwest Florida embayments, there is c onsiderable 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 vari ation in the typical epiphytic load. The role of grazers in determining the ultima te 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 formula tion for including epiphyte loads in estimating light available at the seagrass l eaf surface. This is an importa nt step forward, but the model formulations do not yet appear to have been extensively validated for other systems.

Mesocosm research has shown that differe nt algal components may dominate in the biomass response to nutrient enrichment in sp ite of similar initial conditions, and that temperature, nutrient exposure regime and othe r factors such as graz ing intensity may all

influence the outcome of nutrien t enrichment. The laboratory re sults help support observations that suggest that seagrass losses, even in diffe rent regions of the same limited system, may be caused by competition between seagrasses and di fferent algal components (Short and Burdick 1996). Such results together suggest that if seagrass stress-response models are to be used to7.19 evaluate whether sets of environmental conditions are adequately protective of seagrasses, such models must account for impacts from multiple pathways, and must account for effects of trophic cascades.

7.6 Literature Cited

Ballantine, D. and H. J. Humm. 1975. Benthic algae of the An clote estuary I. Epiphytes of seagrass leaves. Florida Scientist 38:148-162. Bintz, J. C., S. W. Nixon, B. A. Buckley, and S. L. Granger. 2003. Impacts of temperature and nutrients on coastal lagoon plant co mmunities. Estuaries 26:765-776. Borowitzka, M. A. and R. C. Let hbridge. 1989. Seagrass Epiphytes. pp. 458-499. In A. W. D. Larkum, A. J. McComb and S. A. Shepherd (eds). Biology of Seagrasses: A treatise on the biology of seagrasses with special refe rence to the Australian region. Elsevier, Amsterdam. pp. 841. Borum, J. 1985. Development of epiphyte communities on eelgrass ( Zostera marina ) along a nutrient gradient in a Danish es tuary. Marine Biology 87:211-218. Borum, J. 1987. Dynamics of epiphyton on eelgrass ( Zostera marina L.) leaves: Relative roles of algal growth, herbivory, and substratum turnover. Limnology and Oceanography

32:986-992. Borum, J. and S. Wium-Andersen. 1980. Bi omass and production of epiphytes on eelgrass ( Zostera marina L.) in the Øresund, Denmark. Ophelia, Supplement 1:57-64. Borum, J., H. Kaas, and S. Wium-Andersen. 1984. Biomass variation and autotrophic production of an epiphyte-macrophyte community: II. Epiphyte species composition, biomass and production. Ophelia 23:165-179. Brandt, L. A. and E. W. Koch. 2003. Periphyton as a UV-B filter on seagrass leaves: a result of

different transmittance in the UV-B and PAR ranges. Aquatic Botany 76:317-327. Brix, H. and J. E. Lyngby. 1985. Uptake and translocation of phosphorus in eelgrass (Zostera

marina ). Marine Biology 90:111-116. Brush, M. J. and S. W. Nixon. 2002. Direct measurements of light attenuation by epiphytes on eelgrass Zostera marina . Marine Ecology Progress Series 238:73-79. Bulthius, D. A. and W. J. Woelkerling. 1983. Biomass accumulation and shading effects of epiphytes on leaves of the seagrass Heterozostera tasmanica , in Victoria, Australia. Aquatic Botany 16:137-148. Burkholder, J. M., K. M. Mason, and H. B. Glas gow Jr. 1992. Water column nitrate enrichment promotes decline of eelgrass Zostera marina : evidence from seasonal mesocosm experiments. Marine Ecology Progress Series 81:163-178. Burkholder, J. M., H. B. Glasgow Jr., and J. E. Cooke. 1994. Comparative effects of watercolumn nitrate enrichment on eelgrass Zostera marina , shoalgrass Halodule wrightii , and widgeongrass Ruppia maritima . Marine Ecology Progr ess Series 105:121-138. Caine, E. A. 1980. Ecology of two littoral sp ecies of caprellid amphipods (Crustacea) from Wshington, USA. Ma rine Biology 56:327-335. Cambridge, M. L., A. W. Chiffings , C. Brittan, L. Moore, and A. J. McComb. 1986. The loss of seagrass in Cockburn Sound, Western Australia . II. Possible causes of seagrass decline. Aquatic Botany 24:269-285.7.20 Cebrian, J., S. Enriques, M. Fortes, N. Agaw in, J. E. Vermaat, and C. M. Duarte. 1999. Epiphyte accrual on Posidonia oceanica (L.) delile leaves: implication for light absorption. Botanica Marina 42:123-128. Dixon, L. K. 2000. Establishing lig ht limitations for the seagrass Thalassia testudinum : An example from Tampa Bay, Florida. pp. 9-32. In: S. A. Bortone (ed)., Seagrasses: Monitoring, Ecology, Physiology and Manage ment. CRC Press, Boca Raton. pp. 318. Dixon, L. K. and G. Kirkpatrick. 1995. Light at tenuation with respect to seagrasses in Sarasota Bay, Florida. Sarasota Bay National Estuary Program. Mote Marine Laboratory Report no. 407. pp. 53. Drake, L. A., F. C. Dobbs, and R. C. Zimme rman. 2003. Effects of epiphyte load on optical properties and photosynthetic potential of the seagrasses Thalassia testudinum Banks ex König and Zostera marina L. Limnology and Oceanography 48:456-463. Frankovich, T. A. and J. W. Fourquean. 1997. Seagrass epiphyte loads along a nutrient availability gradient, Florida Bay, USA. Marine Ecology Progress Series 159:37-50. Hall, M. O. 1988. Dynamics and interactions of epiphytic macroalg ae and meiofauna on the

seagrass Thalassia testudinum . Ph. D. dissertation, University of South Florida, Tampa, pp. 170. Hall, M. O. and N. J. Eiseman. 1981. The seagra ss epiphytes of the Indian River, Florida. I. Species list with descripti ons and seasonal occurrences. Botanica Marina 24:139-146. Hanisak, M. D. 2001. Photosynthetically Ac tive Radiation, Water Quality and Submerged Aquatic Vegetation in the Indian River Lagoon. Final Report for Contract No. 93W199 to the St. Johns River Water Management District, Palatka, Florida. pp. 502 . Harden, S. W. 1994. Light requirements, epiphyt e load and light reduction for three seagrass species in the Indian River Lagoon, Fl. M. S. thesis, Florida Institute of Technology, Melbourne Fl. pp. 130. Harlin, M. M. 1975. Epiphyte - host relations in seagrass communities. Aquatic Botany 1:125-

131. Harlin, M. M. 1980. Seagrass epiphytes. pp. 117-151. In: R. C. Phillips and C. P. McRoy (eds). Handbook of Seagrass Biology: An Ecosystem Pe rspective. Garland STPM Press, New York. pp. 353. Heijs, F. M. L. 1985. The seasonal distribution and community structure of the epiphytic algae

on Thalassia hemprichii (Ehrenb.)Aschers. from Papua New Guinea. Aquatic Botany

21:295-324. 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. Aquatic Botany 22:83-88. Howard, R. K. and F. T. Short. 1986. Seagra ss growth and survivorship under the influence of 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. Jensen, P. R. and R. A. Gibson. 1986. Pr imary production in three subtropical seagrass communities: A comparison of four autotrophi c components. Florida Scientist 49:129-

141.7.21 Jernakoff, P., A. Brearley, and J. Nielsen. 1996. Factors affecting graze r-epiphyte interactions in temperate seagrass meadows. Oceanography and Marine Biology Annual Reviews

34:109-162. Kita, T. and E. Harada, 1962. Studies on epi phytic communities. I. Abundance and distribution of microalgae and small animals on Zostera blades. Publications of the Seto Marine Biological Laboratory 10:245-257. Kopp, B. S. 1999. Effects of nitrate fe rtilization and shading on physiological and biomechanical properties of eelgrass ( Zostera marina L.). Ph. D. Dissertation, University of Rhode Island, pp. 187. Lapointe, B. E., D. A. Tomasko, and W. R. Matzie. 1994. Eutrophicat ion and trophic state classification of seagrass communities in th e Florida Keys. Bulletin of Maine Science

54:696-717 Lewis, J. B. and C. E. Hollingworth . 1982. Leaf epifauna on the seagrass Thalassia testudinum

. Marine Biology 71:41-49. Lin, H.-J., S. W. Nixon, D. I. Ta ylor, S. L. Granger, and B. D. Buckley. 1996. Responses of epiphytes on eelgrass, Zostera marina L., to separate and combined nitrogen and phosphorus enrichment. Aquatic Botany 52:243-258. Moore, K. A. and R. L. Wetzel. 200 0. Seasonal varia tions in eelgrass ( Zostera marina L.) responses to nutrient enrichment and redu ced light availability in experimental ecosystems. Journal of Experiment al Marine Biology and Ecology 244:1-28. Morgan, M. D. and C. L. Kitting. 1984. Productivity and utilization of the seagrass Halodule wrightii and its attached epiphytes. Li mnology and Oceanography 29:1066-1076. Murray, L., R. B. Sturgis, R. D. Bartleson, W. Severn, and W. M. Kemp. 2000. Scaling submersed plant community responses to expe rimental nutrient enrichment. pp. 241-257. In: S. A. Bortone (ed). Seagrasses: Mon itoring, Ecology, Physiology and Management. CRC Press, Boca Raton. pp. 318. Nagle, J. S. 1968. Distribution of the epibiota of macroepibenthic plants. Contributions in

Marine Science 13:105-144. Neckles, H. A., R. L. Wetzel, and R. J. Orth. 1993. Relative effects of nutrient enrichment and grazing on epiphyte-macrophyte ( Zostera marina L.) dynamics. Oecologia 93:285-295. Nelson, T.A. and J.R. Waaland. 1997. Seasona lity of eelgrass, epiphyt e and grazer biomass and productivity in subtidal eelgrass meadows subjected to moderate tidal amplitude. Aquatic Botany 56:51-74. Neverauskas, V. P. 1987. Accumulation of peri phyton biomass on artifici al substrates deployed near a sewage sludge outfall in South Austra lia. Estuarine, Coastal and Shelf Science

25:509-519. Neverauskas, V. P. 1988. Response of a Posidonia community to prolonged reduction in light. Aquatic Botany 31: 361-366. Orth, R. J. and J. van Montfrans. 1984. Epiphyt e-seagrass relationships with an emphasis on the role of micrograzing: a review. Aquatic Botany 18:43-69. Penhale, P. and W. O. Smith, Jr. 1977. Excr etion of dissolved or ganic carbon by eelgrass ( Zostera marina ) and its epiphytes. Limnol ogy and Oceanography 22:400-407. Phillipart, C. J. M. 1995. Effect of periphyton grazing by Hydrobia ulvae on the growth of

Zostera noltii on a tidal flat in the Dutch Wadde n Sea. Marine Biology 122:431-437.7.22 Phillips, R. C. 1960. Observations of the ecolo gy and distribution of the Florida seagrasses. Florida. State Board of Conservation Marine Laboratory Professional Paper Series No. Two. 72 pp. Rice, J. D., R. P. Trocine, and G. N. Wells. 1983. Factors influencing seagrass ecology in the Indian River Lagoon. Florida Scientist 46:276-286. Sand-Jensen, K. 1977. Effect of epiphytes on eelgrass photosynthesis. Aquatic Botany 3:55-63. Sand-Jensen, K. and J. Borum. 1984. Epiphyte shading and its effect on photosynthesis and diel

metabolism of Lobelia dortmanna L. during the spring bloom in a Danish lake. Aquatic Botany 20:109-119. Sand-Jensen, K. 1990. Epiphyte sh ading: its role in resulting depth distribution of submerged aquatic macrophytes. Folia Geobotanica et Phytotaxonomica, Praha. 25:315-320. Sand-Jensen, K. and N. P. Revsbach. 1987. P hotosynthesis and light adaptation in epiphytemacrophyte associations measured by oxyge n microelectrodes. Limnology and Oceanography 32:452-457. Schanz, A., P. Polte, and H. Asmus. 2002. Cascading effects of hydrodynamics on an epiphytegrazer system in intertidal seagrass beds of the Wadden Sea. Marine Biology 141:287-

297. Short, F. T. and D. M. Burdick. 1996. Quan tifying eelgrass habitat loss in relation to housing development and nitrogen loading in Waquo it Bay, Massachusetts. Estuaries 19:730-

739. Short, F. T., D. M. Burdick, and J. E. Kaldy III. 1995. Mesocosm experiments quantify the effects of eutrophication on eelgrass, Zostera marina . Limnology and Oceanography

40:740-749. Silberstein, K., A.. W. Chiffi ngs, and A. J. McComb. 1986. The loss of seagrass in Cockburn Sound, Western Australia. III. The e ffect of epiphytes on productivity of Posidonia

australis Hook. F. Aquatic Botany 24:355-371. Taylor, D. I., S. W. Nixon, S. L. Granger, a nd B. A. Buckley. 1999. Responses of coastal lagoon plant communities to levels of nutrient enrichment: a mesocosm study. Estuaries

22:1041-1056. Tomasko, D. A. and B. E. Lapointe. 1991. Productivity and biomass of Thalassia testudinum as related to water column nutri ent availability and epiphyte le vels: field observations and experimental studies. Marine Ec ology Progress Series 75:9-17. Tomasko, D. A., C. J. Dawes, and M. O. Hall. 1996. The effects of anthropogenic enrichment on turtle grass ( Thalassia testudinum ) in Sarasota Bay, Florida. Estuaries 19:448-456. Trocine, R. P., J. D. Rice, and G. N. Wells . 1981. Inhibition of s eagrass photosynthesis by ultraviolet-B radiation. Plant Physiology 168:74-81. 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. 1. Algal growth and effects on production of plants and associat ed communities. Marine Ecology Progress

Series 23:179-191. U.S. EPA. 2003. Ambient water quality crit eria for dissolved oxygen, water clarity and chlorophyll a for the Chesapeake Bay and its tidal tri butaries. U.S. EPA, Office of Water and Office of Science and Technology. EPA 900-R-03-002.7.23 van Katwijk, M. M., L. H. T. Vergeer, G. H. W. Schmitz, and J. G. M. Roelofs. 1997. Ammonium toxicity in eelgrass Zostera marina . Marine Ecology Progress Series

157:159-173. Van Montfrans, J., R. L. Wetzel, and R. J. Orth . 1984. Epiphyte-grazer re lationships in seagrass meadows: consequences for seagrass gr owth and production. Estuaries 7:289-309. Wetzel, R. L. and H. A. Neckles. 1986. A model of Zostera marina L. photosynthesis and growth: simulated effects of selected physical-chemical variables and biological interactions. Aqua tic Botany 26:307-323. Williams, S. L. and C. P. McRoy. 1982. Seagra ss productivity: the effect of light on carbon uptake. Aquatic Botany 12:321-344. Williams, S. L. and M. H. Ruckelshaus. 1993. Effects of nitrogen availability and herbivory on eelgrass ( Zostera marina ) and epiphytes. Ecology 74:904-918.8.1 

## 8.0 Macroalgal Interactio ns 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 survey s suggesting negative effects of macroalgae on

Zostera and Thalassia ? 2. Is there corresponding evidence from fiel d manipulations or la boratory 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 (princip ally 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 abundan ce 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 fo llowing eutrophication renders the negative effect of marine eutrophication on seagrass stands an indisputable fact, and indicates that it most likely is the main cause of seagrass decline worldwide.” Although ther e is some evidence that high levels of nitrate and ammonium can be directly toxic to seagrasses (Bur kholder et al. 1992, 1994; Van Katwijk et al. 1997), most researchers have attributed this decline to light reduction caused by excess nutrient stimulation of phytopl ankton, 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 c ould 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 dissolve d sulfide, a known phytotoxin, in the pore water (Goodman et al. 1995).8.2 Table 8.1. Relationships between macroalgal abundance and the distribution or abundance of
Zostera marina, Z. noltii, and Thalassia testudinum , in chronological order.

Algal Taxon Effect Location Reference

Zostera marina

Ulva Zostera suffocated Netherlands Coast den Hartog and Polderman (1975)

Chaetomorpha macroalgae replaced Zostera North Sea Nienhuis (1983)

macroalgae Zostera declined Coast of Poland Plinski and Florczyk (1984)

macroalgae no effect Coast of Ireland Whelan and Cullinane (1985)

Ulva Zostera uprooted, buried Northeast Pacific Kentula and McIntire (1986)

Ulva macroalgae replaced Zostera Venice Lagoon Sfriso et al. (1989)

macroalgae macroalgae replaced Zostera Coast of Denmark Funen County Council (1991)

macroalgae Zostera declined Coast of Poland Kruk-Dowgiallo (1991)

macroalgae Zostera declined Coast of Poland Ciszewski et al. (1992)

macroalgae macroalgae replaced Zostera Northwest Atlantic Valiela et al. (1992)

Ulva Zostera declined Northwest Atlantic Short et al. (1993)

Cladophora Zostera declined Coast of Denmark Thybo-Christesen et al. (1993)

Ulva Zostera bed suffocated SW Coast of England den Hartog (1994)

macroalgae macroalgae replaced Zostera Northwest Atlantic Lyons et al. (1995)

Ulva, Chaetomorpha Zostera declined North Sea Nienhuis (1996)

macroalgae macroalgae replaced Zostera Coast of Portugal Oliveira and Cabecadas (1996)

macroalgae macroalgae replaced Zostera Baltic Sea Schramm (1996)

Cladophora,

Gracilaria decreased meadow area Northwest Atlantic Short and Burdick (1996)

Ulva shading killed Zostera Venice Lagoon Coffaro and Bocci (1997)

Ulva macroalgae replaced Zostera Coast of Portugal Flindt et al. (1997)

Cladophora,

Gracilaria shading killed Zostera Northwest Atlantic Hauxwell et al. (1998)

macroalgae macroalgae replaced Zostera Temperate Zone Raffaelli et al. (1998)

macroalgae loss of seagrass Northw est Atlantic Bricker et al. (1999)

macroalgae macroalgae replaced Zostera Northwest Atlantic Valiela et al. (2000a)

macroalgae macroalgae replaced Zostera Northwest Atlantic Bowen and Valiela (2001)

Cladophora

,

Gracilaria removal increased Zostera Northwest Atlantic Deegan et al. (2002)

macroalgae Zostera declined Northwest Atla ntic Hughes et al. (2002)

macroalgae macroalgae replaced Zostera Coast of Denmark Nielsen et al. (2002)

Cladophora

,

Gracilaria macroalgae replaced Zostera Northeast Atlantic Hauxwell et al. (2003)

macroalgae no displacement Northeast Pacific Kentula and DeWitt (2003)

Ulva no effect Northeast Pacific Thom et al. (2003)

macroalgae macroalgae replaced Zostera Adriatic Coast of Ital y Curiel et al. (2004)8.3 Zostera noltii

Ulva Zostera suffocated Netherlands Coast den Hartog and Polderman (1975)

Ulva macroalgae replaced Zostera Venice Lagoon Sfriso et al. (1989)

macroalgae little effect Dutch Wadden Sea Philippart and Dijkema (1995)

Ulva macroalgae replaced Zostera Spanish Atlantic Coast Niell et al. (1996)

Ulva, Chaetomorpha Zostera declined North Sea Nienhuis (1996)

Ulva macroalgae replaced Zostera Portuguese Atlantic Oliveira and Cabecadas (1996)

Ulva macroalgae replaced Zostera Northwest Atlantic Flindt et al. (1997)

Ulva macroalgae replaced Zostera Portuguese Atlantic Pardal et al. (2000)

Ulva macroalgae replaced Zostera Portuguese Atlantic Cardoso et al. (2004)

Ulva macroalgae replaced Zostera Portuguese Atlantic Patricio et al. (2004)

Thalassia testudinum

macroalgae decreased productivity Gulf of Mexico Cowper (1978)

Chaetomorpha,

Acetabularia uprooted; stopped

recolonization Mexican Caribbean Merino et al. (1992)

Eucheuma possible competition Gulf of Mexico Perez-Enriquez (1996)

macroalgae no effect Gulf of Mexico Bell and Hall (1997)

macroalgae possible competitor Columbian Caribbean Angel and Polania (2001)

macroalgae potential threat Gulf of Mexico Kopecky and Dunton (2006)

Nienhuis (1983) observed that eelgrass abundance in a sout hwest Netherlands’ estuary dropped by about 50% between 1978 and 1980. He listed a number of po ssible causes including macroalgal competition, and suggested that the most plausible explanation for the decrease was an increase in organic matter de position 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 toxi c substances,” killing the root and rhizome system. Two potential phytotoxi ns which could have been pr oduced in high concentrations

under such circumstances are dissolved sulfid e (Goodman et al. 1995) and ammonium (Van Katwijk et al. 1997; 2000).

A mixed bed of Z. marina and Z. noltii in Langstone Harbor, England was completely destroyed by a thick blanket of Ulva spp., due to prolonged anaerobic conditions under the algal mat with corresponding high levels of sulfide and ammonia (den Hartog 1994). However, Whelan and Cullinane (1985) found no significant interaction between macroalgae and eelgrass

in southwest Ireland.

Although there was no evidence of causality, Thybo-Christesen et al. (1993) noted a significant change in the vegetation in Danish coastal waters, with the appearance of filamentous algal mats being associated with a substantial decrease in the abundance of eelgrass. Working in Venice Lagoon, Sfriso et al. (1989) found that high nutrient loads in the lagoon in the 1970s were followed in the 1980s by a major shift in macrophyte composition. Z. marina and Z. noltii beds were replaced by “nitrophile species,” principally Ulva rigida . In some shallow areas of the8.4 lagoons, this macroalga occupied 100% of the bottom at an abundance exceeding 10 kg/m 2 wet weight, leading to periodic o ccurrences of water column anoxi a and extensive mortality of

macrofauna.

Flindt et al. (1997) compared two other Eur opean 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 atmosphe ric deposition, eelgrass meadows also were substantially reduced. However, in this case the direct cause appears to have been increased biomass of phytoplankton and ep iphytic 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 ava ilability and water velo city influenced the structure of the primary producer community, and th at competition for light was a major factor in the interaction between the macroalgae and eelgrass.

Losses of eelgrass and other seagrass in rece nt decades also have been reported and extensively studied in the U.S.A. Orth and Moore (1983) analyzed ex tensive data sets on submerged aquatic vegetation (SAV) from Chesap eake Bay. They found th at, 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 aff ecting the quantity and qual ity of light reaching the plants. However, they did not discuss a specifi c mechanism (e.g., macroalgal increases) for such

an effect.

Valiela et al. (1992) conducted an extensive study of the sources a nd effects of nutrient enrichment in Waquoit Bay, Massachusetts. Th ey concluded that increased development of watersheds there led to increased groundwater nu trient concentrations, which in turn led to increased abundances of macroalgae and major reductions in distribut ions of eelgrass. Continuing this research, Lyons et al. (1995) c oncluded that, in this bay, the abundance of macroalgae increased linearly with nitrogen lo ading 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 se rve 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 Re serve, relating the loss in eelgrass areal extent to housing develo pment and groundwater loadings of n itrogen in the sub-basins of the

watershed. The direct effect of these loadi ngs varied from basin to basin, but included stimulation of phytoplankton, epi phytic 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 th is work. The increas e of nutrient loading

in estuaries was found to increa se macroalgal nitrogen uptake ra te, tissue nitrogen content,8.5 photosynthetic rate, and growth rate. Although th e effects of such loadings are mitigated by fringing salt marshes and higher rates of tidal ex change, high nitrogen loadings and resultant macroalgal blooms can significantly alter estuarin e 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 successf ul mathematical model of the watershed nitrogen loadings of Waquoit Bay. Hauxwell et al. (2003) extended th eir studies in Waquoit Bay, showing an exponential decr ease in eelgrass shoot density and bed area as nitrogen loads increased. They recommended that these variab les 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 in creased growth and standing stock of algae (water column algae, epiphytes, and macroalgae), causi ng light limitation of eelgrass. A major effect, however, was severe light limitation to newly recruiting s hoots by shading from macroalgal canopies < 15 cm in height.

There are few reports of the interaction be tween macroalgae and eelgrass on the Pacific Coast of the U.S.A. Phillips (1984) commented th at large masses of loose macroalgae such as

Ulva spp. commonly occur in seagra ss meadows, especially where tidal currents are sluggish. In Netarts Bay on the Oregon coast, Kentula and McIntire (1986) found a decrease in shoot net primary production of Z. marina in mid-summer concurrent with a decrease in insolation and a rapid increase in the biomass of U. prolifera . They observed that as this macroalga drifted through the eelgrass meadow it beca me 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. Sim ilarly, Thom et al. ( 2003) did not find clear evidence of a decline in eelgra ss 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 co mpeted for light with seagrasses, including T. testudinum, in Redfish Bay, Texas, with the algae havi ng substantially higher growth rates than the seagrasses at irradiances less that 45% of surface irradiance. Merino et al. (1992) studied the relationship between T. testudinum and macroalgal communities (dominated by the genera Chaetomorpha and Acetabularia ) on the Mexican Caribbean coast. They reported that as the algae became more common, the seagrass communities declined. In particular, the algae formed mats covering the bottom communities; then, as oxygen bubbles formed within the mats causing them to float, the remaining seagrass spr outs 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 this8.6 case the author suggested that these seagrasses might be affecting the distribution of the seaweed. In contrast, in Tampa Bay, Florid a, 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 Caribb ean. 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 macroalg ae. 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 m acroalgae but not the eelgrass, while phosphate additions stimulated growth of eelgrass but not green macroalgae. None of the nutrient supplements had a significant eff ect on epiphytic algae or phytopl ankton 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 eelgr ass was significantly increased with inorganic nutrient additions, wh ich 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 diffe rent algal forms (phytoplankton, epiphytes, and macroalgae) occurred in different replicate treatments. However, the enrichment effects on eelgrass shoot density, biom ass, and leaf length were simila r for all replicat es. In each case, the negative effect of al gae 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 vari ous combinations of nutrients. They reported that in the controls phytopla nkton levels remained low while macroalgae and epiphytes were abundant, but in the nutrient- enriched mesocosms phytoplankt on blooms dominated. These mesocosm studies do not appear to have produced definitive results regarding the effect of macroalgae on eelgrass. This point was emphasi zed by Raffaelli et al . (1998), who commented on the absence of controlled manipulative fi eld experiments to explore such effects.

Subsequently, Hauxwell et al. (1998) observe d that benthic algal growth rates and biomass increased with nitrogen load in three Massachusetts estuaries, while abundance of8.7 grazers decreased, resulting in ne gligible top-down cont rol of macroalgal biomass. Then, using macroalgal enclosure or exclusion field experime nts, Hauxwell et al. (2001) compared eelgrass productivity in two estuaries that had a six-fold difference in nitrogen loading rate. The authors concluded that macroalgal cover was the primary cause of eelgrass loss in the high-nitrogen estuary. Upon removal of macroalgae, shoot density, summer growth, and summer aboveground net production all increased signi ficantly. 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 whic h eelgrass declined. Working in Washington State, Nelson and Lee (2001) conducted similar fi eld 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) conduc ted laboratory and field expe riments 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 mobiliza tion 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, consider able gaps resulted in the seagrass canopy.

In 1999-2000, Sullivan (unpub.) manipulated green macroalgae densities in 2-m 2
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 NH 4 + and PO 4 - increased with algae de nsity in all habitats, while porewater nitrite and nitrate differed but di d not respond to algal density. Water column nutrient and oxygen concentrations within algal canopies differ ed from open water levels. However, only the water column oxygen concen trations responded to algal manipulations. Macroalgal density affected seagrass density, wi th algae additions reducing shoot density relative to non-enclosure controls (Figure 8.1).8.8 Random Samples Controls Enclosure Controls Algae Removal Algae AdditionShoots m-2

0 5 10 15 20 25

Figure 8.1. Effects of experime ntal benthic green macroalgal ( Ulva spp.) manipulations on shoot density of Zostera marina in experimental enclosures as compared to controls, for Yaquina Bay, Oregon (G. Sullivan, unpub. data , The Wetlands Institute, Chicago IL). Algae addition treatment is significantly di fferent from all other treatments except the enclosure controls (one-way ANOVA, squa re 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 leve ls at a eutrophic and a mesotrophic/oligotrophic site in a Bermuda lagoon. Nutrient enrichment caused an increase in percent cover of the filame ntous, mat-forming macroalga Spyridea hypnoides and a decline in the percent cover and above-ground biomass of T. testudinum at the eutrophic site. Holmquist (1997) manipulated algal mats ( Laurencia poiteaui ) over study plots of T. testudin um in southwestern Florida Bay. The algal canopy height s 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 increa sed to only about 25% of the initial density. Macia (2000) manipulated drif t algae and sea urchins ( Lytechinus variegatus ) within cages containing T. testudinum in Biscayne Bay, Florida. Unde r 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 increase d grazing pressure there was a synergistic effect of grazing and drift algae on seagrass shoot de nsity. Davis and Fourqur ean (2001) manipulated densities of the rhizophytic algae Halimeda incrassata in plots of T. testudinum in the upper8.9 Florida Keys. They reported that the additi on of the macroalgae had no significant impact on seagrass growth, but that the removal of the m acroalgae 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 signi ficant effects of ma nipulated 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 estuarin e intertidal zones of the Pacific Northwest and elsewhere. The accumulation of green macroalg ae within and upslope of eelgrass meadows may help to retain water on the mudflats during period s of daylight low-tide intervals, thus reducing desiccation of the eelgrass plants . Boese et al. (2003) have docu mented desiccation damage in intertidal eelgrass and suggested that this may be a limiting factor for upper intertidal distribution. In Yaquina Bay, Oregon, acute desiccation stre ss is often observed in late spring and early summer when daylight spring-tides, sun ny and/or windy weather combine. In contrast, in the late summer and fall, th e presence of large amounts of m acroalgae tends to cover exposed sheaths which may provide some protection. Annual eelgrass shoots may not be as susceptible to

desiccation stress since their sh eaths 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, includ ing that of macroalgae, which interferes with the physiology (photosynthesis, respiration, reprodu ction, 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 (suc h as those provided by Hauxwell et al. 2001), for different water body char acteristics (temperature, current velocity, turbidity, grazing pressure, etc.) are needed. Similarly, relationships between macroalgal

abundance and the causative anthro pogenic activity (e.g., normali zed nitrogen load rate) are

needed to recommend corrective actions.

8.6 Conclusions

The studies summarized here generally id entify 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 th e eelgrass, thus reducing its photosynthesis. However, a secondary mechanism is that excessi ve 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 drifti ng macroalgae becoming entangled8.10 with eelgrass and uprooting it, or causing its bur ial 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 anth ropogenic 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 substantia l body of evidence that increases in nitrogen loadings to estuaries are accompanied by increase s in macroalgal abundances and decreases in seagrass distributions. This suggests that, conceptually , macroalgal abundance should be included in estuarine nutrient loading criteria in tended to protect seagrasses in those systems. Hauxwell et al. (2001) provided qua ntitative 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 th at 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 dist ribution 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 relationshi p between nutrient loading and macroalgal abundance for that

system.8.11 8.7 Literature Cited

Angel, I. F. and J. Polania. 2001. Structure a nd distribution of seagrass meadows in San Andres Island, Columbia Caribbean. Boletin Ec otropica: Ecosistemas Tropicales 35:1-24. Armitage, A. R., T. A. Frankovich, K. L. Heck, Jr., and J. W. Fourqurean. 2005. Experimental nutrient enrichment causes complex changes in seagrass, microalgae, and macroalgae community structure in Florid a Bay. Estuaries 28:422-434. Bell, S. S. and M. O. Hall. 1997. Drift macr oalgal abundance in seagra ss beds: investigating large-scale associations with physical and biotic attribut es. Marine Ecology Progress

Series 147:277-283. Bintz, J. C., S. W. Nixon, B. A. Buckley, and S. L. Granger. 2003. Impacts of temperature and nutrients on coastal lagoon plant co mmunities. Estuaries 26:765-776. Boese, B. L., K. E. Alayan, E. F. Gooch, and B. D. Robbins. 2003. Desiccation index: a measure of damage caused by adverse aer ial exposure on intertidal eelgrass ( Zostera

marina ) in an Oregon (USA) estuary. Aquatic Botany 76:329-337. Bowen, J. L. and I. Valiela. 2001. The ecologi cal effects of urbanization of coastal watersheds: historical increases in nitrogen loads and eutrophication of Waquoit Bay estuaries.

Canadian Journal of Fisheries and Aquatic Sciences 58:1489-1500. 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 Sc ience, Silver Spring, MD, 71 pp. Brun, F. G., J. J. Vergara, G. Navarro, I. Hernán dez, and J. L. Pérez-Lloréns. 2003a. Effect of shading by Ulva rigida canopies on growth and car bon balance of the seagrass Zostera

noltii . Marine Ecology Progr ess Series 265:85-96. Brun, F. G., I. Hernandez, J. J. Vergara, and J. L. Pérez-Lloréns. 2003b. Growth, carbon allocation and proteolytic activity in the seagrass Zostera noltii shaded by Ulva canopies. Functional Plant Biology 30:551-560. Burkholder, J. M., K. M. Mason, and H. B. Gl asgow Jr. 1992. Water-column nitrate enrichment promotes decline of eelgrass Zostera marina : evidence from seasonal mesocosm experiments. Marine Ecology Progress Series 81:163-178. Burkholder, J. M., H. B. Glasgow Jr, and J. E. Cooke. 1994. Comparative effects of watercolumn nitrate enrichment on eelgrass Zostera marina , shoalgrass Halodule wrightii , and widgeongrass Ruppia maritima . Marine Ecology Progr ess Series 105:121-138. Cardoso, P. G., M. A. Pardal, A. I. Lillebø, S. M. Ferreira, D. Raffaelli, and J. C. Marques. 2004. Dynamic changes in seagrass assembla ges under eutrophication and implications for recovery. Journal of Experime ntal Marine Biology and Ecology 302:233-248. Ciszewski, P., L. Kruk-Dowgiallo, and L. Zmudzins ki. 1992. Deterioration of the Puck Bay and biotechnical approaches to its reclamation. In: Proceedings of the 12th Baltic Marine Biologists Symposium: physiological and bioc hemical strategies in Baltic organisms; new approaches in ecotoxicological resear ch; changes in Baltic ecosystems, 1991, Helsingor, Denmark, pp. 43-46.8.12 Coffaro, G. and M. Bocci. 1997. Resources competition between Ulva rigida and Zostera

marina : a quantitative approach applied to th e Lagoon of Venice. Ec ological Modelling

102:81-95. Cowper, S. W. 1978. The drift algae community of seagrass beds in Redfish Bay, Texas.

Contributions in Marine Science 21:125-132. Cummins, S. P., D. E. Roberts, and K. D. Zimm erman. 2004. Effects of the green macroalga Enteromorpha intestinalis on macrobenthic and seagrass assemblages in a shallow coastal estuary. Marine Ec ology Progress Series 266:77-87. Curiel, D., A. Rismondo, G. Bellemo, and M. Marzocchi. 2004. Macroalgal biomass and species variations in the Lagoon of Venice (Northern Adriatic Sea, Italy): 1981-1998.

Scientia Marina 68:57-67. Davis, B. C. and J. W. Fourqurean. 2001. Competition between the tropical alga, Halimeda

incrassata, and the seagrass, Thalassia testudinum . Aquatic Botany 71:217-232. Deegan, L. A., A. Wright, S. G. Ayvazian, J. T. Finn, H. Golden, R. R. Merson, and J. Harrison. 2002. Nitrogen loading alters seagrass ecosystem structur e and support of higher trophic levels. Aquatic Conservation: Marine and Freshwater Ecosystems 12:193-212. Den Hartog, C. 1994. Suffocation of a littoral Zostera bed by Enteromorpha radiata . Aquatic Botany 47:21-28. Den Hartog, C. and P. J. G. Polderman. 1975. Ch anges in the seagrass populations of the Dutch Waddenzee. Aquatic Botany 1:141-147. Flindt, M. R., L. Kamp-Nielsen, J. C. Marques, M. A. Pardal, M. Bocci, G. Bendoricchio, J. Salomonsen, S. N. Nielsen, and S. E. Jørgen sen. 1997. Description of the three shallow estuaries: Mondego River (Portugal), Ro skilde Fjord (Denmark) and the Lagoon of Venice (Italy). Ecologi cal Modelling 102:17-31. Funen County Council. 1991. Eutrophication of Coastal Waters. Coastal Water Quality Management in the County of Funen, Denmark, 1976-1990. Department of Technology and Environment, Denmark, 287 p. Goodman, J. L., K. A. Moore, and W. C. Denni son. 1995. Photosynthetic responses of eelgrass ( Zostera marina L.) to light and sediment sulfide in a shallow barrier island lagoon. Aquatic Botany 50:37-47. Harlin, M. M. and B. Thorne-Miller. 1981. Nu trient enrichment of s eagrass beds in a Rhode Island coastal lagoon. Marine Biology 65:221-229. Hauxwell, J., J. McClelland, P. J. Behr, and I. Valiela. 1998. Relative importance of grazing and nutrient controls of macroalgal bioma ss in three temperate shallow estuaries.

Estuaries 21:347-360. Hauxwell, J., J. Cebrián, C. Furlong, and I. Va liela. 2001. Macroalgal canopies contribute to eelgrass ( Zostera marina ) decline in temperate estuar ine ecosystems. Ecology 82:1007-

1022. Hauxwell, J., J. Cebrián, and I. Valiela. 2003. Eelgrass Zostera marina loss in temperate estuaries: relationship to land-derived nitr ogen loads and effect of light limitation imposed by algae. Marine Ec ology Progress Series 247:59-73. Hemminga, M. A. and C. M. Duarte. 2000. S eagrass Ecology. Cambridge University Press, Cambridge, United Kingdom. 298 pp.8.13 Holmquist, J. G. 1997. Disturbance and gap form ation in a marine benthic mosaic: influence of shifting macroaglal patches on seagrass struct ure and mobile invertebrates. Marine Ecology Progress Series 158:121-130. Hughes, J. E., L. A. Deegan, J. C. Wyda, M. J. Weaver, and A. Wright. 2002. The effects of eelgrass habitat loss on estuarine fish communities of southern New England. Estuaries

25:235-249. Irlandi, E. A., B. A. Orlando, and P. D. Biber. 2004. Drift algae-epiphyt e-seagrass interactions in a subtropical Thalassia testudinum meadow. Marine Ecol ogy Progress Series 279:81-

91. Kentula, M. E. and C. D. McIntire. 1986. The autecology and production dynamics of eelgrass ( Zostera marina L.) in Netarts Bay, Oregon. Estuaries 9:188-199. Kentula, M. E. and T. H. DeWitt. 2003. Abundance of seagrass ( Zostera marina L.) and macroalgae in relation to the salinity-temper ature gradient in Ya quina Bay, Oregon,

USA. Estuaries 26:1130-1141. Kopecky, A. L. and K. H. Dunton. 2006. Variab ility in drift macroalgal abundance in relation to biotic and abiotic factors in two seagrass dominated estuaries in the western Gulf of

Mexico. Estuaries and Coasts 29:617-629. Kruk-Dowgiallo, L. 1991. Long-term changes in the structure of underwater meadows of the Puck Lagoon. Acta Ichthyologica et Piscatoria (Supplement) 21:77-84. Lyons, J., J. Ahern, J. McClelland, and I. Valiela. 1995. Macrophyt e abundances in Waquoit Bay estuaries subject to diffe rent nutrient loads and the pot ential role of fringing salt marsh in groundwater nitrogen intercepti on. Biological Bulletin 189:255-256. Maciá, S. 2000. The effects of sea urchin grazing and drift algal blooms on a subtropical seagrass bed community. Journal of Expe rimental Marine Biology and Ecology 246:53-

67. McGlathery, K. J. 1995. Nutrient and grazing in fluences on a subtropical seagrass community. Marine Ecology Progress Series 122:239-252. Merino, M., A. González, E. Reyes, M. Gallegos , and S. Czitrom. 1992. Eutrophication in the lagoons of Cancún, México. Science of th e Total Environment (Supplement) 1992: 861-

870. Nelson, T. A. and A. Lee. 2001. A manipulativ e experiment demonstrates that blooms of the macroalga Ulvaria obscura can reduce eelgrass shoot de nsity. Aquatic Botany 71:149-

154. Niell, F. X., C. Fernández, F. L. Figueroa, F. G. Figueiras, J. M. Fuentes, J. L. Pérez-Llorens, M.J. Garcia-Sánchez, I. Hernández, J. A. Fernández, M. Espejo, J. Buela, M. C. GarciaJiménez, V. Clavero, and C. Jiménez. 1996. Spanish Atlantic coasts. In : W. Schramm and P.H. Nienhuis (eds.). Marine Benthic Ve getation: Recent Changes and the Effects of Eutrophication. Springer-Verlag, Berlin, pp. 265-281. Nielsen, S. L., K. Sand-Jensen, J. Borum, and O. Geertz-Hansen. 2002. Depth colonization of eelgrass ( Zostera marina ) and macroalgae as determined by water transparency in Danish

coastal waters. Estuaries 25:1025-1032. Nienhuis, P. H. 1983. Temporal and spatial patterns of eelgrass ( Zostera marina L.) in a former estuary in The Netherlands, dominated by human activities. Marine Technological Society Journal 17(2):69-77.8.14 Nienhuis, P. H. 1996. The North Sea coasts of Denmark, Germany and The Netherlands. In : W. Schramm and P.H. Nienhuis (eds.). Marine Benthic Ve getation: Recent Changes and the Effects of Eutrophication. Spri nger-Verlag, Berlin, pp. 187-221. Oliveira, J. C. and G. Cabeçadas. 1996. Portugal. In : W. Schramm and P.H. Nienhuis (eds.). Marine Benthic Vegetation: Recent Changes a nd the Effects of Eutrophication. SpringerVerlag, Berlin, pp. 283-292. Orth, R. J. and K. A. Moore. 1983. Chesap eake Bay: an unprecedented decline in submerged aquatic vegetation. Science 222:51-53. Pardal, M. A., J. C. Marques, I. Metelo, A. I. Lillebø, and M. R. Flindt. 2000. Impact of eutrophication on the life cycle, population dynamics and production of Ampithoe valida (Amphipoda) along an estuarine spatial gradie nt (Mondego estuary, Po rtugal). Marine Ecology Progress Series 196:207-219. Patrício, J., R. Ulanowicz, M. A. Pardal, and J. C. Marques. 2004. Ascendency as an ecological indicator: a case study of estu arine pulse eutrophication. Estuarine, Coastal and Shelf

Science 60:23-35. Penhale, P. A. and R. G. Wetzel. 1983. St ructural and functional adaptations of eelgrass ( Zostera marina L.) to the anaerobic sediment envir onment. Canadian Journal of Botany

61:1421-1428. Perez-Enriquez, R. 1996. Summ er biomass, density and weight assessment of the red seaweed Eucheuma isiforme (Rhodophyta, Gigartinales) at Dz ilam, Yucatan State, Mexico.

Botanica Marina 39:251-253. Philippart, C. J. M. and K. S. Dijkema. 1995. Wax and wane of Zostera noltii Hornem. in the Dutch Wadden Sea. Aquatic Botony 49:255-268. Phillips, R. C. 1984. The Ecology of Ee lgrass Meadows in the Pacific Northwest: A Community Profile. U.S. Fish and Wildlife Service, FWS/OBS-84/24, 85 pp. Plínski, M. and I. Florczyk. 1984. Cha nges in the phytobenthos resulting from the eutrophication of the Puck Bay. Limnologica 15:325-327. Raffaelli, D. G., J. A. Raven, and L. J. Poole. 1998. Ecological impact of green macroalgal blooms. Oceanography and Marine Biol ogy: an Annual Review 36:97-125. Schramm, W. 1996. The Baltic Se a and its transition zones. In : W. Schramm and P.H. Nienhuis (eds). Marine Be nthic Vegetation: Recent Ch anges and the Effects of Eutrophication. Springer-Verlag, Berlin, pp. 131-163. Sfriso, A., B. Pavoni, and A. Marcomini. 1989. Macroalgae and phytoplankton standing crops in the central Venice Lagoon: primary production and nutrient balance. Science of the

Total Environment 80:139-159. Short, F. T., D. M. Burdick, J. Wolf, and G. E. Jones. 1993. Eelgrass in Estuarine Research Reserves along the East Coast, U.S.A., Part I: Declines from Pollution and Disease and Part II: Management of Eelgrass Meadows. U.S. Department of Commerce, National Oceanic and Atmospheric Administration, Nati onal Estuarine Research Reserve Program and Coastal Ocean Program, Silver Spring, MD, 107 p. Short, F. T., D. M. Burdick, and J. E. Kaldy III. 1995. Mesocosm experiments quantify the effects of eutrophication on eelgrass, Zostera marina. Limnology and Oceanography

40:740-749.8.15 Short, F. T. and D. M. Burdick. 1996. Quan tifying eelgrass habitat loss in relation to housing development and nitrogen loading in Waquo it Bay, Massachusetts. Estuaries 19:730-

739. Taylor, D. I., S. W. Nixon, S. L. Granger, B. A. Buckley, J. P. McMahon, and H.-J. Lin. 1995. Responses of coastal lagoon plant communities to different forms of nutrient enrichment - a mesocosm experiment. Aquatic Botany 52:19-34. Thom, R. M., A. B. Borde, S. Rumrill, D. L. Woodr uff, G. D. Williams, J. A. Southard, and S. L. Sargeant. 2003. Factors influencing sp atial and annual variability in eelgrass ( Zostera

marina L.) meadows in Willapa Bay, Washington, and Coos Bay, Oregon, estuaries.

Estuaries 26:1117-1129. Thybo-Christesen, M., M. B. Rasmussen, and T. H. Blackburn. 1993. Nutrient fluxes and growth of Cladophora sericea in a shallow Danish bay. Marine Ecology Progress Series

100:273-281. Valiela, I., K. Foreman, M. LaMontagne, D. Hersh, J. Costa, P. Peckol, B. DeMeo-Andreson, C.

D = Avanzo, M. Babione, C.-H. Sham, J. Brawley, and K. Lajtha. 1992. Couplings of

watersheds and coastal waters: sources and consequences of nutri ent enrichment in Waquoit Bay, Massachusetts. Estuaries 15:443-457. Valiela, I., J. McClelland, J. Hauxwell, P. J. Behr , D. Hersh, and K. Foreman. 1997. Macroalgal

blooms in shallow estuaries: controls and ecophysiological and ecosystem consequences. Limnology and Oceanography 42:1105-1118. Valiela, I., G. Tomasky, J. Hauxwell, M. L. Co le, J. Cebrián, and K. D. Kroeger. 2000a. Operationalizing sustainability: management and risk assessment of land-derived nitrogen loads to estuaries. Ecological Applications 10:1006-1023. Valiela, I., M. Geist, J. McClel land, and G. Tomasky. 2000b. N itrogen loading from watersheds to estuaries: verification of the Waquoit Ba y Nitrogen Loading Model. Biogeochemistry

49:277-293. Van Katwijk, M. M., L. H. T. Vergeer, G. H. W. Schmitz, and J. G. M. Roelofs. 1997. Ammonium toxicity in eelgrass Zostera marina . Marine Ecology Progress Series

157:159-173. 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 Wa dden Sea for restoration of Zostera marina beds. Helgoland

Marine Research 54:117-128. Whelan, P. M. and J. P. Cullinane. 1985. The algal flora of a subtidal Zostera bed in Ventry Bay, southwest Ireland. Aquatic Botany 23:41-51.9.1 

## 9.0 The Effects of Temperature a nd Desiccation on the Seagrasses

Zostera marina and Thalassia testudinum

David T. Specht and Bruce L. Boese

9.1 Background

Temperatures elevated above the ambient ra nge have long been recognized as having

deleterious or fatal effects on marine or ganisms (e.g. Sachs 1864, working with Vallisneria and Ceratophyllum; Mayer 1914). Environmental temperatur e influences the ra te of critical biological processes of organisms in general (B illings 1952; Vernberg 19 78), including processes such as photosynthesis and respiration in seagra sses. This review examines and summarizes work characterizing the effect s of natural and anthropogenica lly 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 i nvolving temperature effects on seagrasses, emphasizing that the literature is rife with conflicting views on the relative importance of temperature and irradiance, attributable to the f act 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 ma jor 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) documente d the dramatic effect of heated power plant effluent dilution on the lo cal 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 “s evere” damage. In the zone at ~1-2 ° C above ambient temperature, there was elevat ed productivity (see also Thorhaug, Blake and Schroeder 1978).

Aerial exposure and resulting desiccation stre ss is probably the most important factor limiting the upper intertidal distribution of seagra ss species. Although this is not directly an anthropogenic stressor, alterations in estuarine bathymetry which in crease elevations in intertidal areas would reduce seagrass populations by incr easing the frequency a nd duration of aerial exposure. For example, high rates of sediment ation from logging in coastal watersheds with steep slopes such as found in the Pacific Nort hwest 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 serv e to restrict potential9.2 distribution of eelgrass to sha llower 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 comprehens ively examine the effects of temperature on characteristics of populations of Zostera marina L., including growth ra te, onset of reproduction, senescence, and germination of seeds. Based on field observations, he proposed an optimal temperature window of 10 - 20 E 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, an d 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 ev ents along a latitudinal gradient from North Carolina to Canada indicated that reproduction occurred earlier in the south a nd 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 temperat ure 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, th e 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 characteri stic 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 pr obably limited the northern distribution of

Thalassia testudinum in Florida, although there were regi onal differences. In the Gulf of Mexico, T. testudinum is apparently capable of enduring a warm temperate climate, whereas along Florida's east coast exposur e 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/r hizome 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 fl owering 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 δ 13 C values indicates that seagrasses tend to become more 13 C depleted from tropical (warmer) to temperat e (cooler) regions (Hemminga and Mateo 1996).9.3 Variation (in order of relative importance) of source of carbon, irradiance and temperature on

CO 2 availability in seawater may partly explain th is 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 E C, the two isolates maintained growth differences. Their P

max data under experimental culture indicated that increased photosyntheti c 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) re ciprocally transplanted Z.

marina over temporal, spatial and ecogenetic gradients in North America; they demonstrated that initial leaf widths of transp lanting stock are not a good pred ictor of success. Puget Sound populations, coming from the mid-range of west-c oast geographic and genetic distribution, and low temperature variance habitat, exhibited broa d adaptive tolerance; Alas kan stock, originating in a highly variable thermal ha bitat, 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 reduci ng weight proportionally and respiration of non-photosynthetic tissues. Pollard and Greenway (1993) showed that Z. capricorni from warm, turbid Australian waters had higher than expect ed photosynthetic efficien cies than the same species from a relatively high light environm ent; plants accommodated to low-light turbid waters, at high water temperatures (29-33 ° C) reached H sat with small incremental light increases. One quarter of gross pr oduction 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-surfac e 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 i ndicate that the increase in temperature in the 20 th century is likely to have been the largest of any century during th e past 1,000 years. The increase in sea-surface temperature has been about one half that of the air temper ature increase (~0.2 ° C per decade) between 1950 and 1993. Tide gauge data show th at global average sea level rose between 0.1 and 0.2 m during the 20 th century, and that the global ocean h eat 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 20 th Century mean of ~11.7 ° C. The past nine years (1997-2006) are among the 25 warmest years on reco rd, 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, cha nging 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 geogra phic distribution of seagrasses, and will result9.4 in changes in the patterns of sexual reproduction. Another proposed effect of global warming is an increase in UV. However, Z. capricorni in Australia has been shown to acclimate readily to elevated UV and PAR by production of UV-blocki ng agents and, despite the enhancement of inhibitory effects of high PAR and UV-B by temper ature stress, the relatively small temperature changes associated with climate change are unlike ly 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 latit udinal gradients in North America (Phillips et al. 1983b). The time and temperature at which onset of reproductive phases occur differs between populations apparen tly representing different genotypes on the east and west coasts at the same latitude (see al so Gambi 1988; Phillips and Lewis 1983; Buia and Mazzella 1991; Inglis and Lincoln-Smith 1998; Ph illips 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 temper atures in the range of 10-18 E C cause delayed flowering, with subse quent dehiscence of immature fr uit (Phillips 1960; Moffler and Durako 1987; McMillan 1982). Thus there tends to be increasi ng failure of sexual reproduction as one approaches the northe rn 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 Th alassia 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 sexua l reproduction in response to temperature at the geographic extremes of Zostera distribution than in the temperate central range, where temperatures are relatively mode rate. 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 be ds at higher latitudes, which may reflect success of seed production (and, hence, successful se xual reproduction), due to seasonal temperature effects (see also Davis et al. 1999; Witz 1994; W itz and Dawes 1995; Kirste n 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 la ter flowering, resulting in

loss of immature fruits in res ponse to rapidly rising late spri ng temperatures. They credit Gessner (1970) with the general concept that se xual reproduction was less likely to be successful when aquatic plants encounter less-than-optimal temperatures - genetic va riance 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 E C
in Rhode Island, 10-20 ° C in Halifax, N.S., while Puget S ound temperatures varied only from 8-9.5 10 ° C over two sites. Distinguishable primordi al 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 microscopica lly 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 ma y 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 sola r-induced diurnal temperat ure cycling. Burke et al. (1996), discussing carbohydrate rese rves in eelgrasses, cite the agreement of their data on east

coast Z. marina non-structural carbohydrate (nsc) reserve drawdown as fall progresses to winter with San Francisco Bay population data (Zimmerm an et al. 1995), conditioned by high-turbiditycaused low light levels. Note that many author s cited elsewhere in this review observe the initiation and onset of gr owth of flowering buds in Zostera at very low temperatures in the dead of winter, precisely when nsc rese rves would be at their lowest in most habitats. Zimmerman et al. (1995) also notes that plants with the highest level of nsc rese rves show the highest level of asexual reproduction (i.e., transplant success = ve getative growth) at winter/spring onset of growth; there is a significant correlation of high nsc and growth with low turbidity and consequent higher light availabi lity. 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 anthes is may be suppressed at lower temperatures. Ramage and Schiel (1998) noted th at in New Zealand populations of Z. novazelandica in intertidal platforms, flowering s hoots were more numerous in the low intertidal zone than in upper zones, and two times more in patches bord ering 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). S ilberhorn 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 t hose changes could be attributed to temperature-induced changes in seed production: low water temperatur es 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 pr oduction of vegetative9.6 shoots. This vigorous growth tended to exhaus t 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) determ ined 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 temperatur es would be severely compromised.

A number of authors addressed the phe nomenon 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 (S onora, Mexico, French Mediterranean), with seasonal temperature variation, exhibited salinity-mediated dormancy, or showed no dormancy at

all.

Conclusions regarding temper ature effects on germinati on have changed over time, reflecting broader geographic investigations. Ph illips and Menez (1988) determined that water temperature, and not salinity, was the primary germination control. Hootsmans et al. (1987) experimentally determined that Z. marina seedling survival peaked at 10 ° C and 10-20 o /

oo , Z.

noltii survived best at 10 ° C and 1 o /

oo , and both species showed maximal germination at 30 ° C and 1‰ salinity. In contrast, Moore et al. (1993) and Brenchle y and Probert (1998)

demonstrated that Z. marina and Z. capricorni germination was highest at low temperatures and under anaerobic conditions, and lowest under aerobic conditions.

El Niño events may have local or regional effects on seagrasses. Seddon et al. (2000) reported that Z. muelleri and Z. mucronata (= Z. muelleri subsp. mucronata ) were among a

number of intertidal and shallow subtidal Australi an seagrasses that were drastically affected by high water temperatures associated with the 199 3 El Niño event. Nelson (1997) concluded that intertidal eelgrass plants would tend to declin e as a result of incr eased photoinhibition and desiccation due to increased temp erature and light during El Niño episodes. Thom et al. (2003)

found that warmer winters and cooler summ ers associated with the transition from el Niño to la

Niña ocean conditions corresponded with an incr ease in eelgrass abundance and flowering.

Harrison (1982a) attri buted a decline of Z. marina in the upper portion of a tidal flat drainage channel to impacts of warmed wate r receding with the ebbi ng 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 re productive activities before the putative lethal upper limit for the species is reached (30 ° C), which suggests that th e ultimate response to high water temperature is to behave as a true annual.

Burkholder et al. (1992) found a dverse effects of water colu mn nitrate enrichment in Z.

marina in mesocosms were exacerbated by increasing or high temperatures. The meristematic portion of the shoot disintegrated after several we eks of exposure when water temperatures were

held at 4 ° C above the 10 year (local) mean. Touc hette 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 is9.7 shown to allow significant incr ease 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 lim its of distribution, exhibited morphological and physiological symptoms of decline. Significantly reduced shoo t density, leaf and root production, and altered internal C and N content, support the premise that in more sout herly 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 te mperature regimes in culture, indicating some characte ristics of seagrass phenology can be modified by temperature.

Drew (1979) found that for a range of seagrass species, including Z. marina , lightsaturated gross photosynthetic rates increased in direct proportion to temp erature increase up to a point between 30-35 E C, above which thermal damage caused a rapid reduction. Noting that respiration rates were not so dramatically aff ected, he concluded that effects on gross and net photosynthetic rates differed only in slope. Bulthuis (1985, 1987) comprehensively reviewed temperature effects on photosynthe sis and growth of seagrasses. He found general agreement that the photosynthetic capacity of seagrasses (mos t seagrass genera repres ented) 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 experien ced 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 respira tion 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 re spiration and led to a reduction in P:R ratios. Burke et al. (1996), Th ayer 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 belo w 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 car bon reserves had been exhausted the previous summer under high temperature/high light conditions. Zimmerm an et al. (1995) note that daily H sat 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) conc luded that leaf productivity in Z. marina is limited by temperature, in that increas ing 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 acclima tion, seasonal changes in ambient temperature may not significantly aff ect light-saturated photosynthesis (H sat ) requirements and9.8 whole-plant C balance. Rapid mortality at hi gh temperatures during summer may result instead from thermal disrup tion of metabolism.

The impact of temperature on photosynthesis and respiration may also be affected by both salinity variations and pr eadaptation. Biebl and McRoy (1971) showed that both subtidal

and intertidal forms of Z. marina from Alaska maintained plasmatic resistance for short term (24 hr.) exposures over a range of salinities (0-3. 0x, i.e., 1.0x = 31, 3.0x = ~93) and between -6 ° C

and 34 ° C. Within these limits, photosynthesis increased with temperature in th e 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 te mperature-respiration re lationship toward the consumption of more oxygen in summer for any gi ven 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 distribu tion 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 distribut ion of the species (Evans et al. 1986). Dennison (1987) showed that southern populations of Z. marina exhibit bimodal seas onal patterns of net photosynthesis, due to high respiration duri ng summer months, while northern populations exhibit a more unimodal seasonal pattern, presum ably due to temperature differences between southern and northern populations.

Pérez-Lloréns and Niell (1993, 1994) reporte d 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), suggesti ng 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 main tenance of ionic condition (Schroeder 1975), and that physiologi cal changes such as chlorophyll a fluorescence values will occur prior to observable morphometric cha nges (Byron and Fourqurean 2004). Capone and Taylor (1980) found that C 2 H 2 (acetylene) reduction rates were halved by a decrease of 10 E C

and that rates of N 2 fixation varied ~20-fold, being maxi mal in late summer and minimal in January. Zieman (1970) determined that net le af productivity was signi ficantly temperature dependent, with optimal growth occurring from 23-31 ° C (see also Barber and Behrens 1985).

Borum et al. (2005) and R udnick et al. (2005) review ed 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 el evated above the seasonal ambient level.9.9 Sand-Jensen and Borum (1983) ( Z. marina in Danish waters) and Kerr and Strother (1985, 1989) ( Z. muelleri in Australian waters) concl uded that leaf productivity and photosynthetic ability respectively ei ther were not affected by wa ter temperature, or could be maintained under extreme conditions; secondary mechanisms such at those affecting dark respiration rates and rates of mi neralization 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 temperat ure, while those subjected to high light intensity increased production of phenols. Phenolic com pounds 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 experime nts with eelgrass (Bintz et al. 2003) which combined nutrient enrichment w ith sustained temperature elevation to 4 ° C above a 9-year mean caused significant declines in num ber 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 1997-1 998 ENSO event foretell the possibility of widespread long-range habitat c onversions if average water temp eratures 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 eelg rass 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 le af 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 im portant co-factor (F elger and McRoy 1974).

Poumian-Tapia and Ibarra-Obando (1999) demons trated that seasona l changes in aboveground biomass and leaf area index (LAI) of Z. marina from a Mexican coastal lagoon were associated with water temperature.9.10 At the other end of the temperature sp ectrum, Harrison and Mann (1975) discounted temperature as an important factor in eelgrass growth for a subtid al population in Nova Scotia, as most of the vegetative growth occu rred 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 an d Borum (1984) also conc luded that Setchell’s growth regulation by temperat ure theory was not supported by their observations of Z. marina populations in Denmark, observing th at ~75% of seasonal variance in leaf growth rate could be

attributed to variation in surf ace 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 uppe rmost part of the leaf canopy.

A recent comprehensive review of eelgrass research in San Francisco Bay, CA. (WyllieEcheverria and Fonseca 2003) suppo rted the original proposals of Setchell (1922, op. cit.) regarding the effects of temperat ure on the biotic responses of eel grass (e.g., Phillips et al. 1983; Merkel and Associates 1999 (cited in Wyllie-E cheverria and Fonseca 2003); and Zimmerman et al. 1995). They conclude that tr ansplantation and restoration effo rts pay particular attention to season and temperature regimes in planning and scheduling, as many wi ntertime transplants were significantly inhibited by a combinati on of high turbidity and low temperatures.

Kirkman et al. (1982) (New South Wales, Aust ralia) demonstrated a closer relationship

for Z. capricorni productivity to water temperature than to solar radiation. The partial correlation coefficients between growth and wate r 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 temperatur e not accounted for by solar radiation or solar radiation lagged by 1 month.

Marbà et al. (1996) found leaf a nd shoot growth in Mediterranean Z. marina populations to be associated primarily with average irradiance (R 2 = 0.43 leaf, 0.70 shoot), temperature variation was secondary (R 2 = 0.57 leaf, 0.10 shoot); Z. noltii , growing in the same area, but

elevated in the intert idal with respect to Z. marina , was co-dependent on temperature (R 2 = 0.62 leaf, 0.32 shoot) and light (R 2 = 0.64 leaf, 0.37 shoot), althoug h the light relationship lagged by

~1 month.

Harrison (1982a), Phillips a nd 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 defo liation, 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 especial ly 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 in fluence. Biomass measures (shoots per unit area, plant size) are greatest near the ocean, tape ring off proceeding upstream into warmer water.9.11 During the summers of 1998-99, the overlying water column downstream was, on average, colder, more saline, less turbid, had more ava ilable light and higher dissolved inorganic nitrogen (DIN) than upstream. Temperature was negativ ely (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 p oorly 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) reporte d that leaf growth rates for T. testudinum in Biscayne Bay, Florida (data collected in 1969-70), was highe st when the temperature was between 28-31 E C and salinities ~30, and at stressed habitats, the minimum growth rates were when salinity was low (13-15) and temperatur e 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 shoul d be at a depth below the MLW (Koch 2001). This relationship has clearly been shown in tropical seagrasses where aerial exposure associated with extreme

tides results in seasonal losses of above-ground biomass (Vermaat et al. 1993; Erftemeijer and Herman 1994; De Iongh et al. 1995 ; Stapel et al. 1997). In an extreme case the upper margin of

a Zostera noltii bed was described as A 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 regrow th 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 seagra ss abundance and morphology across the tidal gradient (Bayer 1979; Jacobs 1979; Kentula and McIntire 1986; Keddy 1987; Koch and Beer 1996). Thalassia testudinum starts to become exposed at about a 15-cm water depth, because the blades are somewhat rigid (Phillip s 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 combin e to enhance desiccation of tidally-exposed plants (Straw n 1953, 1961; Phillips 1960). De siccation 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 su mmer temperatures) (P hillips 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 de nsity than subtidal popula tions. In these central ranges it is often found in three zones which ha ve been defined by plant growth characteristics9.12 and related to the degree of tidal exposure. For example Bayer (1979) and Boese et al. (2003) found and defined these zones in Yaquina Bay (New port, OR, semi-diurnal tide, range ~2.4 m) as: 1) a subtidal and lower intertidal pere nnial zone (below 0.25 m MLLW) which consists mainly of perennial shoots which grow vegetati vely from below-ground rhizomes which persists throughout the year, 2) a transi tion zone between 0.25-0.75 MLLW consisting of annual shoots and perennial patches, and 3) an upper inte rtidal zone (0.75 to 1.5 m MLW) which is characterized by annual shoots whic h grow from seeds and are abse nt in winter. Roughly the same growth pattern is representative of eelgr ass 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-inte rtidal zone of annual plants (van Katwijk et al. 2000).

Occasionally Z. marina annuals and perennials are found above these zones, and are associated with micro-topographical features (e .g. tide pools, drainage channels) which would tend to retain water longer during ebb tides. Thus it appears that the upper limit for Z. marina is controlled not directly by tidal elevation but by the duration of water coverage (Kentula and McIntire 1986; Jacobs 1979). It is also possibl e that eelgrass biomass can structurally retard water loss by trapping water during ebb tides as ha s 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 de siccation. 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 seagra ss may reduce the desic cating effects of aerial exposure. Penhale and Smith (1977) noted that Z. marina plants, which had epiphytes removed, lost five times the amount of dissolved organi c carbon following exposure to 1 h of laboratory desiccating conditions, when compar ed to plants with encrusting epiphytes. This suggested to them that the encrusted epiphytes trapped wate r interstitially during the receding tide, which diminished or prevented desicca tion damage during aerial exposure.

The zonal differences between Z. marina annual and perennial form s appear to be due to morphological differences. van Katw ijk 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 sh eaths that could not lie flat on the sediment. These upright sheaths tended to desiccate rapi dly when exposed. These observations were confirmed by Boese et al. (2003), who noted th e 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 photosynthes is at lower leaf water content than Z. marina (Leuschner et al. 1998). Simila r m orphological differences have been noted in tropical seagrass species with tolerance to high temperature and aerial expos ure, correlating with smaller and narrower leaves (McMillan 1984). Although Z. marina is less tolerant of desiccation, exposed leaves are capable of photosynthesi s as long as leaves are moist, and are also able to recover

from mild desiccation when leaves are re-wetted (Leusc hner and Rees 1993).9.13 9.6.2 Effect of tidal amplitude The interaction between tidal exposure and light availability are the principal forcing factors determining the depth rang e 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 tolera nce of desiccation, the minimum depth (Z min ) 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 wa ter 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 (Z

max ) is due to light limitation, then no seagrass can survive if Z

max ≤ Z min . Thus as the tidal range increases, the depth range where seagrasses can occur is linearly decreased (Koch 2001).

This theoretical relationship is supported by the results of Koch and Beer (1996). That study attempted to explain the disappearance of Z. marina from the western portions of Long Island Sound while it persisted in the eastern por tions. Light attenuati on had increased due to eutrophication and the greater tide range in the western portion of the sound appeared to narrow the depth range within which Z. marina could grow, making this shallow population vulnerable to storm events (Koch and Beer 1996). Koch and Beer (1996) concluded that regulatory management of seagrass should take in to account tides as well as light.

9.6.3 Effects of desiccation At the population level, aerial exposure and associated desicca tion stress affects Z.

marina in a variety of ways. With increasing intertidal exposure, shoot density is reduced, leaves become narrower, canopy height is reduced, annual plants become more prevalent and flowering increases; these effects are associated with pe rennial eelgrass meadows becoming progressively more patchy with increasing tide height (Bayer 1979; Jacobs 1979; Kentula and McIntire1986; Harrrison 1982a; Keddy 1987; Ruckleshaus 1994; Koch and Beer 1996; Boese and Robbins unpublished data).

At the individual shoot level, desiccation reduces the plant = s photosynthetic ability (Leuschner and Rees 1993), and leaves become physically damaged (van Katwijk et al. 2000; Boese et al. 2003). The progression of damage is an initial drying of the leaf in and immediately adjacent to the exposed sheaths, followed by gradual loss of pigment over a 7- 14 d interval. These non-pigmented areas often occur in bands across individual leaves which probably affect translocation of photosynthetic products, and we aken the leaf making it more vulnerable to breakage. Desiccation-induced breakage can co ntribute to the reduced canopy height often observed in high intertidal pl ants (Boese et al. 2003).

Desiccating events are episodic, and often associ ated with daylight sp ring tides during the summer, as illustrated in the seasonal losses of intertidal seagrass that occur in the tropics (Vermaat et al. 1993; Erftemeijer and Herman 1994; De Iongh et al. 1995; Stapel et al. 1997; van Lent et al. 1991). Similar episodic event ha ve been observed in Ya quina Bay, Oregon during9.14 daylight low tides in the spring and early summ er, 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 inte rtidal seagrass populat ions and how desiccation stress directly affects individual shoots and populations. Development of plant and population

level models of the effects of e xposure on eelgrass would be useful.

With respect to temperature effects, knowledge of site-specific temperature A windows @ regarding vegetative growth, the initiation of reproduction (anthesi s, pollination, seed set and dispersal), and upper thresholds fo r inhibition of growth, leaf loss and dark respiration seem to be the criteria that would be required for any rest oration or preservation e ffort. Elevated levels of nitrogen, either as nitrate or ammonia, in combination with ev en slightly elevated above norm water temperatures could be held as a violation of acceptable habitat conditions. Additional multifactorial, mesocosm experiments with vary ing temperature and nitrogen concentrations could help define acceptable boundaries from a regulatory standpoint.

9.8 Literature Cited

Adams, J. B. and G. C. Bate. 1994. The tolerance to desiccation of the submerged macrophytes Ruppia cirrhosa (Petagna) Grande and Zostera capensis Setchell. Journal of Experimental Marine Bi ology and Ecology 183:53-62. Barber, B. J. and P. J. Behrens. 1985. Effects of elevated temperature on seasonal in situ leaf productivity of Thalassia testudinum Banks ex König and Syringodium filiforme Kützing. Aquatic Botany 22:61-69. Bayer, R. D. 1979. Intertidal zonation of Zostera marina in the Yaquina estuary, Oregon. Syesis 12:147-154. Beardall, J., S. Beer, and J. A. Raven. 1998. Bi odiversity of marine plants in an era of climate change: Some predictions based on physiolo gical performance. Botanica Marina

41:113-123. Biebl, R. and C. P. McRoy. 1971. Plasma tic resistance and rate of respiration and photosynthesis of Zostera marina at different salinities a nd temperatures. Marine Biology 8:48-56. Billings, W. D. 1952. The environmental complex in relation to plant growth and distribution. Quarterly Review of Biology. 27:251-265. Bintz, J. C., S. W. Nixon, B. A. Buckley, and S. L Granger. 2003. Impacts of temperature and nutrients on coastal lagoon plant co mmunities. Estuaries 26:765-776. Boese, B. L., K. Alayan, E. Gooch, and B. D. Robbins. 2003. Desi ccation index: A direct measure of the effects of tidal exposure on eelgrass ( Zostera marina ) in an Oregon estuary. Aquatic Botany 76:329-337.9.15 Borum, J. 1980. Biomasse og productionsforhold hos ålegræs ( Zostera marina L.) og det tilknyttede epifytsamfund. M.S. Thesis, Univ ersity of Copenhagen. 114 pp. (cited in Wium-Andersen and Borum, 1984) Borum, J., O. Pedersen, T. M. Greve, T. A. Fr ankovich, J. C. Zieman, J. W. Fourqurean, and C. J. Madden. 2005. The potential role of pl ant oxygen and sulphide dynamics in die-off events of the tropical seagrass, Thalassia testudinum . Journal of Ecology 93:148-158. Brenchley, J. L. and R. J. Probert. 1998. S eed germination responses to some environmental factors in the seagrass Zostera capricorni from eastern Australia. Aquatic Botany

62:177-188. Buia, M. C. and L. Mazzella. 1991. Reproduc tive phenology of the Me diterranean seagrasses

Posidonia oceanica (L.) Delile, Cymodocea nodosa (Ucria) Aschers., and Zostera noltii Hornem. Aquatic Botany 40:343-362. Bulthuis, D. A. 1985. The effects of temperat ure on photosynthesis and gr owth of seagrasses.

Estuaries 8:14A. Bulthuis, D. A. 1987. Effect s of temperature on photosynthesis and growth of seagrasses. Aquatic Botany 27:27-40. Burke, M. K., W. C. Dennison, and K. A. Moore. 1996. Non-structural carbohydrate reserves of eelgrass Zostera marina . Marine Ecology Progress Series 137:195-201. Burkholder, J. M., K. M. Mason, and H. B. Glas gow Jr. 1992. Water column nitrate enrichment promotes decline of eelgrass Zostera marina : Evidence from seasonal mesocosm experiments. Marine Ecology Progress Series 81:163-178. Byron, D. A. and J. W. Fourqurean . 2004. Fluorescence responses of Thalassia testudinum to variations in salinity and temperature in Florida Bay. Abstract: 33 rd Annual Marine Benthic Ecology Meeting, Mobile, AL, March 25-28, 2004. Capone, D. G. and B. F. Taylor. 1980. N 2 fixation in the rhizosphere of Thalassia testudinum

. Canadian Journal of Microbiology 26:998-1005. Churchill, A. C. and M. I. Riner. 1978. Anthesis and seed production in Zostera marina L. from Great South Bay, New York, US . Aquatic Botany 4:83-93. Davis, J. L., D. L. Childers, and D. N. K uhn. 1999. Clonal variation in a Florida Bay Thalassia

testudinum meadow: molecular genetic assessmen t of population structure. Marine Ecology Progress Series 186:127-136. Dawes, C. J., R. C. Phillips, and G. Morrison. 2004. Seagrass Communities of the Gulf Coast of Florida: Status and Ecology. Florida Fish and Wildlife Conservation Commission, Fish and Wildlife Research Institute and the Tampa Bay Estuary Program. St. Petersburg, FL. iv + 74 pp. de Cock, A. W. A. M. 1980. Fl owering, pollination and fruiting in Zostera marina L. Aquatic Botany 9:201-220. de Cock, A. W. A. M. 1981. Influence of temperature and varia tions in temperature on flowering in Zostera marina L. under laboratory conditions . Aquatic Botany 10:125-131. de Iongh, H. H., B. J. Wenno, and E. Meelis. 19 95. Seagrass distribution and seasonal biomass changes in relation to dugong grazing in the Moluccas, East Indonesia. Aquatic Botany 50:1-20, 1995.9.16 den Hartog, C., and J. Kuo. 2006. Taxonomy an d Biogeography of Seagrasses. pp. 1-23. In: Larkum, A.W. D., R.J. Orth and C.M. Duarte (Eds.) Seagrasses: Biology, Ecology and Conservation. Springer-Verlag, Berlin. Dennison, W. C. 1987. Effects of light on seagrass photosynthesis, growth and depth distribution. Aquatic Botany 27:15-26. Drew, E. A. 1979. Physiological aspects of pr imary production in seagrasses. Aquatic Botany

7:139-150. Erftemeijer, P. L. A. and P. M. J. Herman. 1994. Seasonal changes in environmental variables, biomass, production and nutrient contents in tw o contrasting tropical intertidal seagrass beds in South Sulawesi, Indonesia. Oecologia 99:45-59. Evans, A. S. 1983. Growth and photosynthetic responses to temperatur e of two populations of

Zostera marina . Biological Bulletin (Woods Ho le) 165:508. (Abstract, General Scientific Meetings of the Marine Biologi cal Laboratory, August 16-19, 1983, Woods Hole, MA.) Evans, A. S., K. L. Webb, and P. A. Penhale. 1986. Photosynthetic temperature acclimation in coexisting seagrasses, Zostera marina L. and Ruppia maritima L. Aquatic Botany

24:185-197. Felger, R. and C. P. McRoy. 1974. Seagrasses as potential food plants. In : Somers, G. F. (Ed.) Seed-Bearing Halophytes as Food Plants. Proc eedings of a Conference, University of Delaware, June 10-11, 1974. Univ. De laware Sea Grant DELU-W-74-001, NOAA Office of Sea Grant, U.S. Dept Comm erce, Grant no. 2 35223; DEL SG 3 75. Fletcher, S. W. and W. W. Fletcher. 1995. Factors affectin g changes in seagrass distribution and diversity patterns in the Indian-River-Lagoon complex between 1940 and 1992.

Bulletin of Marine Science 1:49-58. Gambi, M.C. 1988. Flowering in a Zostera marina bed of San Juan Island (Washington, U.S.A.) during winter. Aquatic Botany 30:267-272. Gessner, F. 1970. Temperature: Plants. Pp. 363-406 in Kinne, O. (ed.). Marine Ecology: A Comprehensive, Integrated Treatise on Life in Oceans and Coastal Waters. Volume 1.

Environmental Factors. Part 1. Wile y-Interscience, London, England. 681 pp. 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. The Netherlands Journal of Sea Research 25:395-404. Glynn, P. W. 1968. Mass mortalities of echinoids and other reef flat organisms coincident with midday low water exposures in Puer to Rico. Marine Biology 1:226-243 Harrison, P. G. 1982a. Spatial and temporal patt erns in abundance of tw o intertidal seagrasses,

Zostera americana den Hartog and Zostera marina L. Aquatic Botany 12:305-320.

Harrison P. G. 1982b. Seasonal a nd year-to-year variations in mixed intertidal populations of Zostera japonica Aschers. & Graebn. and Ruppia maritima L. S.L. Aquatic Botany.

14:357-371. Harrison, P. G. and K. H. Mann. 1975. Chemi cal changes during the s easonal cycle of growth and decay in eelgrass ( Zostera marina ) on the Atlantic coast of Canada. Journal of the

Fisheries Research Board of Canada 32:615-621. Hemminga, M.A. and M.A. Mateo. 1996. Stable carbon isotopes in seagra sses: Variability in ratios and use in ecological studies. Marine Ecology Progress Series 140:285-298.9.17 Holmquist, J. G., G. V. N. Powell, and S. M Sogard. 1989. Sediment, water level and water temperature characteri stics of Florida Bay = s grass-covered mud banks. Bulletin of

Marine Science 44:348-364. Hootsmans, M. J. M., J. E. Vermaat, and W. van Vierssen. 1987. Seed-bank development, germination and early seedling survival of two seagrass species from the Netherlands:

Zostera marina L. and Zostera noltii Hornem. Aquatic Botany 28:275-285. Ibarra-Obando, S. E., C. -F. Boudouresque, and M. Roux. 1997. Leaf dynamics and production

of a Zostera marina bed near its southern distributi onal limit. Aquatic Botany 58:99-112. Inglis, G. J. and M. P. Lincoln-Smith. 1998. Synchronous flowering of estuarine seagrass meadows. Aquatic Botany 60:37-48. Irlandi, E., B. Orlando, S. Macia, P. Biber, T. Jo nes, L.Kaufman, D. Lirman, and E. T. Patterson.

2002. The influence of freshwater runoff on bi omass, morphometrics, and production of

Thalassia testudinum . Aquatic Botany 72:67-78. IPCC. 2001. Climate Change 2001: The Scientif ic Basis. Contributi on of Working Group I to the Third Assessment Report of the Inte rgovernmental Panel on Climate Change [Houghton. J.T., Y. Ding, D.J. Griggs, M. N oguer, P.J. van der Linden, X. Dai, K. Maskell, and C.A. Johnson (eds.)]. Cambri dge University Press, Cambridge, U.K. and New York, NY, USA, 881 pp. http:// www.grida.no/climate/ipcc_tar/wg1/005.htm. Jacobs, R. P. W. M. 1979. Distribution and as pects of the production and biomass of eelgrass,

Zostera marina L., at Roscoff, France. Aquatic Botany 7:151-172. Johnson, M. R., S. L. Williams, C. H. Lieber man, and A. Solbak. 2003. Changes in the abundance of the seagrasses Zostera marina L. (eelgrass) and Ruppia maritima L. (widgeongrass) in San Diego, California, followi ng an El Niño event. Estuaries 26:106-

115. Keddy, C. J. 1987. Reproduction of annual eelg rass: Variation among habitats and comparison with perennial eelgrass ( Zostera marina L.). Aquatic Botany 27:243-256. Kentula, M. E. and C. D. McIntire. 1986. The autecology and production dynamics of eelgrass ( Zostera marina L.) in Netarts Bay, Oregon. Estuaries 9:188-199. Kentula, M. E. and T. H. DeWitt. 2003. Abundance of seagrass ( Zostera marina L.) and macroalgae in relation to the salinity-temper ature gradient in Ya quina Bay, Oregon,

USA. Estuaries 26:1130-1141. Kerr, E. A. and S. Strother. 1985. Effect s of irradiance, temperature and salinity on photosynthesis of Zostera muelleri . Aquatic Botany 23:177-183. Kerr, E. A., and S. Strother. 1989. S easonal changes in leaf growth rate of Zostera muelleri

Irmisch ex Aschers. in south-eastern Aust ralia. Aquatic Botany 33:131-140. Kirkman, H., I. H. Cook, and D. D. Reid. 1982. Biomass and growth of Zostera capricorni Aschers. in Port Hacking, N.S.W., Australia. Aquatic Botany 12:57-67. Kirsten, J. H., C. J. Dawes, and B. J. Cochrane. 1998. Randomly amplified polymorphism detection (RAPD) reveals high genetic diversity in Thalassia testudinum Banks ex König (Turtlegrass). Aquatic Botany 61:269-287. Koch, E. W. 2001. Beyond light: physical, geolog ical, and geochemical parameters as possible submersed aquatic vegetation requi rements. Estuaries 24:1-17. Koch, E. W. and S. Beer. 1996. Ti des, light and the distribution of Zostera marina in Long Island Sound, USA. Aquatic Botany 53:97-107.9.18 Komar, P. D. 1997. Sediment Accumulati on in Tillamook Bay, Oregon, a Large DrownedRiver Estuary. Report for the Tillamook Bay National Estuary Project, May, 1997. 23 pp. (<http://www.tbnep.org/re ports/sedimentation.html>). Larkum, A. W. D., A. J. McComb , and S. A. Shepherd. 1989. Bi ology of Seagrasses. A treatise on the biology of seagrasses with special re ference to the Austra lian region. Aquatic Plant Studies 2. Elsevier, New York. Pp. 656-659. Leuschner, C. and U. Rees. 1993. CO 2 gas exchange of two intertidal seagrass species, Zostera

marina L. and Zostera noltii Hornem., during emersion. Aq uatic Botany 45:53-62. Leuschner, C., S. Landwehr, and U. Mehlig. 1998. Limitation of carbon assimilation of

intertidal Zostera noltii and Z. marina by desiccation at low ti de. Aquatic Botany 62:171-

176. Linton, D., and T. Fisher (Eds.). 2004. Ca rribean Coastal Marine Productivity Program: 1993-

2003. CARICOMP. ISBN 976 610 624 X. Loques, F., G. Caye, and A. Meinesz. 1990. Germination in the marine phanerogam Zostera

noltii Hornemann at Golfe Juan, French Mediterranean. Aquatic Botany 38:249-260. Lutova, M. I. and N. L. Feldman. 1981. Genot ypic temperature adaptations of cellular functions and proteins in two Zostera species. Oecologia 49:92-95. Macauley, J. M., J. R. Clark, and W. A. Price. 1988. Seasonal change s in the standing crop and chlorophyll content of Thalassia testudinum Banks ex Konig and its epiphytes in the Northern Gulf of Mexico. Aquatic Bo tany 31:277-287. EPA/600/J-88/280, (ERL,GB 611). (Available from NTIS, Springfield, VA: PB89-209803). Marbà, N., J. Cebrian, S. Enríquez, and C.M. Duarte. 1996. Growth patterns of western Mediterranean seagrasses: Species-specific responses to seasonal forcing. Marine Ecology Progress Series 133:203-215. 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. Mayer, A. G. 1914. The effects of temperat ure upon tropical marine organisms. Carnegie Institute of Washington, Papers of the Tortugas Laboratory, 6:1-24. McManus, J., P. D. Komar, G. Bostrom, D Colber t, and J. J. Marra. 1998. Sediment Sources and the History of Accumulation in Tillamook Bay, Oregon. Final Report, The Tillamook Bay National Estuary Project Se dimentation Study. 59 pp., Appendix. (<http://www.tbnep.org/reports/sedimentation.html>). McMillan, C. 1978. Morphogeographic variation unde r controlled conditions in five seagrasses,

Thalassia testudinum

, Halodule wrightii

, Syringodium filiforme

, Halophila engelmannii

, and Zostera marina . Aquatic Botany 4:169-189. McMillan, C. 1982. Isozymes in seag rasses. Aquatic Botany. 14:231-243. McMillan, C. 1983. Seed germination for an annual form of Zostera marina from the Sea of Cortez, Mexico. Aquatic Botany 16:105-110. McMillan, C. 1984. The distribu tion of tropical seagrasses with relation to thei r tolerance of high temperatures. Aqua tic Botany 19:369-380, 1984. McMillan, C. and F. N. Moseley. 1967. Salinity tolerances of five marine spermatophytes of Redfish Bay, Texas. Ecology 48:503-506.9.19 McRoy, C. P. 1970. On the biology of eelgras s in Alaska. PhD Disse rtation, Univ. Alaska, College, AK. Pp. 132-156. 71-25,292 - Un iversity Microfilms, Ann Arbor, MI. Meling-Lopez, A. E. and S. E. Ibarra-Obando. 1999. Annual life cycles of two Zostera marina L. populations in the Gulf of California: co ntrasts in seasonality a nd reproductive effort. Aquatic Botany 65:59-69.

Merkel and Associates. 1999. Richmond Harbor Navigation Improvement Project, Postdredging eelgrass survey. Prepared for Tetr a Tech, Inc. San Francisco, CA. 22 pp. (cited in Wyllie-Echeverria and Fonseca 2003). Moffler, M. D. and M. J. Durako. 1987. Re productive biology of th e tropical-subtropical seagrasses of the southeastern United States . Pp. 77-88 in Durako, M. J., R. C. Phillips, and R. R. Lewis III (eds.). Proceedings of the Symposium on Subtropical-Tropical Seagrasses of the Southeastern United States. Florida Marine Research Publications No.

42. Florida Marine Research Institut e, St. Petersburg, Florida. 209 pp. Moore, K. A., R. J. Orth, and J. F. Nowak. 1993. Environmental regulation of seed germination

in Zostera marina L. (eelgrass) in Chesapeake Bay: effects of light, oxygen and sediment burial. Aquatic Botany 45:79-91. Nelson, T. 1997. Interannual variance in a su btidal eelgrass community. Aquatic Botany

56:245-252. Nienhuis, P. H. 1983. Temporal and spatial patterns of eelgrass ( Zostera marina L.) in a former estuary in the Netherlands, dominated by hu man activities. Marine Technology Society

Journal 17:69-77.

NOAA. 2006. 2006 Annual Clim ate Review. U.S. Summary. <http://www.ncdc.noaa.gov/oa/climate/re search/2006/ann/us-summary.html> Olesen, B. and K. Sand-Jensen. 1993. Seasonal acclimatization of eelgrass Zostera marina growth to light. Marine Ec ology Progress Series 94:91-99. Orth, R. J. and K. A. Moore. 1983. Seed germination and seedling growth of Zostera marina L. (eelgrass) in the Chesapeake Ba y. Aquatic Botany 15:117-131. Orth, R .J. and K. A. Moore. 1986. Seasonal an d year-to-year variations in the growth of

Zostera marina L. (eelgrass) in the lower Ches apeake Bay. Aquatic Botany 24:335-341.

Penhale P.A. and W. O. Smith Jr. 1977. Excretion of dissolved organic carbon by eelgrass ( Zostera marina ) and it epiphytes. Limnology and Oceanography 22:400-407. Pérez-Lloréns, J. L. and F. X. Niell. 1993. Temperature and emergence effects on the net photosynthesis of two Zostera noltii Hornem. morphotypes. Hydrobiologia 254:53-64. Pérez-Lloréns, J. L. and F. X. Niell. 1994. Photosynthesis in air: Comparative responses to different temperatures of two morphotypes of Zostera noltii Hornem. from Palmones River estuary (southern Spain). Verha ndlungen der Internationalen Vereinigung für Limnologie 25:2265-2269. Phillips, R. C. 1960. Observations on the ecolo gy and distribution of the Florida seagrasses. Professional Paper Series. Florid a Board of Conservation 2:1-72. Phillips, R. C. 1972. Ecological life history of Zostera marina L. (eelgrass in Puget Sound, Washington. PhD Thesis, University of Washington, Seattle. 73-3771 University Microfilms, Ann Arbor, MI, pp. 32 & 121. Phillips, R. C. and T. W. H. Backman. 1983. Phenology and reproductive biology of eelgrass ( Zostera ma rina L.) at Bahia Kino, Sea of Cortez, Mexico. Aquatic Botany 17:85-90.9.20 Phillips, R. C., W. S. Grant, and C.P. McR oy. 1983a. Reproductive strategies of eelgrass ( Zostera marina L.). Aquatic Botany 16:1-20, 1983. Phillips, R. C. and R. L. Lewis III. 1983. Infl uence of environmental gr adients on variations in leaf widths and transplant success in Nort h American seagrasses. Marine Technology Society Journal 17:59-68. Phillips, R. C., C. McMillan, and K.W. Bridges. 1983b. Phenology of eelgrass, Zostera marina L., along latitudinal gradients on Nort h America. Aquatic Botany 15:145-146. Phillips, R. C. and E. G. Menez. 1988. Seag rasses. Smithsonian Contributions to the Marine Sciences - Number 34. Smithsonian Institution Press, Washington, D.C., pp. 15-18. Pollard, P. C. and M. Greenway. 1993. P hotosynthetic characteristics of seagrasses ( Cymodocea serrulata

, Thalassia hemprichii and Zostera capricorni in a lowlight environment, with a comparison of leaf-m arking and lacunal-gas measurements of productivity. Australian Journal of Mari ne and Freshwater Re search 44:127-139. Poumian-Tapia, M. and S. E. Ibarra-Obando. 1999. Demography and biomass of the seagrass

Zostera marina in a Mexican coastal lagoon. Estuaries 22:837-847. Powell, G. V.N. and F. C. Schaffner. 1991. Water trapping by seagrasses occupying bank habitats in Florida Bay. Estuarine, Coastal and Shelf Science. 32:43-60. Ramage, D. L. and D. R. Schiel. 1998. Reproduction in the seagrass Zostera novazelandica on intertidal platforms in southern New Zealand. Marine Biology 130:479-489. Robblee, M. B., T. R. Barber, P. R. Carlson, M. J. Durako, J. W. Fourqurean, L. K. Muehlstein, D. Porter, L. A. Yarbro, R. T. Zieman, and J. C. Zieman. 1991. Mass mortality of the tropical seagrass Thalassia testudinum in Florida Bay (USA). Marine Ecology Progress

Series 71:297-299. Ruckleshaus, M. H. 1994. Ecological and gene tic factors affecting popu lation structure in the marine angiosperm, Zostera marina L. PhD Dissertation, Un iversity of Washington,

Seattle. Rudnick, E. T., P. B. Ortner, J. A. Browder, and S. M. Davis. 2005. A conceptual ecological model of Florida Bay. Wetlands 25:870-883. Sachs, J. 1864. Ueber die obere Temperturg ränze der Vegetation. Flora N. R. 22:139-215. 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. Schiel, D. R., J. R. Steinbeck, and M. S. Fo ster. 2004. Ten years of induced ocean warming causes comprehensive changes in marine benthic communities. Ecology 85:1833-1839. Schroeder, P. B. 1975. Thermal stress in Thalassia testudinum . Ph.D. Dissertation, University of Miami, Coral Gales, Florida. 157 p. Seddon, S., R. M. Connolly, and K. S. Edyvane. 2000. Large-scale seag rass dieback in northern Spencer Gulf, South Australia. Aquatic Botany 66:297-310. Setchell, W. A. 1929. Morphological and phenological notes on Zostera marina L. University

of California Publications in Botany, 14:389-452, 59 fig. Short, F. T. and H. A. Neckles. 1999. The effects of global climate change on seagrasses. Aquatic Botany 63(3-4):169-196. Silberhorn, G. M., R. J. Orth, and K. A. M oore. 1983. Anthesis and seed production in Zostera

marina L. (eelgrass) from the Chesap eake Bay. Aquatic Botany 15:133-144.9.21 Stapel, J., R. Manuntun, and M.A. Hemminga. 1997. Biomass loss and nutrient redistribution in

an Indonesian Thalassia hemprichii seagrass bed following seasonal low tide exposure during daylight. Marine Ecol ogy Progress Series 148:251-262. Strawn, K. 1953. A study of the dwarf-seahorse, Hippocampus regulus Ginsburg, at Cedar Key, Florida. M.S. Thesis, University of Florida, Gainesville, Florida. 114 pp. Strawn, K. 1961. Factors influencing the zona tion of submersed monocotyledons at Cedar Key, Florida. Journal of Wildlife Management 25:178-189. Styllas, M. N. 2001. Sediment accumulation and human impacts in Tillamook Bay, Oregon. Thesis, Master of Science in Oceanography, Oregon State University, Corvallis, Oregon. (<http://www.tbnep.org/reports/sedimentation.html>). Thayer, G. W., D. A. Wolfe, and R. B. Williams. 1975. The impact of man on seagrass systems.

American Scientist 63:288-296. Thom, R. M., A. B. Borden, S, Rumrill, D. L. Woodruff, G. D. Williams, J. A. Southard, and S. L. Sargeant. 2003. Factors influencing sp atial and annual variab ility in eelgrass ( Zostera

marina L.) meadows in Willapa Bay, Washington, and Coos Bay, Oregon, estuaries.

Estuaries 26:1117-1129. Thorhaug, A., N. Blake, and P. B. Schroeder. 19 78. The effect of heat ed effluents from power plants on seagrass ( Thalassia ) communities quantitatively comparing estuaries in the subtropic to the tropics. Mari ne Pollution Bulletin 9:181-187. Thorhaug, A., D. A. Segar, and M. A. Roessler. 1973. Impact of a power plant on a subtropical

estuarine environment. Marine Pollution Bulletin 4:166-169. Tomasko, D. A. and M. O. Hall. 1999. Productivity and biomass of the seagrass Thalassia

testudinum along a gradient of freshwater influe nce in Charlotte Harbor, Florida.

Estuaries 22:592-602. Touchette, B. W. and J. M. Burkholder. 2002. Seasonal variations in carbon and nitrogen constituents in eelgrass ( Zostera marina L.) as influenced by in creased temperature and

water-column nitrate. Botanica Marina 45:23-34. Touchette, B. W., J. M. Burkholder, and H. B. Glasgow Jr. 2003. Variations in eelgrass ( Zostera marina L.) morphology and internal nutrient composition as influenced by increased temperature and water column nitrate. Estuaries 26(1):142-155. van Katwijk M. M., D. C. R. Hermus, D. J. de Jong, R. M. Asmus, and V. N. de Jong. 2000. Habitat suitability of the Wa dden Sea for restoration of Zostera marina beds. Helgolander Marine Research. 54:117-128. van Lent, F., P. H. Nienhuis, and J. M. Verschuure. 1991. Production and biomass of the

seagrasses Zostera noltii Hornem. and Cymodocea nodosa (Ucria) Aschers. at the Band d'Arguin (Mauritania, NW Africa) a prelimin ary approach. Aquatic Botany 41:353-367. Vermaat, J. E., J. A. J. Beijer, M. J. M. Hoot smans, and C. J. M. Philippart. 1993. Leaf dynamics and standing stocks of intertidal Zostera noltii Hornem. and Cymodocea nodosa (Ucria) Ascherson on the Band d'Arguin (Mauritania). Hyrobiologia 258:59-72. Vernberg, F. J. 1978. Multiple-Factor and Synergistic Stresses in Aquatic Systems. Pp. 726- 747, In: Thorp, J. H. and J. W. Gibbons (E ds.). Energy and Environmental Stress in Aquatic Systems. Selected papers from a Symposium held at Augusta, Georgia, November 2-4, 1977. Technical Information Center, U. S. Department of Energy.

CONF-771114.9.22 Vergeer, L. H. T., T. L. Aarts, and J. D. de Groot. 1995. The > wasting disease = and the effect of abiotic factors (light intensity, temper ature, salinity) and infection with Labyrinthula

zosterae on the phenolic content of Zostera marina shoots. Aquatic Botany 52:35-44. Verhagen, J. H. G. and P. H. Nienhuis. 1983. A simulation model of production, seasonal changes in biomass and di stribution of eelgrass ( Zostera marina ) in Lake Grevelingen. Marine Ecology Progress Series 10:197-195. Witz, J. J. A. 1994. Flowering frequency, sex distribution, density and age of short shoots in

three west-central Florida Thalassia testudinum Banks ex König (Hydrocharitaceae) meadows. M.S. Thesis. University of South Florida, Tampa, Florida. Witz, M. J. A., and C. J. Dawes. 1995. Flowering and short shoot age in three Thalassia

testudinum meadows off west-central Florid a. Botanica Marina 38:431-436. Wium-Andersen, S. and J. Borum. 1984. Biom ass variation and autotr ophic production of an epiphyte-macrophyte community in a coas tal Danish area: I. Eelgrass ( Zostera marina L.) biomass and net production. Ophelia 23:33-46. Wyllie-Echeverria, S. and M. Fonseca. 2003. Eelgrass ( Zostera marina L.) research in San Francisco Bay, California from 1920 to the pr esent. Literature Review. NCCOS S.F. Bay Eelgrass ( Zostera marina ) Project. February, 2003; pp. 20; <http://response.restoration.no aa.gov/cpr/watershed/sanfranc isco/sfb_html/resources/bibs

.html>. Zieman, J. C. 1970. The effects of a thermal effluent stress on the sea-grasses and macroalgae in the vicinity of Turkey Point, Biscayne Bay, Florida. Ph.D. dissertation. University of Miami, Coral Gables, FL. 129 pp. Zieman, J. C. 1975. Seasonal variation of Turtle Grass, Thalassia testudinum König, with reference to temperature and salinity effects. Aquatic Botany 1:107-123. Zieman, J. C. and R. G. Wetzel. 1980. Producti vity in Seagrasses: Methods and Rates. pp. 101- 279. In: Phillips, R.C. and C.P. McRoy (Eds.) Handbook of Seagrass Biology: An Ecosystem Perspective. Garland STPM Press, New York. Zimmerman, R. C., J. L. Reguzzoni, S. Wyllie-Echever ria, M. Josselyn, and R. S. Alberte. 1995.

Assessment of environmental suitability for growth of Zostera marina L. (eelgrass) in San Francisco Bay. Aquatic Botany 39:353-366. Zimmerman, R. C., R. D. Smith, and R. A. Albe rte. 1989. Thermal acclimation and whole-plant

carbon balance in Zostera marina (eelgrass). Journal of E xperimental Marine Biology and Ecology 130:93-109.10.1 

## 10.0 The Effects of Bi oturbation and Bioirri gation 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, feed ing activities (e.g., bioturbation), or their ventilation of burrows and tubes (e.g., bioirriga tion). Compared to othe r limiting or facilitating factors, little mention is made of bioturbation or bioirr igation in the seagrass literature prior to the early 1980’s (i.e., prior to Suchanek 1983). Since then, however, seve ral studies (see below) and one review (Short and Wyllie-Echeverria 199 6) present compelling evidence that these processes may profoundly affect the local di stribution, abundance, and productivity of

seagrasses.

Generally speaking, biotur bation negatively affects seag rasses (although it may do the reverse in some circumstances), whereas bioirrig ation has positive or neutral effects. These processes are ecologically distinct from other impo rtant plant-animal interactions that can affect seagrasses, such as herbivory and epifaunal colo nization of seagrass leaves . Bioturbation is the biogenic transport of particulate matter (inorganic and organic) within the sediment column, resulting from burrowing, excavation, and f eeding activities of in faunal and epifaunal invertebrates, demersal fish and marine mammals , and birds. Bioturbatio n includes the vertical and horizontal mixing of particulate matter and associated porewater within the sediment column, the deposition of sediment from dept h 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 pumpi ng 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 respiratio n, 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 bioi rrigation to seagrasses, while less studied than the negative effects of bioturba tion, potentially include oxygena tion of sediments, import of nitrogen-rich particulate matter into the sediment (via suspension feeding), stimulation of organic matter remineralization, and enhancement of oxi dation or removal of toxic substances in sediments (i.e., hydrogen sulfide, ammonia) (s ee reviews by Aller 1988 ; Kristensen 1988; Pearson 2001). Bioturbation can have similar imp acts on sediment geochemistry as bioirrigation (i.e., enhancement of organic matter remineraliz ation, 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 bioturbati on/bioirrigation, nutri ent enrichment and seagrass condition, although hypotheses regarding cons equences of such interactions can be proposed. Both10.2 bioturbation and bioirrigation can reduce the adve rse impacts of eutrophication by increasing the rates of organic matter burial, below-gr ound decomposition, and nutrient cycling; by oxygenating sediment porewater; and by increasi ng the flux of dissolved metabolites and nutrients from sediments into the water column. Bioturbation can potentially increase the adverse effects of nutrient enrichment by decreasi ng water column light leve ls (i.e., by increasing turbidity as a result of sediment resuspensi on) or by acting as an independent stressor on nutrient-stressed plants. These in teractions, their relevance to s eagrass 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 determ ine the importance (and the dire ction 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 s uggest 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, gr owth, and survival by burial or uprooting of seeds, seedlings, shoots and patche s 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 bi oturbators and seagrasses, and obs ervations of burial of seagrass shoots by excavated sediments wh ere populations overlapped. Stronger evidence was provided in several studies featuring field or laboratory experiments in whic h seagrasses were transplanted into sediments containing different densities of biotur bators, 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 th e mechanisms by which bioturbation disturbs

seagrasses.

10.2.1 Burrowing Shrimp Bioturbation The most frequently reported bioturbato r-seagrass interactions are those involving thalassinid burrowing shrimp (Arthropoda: Decapoda : Thalassinidae) as the bioturbators. Burrowing shrimp have been reported to various ly 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), popul ation distribution (Suchanek 1983; Harrison 1987; Pranovi et al.1996; Sieb ert and Branch 2005; Dumbauld and WyllieEcheverria 2003), and community st ructure (Duarte et al. 1997). Post-larval stages of burrowing shrimp live in extensive burrow galleries excavat ed in estuarine and marine sediments, have prodigious rates of sediment turnover (reviewe d by Rowden and Jones 1993), and are important ecosystem engineering species in many coastal sy stems because of their influence on benthic10.3 Figure 10.1. Illustration of some of the adverse effects of bioturbation on seagrasses (adapted from Suchanek 1983, Figure 4). Arrows show paths of sediment subduction, advection and resuspension as result of sediment reworking by burrowing shrimp.

Table 10.1 Summary of sources of significant bioturbation impacts for different seagrasses

Bioturbation Source Seagrass Species Reference

Burrowing shrimp Zostera marina

, Z. japonica Thompson and Pritchard 1969

Z. marina

, Z. japonica Harrison 1987

Z. japonica Dumbauld and WyllieEcheverria 2003

Z. japonica, Z. capricorni Berkenbusch et al. 2007

Z. noltii Pranovi et al. 1996

Z. capensis Siebert and Branch 2005, 2006

Z. capensis Angel et al. 2006

Posidonia oceanica Molenaar and Meinesz 1995

Halodule uninervis Duarte et al. 1997

Thalassia testudinum Suchanek 1983

T. hemprichii Duarte et al. 1997

Syringodium isoetifolium Duarte et al. 1997 current

plume of ejected sediment covering

seagrasses

burrow under-cutting

seagrasses feeding pit excavating

seagrasses burrow mounds from
infaunal bioturbator, such as burrowing shrimp10.4 Cymodocea rotundata Duarte et al. 1997

C. serrulata Duarte et al. 1997

Stingrays Z. marina Orth 1975

Z. marina Merkel 1990

Z. marina

, H. wrightii Townsend and Fonseca 1998

H. wrightii Fonseca et al. 1994

T. testudinum Ogden 1980

T. testudinum Zieman 1982

T. testudinum Valentine et al. 1994

Syringodium sp. Fonseca et al. 1998

Crabs Z. capricorni (=novazelandica) Woods and Schiel 1997

Z. marina Davis and Short 1997

Z. marina Davis et al. 1998

T. testudinum Valentine et al. 1994

Polychaetes Z. noltii Phillipart 1994

Z. marina Luckenbach and Orth 1999

Z. marina Davis and Short 1997

Zostera spp. Hughes et al. 2000

Echinoderms Z. marina Backman 1984

community structure and geochemical proc esses (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 publis hed study that connected burrowing shrimp bioturbation with a decrease in seagrass c ondition and abundance. He noted a negative correlation in the spatial distribut ion 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 percen t 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 de nsity steadily decline in abundance over a five month period whereas turtlegrass transplanted into areas with low shrimp densities were “lush10.5 Figure 10.2. Effects of bi oturbating burrowing shrimp ( Callianassa spp.) on turtle grass ( Thalassia testudinum ) on subtidal sand flats on St. Cr oix island (US Virgin Islands); from Suchanek (1983). A. Field data show ing inverse relationships between burrowing shrimp abundance and seagrass percent cover at Tague Bay (solid line) and Great Pond Bay (dashed line), and with seagrass productiv ity (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 tran splanted into areas with high densities (experimental; dashed line) or low densitie s (control; solid line) of burrowing shrimp. Reproduced with permission from Journal of Marine Research and T.H. Suchanek. 10

0 20 30 40 50 60

OCT NOV DEC JAN FEB MAR

1980 1981 C o n t r o l B l a d e s

E

x

p

t ’

L B l a

d

e

s

C o n t ro l S h o o t s

Ex p

t

’ l Sh o

o

t s# Blades or Shoots 

## A

100

0

80

60

40

20

0

2 4 6 8 12 10 14 16 0 5

4

3

2

1 Great Pond Bay - Productivity

Great Pond Bay - %Cover

Tague Bay - %CoverMax. %CoverSeagrass & AlgaeMax. ProductivitySeagrass & Algae (g dry wt)

Callianassa Mounds m -2



## B

10

0 20 30 40 50 60

OCT NOV DEC JAN FEB MAR

1980 1981 C o n t r o l B l a d e s

E

x

p

t ’

L B l a

d

e

s

C o n t ro l S h o o t s

Ex p

t

’ l Sh o

o

t s# Blades or Shoots

10

0 20 30 40 50 60

OCT NOV DEC JAN FEB MAR

1980 1981 C o n t r o l B l a d e s

E

x

p

t ’

L B l a

d

e

s

C o n t ro l S h o o t s

Ex p

t

’ l Sh o

o

t s# Blades or Shoots 

## A

100

0

80

60

40

20

0

2 4 6 8 12 10 14 16 0 5

4

3

2

1 Great Pond Bay - Productivity

Great Pond Bay - %Cover

Tague Bay - %CoverMax. %CoverSeagrass & AlgaeMax. ProductivitySeagrass & Algae (g dry wt)

Callianassa Mounds m -2 100

0

80

60

40

20

0

2 4 6 8 12 10 14 16 0 5

4

3

2

1 Great Pond Bay - Productivity

Great Pond Bay - %Cover

Tague Bay - %CoverMax. %CoverSeagrass & AlgaeMax. ProductivitySeagrass & Algae (g dry wt)

Callianassa Mounds m -2



## B10.6 and healthy” (Figure 10.2.B). Suchanek (1983) proposed that burial by excavated sediments, or shading due to either turbidity or deposition of fine-grained resuspended sediment onto leaves were the likely mechanisms causing decreased turtlegrass productivity and survival. He concluded that burrowing shrimp could severely limit T. testudinum distribution and negatively affect the habitat and energeti c value of Caribbean seagrass beds. Whereas Suchanek (1983)

characterized the di rect effects of Callianassa spp. on turtlegrass, Wanless et al. (1988) reported

an indirect effect of bi oturbation by these shrimp on T. testudinum , positing that hurricane transported sediments from shrimp’s mounds (whi ch had a characteristic sediment grain size distribution) contributed to the smothering of seagrasses.

In subtidal sandy habitats (specifically the French Mediterranean), Molenaar and Meinesz (1995) reported that populations of the shrimp C. tyrrhena reduced survival of Posidonia

oceanica transplanted by ~40% relative to sands not inhabited by burrowing shrimp. They proposed that the transplants were buried by sediments excavated by C. tyrrhena , which is a prodigious bioturbator; at densities of 16 shrimp m -2 , they deposit ~600 g m -2 d -1 of sand on the seafloor (Ott 1976 in Molenaar and Meinesz 1995) . Pranovi et al. (1996, F. Pranovi, personal communication) observed a sharp decrease in the presence of the seagrass, Z. noltii , in Venice Lagoon (Italy) during the early 1990 ’s and a coincident increase in the abundance of the mud shrimp, Upogebia pusilla ; this pattern was also found in spat ial surveys conducted in the lagoon. Bioturbation by the shrimp was proposed as a mech anism to explain the di stribution patterns, but no experiments were conducted to test the hypothesis (F. Pranovi, personal communication).

Callianassid burrowing shrimp can limit the intertidal distribution of seagrasses. In Langebaan Lagoon (west coast of South Africa), Branch et al. (2003) and Siebert and Branch (2005) observed that the seagrass, Z. capensis , was relegated to the upper intertidal when cooccurring on tide flats with the ghost shrimp C. kraussi . Using an elegant field experiment, Siebert and Branch (2006) demonstrated that patches of seagrass increased in area when transplanted into lower intertidal areas from which the shrimp had been extirpated. However, if high densities of ghost shrimp were present, the s eagrass 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 a nd 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 callian assid shrimp did not colonize areas immediately down-slope of the seag rass 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 tran splanted ghost shrimp ( C. filholi ) or small patches of seagrass ( Z. capricorni ) into field enclosures containing the other species (i.e., seagrass or shrimp, respectivel y), and compared the a bundances of manipulated shrimp and seagrass relative to untransplanted, encl osed 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-domin ated enclosures, transplanted ghost shrimp abundances declined whereas the seagrass abundance was unchange d relative to the10.7 unmanipulated control treatment. They inferred that the density of the root mat prevented the ghost shrimp from burrowing into the sediment, as reported or inferred in other studies (i.e., Brenchley 1982; Harrison 1987; Siebert and Branch 2006).

Interactions between intertidal burrowing shri mp and seagrasses in NE Pacific estuaries have been the subject of severa l studies, particularly focuse d 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 inte ractions, however, were not unidirectional. In British Columbia (Canada), Harrison (1987) observed ne gative correlations in

the abundances of Z. marina, Z. japonica and ghost shrimp, and proposed that ghost shrimp bioturbation could prevent the seagrasses from colonizing unveget ated 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 s eagrasses, but that the seagrasses also could limit the shrimp’s distribution if the wate r column was sufficiently clear to allow earlylateral growth of the seagrasses in early Spring (see section 10.4). Dumbauld and WyllieEcheverria (2003) reported that burrowing shrimp bioturbation was an important factor determining recruitment success for Z. japonica seedlings in intertidal sediments of Willapa Bay estuary (Washington, USA). They observed that , following the eradication of burrowing shrimp (predominantly the ghost shrimp, N. californiensis ) by pesticide application, Z. japonica and Z.

marina often colonized intertidal habitats, whereas seagrass seedlings were sparse or absent from ghost shrimp dominated areas. They measured no significant difference in vertical distribution, viability or germination success of Z. japonica seeds among experimental plots (shrimpdominated sediments treated, or not, with pestic ide to kill the shrimp), but decreased growth and

loss of Z. japonica seedlings in plots containing ghost shri mp. 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 germ inated below that depth. Dumbauld and WyllieEcheverria (2003) concluded that “seedling survival is important for recruitment from the seed bank and while shrimp may cause some loss and decreased germination success, the effect of bioturbation on seedling survival, in part due to the coincident timing of shrimp activity and sprouting in early spring, is more important at the population level.” (pg. 37).

Thompson and Pritchard (1969) noted that the upper in tertidal limit of Z. marina in Oregon matched the lower limit of dense ghost sh rimp populations and that seasonal expansion of either species of eelgrass on sand flats was co rrelated with declines in the abundance of ghost

shrim p. In contrast to Dumbauld and Wyllie-E cheverria (2003), Thompson and Pritchard (1969) found that ghost shrimp dens ities declined in experime ntally created plots of Z. japonica

. Berkenbusch et al. (2007) reported much the same result from a field study in which ghost shrimp or small patches of Japanese eelgrass were reciprocally transplanted into field enclosures containing the other species (i.e., Z. japonica or shrimp, respectively). In all cases, the patches

of Z. japonica grew whereas the abundance of N. californiensis declined. They suggested that a combination of rapid growth by the eelgrass and relatively low bioturbation by N. californiensis10.8 (as compared to it’s New Zealand counterpart, C. filholi ) allowed the invasive eelgrass to be the superior competitor in these experiments. In laboratory experiments, Brenchley (1982) found

that the root and rhizome mat of Z. marina greatly inhibited burrowing by the burrowing shrimp N. californiensis and Upogebia pugettensis . For presumably the same reasons, Harrison (1987)

observed that N. californiensis were largely unsuccessf ul at colonizing dense Z. marina patches on a British Columbia tide flat. The key di fference 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 br ought into conflic t with the ghost shrimp. Synthesizing the results of expe rimental studies on in teractions between N. californiensis and Zostera spp. in NE Pacific estuaries, it appears that bioturbation by ghost shrimp can suppress the germination of eelgrass s eeds and the survival of seedlings (at least for Z. japonica and probably for Z. marina ), but that established patc hes of adult eelgrass are resistant to adverse impacts from ghost shri mp probably because the root and rhizome mat inhibits burrowing by the shrimp.

Thom et al. (2003) found burrowing sh rimp 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 thes e 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 Calif ornia, 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 tu rnover (which are lower for U. pugettensis ) and thus less disturbance to eelgrass by mud shrimp (T. H. DeWitt, personal observation), different capabilities to burrow through the eelgrass root and rhizome mat (Brenchley 1982), or increased susceptibility of N. californiensis to predation in lower intertidal and subtidal eelgrass habitats (Posey 1986). In South Africa, Siebert and Branch (2005, 2006) found high densities of the mud shrimp U. africana coexisting with Z. capensis in the upper intertidal where both species apparently found refuge from inte ractions 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 n eed to constantly excava te and ingest organicrich sediments as do deposit-feeding Callianassids.

In contrast to the suppression of populations of seagrasses in single-species meadows, burrowing shrimp bioturbation can increase seag rass species diversity within mixed-species meadows. Cumulative disturbance by burrowing shri mp to several species of tropical seagrasses in the Silaqui and Santiago Islands (Philippines) wa s estimated to be greater than the disturbance caused by sediment transport associated with hur ricanes and typhoons (Duart e et al. 1997). They conducted a field experiment to measure the eff ects of episodic deposition of sediment on the growth, survival, and demogra phy of seven seagrass species ( Thalassia hemprichii, Enhalus10.9 acoroides, Cymodocea rotundata, C. serrulata, Halodule uninervis, Syringodium isoetifolium,

and Halophilla ovalis , listed by relative abunda nce). 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 tr eatments; and one ( H. ovalis ) opportunistically increased in abunda nce in most treatments. Diffe rential 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 sc ale from sub-meter (burrowing shrimp) to 10 3 m (hurricanes and typhoons), with the frequency of occurrence being approximately inverse of their size (i.e., m onthly 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 s cale 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 sustai n higher seagrass specie s diversity than under conditions of very low or very high disturbance (sensu the interm ediate disturbance hypothesis; Sousa 1984).

In summary, disturbance from burrowing sh rimp bioturbation (particularly Callianassid ghost shrimp) significantly disturbs several specie s 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 suscep tible to burrowing shrimp bioturbation than are adult life stages. The interaction between s eagrasses and burrowing shrimp is not, however, always tilted in favor of the bioturbator. De nse root mats have been shown or suggested to prevent Callianassid ghost shri mp from burrowing into sediment s 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 seag rass growth (see section 10.4).

10.2.2 Stingray Bioturbation Stingrays (Vertebrata: Chondrichthyes: Mylioba tiformes) forage for benthic invertebrate prey by excavating sediments using jets of wate r blown from the mouth and plunger-like suction created with the pectoral disc (M artin 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 larg e areas of one Z. marina meadow (ca. 4 km 2 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 ( Aetobatis10.10 narinari ) were observed to damage turtlegrass ( T. testudinum ) habitat while feeding (Ogden 1980; Zieman 1982 in Valentine et al.1994).

Valentine et al. (1994) documented the pers istence 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 stingr ays. 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 belo w-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 persis tent bare patches in these T.

testudinum meadows.

By contrast, Townsend and Fonseca (1998) de termined that large bioturbation pits, possibly created by stingrays, were deeper (mean 4 cm, maximum 12 cm) than the rhizome depth of the predominant seagrasses ( Z. marina and Halodule wrightii ; 1 to 5 cm) in mixed-species seagrass meadows (North Carolina, USA). Many smaller pits, possibly created by crabs, were also deeper than the rhizomes . Therefore bioturbators coul d 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 impor tant process creating pa tchiness in seagrass beds and disrupting linkages among seagrass patc hes, increasing the risk of erosion of the seagrass patches by currents and waves (Fonsec a and Bell 1998). Thus, bioturbation may be important for generating and maintaining landsca pe-scale distribution patterns of seagrasses.

In addition to affecting extant populations of seagrasses, bioturbation by stingrays has been identified as an important source of dist urbance 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 (Califor nia, USA) (Merkel 1990). Transplanted eelgrass protected by stingray exclus ion 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, US A), and cages constructed of 2.5 cm mesh galvanized chicken wire improved H. wrightii survival to 60% relative to <1% survival in uncaged controls (Fonseca et al.1994). Stingray s’ impacts to transplant ed seagrass can occur very rapidly; Fonseca et al. (1998) reported 100% loss of Halodule sp. and Syringodium sp. transplants within 24 h of planting where stingray-exclusion cages were not used.

10.2.3 Crab Bioturbation A handful of studies have reported that bioturbation by crabs a nd 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. Valentine10.11 et al. (1994) determined that bioturbation by stone crabs ( Menippe spp.) was the most likely cause of open patches in T. testudinum meadows in St. Joseph Bay, although stingrays had at first appeared to be the culpri ts (see above). Stone crabs we re 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 t hus enlarging the area of seagra ss disturbed by the crabs. As Valentine et al. (1994) point out, stone crab burr owing 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 ha ve 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 l onger in seagrass patche s than on open mudflats, possibly because the interconnecting rhizomes a nd roots reduce cave-ins of burrow walls. As with stone crabs, burrowing by M. hirtipes appears to disrupt the sedi ment-binding properties of

Z. novazelandica ’s rhizome-root mat, dislodging the seag rass, requiring the cr abs to extend their burrows further into the seagrass patch, etc., se tting 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, crea ting a greater surface area for further erosion by waves and currents. The area immediately su rrounding the burrow is rapidly undermined and torn away, ultimately accelerating the erosion of th e 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 poi nts 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 bioturba tion 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 M iller (2006) reported that bioturbation by green crabs was responsible for the near ly 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 shea th 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 w ithin 1 wk10.12 of exposure to the crabs (Davis et al.1998). Cages alone were not su ccessful at keeping C.

maenas away from transplants, but using crab trap s in addition to exclusion cages improved the survival rate of transplanted eelgrass (Davis and Short 1997).

Bioturbation associated with horseshoe crabs ( Limulus polyphemus ) foraging activity in Great Bay Estuary (New Hampshire, USA) uprooted unprotected established plants and transplanted shoots of Z. marina (Davis and Short 1997); wire mesh cages successfully protected transplants from disturbance by horseshoe cr abs. 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 bi oturbation by these crab s significantly affects seagrasses distribution or surv ival 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 polychaet e bioturbation have also been reported, and will

be discussed in a later section. Phil lipart (1994) reported that populations of Z. noltii and the lugworm, Arenicola marina , maintained non-overlapping distribu tions on tide flats of the Dutch Wadden Sea (Terschelling, The Netherlands). The sp ecies’ distributions met at an abrupt border, even though both had similar substrate and emer sion limitations. Seagrass shoots transplanted into tide flat plots containing hi gh lugworm density (ca. 68 worms m -2 ) soon had sediments deposited upon them from the worms’ fecal cas tings 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 su rvival 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 tr eatments 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 eff ects on the dispersal and survival of Z. marina seeds in Chesapeake Bay (Virginia, USA) (Luckenbach an d 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, a nd seeds became passively trapped within small biogenic depressions or pits. Subsequently, many trapped seeds were buried by sediments10.13 reworked (defecated) by C. torquata . Trapping reduced the dispersal of eelgrass seeds, which potentially slows the expansion of seag rass 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 e xperiments, shoots of Z.

noltii that were protected from N. diversicolor had higher biomass, higher survival, and less damage to roots than unprotected transplant s (Figure 10.3). Worms apparently damaged Z. noltiiroots by burrowing and leaves by herbivory. However, the authors noted that bioturbationinduced damage to roots may have been an expe rimental design artifact, and thus the role of bioturbation in this worm-seagrass interaction is uncertain.

Figure 10.3. Results of two laboratory experiment s showing decreased survivorship of seagrass ( Zostera noltii ) planted into sediments containing polychaetes ( Nereis diversicolor ; open circles and triangles, and dashed or dotte d line) or devoid of worms (controls; filled circles and triangles, and solid lines) (modi fied from Figure 6 of Hughes et al. 2000).

Davis and Short (1997) noted that bioturbati on 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, af ter which the leaves were buried by bioturbated sediments leading to shoot death. No experime nts were conducted to test the hypothesis that eelgrass shoot death was clearly caused by activities of N. virens , although their natural history observations presented a plausible mechanism. 

### 1.5Survivorship (log)



### 0 0.5 1 2

012 5 34 6789 Days Polychaetes present (Nov. 1993 expt.)

Polychaetes present (Feb. 1994 expt.)

Control (Nov. 1993 expt.)

Control (Feb. 1994 expt.) 

### 1.5Survivorship (log)



### 0 0.5 1 2

012 5 34 6789 Days Polychaetes present (Nov. 1993 expt.)

Polychaetes present (Feb. 1994 expt.)

Control (Nov. 1993 expt.)

Control (Feb. 1994 expt.)10.14 10.2.5 Echinoderm Bioturbation Sea cucumbers and sand dollars (Echi nodermata: Holothuroidea and Echinoidea, respectively) have been obser ved to disturb seagrasses duri ng their burrowing and feeding activities. However, only two studies have experimentally examined whether bioturbation by echinoderms can affect seagrass growth, survival, or populations. Burrowing sea cucumbers ( Holothuria arenicola ) in the Bahamas were observed to smother turtlegrass ( T. testudinum ) with defecated sediments (Mosher 1980, in Valentine et al. 1994); however, whether this adversely affected the seagrass was not discussed. In field experiments, Backman (1984) demonstrated that dense populations of sand dollars ( Dendraster excentricus ) uprooted eelgrass ( Z. marina ) and inhibited colonization of unvegetated ha bitats in Puget Sound, Washington.

Sand dollars ( Mellita quinquiesperforata ) were frequently observed in unvegetated patches within subtropical turt legrass meadows of St. Joseph Ba y (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 th e 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 perimete r 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 di fference 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 burrowi ng stone crabs was responsible for erosion of the edges of turtlegrass beds a nd therefore the biogenic creation of open space in those seagrass meadows (see above).

Although there are few studies of echinode rm 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 s eagrass 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 echinode rms is more likely to occur along open coasts

rather than in estuaries.

10.2.6 Mechanisms of Bioturbation-Indu ced Disturbance to Seagrasses Relatively little is known about the actua l mechanism by which the seagrasses are harmed, compared to observations and experi ments that demonstrat e the net effects of bioturbators on seagrass shoot growth and survival. The most commonly reported mechanisms

in the studies reviewed above we re burial of seagrass shoots a nd seeds, uprooting of shoots and patches, undermining edges of seagrass patches, damaging roots or rh izomes, and shading by deposition of resuspended sediments onto leaves. Indirect mechanisms by which bioturbators were reported to damage seagrasses include redu cing water column light availability because of increased turbidity from resuspended fine sediments (Suchanek 1983), burial by storm-10.15 transported sediments previously excavated by bioturbators (Wan less 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 st udies were the magnitudes of th e disturbing process measured (e.g., burial rate, burrowing rate, sediment deposition rate, light reduction, et c.). In all of the studies reviewed above, the mechanism of distur bance 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 ro ots and rhizomes, sediment on leaves), biogenic structures in the environment (i.e., burrows, mo unds, and pits), or feeding habits of the bioturbators (i.e., absence of he rbivory). Transplant experime nts 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 caus e the growth and surviv al responses seen in the seagrasses (but see Duarte et al. 1997). Furt hermore, few of the studi es quantified the spatial scale over which the disturbance could be expect ed to occur. Ideally, this would require a comparison of the landscape and regional distributi ons of the bioturbator species, of the seagrass species, and of the habitat suitable for sust ainable seagrass growth, and experimental demonstration of the responses of seagrasses to bioturbation at random lo cations throughout that landscape. No study has put it al l together, measuring the natura l rates and magn itudes of the causal process of disturbance (i.e ., burial, etc), measuring the res ponses of appropriate seagrass life history stages to those leve ls of disturbance, and determin ing 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 biotur bator impacts to seagra ss populations and the functions of seagrass habitats, as stated in mo st of the papers (for example, “bioturbation by

Callianassa has a direct effect on seagrass beds but may indirectly influence a multitude of faunal relationships both within grass beds and in nearby communities such as coral reefs, mangroves, and the deep sea”, Suchanek 1983, p. 296).

10.3 Role of Bioturbation and Bioirrigat ion in Enhancing Seagrass Populations

Bioturbation and bioirrigation can benefit s eagrass populations and communities in two ways: 1) modification of sediment geochemistry to create conditions favorable for seagrass growth, and 2) facilitatio n of seagrass recruitment and coloni zation. 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 seagra ss meadows can be expected to

also increase.

10.3.1 Modification of Se diment Geochemistry Bioirrigation and bioturbation increase th e exchange of solutes and organic matter

between the water column and the sediment co lumn, strongly affecting sedimentary microbial processes such as organic matter decomposition, ammonification, nitrific ation-denitrification, and sulfide oxidation-reduction (A ller 1988; Kristensen 1988; Hopkinson et al. 1999; Fukukawa10.16 2001)

. To the extent that these sediment geochemical processes limit seagrass growth and survival (Chapter 6), bioirrigation and biot urbation can stimulate processes favorable for seagrass growth (e.g., fertilization), reduce conditions harmful to seagrass survival (i.e., eutrophication, hypoxia), and redu ce toxic substances within the rhizosphere (i.e., hydrogen sulfide, ammonia).

The role of bioturbating a nd bioirrigating benthic fauna in elevating sediment nutrient concentrations is well documented (Aller 1988 ; Kristensen 1988; Sc haffner et al.1992; Hopkinson et al. 1999; Felder 2001; Eyre and Ferguson 2002; Webb and Eyre 2004). Bioturbation and bioirrigation transfer organi c matter from the sediment surface and water column into the sediment column, enhancing ae robic 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 increas ing organic matter concentr ation, at least up to some threshold where ingestion rate is maximized or organic matter loadi ng causes toxicity (i.e., hypoxia or toxic contaminants as sociated 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 seag rass productivity or growth. Thayer and Fonseca (1984) suggested that under conditions of nutrient limitation, “nitrogen availability [for eelgrass, Z. marina ]...may be limited by the requirements of the heterotrophic community responsible for the decomposition of orga nic matter in sediments” (p. 31). Phillipart (1994) suggested that low densiti es 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 Re usch and Williams (1998) demonstrated that the mussels, Mytilus edulis and Muscalista senhousia , respectively, caused porewater concentrations of ammonium and phosphate to increase dramatic ally, and both papers suggested that these suspension feeders fertilized beds of Z. marina . Peterson and Heck (1999, 2001a, 2001b) demonstrated that the suspension feeding mussel ( Modiolus americanus ) enhanced productivity of turtlegrass ( T. testudinum ) through fertilization of sediments under nutr ient-limited conditions in St. Joseph Bay (Florida, USA). The mechan ism by which mussels fertilized seagrass beds was basically the same in all five papers: muss els ingested nitrogen-rich seston and transferred this material to the sediment surface via deposition of feces and pseudofeces; particulate organic nitrogen in the fecal matter was microbially re mineralized into dissolved inorganic nitrogen through ammonification and nitrific ation; and the dissolved inor ganic nitrogen was absorbed by seagrass roots and transformed into new tissue. Suspension feeding and deposit feeding benthic invertebrates (i.e., bioirrigators and small-scal e bioturbators) are common in seagrass beds, sometimes in great abundance (O rth et al. 1984; Beal 1994; Bachel et et al. 2000; Bowden et al. 2001), and it is highly likely that they also fertil ize 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 (A ller 1988, 1994; Fukukawa 2001; Meil e et al. 2001). Feeding and respiratory currents generated by burrowing in fauna exchanges burrow water with overlying10.17 water. Typically, overlying water has higher concentrations of dissolved oxygen and lower concentrations of nutrients, hydrogen sulfide, CO 2 and other porewater constituents than burrow water. Sediment porewater adjacent to the bu rrows diffusively exchanges with burrow waters (Aller 1988; Kristensen 1988), and t hus ventilation of burrows results in increased porewater oxygen concentrations, and decrea sed concentrations of nutrien ts, hydrogen sulfide, etc. (Kristensen 1988; Aller 1994; Furukawa 2001). El evated levels of a mmonium and hydrogen sulfide have adverse effects on seagrasses as discu ssed in Chapters 3 and 6. Thus, the presence of bioirrigators has the potential to increase th e survival and growth of seagrasses by reducing

toxic concentrations of those compounds in sediment porew ater. Little research has been

conducted on this interaction be tween bioirrigators and seagra sses. Using a physiologicallybased stress-response model for Thalassia testudinum , Eldridge and Kaldy (2004) suggests that bioirrigation has a very large impact on turt legrass 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 seagra ss 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 he rbivory (Luckenbach and Orth 1999), by trapping seeds in shallow-water sedime nts that might otherwise be transported by currents to deep waters (i.e., below the species’ p hotic 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 a bove, bioturbation can oxygenate and remediate sediments that are anoxic or hypoxic due to high organic matter loading (e.g., eutrophication) and are toxic to seagrass es (van Katwijk et al. 1997; Koch and Erskine 2001), possibly resulting in geochemical condi tions tolerable to seagrass colonization, particularly if remineralization and burial ex ceed organic matter loading (Aller 1988; Fukukawa 2001). Bioturbation can also facilitate restora tion of sediment texture altered by episodic

sedimentation. Norkko et al. (2002) reported that bi oturbation facilitated the recovery of macrofaunal communities following the experiment al deposition of terrigenous clay onto tide flats, simulating massive sediment runoff associat ed with poor land management. Duarte et al. (1997) noted that sediment reworking by burrowi ng shrimp reduced the thickness of sediment layers experimentally deposited on seagrass patches; this may have benefite d seagrasses in those patches to some extent given th at several of the seagrass spec ies 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 bioturbati on can create open space within areas monopolized by dominant space competitor(s). Those areas ar e then available for colonization by weaker10.18 competitors. This is an example of the intermed iate disturbance hypothesis, which predicts that in communities structured by competition, bi odiversity is greatest under conditions of intermediate disturbance, and reduced under conditions of low di sturbance (where the dominant competitor(s) dominate all resources) or high di sturbance (where only the most disturbancetolerant species persist) (Sousa 1984; Dial and Roughgarden 1998). As discussed in section 10.2.1, Duarte et al. (1997) observed that bioturbation by burro wing shrimp could be partially responsible for maintaining high seagrass diversity in Philippine seag rass meadows by creating open patches within dense Thalassia hemprichii stands that could be colonized by more opportunistic species of seagrass. A similar scen ario may maintain dive rsity in mixed-species seagrass beds of south Florida (B.J. Peterson, personal communication).

10.4 Effects of Interactions Between Bioturba tion/Bioirrigation and Water Quality on Seagrasses

Previous sections have desc ribed how bioturbation can act to limit or facilitate the growth of seagrass populations; a third s eagrass response could be “no eff ect”. (To the best of my knowledge, bioirrigation only has neutral or facil itating 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, specie s with low sediment reworking rate s, or by low population densities of large bioturbators) would be likely to have neut ral direct effects or co uld facilitate seagrass condition by stimulating nutrient cycling and or ganic matter remineralization. High intensity bioturbation (such as that cause d 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. Variat ion in seasonality of sediment reworking rates among bioturbators could have important conseq uences for seagrasses. Seagrasses whose growth rates were seasonally synchronous with th e reworking rates of bioturbators would have a better chance of growing away (vertically or hor izontally) from the disturbance. Seagrasses encountering bioturbators that had no seasonal va riation in sediment reworking rates would be less able to grow away from the disturban ce during low growth s easons (i.e., winter).

Other factors (such as water quality) that limit seagrass condition could affect the outcome of interactions between bioturbators, bi oirrigators, and seagrasses. Little is known

about the effects of interactions between biotur bation/bioirrigation and water quality on seagrass growth or survival. The issue has been addr essed in only three stud ies. Phillipart (1994) suggested that, following th e historical decline of Zostera spp. in the Wadden Sea due to disease and eutrophication, expansion of populations of bioturbating lugworms may have prevented recovery of seagrass populations. Similarly, H ughes et al. (2000) suggested that expansion of populations of the bioturbating polychaete Nereis diversicolor may have hindered natural and

intentional restoration of Zostera spp. populations in south-east England following their decline due to wasting disease and eutr ophication. In these two cases, the bioturbator-water quality interaction is temporally out of phase, with bioturbation impacts expressed after impacts of poor10.19 water quality and wasting disease (which itself may be linked to poor water quality, Hawkins et al. 1999). One could argue that it is not valid to consider th ese cases as examples of the interaction; however, the net detrimental effect to seagrass p opulations (if true) would be likely a consequence of both sources of disturbance acting sequentially.

Over a 15 year period, Harrison (1987) observe d an increase in the si ze of seagrass beds ( Zostera marina and Z. japonica .) on a British Columbia (Canada) tide flat, and a decrease in the area dominated by burrowing shrimp ( Neotrypaea californiensis ), following the construction of a breakwater. The breakwater greatly reduced currents over the tide flat, and water clarity increased substantially. Transp lant 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 bioturbati on activity was seasonal, being greatest in the summer and early fall, and least in the winter and early spring, pr obably in response to temperature (Fritz 2002). Harris on (1987) proposed that the increas e in water clarity allowed the seagrass patches to start growi ng 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 burro w through the seagrass bed. Lateral expansion of the seagrass bed stopped as shrimp bioturbatio n 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 intera ction hypothesis is that when water clarity was reduced, burrowing shrimp biot urbation limited growth of Zostera spp. patches. No experiments

were conducted to test this inte raction hypothesis (other than the previously discus sed transplant experiments), and it is possible that the seagra ss bed expansion occurred only because of the increase in water clarity. In any case, the hypot hesis illustrates one way in which water quality and bioturbation may interact to limit seagrass populations.

Bioturbation and bioirrigation may help to offset adverse impact s of poor water quality (such as eutrophication and sedimentation) that c ould 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 rest oration 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 cap acity of populations of benthic suspension feeders to daily filter large pr oportions (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 Gr iffen et al. 2004). Other investigators have speculated that reduction of populat ions of suspension feeders coul d contribute to an increase in turbidity and, consequently, re duce populations of subm erged aqua tic vegetation (SAV). Newell (1988) and Ulanowicz and Tuttle (1992) sugge sted that over harvesting of oyster ( Crassostrea10.20 virginica ) populations may have contribu ted 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 cause d in part by changes in water clarity due to filter feeding by increasing a nd then declining populations of the invasive Asiatic bivalve, Corbicula fluminea . Only one study has explicitly examined whether benthic suspension feeders could remove sufficient particulat e material to increase the light field enough to affect seagrass growth and survival. Newell and Koch (2004) de veloped a model to test whether the filtering capacity of oysters ( C. virginica ) and hard clams ( Mercenaria mercenaria ) at various population

densities could affect Ruppia maritima growth in Chesapeake Bay. They determined that filtering by high densities of oysters, as occurred historically, could have reduced turbidity sufficiently to enhance seagrass growth. However, hard clams had much lower filtration rates than oysters, and in their model, high population densities of the clams were unable to reduce turbidity enough to benefit R. maritima

.

Clearly not all benthic suspension feeders have the capability to substantially reduce turbidity, and the best example of an organism that does have this ability (i.e., oysters) is not a bioturbator/bioirrigator . Newell and Koch (2004) conclude d that the potential for benthic suspension feeders to reduce turbidity and enha nce seagrass growth is limited to those invertebrates that 1) are able to sustain high filtration rates un der 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 bioi rrigating infauna that does have the potential to significantly reduce turbid ity are suspension-feed ing upogebiid burrowing shrimp. Populations of Upogebia pusilla (Dworschak 1981) and U. pugettensis (DeWitt et al. 2004; Griffen et al. 2004) are estimated to da ily 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), th en populations of Upogebid burrowing shrimp may be able to reduce turbid ity 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 am ong bioturbating and bioi rrigating crustaceans (amphipods, thalassinids) and bivalves. Furthe rmore, particles entrained in water passing through U. pugettensis burrows (e.g., during bioirrigati on) can become trapped on the burrow walls and thus removed from the water column (G riffen et al. 2004). Those authors estimate that ~40% of particle removal by the suspension- feeding and bioirrigating burrowing shrimp, U. pugettensis , is caused by burrow-wall entrapment. While this process likely occurs for many other bioirrigating infauna, it has not been studie d for other species. But, to the extent that U. pugettensis populations have the potential to signif icantly reduce turbidity, then burrow-wall

entrapm ent of particles is an important m echanisms by which bioi rrigating infauna (or Upogebiids at least) can reduce turbidity sufficiently to benefit seagrasses.10.21 10.5 Implications for Seagrass Protection and Restoration

Narrative or numerical criteria for protecti on 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 aff ect seagrasses. As discussed in Section 10.4, bioturbators can potentially increase the advers e effects of nutrient enrichment by decreasing water column light levels (i.e., by increasing turbid ity as a result of sediment resuspension) or by acting as an independent stressor on nutrient-stressed plants. Howe ver, these interactions have been specifically examined in only a very few studies and consequently are poorly understood. An initial step to account for adverse interacti ons between water quality and bioturbators would be development of guidelines to identify the magn itude of risk of seagrass loss caused by the presence of bioturbators. Those guidelines c ould 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 th e population density of bioturbators as lower, equal, or higher than the densit ies reported to damage seagrass es. Short et al. (2002) recommend using similar information as part of their Tran splant Suitability Index to rank sites for their potential for successful seagrass ( Z. marina ) restoration. Determination of an appropriate safety margin to account for bioturbator-s tressor interactions will require additional information on how those interactions magnify adverse effects of the stressor alone.

To complicate matters, bioirrigators and biot urbators can reduce some adverse effects of poor water quality, particularly those caused by nu trient enrichment (i.e ., reducing turbidity or phytoplankton concentration; oxygenating sediments; flushing ammonia, H 2 S, or other toxins from sediments; see Sections 10.3 and 10.4). Thus , the presence of bioirriga tors and bioturbators in or near seagrass beds may help explain s eemingly 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 th e 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, manipulativ e experiments have demonstrated that other

bioturbators taxa can kill tran splanted 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 rest oration 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 ex cluding burrowing shrimp or polychaetes because these organisms are relatively small bodied and li ve underground. Pesticide application to kill polychaetes or burrowing shrimp can enhance su rvival of seagrasses (de Deckere et al. 2001;10.22 Dumbauld and Wylie-Echeverria 2003), although us e of broad-spectrum pesticides to control indigenous infauna is controve rsial (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 matte r concentrations by virtue of their oxygenating sediments and increasing the efflux of toxic hydrogen sulfide or ammonia from sediment porewater, and thereby increase th e probability of seagrass growth or survival. Also, suspension-feeding, bioirrigating benthic fa una 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 coul d enhance restoration of seagrass beds. While their example was for epibenthic oysters, the pr inciple holds for bioirr igating or bioturbating infauna such as Upogebiid burrowing shrimp or ot her 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 limiti ng factors or facilitators of seagrass population growth. The three critical knowledge gaps c oncerning the negative and positive effects of bioturbation and bioirrigation on seagrass populations are: 1) effects of bioirriga tion on seagrasses, particularly with respect to oxygenating root zone sediments, 2) interactions between bioturbation/bioirrigation and abiotic limiting factor s (especially anthropogenic stressors), and 3) scaling the effects of bioturbator/bi oirrigator populations to seagrasses,.

Recent models of seagrass-sediment interact ions (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 pote ntial to increase the growth and survival of seagrasses by enhancing the oxidation of thos e toxic metabolites. While these geochemical processes have been empirically demonstrated for unvegetated sediments, their impacts on seagrasses have not been investigated e xperimentally 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 a nd poor water quality might have

an increased adverse effect on seagrass populations, compared to th e effect of either stressor alone. Other research suggests th at the populations of bioirrigator s or bioturbators might reduce

one or more adverse effects of nutrient enrichme nt to seagrass populations (see Section 10.3). Further investigation of these interactions, particularly for be nthic invertebrate bioturbators and bioirrigators that are common in habitats that suitable for seagrasses. The potential for bioturbation or bioirrigation to reduce adverse e ffects of certain pollutants may be valuable for protection or restorati on of seagrass populations.10.23 Little is known about scali ng the effects of bioturbator or bioirrigator populations to seagrasses. While several investigators have discussed the potential for dense populations of specific bioturbators or bioirrigat ors to affect seagrasses, the de nsity-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 th ose 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 larg er area than that which the animals occupy. Related to this is the question of the geographic ubiquity of bioturbator-seag rass and bioirrigatorseagrass interactions. Whereas study sites inve stigating 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 bi oturbator or bioirrig ator populations scale relative to seagrass populations will be critical for es timating the risks or benefits that seagrass populations obtain from the presence of those animals.10.24 10.7 Literature Cited

Allanson, B. R., D. Skinner, and J. Imberger. 1992. Flow in prawn burro ws. Estuarine, Coastal

and Shelf Science 35:253-266. Aller, R. C. 1988. Benthic fauna and biogeochemical processes in marine sediments: the role of burrow structures. pp. 301-338. In: T. H. Blackburn and J. Sorensen (eds.), Nitrogen Cycling in Coastal Environments, John Wiley and Sons. Aller, R. C. 1994. Bioturbation and remineralizat ion of sedimentary orga nic matter: effects of redox oscillation. Chemical Geology 114:331-345. Angel, A., G. M. Branch, R. M. Wanless, and T. Siebert. 2006. Causes of rarity and range restriction of an endangered endemic limpet, Siphonaria compressa . Journal of Experimental Marine Bi ology and Ecology 330:245-260. Bachelet, G., X. De Monta udouin, I. Auby, and P. Labourg. 2000. Seasonal changes in macrophyte and macrozoobenthos assemblages in three coastal lagoons under varying degrees of eutrophication. ICES Jour nal of Marine Science 57:1495-1506. Backman, T. W. H. 1984. Phenotypic expressions of Zostera marina L. ecotypes in Puget Sound, Washington. Ph.D. dissertation, University of Washington, Seattle, Washington. pp. 178. Beal, B. F. 1994. Effects of seagrass ( Zostera marina ) cover on growth and survival of softshell clams ( Mya arenaria ). Journal of Shellfish Research 13:311-312, 1994. Berkenbusch, K. and A. A. Rowden. 2003. Ecosystem engineering – moving away from ‘just-so’ stories. New Zealand Journal of Ecology 27:67-73. Berkenbusch, K. and A. A. Rowden. 2007. An examin ation of the spatial a nd temporal generality of the influence of ecosystem engineers on the composition of asso ciated assemblages. Aquatic Ecology 41:129-147. Berkenbusch, K., A. A. Rowden, and T.E. Myer s. 2007. Interactions between seagrasses and burrowing ghost shrimps and their influen ce on infaunal assemblages. Journal of Experimental Marine Biol ogy and Ecology 341: 70-84. Bigley, R. E. 1981. The population biol ogy of two intertid al seagrasses, Zostera japonica and Ruppia maritime at Roberts Bank, British Columbia. M.S. Thesis. University of British Columbia, Vancouver, BC, Canada. pp. 205. Bowden, D. A., A. A. Rowden, and M. J. Attrill. 2001. Effect of patch size and in-pat ch location on the infaunal macroinvertebrate assemblages of Zostera marina seagrass beds. Journal of Experimental Marine Bi ology and Ecology 259:133-154. Brenchley, G. A. 1982. Mechanisms of spatia l competition in marine soft-bottom communities. Journal of Experimental Mari ne Biology and Ecology 60:17-33. Chesher, R. H. 1969. Contri butions to the biology of Meoma ventricosus (Echinoidea: Sptangoida). Bulletin of Marine Science 19:72-110. Dame, R. F. and T. C. Prins. 1998. Bivalve ca rrying capacity in coasta l ecosystems. Aquatic Ecology 31:409–421. Davis, R. C. and F. T. Shor t. 1997. Restoring eelgrass, Zostera marina L., habitat using a new transplanting techniqu e: The horizontal rhizome method. Aquatic Botany 59:1-15. Davis, R. C., F. T. Short, and D. M. Burdick. 1998. Quantifying the effects of green crab damage to eelgrass transplants. Restoration Ecology 6:297-302.10.25 de Deckere, E. M. G. T., T. J. Tolhurst, and J. F. C. de Brouwer. 2001. Destabilization of cohesive sediments by infauna. Estuarine, Coastal and Shelf Science 53:665-669. DeWitt, T. H., A. F. D’Andrea, C. A. Brown, B. D. Griffen, and P. M. Eldridge. 2004. Impact of burrowing shrimp populations on nitrogen cyc ling and water quality in western North American temperate estuaries. pp. 107-118. In: A. Tamaki (ed.), Proceedings of the Symposium on Ecology of Large Bioturbators in Tidal Flat s and Shallow Sublittoral

Sediments – from Individual Behavior to Thei r Role as Ecosystem Engineers. University of Nagasaki, Japan. Dial, R. and J. Roughgarden. 1998. Theory of mari ne communities: The intermediate disturbance hypothesis. Ecology 79:1412-1424. Domning, D.P. 2001. Sirenians, seagrasses, a nd Cenozoic ecological cha nge in the Caribbean. Palaeogeography, Palaeoclimat ology, Palaeoecology 166:27-50. Duarte, C. M., J. Terrados, N. S. R. Agawin, M. D. Fortes, S. Bach, and W. J. Kenworthy. 1997. Response of a mixed Philippine seagrass meadow to experimental burial. Marine Ecology Progress Series 147:285-294. Dumbauld, B. R. and S. Wyllie-Echeverria. 2003. The influence of burrowing thalassinid shrimps on the distribution of intertidal seagrasses in Willapa Bay, Washington, USA. Aquatic Botany 77:27-42. Dworschak, P. C. 1981. The pumping rates of the burrowing shrimp Upogebia pusilla (Petagna) (Decapoda: Thalassinidea). Journal of Expe rimental Marine Biology and Ecology 52:25–

35. Dworschak, P. C. 2000. Global diversity in the Th alassinidea (Decapoda). J ournal of Crustacean Biology 20:238-245. Eldridge, P. M. and J. E. Kaldy. 2004. Stre ss response model for th e tropical seagrass Thalassia

testudinum : the interactions of light, temperatur e, sedimentation and geochemistry.

Estuaries 27:923-937. Eldridge, P. M. and J. W. Morse. 2000. A diag enetic model for sediment-seagrass interactions, Marine Chemistry 70:89-103. Eyre, B. D. and A. J. P. Ferguson. 2002. Co mparison of carbon production and decomposition,

benthic nutrient fluxes and de nitrification in seagrass, phyt oplankton, benthic microalgae and macroalgae-dominated warm-tempera te Australian lagoons. Marine Ecology Progress Series 229:43-59. Felder, D. L. 2001. Diversity and ecological sign ificance of deep-burrowi ng macrocrustaceans in coastal tropical waters of the Americas (Deca poda: Thalassinidea). Interciencia 26:440-

449. Feldman, K. L., B. R. Dumbauld, T. H. DeWitt, and D. C. Doty. 2000. Oysters, crabs, and burrowing shrimp: Review of an environm ental conflict over a quatic resources and pesticides in Washington state’s (USA) coastal estuaries. Estuaries 23:141-176. Fonseca, M. S. and S. S. Bell. 1998. The influe nce of physical setting on seagrass landscapes near Beaufort, North Carolina, USA. Ma rine Ecology Progress Series 171:109-121. Fonseca, M. S., W. J. Kenworthy, F. X Courtney, and M.O. Hall. 1994. Seagrass planting in the

southeastern United States: Methods for acceler ating habitat development. Restoration Ecology 2:198-212.10.26 Fonseca, M. S., G. W. Thayer, and W. J. Kenw orthy. 1998. Guidelines for the conservation and restoration of seagrasses in the United St ates and adjacent wate rs. NOAA Coastal Ocean Program Decision Analysis Series No. 12. NOAA Coastal Ocean Office, Silver Spring, MD. 222 pp. Fritz, C. E. 2002. A seasonal st udy of sediment reworking by Neotrypaea californiensi s in Yaquina Bay, Oregon. Masters Thesis, Or egon State University, Corvallis, Oregon. Fukukawa, Y. 2001. Biogeochemical consequen ces of macrofauna burrow ventilation.

Geochemical Transactions. 2:83-91. Garbary, D.J., and A.G. Miller. 2006. Green crabs ( Carcinus maenas ) as the grim reaper: destruction of eelgrass beds in Nova Sco tia. Journal of Shellfish Research 25:728. Gerritsen J., A. F. Holland, and D. Irvine. 1994. Suspension feeding bivalves and the fate of primary production: an estuarine model applie d to Chesapeake Bay. Estuaries 17:403-

416. Griffen, B. D., T. H. DeWitt, and C. Langdon. 2004. Particle removal rates by the mud shrimp Upogebia pugettensis , its burrow, and a commensal clam: effects on estuarine phytoplankton abundance. Marine Ecol ogy Progress Series 269:223-236. Harrison, P. G. 1987. Natural expansion and experimental manipulation of seagrass ( Zostera spp.) abundance and the response of infaunal inve rtebrates. Estuarin e, Coastal and Shelf

Science 24:799-812. Hawkins, S. J., J. R. Allen, and S. Bray. 1999. Restoration of temper ate marine and coastal ecosystems: nudging nature. Aquatic Conservati on: Marine and Freshwater Ecosystems

9:23-46. Hopkinson, Jr., C. S., A. E. Giblin, J. Tucker, and R. H. Garritt. 1999. Benthic metabolism and nutrient cycling along an estuarine salin ity gradient. Estuaries 22:863-881. Huettel, M. 1990. Influence of the lugworm Arenicola marina on porewater nutrient profiles in

sand flat sediments. Marine Ec ology Progress Series 62:241-248. Hughes, R. G., D. Lloyd, L. Ball, and D. Emson. 2000. The effects of the polychaete Nereis

diversicolor on the distribution and transplanting success of Zostera noltii . Helgoland

Marine Research 54:129-136. Jumars, P. A. and R. A. Wheatcroft. 1988. Re sponses of benthos to changing food quality and quantity, with a focus on deposit feeding a nd bioturbation. pp. 235-253. In: W. H. Berger, V. S. Smetacek, and G. Wefer (eds.) . Productivity of the Ocean: Present and Past. Wiley-Interscience, New York. pp. 470. Koch, M. S. and J.M. Erskine. 2001. Sulf ide as a phytotoxin to the tropical seagrass Thalassia

testudinum : interactions with light, salinity and te mperature. Journal of Experimental Marine Biology and Ecology 266:81-95. Kristensen, E. 1988. Benthic fauna and biogeochemical processes in marine sediments: microbial activities and fluxes. pp. 275-299. In: T. H. Blackburn and J. Sorensen (eds.), Nitrogen Cycling in Coastal Environments. John Wiley and Sons, New York. Luckenbach, M. W. and R. J. Orth. 1999. E ffects of a deposit-feedi ng invertebrate on the entrapment of Zostera marina L. seeds. Aquatic Botany 62:235-247. Martin, R. A. 2003. The Amazing Sensory Talent s of Batoids. World Wide Web Publication. www.elasmo-research.org/educatio n/topics/s_batoid_senses.htm10.27 Meile, C., C. M. Koretsky, and P. Van Cappe llen. 2001. Quantifying bi oirrigation in aquatic sediments: An inverse modeling appro ach. Limnology and Oceanography 46:164-177. Merkel, K. W. 1990. Eelgrass transplanting in so uth San Diego Bay, California. pp. 28-42 In: S. Wyllie-Echerverria, R. S. Hoffman and K. W. Merkel (eds.), Proceedings of the California Eelgrass Symposium, Chula Vista, California, May 27 and 28, 1988. Sweetwater River Press, National City, California. Molenaar, H. and A. Meinesz.. 1995. Vegetative reproduction in Posidonia oceanica : Survival and development of transplanted cuttings acco rding to different spacings, arrangements

and substrates. Botanica Marina 38:313-322. Moore, K. A., R. J. Orth, and J. F. Nowak. 1993. Environmental regulatio n of seed germination

in Zostera marina L. (eelgrass) in Chesapeake Bay: Effects of light, oxygen and sediment burial. Aquatic Botany 45:79-91. Mosher, C. 1980. Distribution of Holothuria areicola Semper in the Bahamas with observations on habitat, behavior and feeding activity (E chinodermata: Holoturoidea). Bulletin of

Marine Science 30:1-12. Newell, R. I. E. 1988. Ecological Changes in Chesapeake Bay: Are they the result of overharvesting the American oyster (Crass ostrea virginica), p. 536–546. In M. Lynch (ed.), Understanding the Estuary: Advances in Chesapeake Bay Research. Chesapeake Research Consortium Publication 129, Gloucester Point, Virginia. Download from www.vims.edu/GreyLit/crc129.pdf. Newell, R. I. E. and E. W. Koch. 2004. Modeling seagrass density and distribution in response to changes in turbidity stemming from bivalve fi ltration and seagrass sediment stabilization.

Estuaries 27:793–806. Norkko, A., S. F. Thrush, J. E. Hewitt, V. J. Cummings, J. Norkko, J. I. Ellis, G. A. Funnell, D. Schultz, and I. MacDonald. 2002. Smothering of estuarine sandflats by terrigenous clay:

the role of wind-wave disturbance and bi oturbation in site-dependent macrofaunal recovery. Marine Ecology Progress Series 234:23-41. Ogden, J. C. 1980. Faunal relationships in Cari bbean seagrass beds. , pp. 171-198. In: R. C. Phillips and C. P. McCoy (eds.), Handbook of Seagrass Biology: An Ecosystem Perspective, Garland STPM Press, New York. Orth, R. J. 1975. Destruction of eelgrass, Zostera marina by the Cownose Ray, Rhinoptera

bonasus , in the Chesapeake Bay. Chesapeake Science. 16:205-208. Orth, R. J., K. L. Heck Jr., and J. van Montfrans. 1984. Faunal communities in seagrass beds: A

review of the influence of plant structure and prey char acteristics on predator-prey relationship. Estuaries 7:339-350. Orth, R .J., M. W. Luckenbach, and K. A. Moor e. 1994. Seed dispersal in a marine Macrophyte: Implications for colonization a nd restoration. Ecology 75:1927-1939. Ott, J. A., B. Fuchs, R. Fuchs, and A. Malasek. 1976. Observations on the biology of Callianassa stebbingi Borrodaille and Upogebia litoralis Risso. Senckenbergiana Maritime 8:61-79. Packard, J.M. 1984. Impact of manatees Trichechus manatus on seagrass communities in eastern Florida. Acta Zoologica Fennica 172:21-22. Pearson, T. H. 2001. Functional group ecology in soft-sediment marine benthos: the role of bioturbation. pp. 233-267. In: R. B. Gibs on, M. Barnes and R. J. A. Atkinson. Oceanography and Marine Biology Annual Review s. Vol. 39. Taylor & Francis Ltd..10.28 Peterson B. J. and K. L. Heck, Jr. 1999. The pot ential for suspension feeding bivalves to increase seagrass productivity. Journa l of Experimental Marine Biology and Ecology 240:37–52. Peterson B. J. and K. L. Heck, Jr. 2001a. Posi tive interactions between suspension feeding bivalves and seagrass—a facultative mu tualism. Marine Ecology Progress Series

213:143–155. Peterson B. J. and K. L. Heck, Jr. 2001b. An experimental test of the mechanism by which suspension feeding bivalves elevate seag rass productivity. Marine Ecology Progress

Series 218:115-125. Peterson, B. J., K. L. Heck, Jr., and J. W. F ourqurean. 2003. Organisms as resource providers: a tale of two suspension feeders in seagrass co mmunities. Gulf of Mexico Science 21:120. Phelps, H. L. 1994. The asiatic clam ( Corbicula fluminea ) invasion and system-level ecological change in the Potomac River estuary near Washington, D.C. Estuaries 17:614–621. Phillipart, C. J. M. 1994. Interactions between Arenicola marina and Zostera noltii on a tidal flat

in the Wadden Sea. Marine Ec ology Progress Series 111:251-257. Posey, M. H. 1986. Changes in a benthic community associated with dense beds of a burrowing deposit feeder, Callianassa californiensis . Marine Ecology Progress Series 31:15-22. Pranovi, F., D. Curiel, A. Rismondo, and M. Marz occhi. 1996. Seagrass m eadow distribution in relation to the presence of Upogebia pusilla (Crustacea, Decapoda ) in the Lagoon of Venice. Biologia Marina Mediterranea. 4: 169-171. (Italian with English abstract) Preen, A. 1995. Impacts of dugong foraging on seagra ss habitats: observational and experimental evidence for cultivation grazing. Marine Ecology Progress Series 124:201–213. Prins, T. C., A. C. Smaal, and R. F. Dame. 1997. A review of the feedbacks between bivalve grazing and ecosystem processes. Aquatic Ecology 31:349-359. Reusch, T. B. H., A. R. O. Chapman, and J. P. Gröger. 1994. Blue mussels Mytilus edulis do not interfere with eelgrass Zostera marina but fertilize shoot grow th through biodeposition. Marine Ecology Progress Series 108:265–282. Reusch, T. B. H. and S. L. Williams. 1998. Variable responses of native eelgrass Zostera marina to non-indigenous bivalve Musculista senhousia . Oecologia 113:428–442. Rowden, A. A. and M. B. Jones. 1993. Critical evaluation of sediment turnover estimates for Callianassidae (Decapoda: Thalassinidea). J ournal of Experimental Marine Biology and Ecology 173:265-272. Schaffner, L. C., P. Jonsson, R. J. Diaz, R. Rosenberg, and P. Gapcynski. 1992. Benthic communities and bioturbation history of estuarine and coastal systems: Effects of hypoxia and anoxia. pp. 1001-1016. In: R. A. Vollenweider, R. Marchetti, and R. Viviani (eds), Marine Coastal Eu trophication, Elsevier, London. Short, F. T., R. C. Davis, B. S. Kopp, C. A. Short, and D. M. Burdick. 2002. Site-selection model for optimal transplantation of eelgrass Zostera marina in the northeastern US. Marine Ecology Progress Series 227:253–267. Short, F. T., K. Matso, H. M. Hoven, J. Whitten, D. M. Burdick, and C. A. Short. 2001. Lobster use of eelgrass habitat in the Piscataqua River on the New Hampshire/Maine border,

USA. Estuaries 24:277-284. Short, F. T. and S. Wyllie-Echeverria. 1996. Natural and human induced disturbances of seagrasses. Environmenta l Conservation 23:17-27.10.29 Siebert, T., and G. M. Bran ch. 2005. Interactions between Zostera capensis

, Callianassa kraussi

and Upogebia africana : deductions from field surveys in Langebaan Lagoon, South

Africa. African Journal of Marine Science 27: 345-356. Siebert, T., and G. M. Branch. 2007. Ecosyste m engineers: interactions between eelgrass Zostera capensis and the sandprawn Callianassa kruassi and their indirect effects on the mudprawn Upogebia africana . Journal of Experimental Marine Biology and Ecology

338: 253-270. Sousa, W. P. 1984. The role of disturbance in natural communities. Annual Review of Ecology and Systematics 15:353-391. Suchanek, T. H. 1983. Control of seagra ss communities and sediment distribution by

Callianassa (Crustacea, Thalassinide a) bioturbation. Journal of Marine Research

41:281-29. Tamaki, A. 2004. Ghost shrimps, snails, and clam s on intertidal sandflats in the Ariake Sound estuarine system: seeking for unified perspe ctives on their population explosions and declines. pp. 107-118. In: A. Tamaki (ed. ), Proceedings of the Symposium on Ecology of Large Bioturbators in Ti dal Flats and Shallow Sublittoral Sediments – from Individual Behavior to Their Role as Ecosystem Engi neers. University of Nagasaki, Japan. Thayer, G. W. 1983. Sediment-mediated biologi cal disturbance and the evolution of marine benthos. pp. 479-625. In: M. J. S. Tevesz and P. L. McCall (eds.). Bi otic Interactions in Recent and Fossil Benthic Communities. Plenum, New York. Thayer, G. W. and M. S. Fonseca. 1984. The ecology of eelgrass meadows of the Atlantic coast: a community profile. US Fish & Wildlife Service, FWS/OBS-84/02. Thom, R. M., A. B. Borde, S. Rumrill, D. L. Woodr uff, G. D. Williams, J. A. Southard, and S. L. Sargeant. 2003. Factors influencing spa tial and annual variability in eelgrass ( Zostera

marina L.) meadows in Willapa Bay, Washington, and Coos Bay, Oregon, estuaries.

Estuaries 26:1117-1129. Thompson, L. D. and A. W. Pritcha rd. 1969. Osmoregulatory capacities of Callianassa and Upogebia (Crustacea: Thalassinidae). Biological Bulletin 136:114-129. Townsend E. C. and M. S. Fonseca. 1998. Bioturbation as a potential mechanism influencing spatial heterogeneity of North Carolina seagrass beds. Marine Ecology Progress Series

169:123-132. Ulanowicz, R. and J. H. Tuttle. 1992. The trophic consequences of oyster stock rehabilitation in Chesapeake Bay. Estuaries 15:298–306. Valentine, J. F., K. L. Heck, Jr., P. Harper, and M. Beck. 1994. Effects of bioturbation in controlling turtlegrass ( Thalassia testudinum Banks ex Koenig) abundance: Evidence

from field enclosures and observations in th e northern Gulf of Mexico. Journal of Experimental Marine Biol ogy and Ecology 178:181-192. van Katwijk, M. M. and D. C. R. Hermus. 2000. Effects of water dynamics on Zostera marina: transplantation experiments in the intertidal Dutch Wadden Sea. Marine Ecology Progress Series 208:107–118. van Katwijk, M. M., L. H. T. Vergeer, G. H. W. Schmitz and J. G. M. Roelofs. 1997. Ammonium toxicity in eelgrass Zostera marina . Marine Ecology Progress Series

157:159-173.10.30 Wanless, H. R., L. P. Tedesco, and K. M. Ty rrell. 1988. Production of subtidal tubular and surficial tempestites by Hurricane Kate, Caicos Platform, British West Indies. Journal of Sedimentary Petrology 58:739-750. Webb, A. P. and B. D. Eyre. 2004. Effect of natural populations of burrowing thalassinidean shrimp on sediment irrigation, benthic metabolism, nutrient fluxes and denitrification. Marine Ecology Progress Series 268:205-220. Wheatcroft, R. A. and W. R. Martin. 1996. Spatial variation in short-term ( 234 Th) sediment bioturbation intensity along an organic-carbon gradient. Journal of Marine Research 54:

763-792. Williams, S. L. 1990. Experimental studies of Caribbean seagrass bed development. Ecological Monographs 60:449-469. Woods, C. M. C. and D. R. Schiel. 1997. Use of seagrass Zostera novazelandica (Setchell, 1933) as habitat and food by the crab Macrophthalmus hirtipes (Heller 1862) (Brachyura: Ocypodidae) on rocky intertidal platforms in southern New Zealand. Journal of Experimental Marine Bi ology and Ecology 214:49-65. Zieman, J. C. 1982. The ecology of seagrasses of south Florida: A community perspective. U.S. Fish and Wildlife Service, O ffice of Biological Services. Washington, D.C. FWS OBS-82 25. 158 pp.11.1 

## 11.0 Research Gaps in Relation to Setting Protective Criteria

Walter G. Nelson

11.1 Overview

The U.S. EPA sponsored a workshop in Balt imore, Maryland in 2003 entitled: Towards a National Basis for Protective Criteria for Submerge d Aquatic Vegetation, in conjunction with the

National Conference on Coastal and Estuarine Hab itat Restoration. A group of more than 60 scientists and resource managers from 21 coasta l states discussed the scientific knowledge base and the challenges in developing a national fr amework for protection of submerged aquatic vegetation. Through a series of regional presen tations on the status of SAV research, followed by break-out discussion groups, seve ral key questions were addressed. The principal questions

were: Does our current state of scientific knowledge allow draf ting 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 fo r the Development of Regional Nutrient Criteria (US EPA, 1998), which proposed an ope rational 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 Ecor egional classification system as the regional framework for development of nutrient crite ria. A variety of guidance documents recommending ambient water quality criteria fo r different water body types (e.g. wetlands, US EPA 2000; lakes and reservoirs, US EPA, 2001) ha ve been developed for various Level III or sublevel III ecoregions. The tradeoffs of devel oping protective criteria at larger, aggregated ecological scales versus individual water bodies is clearly one of accuracy and relevance of a

criterion versus the data collec tion 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 deve lop criteria at the system or sub-system scale. On the other hand, if a criter ion developed at regional scale is not applicable or protective of individual systems or sub-syst ems, 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 mana gement 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 bette r with the public. Establishment of effective criteria at a spatial scale whic h avoids a “Balkanization” of standards may be possible. The guidance document for Chesapeake Bay water qu ality criteria (US EPA 2003) illustrates one possible approach to establishing criteria for seagrasses, where cr iteria within this large and11.2 complex system are determined based on four se parate 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-sy stem), the workshop participants identified a wide range of issu es relating to the knowledge base needed to establish protective criteria. There were questions raised concerni ng the utility of using classical water quality

criteria based on ambient concentr ations 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 principl e 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 cu rrent approaches to establishing light requirements fo r seagrass populations is the abil ity 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 repr esenting the ability of the plant to maintain a positive carbon balance. Additiona l studies are needed on the rate s of translocation of sucrose

from rhizomes to shoots in Zostera marina during times of negative car bon balance to assess the importance of this mechanism of mitigating low li ght 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 am monium is directly toxic to seagrasses needs resolution. Apparently contradict ory 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 populat ions in high salinity areas, (van Katwijk et al. 1999), th e interactions of nut rient effects with sa linity have not been widely studied. Additional information on the effect of increasing the variance in salinity on Z.

marina would be helpful, especially in term s of variation of population response.

11.2.4 Hydrodynamic Factors Hydrodynamic factors clearly influence the su rvival and distributi on of seagrasses, and more data may be needed to define current velocity and wave tole rance ranges for various species and systems. Relative differences in sens itivity of various life stages (e.g. seedlings) to hydrodynamic factors may be needed. Because many aspects of hydrodynamic influences on11.3 seagrasses are determined by natural factors (wind speed, depth), it is most likely that hydrodynamic influences would be c onsidered in terms of whether sufficient safety margins are present in setting water quality criteria. Hydrodynamics consider ations are clearly critical in planning seagrass restor ation projects.

11.2.5 Sediment Characteristics

Grain size of the substratum influences the distribution and health of Zostera marina

, although there remain uncertainties concerning the mechanisms. In general, we need a better understanding of how sediment pore water concentra tion of potentially toxic substances, such as ammonium and dissolved sulfides, interact with th e 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 woul d be adequately protective.

11.2.6 Epiphytes Epiphytes influence availabl e light quality and quantity and may partly determine the ability of seagrasses to grow and survive. Th e role of epiphytic c over in affecting light availability may be an essential element to incl ude in development of management criteria for protection of coastal seagrass beds. The techni cal guidance for ambient wa ter quality criteria for Chesapeake Bay (U.S. EPA 2003) provides an exp licit formulation for in cluding epiphyte loads in estimating light available at the seagrass l eaf surface. Model formulations need to be validated for other systems. Mesocosm research has shown that differe nt algal components may dominate in the biomass response to nutrient enrichment in spite of similar initial conditions, and that temperature, nutrient e xposure regime and other factors such as grazing levels may all

influence the outcome of nutrient enrichment. T hus, models used for development of seagrass criteria must be able to account for impacts fr om multiple pathways, and must account for effects of trophic cascades.

11.2.7 Macroalgal interactions Quantitative relationships between macroalg al canopy height (or corresponding measures of abundance) and specific impacts on eelgrass pl ants (such as those provided by Hauxwell et al. 2001), for different water body characteristics (tem perature, current velocity, turbidity, grazing pressure, etc.) are needed. Similarly, rela tionships between macroalgal abundance and the causative anthropogenic activity (e.g., normalized nitrogen load rate) ar e needed to recommend corrective actions. It seems likely that such re lationships between nutrient loading, macroalgal abundance, and seagrass distributi on are site-specific, but this s upposition needs to be confirmed.

11.2.8 Desiccation and Temperature Impacts For a region such as the west coast, where in tertidal seagrass can be a large percentage of the total population, better knowle dge of desiccation effects on i ndividual shoots and populations is needed to support multi-stressor, seagrass stre ss-response models. Better definition of the interactive effects of temperatur e and nutrient effects is a critical need, both for site specific permitting of heated discharge waters, and to be ab le to set protective criteria under conditions of global warming.11.4 11.2.9 Bioturbation The effects of bioturbation and bioirriga tion 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 th e sediment, e.g. alteration of sediment oxygen levels and sediment resuspension rates, could re sult in criteria either overly conservative or insufficiently protective of seagrasses.

11.3 Conclusions

The current review confirms that there is a great deal of scientific information currently available concerning the responses of Zostera marina and Thalassia testudinum to a wide range

of environmental factors. A main theme that has emerged from the review is that the interactive effects among factors influencing seagrass survival remain relati vely poorly known, especially across broader regional scales. This appears true even for su ch fundamental environmental characteristics as salinity and temperature and thei r interactions in the expression of nutrient or sediment impacts on SAV, although resear ch is beginning to fill this gap.

Clearly of concern is whether current m odeling 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 im pacts 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 influen ces (Chapter 6) suggests that th ere are important influences on plant survival through sediment as sociated mechanisms that may not be adequately captured by water quality criteria alone. The review of eff ects of light limitation (Cha pter 2) suggested there may be advantages to looking for integrative, pl ant 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 possib ility of using such integrative measures, specifically the Nutrient Polluti on 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 fundame ntal research needs which would greatly improve the ability to set protective criteria for se agrasses. However, there is also a considerable scientific knowledge base that can be used in establishing criteria for protection of seagrasses. The concept of protective crit eria is actually one that em bodies principles of adaptive management. Initial criteria may be based on inco mplete data or modeling approaches, but it is11.5 expected that as better data and approaches become available, they will be applied to improve the protective criteria. Various lo calized efforts have made signi ficant progress in establishing seagrass protective criteria, e.g. Tampa Bay, Ch esapeake 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 knowledg e currently available.

11. 4 Literature Cited

Hauxwell, J., J. Cebrián, C. Furlong, and I. Va liela. 2001. Macroalgal canopies contribute to eelgrass ( Zostera marina ) decline in temperate estuar ine ecosystems. Ecology 82:1007-

1022. Lee, K. S., F. T. Short, and D. M. Burdick. 2004. Development of a nut rient 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. Bint z. 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 Deve lopment of Regional Nutr ient Criteria. U.S. EPA, Office of Water. EPA 822-R-98-002. U.S. EPA. 2000. Ambient Water Quality Crit eria Recommendations. Information Supporting the Development of State and Tribal Nutrient Criteria. Wetlands in Nutrient Ecoregion XIII. U.S. EPA, Office of Water, Offi ce of Science and Technology. EPA 822-B-00-

023. U.S. EPA. 2001. Ambient Water Quality Crit eria Recommendations. Information Supporting the Development of State and Tribal Nutrient Criteria. Lakes and Reservoirs in Nutrient Region III. U.S. EPA, Office of Water, O ffice of Science and Technology. EPA 822-B01-008. U.S. EPA. 2003. Ambient Water Quality Crit eria 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. Gasse ling, 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.