[From the U.S. Government Printing Office, www.gpo.gov]
Ecology And Management Of Maine's Eelgrass,
Rockweeds, And Kelps


















Gail S. Wippelhauser, Ph. D.
Maine Natural Areas Program
Department of Conservation

June 1996



















US Department of Commerce
NOAA Coastal Services Center Library
2234 South Hobson Avenue
Charleston, SC 29405-2413

To obtain a copy contact:
Information Manager
Maine Natural Areas Program
Department of Conservation
# 93 State House Station
Augusta, ME 04333
or
Maine State Planning Office
# 38 State House Station
Augusta, ME 04333
Bibliographic citation:

Wippelhauser, G. 1996. Ecology And Management Of Maine's Eelgrass, Rockweeds,
And Kelps. Maine Natural Areas Program, Department of Conservation, Augusta, ME.
73 pp.

Program support from: Maine State Planning Office, Maine Coastal Program















A report of the Maine State Planning Office with financial assistance provided by the
Coastal Zone Management Act of 1972, as amended, administered by the Office of
Ocean and Coastal Resource Management, National Oceanic and Atmospheric
Administration under Award No. NA570Z0303.
M
-11=1






























































I
Program Support from:
Maine Coastal Program
Maine State Planning Office


Ecology And Management Of Maine's Eelgrass,
Rockweeds, And Keips














Gail S. Wippelhauser, Ph. D.
Maine Natural Areas Program
Department of Conservation
June
1996

TABLE OF CONTENTS

EXECUTIVE SUMMARY.I
INTRODUCTION.2
SUMMARY OF THREATS.3
CURRENT PROTECTION.3
Federal.3
State of Maine.4
DEVELOPMENT OF STATE (DP MAINE GUIDELINES                                                                    ................................4
RECOMMENDED STANDARDS FOR EELGRASS, ROCKVVEED AND KELP
HABITAT UNDER NRPA.:5
RECOMMENDED WATER QUALITY STANDARDS
	..............................
....7
RECOMMENDED RESEARCH AND EDUCATION
	..............................
.....7
REFERENCES ..............................................................10

EELGRASS.11..........
INDENTIFICATION AND DESCRIPTION...........................................
DISTRIBUTION.1
LIFE CYCLE.1
POPULATION TRENDS.19
ECOLOGICAL REQUIREMENTS ................................................20
LihL.....................................................2.................2
Temperature...............................................................20
Saiiy...........................................................2...........0c
Water motion..............................................................21
Substrate .................................................................21
Nutrients..................................................................21
ECOLOGICAL FUNCTION .....................................................21
Effect on wvater motion and sediment dynamics ....................................21
Nutrient cycling.............................................................22
Food source...............................................................23
Grazing on live eelgrass......................................................23
Grazing on epiphytes ........................................................23
Consumption of eelgrass-derived detritus.........................................23
Habitat...................................................................24
CONCLUSION ..............................................................25
THREATS..................................................................25
REFERENCES ..............................................................28

ROCKWEEDS..........................I.....................................36
INDENTIFICATION AND DESCRIPTION ..........................................36
DISTRIBUITION..............................................................36
LIFE CYCLE ................................................................37
POPULATION TRENDS.38
ECOLOGICAL REQUIREMENTS ................................................38
LihL.....................................................3.................3
Water ...................................................................39
Temperature...............................................................40)
Saiiy...........................................................4...........0c
Water motion..............................................................41
Substrate .................................................................41
Nutrients .................................................................41
ECOLOGICAL FUNCTION .....................................................42
Nutrient cycling.............................................................42
tUptake of pollutants .........................................................42
Food source...............................................................43

Grazing on live rockwed	.....................................................43
Grazing on epi phytes	........................................................43
Consumption of eeigrass-derived detritus                                                                 .........................................43
Habitat...................................................................44
Commercial uses ...........................................................45
CONCLUSION ..............................................................45
THREATS..................................................................45
REFERENCES ..............................................................48

KELPS ....................................................................56
INDENTIFICATION AND DESCRIPTION ..........................................56
DISTRIBUTION..............................................................56
LIFE CYCLE ................................................................57
POPULATION TRENDS-......................................................59
ECOLOGICAL REQUIREMENTS ................................................60
LihL......................................................6.................0
Temperature...............................................................60
Saiiy..........................................................6............1
Water motion .............................................................61
Substrate .................................................................61
Nutrients..................................................................61
ECOLOGICAL FUNCTION .....................................................62
Effect on wvater motion .......................................................62
Nutrient cycling.............................................................62
Toxic Accumulations.........................................................62
Food source...............................................................62
Grazing on live kelp .........................................................63
Consumption of kelp-derived detritus............................................63
Habitat...................................................................64
Commercial uses ...........................................................64
CONCLUSION ..............................................................66
THRE-ATS..................................................................66
REFERENCES ..............................................................68

1
EXECUTIVE SUMMARY
The purpose of this project, initiated by the Maine Coastal Program, was to determine whether three types
of marine benthic vegetation, namely eelgrass (Zostera marina), kelps (Alaria esculenta, Agarum cribrosum,
Laminaria digitata, L. Iongicruris, L. saccharina, and Saccorhiza dermatodea), and rockweeds (Ascophyllum
nodosum, Fucus vesiculosus, F. evanescens, F. distichus, and F. spiralis), should specifically be protected
in Maine, and to develop management guidelines if protection was warranted. The Maine Natural Areas
Program was contracted by the Maine Coastal Program 1) to conduct a literature search on eelgrass,
rockweeds and kelps to assess ecological requirements, ecological functions, and threats, 2) to convene a
panel of experts to evaluate the importance of these species in Maine, and 3) to develop management
guidelines for eelgrass, rockweed and kelp habitat if appropriate.

The three types of benthic vegetation have different ecological requirements, and are found in different
habitats. Rockweeds grows primarily in the intertidal zone, and can tolerate exposure and desiccation
during low tide. Eelgrass and kelps grow primarily in the subtidal zone, and are intolerant of desiccation.
Rockweeds and kelps are found on rock, boulder, or cobble substratum, while eelgrass grows in mud, sand,
or gravel. Rockweeds and kelps are algae, lacking true roots, stems or leaves. As a result, they must
absorb nutrients from the water column. Eelgrass, a vascular plant, possesses true roots, stems and leaves.
Because it has roots, eelgrass can stabilize and bind soft substrates, and can absorb nutrients from the
substrate. Unlike the kelps, rockweeds and eelgrass have a high light requirement for growth.

These three types of benthic vegetation make similar contributions to the ecosystem, that is, they fulfill
similar ecological functions. They are primary producers, capable of synthesizing organic material using the
sun's energy. They are highly productive and contribute large amounts of organic matter to the nearshore
environment. Relatively few animals consume living eelgrass and kelps. Instead, the organic matter
produced by these three types of vegetation enters marine food webs as detritus. Rockweeds and kelps,
however, become available as detritus more quickly than eelgrass. By absorbing nutrients, including
inorganic waste products released by animals, and producing new organic matter, these three types of
vegetation contribute to nutrient cycling. All three types of benthic vegetation add structural complexity to
the habitat and increase the amount of substrate available for other species of plants and animals. As a
result, species abundance and diversity is high in habitats supporting eelgrass, rockweeds and kelps when
compared to unvegetated habitats. A number of commercially important species have been found to
associate with all three types of benthic vegetation. Finally, eelgrass and kelps have been shown to reduce
water currents by frictional forces. Although there is no evidence, rockweeds probably slows water currents
as well.
The three types of benthic vegetation are vulnerable to different threats. Eelgrass and kelps face natural
biological threats that cause enormous fluctuations in their abundance. Eelgrass is sporadically decimated
by a wasting disease for which there is no cure, and kelps are consumed by large populations of grazing sea
urchins. These natural biological threats are exacerbated by changes in the physical environment, which
are often of anthropogenic origin. Eelgrass is vulnerable to reductions of light levels, physical disturbance,
and overharvesting. Rockweeds are vulnerable to reductions of light levels, overharvesting, effluent from
pulp mills, thermal discharges, petroleum hydrocarbons, and competition from non-native species. Kelps
are vulnerable to overharvesting, effluent from pulp mills, petroleum hydrocarbons, and competition from
non-native species. Reductions of light levels in most cases result from sediment loading, nutrient loading,
and shade-producing structures.

A panel of scientists, planners, and representatives from regulatory agencies reviewed information on the
three types of marine benthic vegetation, helped identify threats to them, and made management
recommendations. Currently, eelgrass is specifically protected as a special aquatic site only at the federal
level, whereas rockweeds and kelps are not specifically protected at either the federal or state level. The
panel recommends that consideration be given to 1) inclusion of rockweeds and kelps in the Army Corps of
Engineers list of special aquatic sites, 2) development of a State of Maine policy for protection of eelgrass,
rockweed, and kelp habitat that incorporates Natural Resources Protection Act standards and water quality
standards, and 3) support of education initiatives and research regarding these species.

2
INTRODUCTION

The purpose of this project, initiated by the Maine Coastal Program, was to determine whether 12 common
species of marine benthic vegetation, namely eelgrass (Zostera marina), kelps (Alaria esculenta, Agarum
cribrosum, Laminaria digitata, L. Jongicruris, L. saccharina, and Saccorhiza dermatodea), and rockweeds
(Ascophyllum nodosum Fucus vesiculosus, F. evanescens, F. distichus, and F. spiralis) should specifically
be protected in Maine, and to develop management guidelines if protection was warranted.

The project was initiated for a number of reasons. First, benthic vegetation is generally considered to be an
important source of food, in the form of detritus, for the marine animals that live in shallow coastal waters
(Teal 1962; Mann 1973; Newell 1984; Simestad and Wissmar 1985). Second, benthic vegetation adds
structural complexity to the habitat in which it lives, making the habitat attractive to a greater variety of
marine organisms (e.g., MacDonald et al. 1984; Olney and Boehlert. 1988; Heck et al. 1989). Third,
comprehensive information about the ecological requirements and ecological functions of this group of
primary producers has never been compiled in a single document. The Gulf of Maine is an unique body of
water along the eastern coast of the United States. It has a great diversity of substratum and habitat types
(Emery et al. 1965; Brown 1993) that support a diverse group of benthic plants, and it is part of a fauna-rich
zoogeographic zone that extends northward from Cape Cod (Larsen 1979). The relative importance of
'these 12 species of primary producers and their ecological function within the Gulf of Maine ecosystem has
not been addressed previously. Finally, the distribution of eelgrass is being mapped for the Oil Spill
Response Committee, and the ecological information compiled here may be of use to the Committee.

Inclusion of A. nodosum, L. digitata, and Z. marina on the list of priority species compiled by the Gulf of
Maine Council on the Marine Environment (GOMCME) is further indication of their importance. The
GOMCME identified 161 priority species in the Gulf of Maine and ranked them by twelve criteria that
included their environmental, scientific, and commercial importance, abundance, vulnerability to population
decline, and life-history characteristics. Using these criteria, A. nodosum, Z. marina, and L digitata were
ranked 4, 12, and 45, respectively.
Eelgrass, rockweeds, and kelps are all found in nearshore areas, and are particularly susceptible to human
activities. Point and non-point sources of pollution, structures such as piers and docks, dredging and mining
projects, aquaculture activities, harvesting, and fishing practices all have the potential to impact these types
of benthic vegetation to varying degrees. Development of a management policy for these types of benthic
vegetation before more severe losses occur makes sense ecologically and economically.
The Maine Natural Areas Program was contracted by the Maine Coastal Program 1) to conduct a literature
search on eelgrass, rockweeds and kelps to assess ecological requirements, ecological functions, and
threats, 2) to convene a panel of experts to evaluate the importance of these species in Maine, and 3) to
develop management guidelines for eelgrass, rockweed and kelp habitat if they were deemed necessary.
The panel of experts that participated in discussions and reviewed this document included scientists and
representatives of planning, resource, and regulatory agencies. The panel included: Seth Barker
(Department of Marine Resources), Richard Bostwick (Department of Transportation), Dr. Susan Brawley
(University of Maine), Doug Burdick (Department of Environmental Protection), Phil Colarusso (US
Environmental Protection Agency), Pat Corr (Department of Inland Fisheries and Wildlife), Dr. Michelle
Dionne (Wells National Estuarine Research Reserve), Lee Doggett (Casco Bay Estuary Project), Jon
Kuriand (National Marine Fisheries Service), Dr. Linda Mercer (Department of Marine Resources), Josie
Quintrell (State Planning Office), Dr. Fred Short (University of New Hampshire), Laurie Steams (University
of Maine), Dr. Robert Steneck (University of Maine), Laura Taylor (State Planning Office), Dr. Robert Vadas
(University of Maine), Robert Van-Riper (Department of Transportation), and Dr. Gail Wippelhauser
(Department of Conservation).

3
SUMMARY OF THREATS

Threats to eelgrass, rockweeds, and kelps are summarized in the following table. The threats were verified
by a reviewof the literature and by consultation with experts. Other threats may exist, but have not been
documented. With one exception, these threats are the result of human activities. In most cases, a
particular activity can impact more than one type of benthic vegetation.

Threat                                                           Eelgrass       Rockweeds         Kelps
Reduction of light by nutrient loading	/              -J
Reduction of light by sediment or shading	i /
Episodic reduction of population by natural causes	V                                /
Physical disturbance	i
Local overharvesting	/                   / /
Chlorate effluent from pulp mills                        /                                          V
Thermal discharge from power plants	VJ
Reproductive interference by petroleum hydrocarbons	V /	V
Competition from non-native species	a|	V
Heavy metals	V


CURRENT PROTECTION

Federal

The United States Army Corps of Engineers (ACOE) regulates, by issuing permits, the discharge of dredged
or fill material and the placing of pilings into the navigable waters and wetlands of the United States under
Section 404 of the Clean Water Act. The ACOE also regulates temporary (seasonal) and permanent
structures or work in or affecting navigable waters under Section 10 of the Rivers and Harbors Act. Three
federal resource agencies, the Environmental Protection Agency (EPA), the National Marine Fisheries
Service (NMFS) of the National Oceanic and Atmospheric Administration (NOAA), and the U.S. Fish and
Wildlife Service (USFWS) evaluate the proposed projects and assess potential impacts to natural
resources, including habitat.
Especially valuable types of marine habitats, termed "special aquatic sites," are defined under the Clean
Water Act ï¿½404(b)(1) Guidelines (for implementing the Act) as "geographic areas, large or small, possessing
special ecological characteristics of productivity, habitat, wildlife protection, or other important and easily
disrupted ecological values. These areas are generally recognized as significantly influencing or positively
contributing to the general overall environmental health or vitality of the entire ecosystem of a region." Mud
flats, vegetated shallows (permanently inundated, rooted aquatic vegetation), and wetlands are three
specific special aquatic sites that are mentioned in the Guidelines. Eelgrass is a type of rooted aquatic
vegetation, but algae, such as rockweeds and kelps, are not.
The EPA regulates, by issuing permits, the discharge of any pollutant from a point source into the waters of
the United States under Section 401 of the Clean Water Act. Municipal sewage treatment plants, industrial
wastewater, and stormwater originating from certain classes of activities are all regulated under this system.
Nonpoint sources of pollution are not directly regulated at the federal level. However, in accordance with
section 6217 of the Coastal Zone Act Reauthorization and Amendments of 1990, the Maine Coastal
Nonpoint Source Control Program (State Planning Office and Department of Environmental Protection
1995) was submitted to NOAA and EPA. This program, which has not yet been approved by EPA, calls for
the implementation of Best Management Practices in all towns that have land areas draining directly into
coastal waters. Funding for demonstration projects and training programs that address the reduction of non-
point sources of pollution are available from EPA.

4
State of Maine

The Natural Resources Protection Act (NRPA) of Maine requires permits to be issued by the Department of
Environmental Protection (DEP) for any dredging, bulldozing, removal of material, draining or dewatering,
filling, and construction, repair or alteration of any permanent structure that affects coastal wetlands, defined
as all tidal and subtidal lands, and other regulated areas. Structures that are not permanent (not in a fixed
location in on or over the water for a period exceeding seven months each year) are not regulated.
Proposed activities fall into two categories, Permit-By-Rule and Full Permit. Permit-By-Rule are specific
activities (e.g., construction of an individual pile-supported pier) that should not significantly affect coastal
wetlands when carried out according to DEP prescribed standards. Full permit activities have a greater
potential for affecting coastal wetlands, and to these DEP applies the standards of avoidance, minimum
alteration, compensation, and no unreasonable impact. Applicants for a Full Permit must demonstrate that
there will be "no meaningful loss of the wetland's function and values based on the nature of the activity."
The functions to be considered include food chain support, fisheries, wetland plant habitat, aquatic habitat
and wildlife habitat, but there is no standard methodology for assessing functions and values of coastal
wetlands (Doug Burdick pers. comm.).
Federal and state permitting may be more closely aligned in the future. The Army Corps of Engineers has
recently initiated a three-category resource assessment policy for tidal and navigable waters. Category I
activities are those with little potential to affect special aquatic sites, and have no equivalent under NRPA at
present. Category II and Individual Permit activities are equivalent to DEP Permit-By-Rule and Full Permit
activities, respectively.

The DEP regulates, by issuing permits, the discharge of pollutants from point sources into Maine waters
under the Maine Protection and Improvement of Water Act. Under Maine's Water Classification Program
(Title 38, Article 4-A, ï¿½464.470), estuarine and marine waters were placed into one of three classes (SA, SB,
or SC). Each class of water is suitable for specific activities, and has designated standards for dissolved
oxygen, bacterial content, and discharge of pollutants. Class SA waters, for example, "shall be of such
quality that they are suitable for the designated uses of recreation in and on the water, fishing, aquaculture,
propagation and harvesting of shellfish and navigation and as habitat for fish and other estuarine and
marine life." Specific protection for marine benthic vegetation would require defining some of these terms
(Lee Doggett pers. comm.).
Nonpoint sources of pollution are indirectly addressed through state laws that are administered by the DEP
(Site Location of Development Act, Natural Resources Protection Act), and by municipalities (Municipal
Subdivision Law and Land Use Regulation Act). Recommendations for reducing nonpoint sources of
pollution have been made in the Maine Coastal Nonpoint Source Control Program (see CURRENT
PROTECTION - Federal).
The Department of Marine Resources (DMR) regulates impacts on benthic marine plants in two ways. First,
it requires purchase of a license for the harvesting and selling of seaweeds. Second, it requires applicants
requesting suspended, net pen, or bottom aquaculture leases to assess the resources, including the
submerged vegetation beds, in the lease tract. In addition, DMR inspects the proposed aquaculture site
prior to the initial lease agreement and prior to lease renewals.
The Maine Seaweed Council, formed in 1993, is a coalition of seaweed industry representatives, seaweed
farmers, researchers, educators, and govemment representatives. One objective of the Maine Seaweed
Council is to protect and promote the sustainable use of macroalgae harvested from and grown in Maine's
coastal waters. The Council is working to develop industry standards and to educate the public about the
value of this resource. The Council is located at 7 Industrial Park, Brunswick, Maine, 04011.

DEVELOPMENT OF STATE OF MAINE GUIDELINES
The State of Maine needs to develop specific guidelines for the protection of eelgrass, rockweed, and kelp
habitat. Research has shown these types of benthic vegetation are important because they help stabilize

5
sediments and control erosion, contribute to nutrient cycling, provide food to grazers and detritus feeders,
help maintain water quality, and attract numerous species of invertebrates and vertebrates by adding
structural complexity to the water column and substratum. The goals of the protection guidelines should be
to:

* evaluate existing statutes to determine their applicability to eelgrass, rockweed, and kelp habitat,
*	coordinate federal, state and local efforts to protect eelgrass, rockweed and kelp habitat,
*	describe suitable activities, suitable water quality characteristics, and suitable methods for assessing the
functions and values of eelgrass, rockweed and kelp habitat,
*	reduce point and non-point sources of pollution near eelgrass, rockweed or kelp habitat,
*	avoid placement of coastal projects in eelgrass, rockweed or kelp habitat whenever possible,
*	minimize impacts if a coastal project must occur in eelgrass, rockweed or kelp habitat,
ï¿½	compensate for unavoidable impacts due to coastal projects in eelgrass, rockweed or kelp habitat,
*	incorporate into the permitting process consideration of the cumulative and temporal impacts of coastal
projects on eelgrass, rockweed or kelp habitat,
*	address the exclusion of non-permanent structures from the permitting process, and
*	address the problem of introduced or non-native species.

RECOMMENDED STANDARDS FOR EELGRASS, ROCKWEED AND KELP HABITAT UNDER NRPA
The standards outlined below are recommended for Permit-By-Rule and Full Permit activities under NRPA,
that is, activities which include any dredging, bulldozing, removal of material, draining or dewatering, filling,
and construction, repair or alteration of any permanent structure (including aquaculture) that affects
eelgrass, rockweed or kelp habitat.

1. Avoid placing coastal development projects in eelgrass, rockweed and kelp habitat when possible.
Permit-By-Rule and Full Permit activities should not occur in areas with eelgrass, rockweed or kelp habitat.

The applicant, with assistance from resource and regulatory agencies, should determine whether eelgrass,
rockweeds or kelps occurs at the location of proposed project. If one or more of these species is present,
the project should be relocated if possible.
A. Eelgrass habitat in Maine is in the process of being mapped. For activities with little potential for
affecting eelgrass (ACOE Category I), consult the digital files of eelgrass distribution available
through the DEP regional offices (for areas that have been surveyed recently) or the Coastal Marine
Geology Maps (for areas not surveyed) available through the Natural Resources and Information
Mapping Center, Department of Conservation. For activities with a high potential to affect eelgrass
(ACOE Category II or III; NRPA Permit-By-Rule and Full Permit), conduct an on-site, in-water
survey by snorkeling, SCUBA diving, underwater video or other means.
B. Rockweed habitat has not been mapped. For activities with a high potential to affect rockweeds
(ACOE Category II or III; NRPA Permit-By-Rule and Full Permit), conduct an on-site, on-foot survey
of the intertidal zone.
C. Kelp habitat has not been mapped. For activities with a high potential to affect kelps (ACOE
Category II or II; NRPA Permit-By-Rule and Full Permit), conduct an on-site, in-water survey by
snorkeling, SCUBA diving, underwater video or other means.

6
2. Minimize the impact of coastal development projects through site design and mitigate for loss of functions I
and values.

If there is no practical alternative to placing a coastal development project in eelgrass, rockweed or kelp
habitat, then the project should be designed to minimize its impact on the habitats, and mitigation for lost
functions and values of these habitat should be included in the project.

A. The applicant should incorporate the following into the site design as appropriate:
1) The amount of dredging and filling that must occur should be minimized.
2) Structures over eelgrass, rockweeds, or kelps should be supported by piers, and should be as
high as possible, as narrow as possible, and oriented as close to north-south as possible to
reduce shading. Specifically, boat docks should be I m wide and oriented to within 100 of north-
south with the base of the decking at least 3 m above the marine bottom (Burdick and Short in
prep.). If the orientation is beyond the 10ï¿½ limit, then 0.2 m should be added to the height for
each 100 increment; if the dock is wider than 1 m, then 0.2 m should be added to the height for
every 0.5 m increment in width (Burdick and Short in prep.).
3) Current velocity and wave energy in and around eelgrass, rockweed or kelp beds should be
determined prior to initiation of the project, and should be maintained by careful placement of
fill, designing pier structures with maximum spacing between pilings (width of free flow.total
width = 0.86:1) (Colarusso pers. comm.), use of rip rap slopes rather than bulkheads, and
preservation of the natural contour of the shoreline.
Mitigation or replacement for lost functions and values of should be incorporated into the permit. However,
the level of mitigation should depend on the type of impact. Standards should describe when mitigation is
required, how mitigation is to be accomplished, and how mitigation success is to be evaluated.
1) Mitigation for eelgrass involves transplanting eelgrass to another site. Methodology for the
transplanting process has been studied and evaluated in several places on the east coast of the
United States (e.g., Fonseca et al. 1982). The level of mitigation required would depend on the
type of impact. Creation of new substrate for transplanting eelgrass would be required for the
most severe case of mitigation, for example, permanent loss of eelgrass habitat. Transplanting
onto existing substrate, would be required for temporary losses of eelgrass.

2)  No mitigation methodology exists for algal vegetation, and requiring mitigation before a
methodology exists would be inappropriate. Development of a methodology, perhaps as a
small scale pilot project, should be considered. A possible method of mitigation does exist for
kelps. Adult kelps can be removed from their natural substrate, affixed with wires to an
artificial substrate (cement block), and transplanted to a new habitat (described in Bologna and
Steneck 1993). Transplanted kelps would provide habitat for benthic organisms, and would
serve as a source of propagules.

3. Compensate for loss of eelgrass, rockweed or kelp habitat due to coastal projects. Avoidance of
eelgrass, rockweed and kelp habitat and minimization of impact should always be fully pursued before
compensation is considered.
Standards should describe when compensation is required, how compensation is to be accomplished, and
how compensation success is to be evaluated. Compensation standards have been developed at the
federal level for eelgrass, but no compensation standards exist for rockweeds and kelps. The EPA often
requires replacement of eelgrass in a ratio of 3: 1 as compensatory mitigation for projects that destroy
eelgrass, and for losses greater than 3 acres, mitigation must be completed and must be proven successful
before construction is allowed.

7
4. The Department of Environmental Protection should evaluate, as part of the permitting process, the
cumulative and temporal impacts of coastal projects on eelgrass, rockweed, and kelp habitat.

The cumulative and temporal impacts of coastal projects on the three types of vegetation need to be
addressed. Although an individual project, for example a single dock, may have little impact on benthic
vegetation, the cumulative impacts of numerous docks may be sufficient to destroy the vegetation. A
dredging project that will suspend sediments for months will be more damaging to these three types of
vegetation than a dredging project that will be completed in a week.

5. The Department of Environmental Protection should evaluate the impact of non-permanent structures on
eelgrass, rockweed and kelp habitat, and should include in the permitting process the structures that have a
large impact on eelgrass, rockweed and kelp habitat.


RECOMMENDED WATER QUALITY STANDARDS
The standards outlined below are recommended for the purpose of reducing both point and nonpoint
sources of pollution in areas with eelgrass, rockweed or kelp habitat. Water quality has been identified as
important for maintaining healthy populations of these three types of benthic vegetation. High levels of
sediments and nutrients decrease the amount of light that reaches benthic vegetation, stimulate the growth
of species that compete with benthic vegetation, and reduce available oxygen. In addition, a number of
chemicals are toxic to some life history stages of eelgrass, rockweeds and kelps. As the human population
along the coast grows, maintaining good water quality will become increasingly important for the
conservation of these three types of benthic vegetation. The standards recommended below can be
incorporated into state laws, town ordinances, town and regional comprehensive plans, and watershed
management plans.
1. Minimize point sources of pollution by developing water quality requirements for eelgrass, rockweed and
kelp habitat and incorporating them into the State's Water Classification Program. Water quality
requirements should include light attenuation coefficient, secchi depth, total suspended solids, chlorophyll a,
dissolved inorganic nitrogen, dissolved inorganic phosphorus, heavy metals, chlorate, and biological oxygen
demand.
2. Minimize the impact of nonpoint sources of pollution on eelgrass, rockweed and kelp habitat by:
A. Implementing Best Management Practices (available for erosion and sediment control and
stormwater management) in land areas that drain directly to coastal water and that have new or
significant existing sources of nonpoint pollution. Best Management Practices will reduce nonpoint
sources of sediment, nutrients and toxic substances.

B. Eliminating overboard discharges from vessels by increasing the number of pumpout stations
and publicizing their locations.
C. Requiring proper maintenance of septic systems and domestic overboard discharges.


RECOMMENDED RESEARCH AND EDUCATION

Eelgrass

1. Develop educational material for the public that describes the ecological requirements and ecological
functions of eeigrass.

2. Dedicate funds to complete the eelgrass survey/mapping project and to provide for periodic upgrades. By
1997 most of the coast will have been surveyed and the distribution of eelgrass mapped. When the coast

8
has been completely mapped, developers will be able to review sites early in the planning stages. Data
should be provided to each DEP regional office.

3. Develop a GIS data layer of coastal structures (docks, dredge projects, fill projects) that could be
combined with the data layer of eelgrass habitat to assess the scope of cumulative impacts.

4. Develop a monitoring program to assess efficacy of site design standards, mitigation activity, and
compensation activity.
5. Determine whether a minimum distance should be maintained between eelgrass and the footprint of a
development project to avoid indirect impacts such as sediment-induced shading and alteration of the
physical environment. If minimum distances are necessary, develop a research/monitoring program to
determine the minimum buffer distance that should be maintained between eelgrass and a project footprint,
and assess efficacy of the buffer.
6. Research and test eelgrass mitigation techniques in Maine to develop feasible methodologies for
compensating for unavoidable impacts.

7. Research mitigation policies developed by other states that might serve as a model for Maine, determine
the success of these mitigation policies, and develop a mitigation policy for eelgrass in Maine.
8. Conduct seasonal, site-specific surveys to assess the use of eelgrass by marine organisms and wildlife.

9. Develop water quality standards, methods for assessing functions and values, conditions and standards
for mitigation, and conditions and standards for compensation for eelgrass.


Rockweeds

1. Develop educational material for the public that describes the ecological requirements and ecological
functions of rockweeds.

2. Dedicate funds to initiate and complete a rockweed survey/mapping project and to provide for periodic
upgrades.
3. Conduct seasonal, site-specific surveys to assess the use of rockweeds by marine organisms.

4. Conduct research to determine the effects of harvesting protocols on the ability of rockweeds to
recolonize and on the benthic community associated with rockweed habitat.
5. Research mitigation policies developed by other states or countries that might serve as a model for
Maine, determine the success of these mitigation policies, and develop a mitigation policy for rockweeds in
Maine.
6. Develop water quality standards, methods for assessing functions and values, conditions and standards
for mitigation, and conditions and standards for compensation for rockweeds.


Kelps

1. Develop educational material for the public that describes the ecological requirements and ecological
functions of kelps.
2. Dedicate funds to initiate and complete a kelp survey/mapping project and to provide for periodic
upgrades.

9
3. Conduct seasonal, site-specific surveys to assess the use of kelps by marine organismns.

4. Conduct research to determine the effects of harvesfing protocols on the ability of kelps to recolonize and
on the benthic cornmunity associated with kelp habitat.
5. Develop water quality standards, methods for assessing functions and values, conditions and standards
for mifigation, and conditions and standards for compensation for kelps.


General

1. Research methods for preventing intr-oduction of non-native species.

2. If managers feel diversity is important, use comparative techniques and scale to analyze diversity in each
habitat.

3. Develop standards for using each of the three types of benthic vegetation as indicators of vvater quality.

4. Expand study of ecological role of habitat types to include unvegetated habitats.

10
REFERENCES
Duggins, D.O., C.A. Simenstad, and J.A. Estes. 1989. Magnification of secondary production by kelp
detritus in coastal marine ecosystems. Science 245: 170-173.
Emery, K.O., A.S. Merrill, and J.V. Trumbull. 1965. Geology and biology of the sea floor as deduced form
simultaneous photographs and samples. Limnol. Oceanog. 10: 1-21.
Fonseca, M.S., W.J. Kenworthy, G.W. Thayer, and D.Y. Heller. 1982. Transplanting of the seagrasses
Zostera marina and Halodule wrightii for the stabilization of subtidal dredge material. Annual Report of
the National Marine Fisheries Service, Southeast Fisheries Center, Beaufort Laboratory, Division of
Estuarine and Coastal Ecology, Beaufort, N.C. to the US Army corps of Engineers, Coastal
Engineering Research Center, Fort Belvoir, Va. 61 pp.

Heck, K.L., Jr., K.W. Able, M.P. Fahay, and C.T. Roman. 1989. Fishes and decapod crustaceans of Cape
Cod eelgrass meadows: species composition. Estuaries 12: 59-65.

Larsen, P.F. 1979. Statement of Peter F. Larsen before Spencer Nissen, Administrative Law judge in the
matter of Pittston Company/NPDE Permit application No. ME DO22420. U.S. Department of
Environmental Protection Agency, Washington, D.C.

MacDonald, J.S., M.J. Dadswell, R.G. Appy, G.D. Melvin, and D.A. Methven. 1984. Fishes, fish
assemblages, and their seasonal movements in the lower Bay of Fundy and Passamaquoddy Bay,
Canada. Fish. Bull. 82: 121-139.

Mann, K.H. 1973. Seaweeds: their productivity and strategy for growth. Science 182: 975.

Newell, R.C. 1984. The biological role of detritus in the marine environment. In: Fasham, M.J.R. (ed.). Flows
of energy and materials in marine ecosystems: theory and practice. Plenum Press, NewYork. Pages
317-343.
Olney J.E. and G.W. Boehlert. 1988. Nearshore ichthyoplankton associated with seagrass beds in the lower
Chesapeake Bay. Mar. Ecol. Prog. Ser. 45: 33-43.

Simenstad, C.A. and R.C. Wissmar. 1985. 813C evidence of the origins and fates of organic carbon in
esturarine and nearshore food webs. Mar. Ecol. Prog. Ser. 22: 141-152.

State Planning Office and Department of Environmental Protection. 1995. Maine Coastal Nonpoint Source
Control Program. Submitted to the National Atmospheric and Oceanic Administration and the US
Environmental Protection Agency on July 19, 1995. State Planning Office, Natural Resources Policy
Division and Department of Environmental Protection, Bureau of Land and Water Quality, Augusta,
Maine.

Teal, J.M. 1962. Energy flow in the salt marsh ecosystem of Georgia. Ecology 43: 614-624.
Weinstein, M.P. et al. 1980. Mar. Biol. 58: 227-243.

11
EELGRASS

IDENTIFICATION AND DESCRIPTION
Eelgrass (Zostera marina L.) is one of approximately 47 species of seagrasses, flowering plants which
complete their entire life cycle in seawater (Thayer et al. 1984). Like all seagrasses, eelgrass has leaves
and an extensive underground system of roots and rhizomes (Thayer et al. 1984). The epidermis of the
leaves contains a high concentration of chloroplasts and is the principal site of photosynthesis (Sculthorpe
1967; Tomlinson 1980). The roots and rhizomes anchor the plants in the substrate, and like the leaves,
absorb nutrients and gases. In common with its terrestrial relatives, eelgrass has a functioning vascular
system (xylem and phloem tissue) which allows transport of substances through the plant. Eelgrass exhibits
both vegetative expansion (by the underground roots and rhizomes) and sexual reproduction (by seed).
Vegetative reproduction is important for the maintenance of eelgrass meadows (Tomlinson 1980). Seeds
are important for colonizing unvegetated areas, and for genetic adaptation.

An individual plant consists of a rhizome which bears 2-5 linear strap-shaped leaves (Thayer et al. 1984).
The leaves are enclosed at the base by a sheath that forms a stem-like structure. Younger leaves are
subtended by older leaves giving the shoot a laterally compressed appearance. The leaves are
approximately 0.3-0.7 cm wide and are rounded at the tip. Most eelgrass is perennial, but annual
populations have been described in Maine.


DISTRIBUTION
Eelgrass has a wide geographical distribution. It is found on the east coast of North America from Hudson
Bay to the Carolinas (33-650N), in northern Europe from Spain to Finland and the Barents Sea, on the west
coast of the United States from Baja California to Alaska, and along the coasts of Japan, North Korea and
South Korea (Thayer et al. 1979; Thayer et al. 1984). In many areas eelgrass is the dominant species of
submerged aquatic vegetation (Thayer et al. 1984).
Eelgrass grows in shallow coastal water in a variety of substrates in salinities ranging from 0 to 30 ppt and in
water temperatures ranging from less than 0ï¿½C to greater than 30ï¿½C (Thayer et al. 1984). Eelgrass is limited
to relatively shallow water primarily by the availability of light (Thayer et al. 1984). In Maine, eelgrass is
found from the intertidal zone to a depth of approximately 11 m (Short pers. comm).
The geographical distribution and relative abundance of eelgrass along the Maine coast currently is being
mapped for the Oil Spill Response Commitee by Seth Barker of the Department of Marine Resources, and
will be available as a GIS data layer. The distributional maps are based on interpretation of aerial
photographs at a scale of 1:12000 (1 in = 1000 ft) and subsequent site visits to assess the accuracy of the
interpretation (groundtruth). The areas with the greatest probability of experiencing an oil spill were mapped
first. Provisional maps have been completed for Casco Bay, Cobscook Bay, Penobscot Bay, and the
Piscataqua River. Information will soon be available for the mid-coast region (Small Point to Pemaquid
Point), Machias Bay, and the St. Croix River. Photography recently has been acquired for the coast from
York to Cape Elizabeth and Pemaquid Point to Owi's Head, and maps of these areas should be available in
1996. For the areas mapped to date, Casco Bay has the greatest area of eelgrass (Table 1; Figs. 1-6).


LIFE CYCLE
The generalized life cycle of eelgrass can be described in four stages (Setchell 1929). The seed
germinates (stage 1); the first shoot develops, new leaves are added, and roots and rhizomes form (stage
2); there is further development of existing shoots and addition of new ones by vegetative reproduction
(stage 3); some of the oldest shoots develop into erect flowering stalks (stage 4). Growth occurs at the
meristem in the basal area of the leaf, and can involve an increase in the length of the primary shoot, the
number of leaves per shoot, or the number of shoots per original seedling (Thayer et al. 1984). Leaf growth

12
Table 1. Number and area of eelgrass beds by percent cover for seven coastal regions, total area of
eelgrass beds by region, and total area of eelgrass beds for all regions (data provided by Seth Barker). In
the Penobscot Bay region, eelgrass beds were coded by percent cover near Vinalhaven and Northhaven,
but were not coded by percent cover elsewhere. In Machias Bay, a percent cover could not be assigned to
the five uncoded eelgrass beds.
Region          Percent cover   Number   Area by cover              Area by region
(m2)           (mZ)       (acres)  (hectares)
Piscataqua River               0-10	22	151,521.3
10-40	41	170,532.9
40-70	46	518,419.7
70-100	30	317,488.8
1,157,962.6     286        1152
Casco Bay                      0-10	201	3,687,447.8
10-40	393	4,982,218.8
40-70	529	5,513,694.5
70-100	307	14,372,432.3
28,555,793.4    7056        2855
Mid-Coast	0-10	70	757,641.9
(Small Point to	10-40	122	1,026,437.4
Pemaquid Point)	40-70	211	1,423,699.3
70-100	94	995,252.0
4,203,030.6    1038         420
Penobscot Bay                  0-10	31	360,997.0
10-40	63	241,322.6
40-70	88	473,944.5
70-100	132	724,451.9
uncoded	381    10,464,684.1
12,265,380.1    3030        1226
Machias Bay                    0-10	32	2,126,930.2
10-40	86	2,408,884.4
40-70	51	1,786,837.2
70-100	24	243,791.5
uncoded	5	224,714.6
6,791,157.9    1678         679
Cobscook Bay                   0-10	70	548,841.4
10-40	173	1,442,128.2
40-70	208	2,226,190.3
70-100	44	468,833.7
4,685,993.61    1157         468
St. Croix                      0-10	6	35,261.6
10-40	9	114,638.6
40-70	5	47,246.8
70-100	2	24,397.5
221,544.6       54	21

Total                                                            57,880,862.7   14299	6821


FIgLWe~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~
~~~~~~~~~~~~~~~~~~~z 1.Psaa aieai

















lb


3                33'

3                                      3        3
3  '3
3,


3  3
.33
I)                                               -          3  3
'/_ IVLL_ 33 3



.ry
3333                           333
3_3  3

33 3




_33   -            3        3                                           3      3   3'


3.                3	333.3.3
3	3        N    3

3    3        3  3              3	3   3    3




3              3  3	3
3)	4














3                  3           3              3	3	3
3	4



3	33333









_33	.3	33 3
3                3	3       3



33                        4




3  3         3   3   3
















_    B
-I
* -a
-3



































































































































































3              I
/




'S







p
L
I

(IIU



_wit CovEr
7j0-10%
10-40%
*4CY70%
m 70-
100%
s' . .

i
'% IV ;
I . i I.. II ,,
4

251111    3~~~~~~~~~~~~~~~~~~~~~~~~~~~~0




























S
 



Prtland.j

II


I
0
p
v
'i
0

0
0
a
d
0
Fbrout Cove
10-10%
~~40-70 %
-~ 70-IOfD/o
0
N
A
-.k
9 Ki I on-des -th.
3 0
3 6
3      0      3      6      9 KiIon_es

15












3.Midcoast-Snidl  toP  PeointCoe














10-40%
~4070%
.~~70-1000/o
2      0      2      4 liIom-eters
T%ii"

-   % I . " I -1                - I 11 . I . I
V         16
ttonin
-
0
I
.'. -, 4
I
a
0*
p
f%- -%
a
9
P a

0
6
0r
a0            --
a
. 0
r
0
.0
t
9      , 6_206
4
8
U:    Q.    4
8~~~~
I

Figure 5. Machias and Little Matchias
Bays
17













































0~~~~~~~~~~~~~~~~q














F9~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~
. t I
, i
I -.1





I4














2
0

10401O/o

-~ UroodEd
C
1.5            0            1.5
3 K Icantfs
iiiiiiiii

Figure 6. Cobscook Bay and Vicinity             18
























'4070%

19
is fastest just after it emerges and decreases with age. Eventually the oldest leaf is sloughed off and
replaced by a young rapidly growing leaf.
Flowering shoots develop from the vegetative plants in the second season of growth (Setchell 1929), and
the time of flowering is strongly influenced by water temperature. The flowering process requires 30-60
days, and occurs later in the season and lasts longest in the north (Phillips et al. 1983b). Pollination occurs
entirely underwater. Cross pollination is normal (DeCock 1980), but self pollination can occur. Pollination
depends in part on water movement, perhaps explaining the 14-72% range of successful pollination that has
been reported (Churchill and Riner 1978; Orth and Moore 1983a; Kenworthy et al. 1980).
The contribution of seeds to the abundance of eelgrass depends on the abundance of flowering shoots, the
number of seeds produced and the rate of seed germination. All of these can vary widely, although there
may be a trend towards increased flowering in disturbed sites (Thayer et al. 1984). Unlike seeds deposited
in quiescent areas, those deposited in open-water, high-energy habitats with strong currents and
considerable wave action may not become established (Fonseca et al. 1982a). In high energy habitats,
growth may be restricted to vegetative reproduction.
From NewYork to Nova Scotia, eelgrass seeds mature between June and July (Silberhom et al. 1983). In
New York, seeds collected in July germinated from October to December;, seedling growth and development
occurred in autumn and spring but not through the winter (Churchill 1983). Throughout its range, eelgrass
biomass peaks in the summer and sharply declines in winter. In northern latitudes shoots produced in early
spring will reproduce vegetatively through the summer and early fall.
In Nova Scotia, Maine, and Rhode Island an annual form of eelgrass has been described, although
differences in the two forms are ascribed to nongenetic factors (Gagnon et al. 1980). In regions wiere ice
forms in winter, the annual form grows intertidally and in shallow water and the perennial form grows in
deeper water (Robertson and Mann 1984). Beds that recolonize by seed each year may not be evident
early in the growing season.


POPULATION TRENDS
Disease
Between 1930 and 1933 a '"asting disease" destroyed approximately 90% of the eelgrass throughout its
range (Tufin 1942). At the time it was suggested, but never proven, that a marine slime mold was the
causative agent. Bacteria, fungi, commercial fish harvesting, pollution, competing organisms, and unusual
climatic conditions including reduced light levels were also postulated as the causative agent of the wasting
disease (Thayer et al. 1984).
Extensive recolonization by eelgrass over its range did not occur until after 1945 and full recovery, wMen it
occurred, took 30 to 40 years (Rasmussen 1977). Eelgrass in salinities less than 12 to 15 ppt apparently
was immune to the wasting disease and eventually formed the stocks for recolonization (Rasmussen 1973).
In Chesapeake Bay (Orth and Moore 1981, 1982b) and in Cape Ann, Massachusetts (Dexter 1985) large
scale oscillations in abundance, perhaps due to the wasting disease, have occurred since the 1930s. In
other areas, eelgrass still has not returned (Thayer et al. 1984).
In the early 1980s, a new outbreak of disease and localized declines in eelgrass abundance occurred on
both sides of the Atlantic (Short and Mathieson 1985; Short et al. 1986; Short et al. 1988). The cause of the
decline, documented in the laboratory and in mesocosm experiments, was a marine slime mold
(Labyrinthula zosterae) (Muehlstein et al. 1991) which produced symptoms similar to those of the 1930s
wasting disease (Short and Mathieson 1985; Short et al. 1986; Short et al. 1987; Short et al. 1988;
Muehlstein et al. 1988; Muehlstein 1989). Burdick et al. (1993) have developed an index, based on visual
determination, for assessing and monitoring the progress of the disease.

20
It is impossible to determine whether the decline of eelgrass in the 1930s resulted in an overall decline in
commercially important fish and invertebrates because quantitative data was not available on the fauna that
utilize eelgrass beds (Thayer et al. 1984). Biological and geomorphological changes were reported in some
places. After eelgrass beds were lost in Denmark, the substrate became coarser and deposits of fine mud
adjacent to eelgrass beds changed from anoxic to oxidized sediments (Rasmussen 1973, 1977). In Woods
Hole, overall species abundance decreased when eelgrass beds were lost (Stauffer 1937), but similar
changes were not seen in Denmark (Rasumssen 1973, 1977). These seemingly contradictory results may
reflect differences in sampling technique or differences in eelgrass function with geographic location.
In the United States, catastrophic declines were documented for the Atlantic brant (Branta bemicla hrota)
and the bay scallop (Argopecten irradians) following the 1930s decline of eelgrass; a more limited decline
was documented for the Canada goose. Pdor to the wasting disease, brant geese fed almost exclusively on
eelgrass (Cottam 1934; Cottam et al. 1944; Cottam and Monroe 1954; den Hartog 1977). The decline of the
population was exacerbated by hunting and poor reproductive success (Corr pers. comm.). Following the
wasting disease, the abundance of bay scallops declined precipitously in North Carolina and Chesapeake
Bay. All stages of the bay scallop, which do not occur in Maine, feed on eelgrass-derived detritus (Kirby-
Smith 1972; Kirby-Smith and Barber 1974) and the postlarvae settle on eelgrass leaves (Gutsell 1930;
Thayer and Stuart 1974). Recovery to commercially harvestable quantities never occurred in Chesapeake
Bay and required neady 30 years in North Carolina (Orth and Moore 1982b).

ECOLOGICAL REQUIREMENTS
Light
Sunlight is the most fundamental requirement for eelgrass because light energizes the reactions of
photosynthesis. As light passes through water and the substances dissolved and suspended in it, different
wavelengths of light are selectively absorbed, causing the total amount of available light and the range of
available wavelengths to decrease with depth. The depth at which eelgrass can grow is limited by the
amount of light that reaches the leaf canopy.
A number of experiments using shading devices have shown that reductions in light cause significant
decreases in plant density (Backman and Badlotti 1976; Burkholder and Doheny 1968; Short 1975; Short et
al. 1993). The density of plants grown in reduced sunlight (63% reduction) was 5% that of plants grown in
full sunlight after nine months (Backman and Badlotti 1976). Mesocosm experiments conducted in New
Hampshire (Short et al. 1993) demonstrated a logarithmic decline in shoot density and biomass productivity
of eelgrass with a decrease in light intensity after five months. Shading appears to have the greatest effect
on plants growing in minimal light (Dennison and Alberte 1982).
Temperature
Eelgrass tolerates a wide range of temperatures (0C to 30ï¿½C), but temperature extremes in combination
with other factors can have detrimental effects (Thayer et al. 1984) In southern latitudes, exposure and
desiccation are a threat, while in northern areas ice moving with winds and tides can scour the bottom and
uproot plants. In one winter in Rhode Island, shoot density declined from 4000 to 400/m2 due to ice (Short
1975).
Salinity
Eelgrass has been reported growing at salinities from nearly fresh (0 ppt) to full strength seawater (35 ppt) or
higher (Thayer et al. 1984). An optimum salinity has never been determined. However, salinity may affect
seed germination which has been reported to increase as salinity decreases (Lamounette 1977; Phillips et
al. 1983a).

21
Water motion
Several researchers (Phillips 1972; Conover 1964; Short 1975) have noted that the most luxuriant eelgrass
beds are usually in areas of moderate to high current speed. Water currents of increased velocity may
make nutrients more available to the leaves (Conover 1964, 1968), perhaps accounting for the luxuriant
growth. Conover (1964) and Fonseca and Kenworthy (1987) suggested there may be an optimum current
speed between 20-40 cm/sec belowwhich metabolism may be limited by diffusion, and above which growth
may decline as a result of physical disruption. Higher current velocities and wave action can erode
sediments, plants and seeds, prevent deposition, or deposit large amounts of sediment.
Substrate
Eelgrass grows on substrates varying from firm sand to fine mud (Ostenfeld 1908; den Hartog 1970) and
occasionally on cobble in New England (Riggs and Fralick 1975; Thayer et al. 1984). Eelgrass distribution
apparently is not limited by substrate type as long as the roots and rhizomes can penetrate the substrate
(Kenworthy et al. 1982).
Nutrients
Eelgrass can absorb inorganic compounds containing carbon, nitrogen, phosphorus, and trace metals from
the sediments via the roots and from the water column via the leaves (McRoy and Barsdate 1970; McRoy
and Goering 1974; Penhale and Thayer 1980; Brinkhuis et al. 1980; Short 1981; Thursby and Harlin 1982).
The ability to absorb nutrients from either the water column or the sediment contributes to the high
productivity that has been documented for eelgrass.
Eelgrass may be nitrogen limited under certain circumstances (Short 1981, 1983a, b; lizumi et al. 1982). If
nitrogen in the water column is insufficient, eelgrass may have to compete with microorganisms for nitrogen
in the sediments (lizumi et al. 1982; Thayer et al. 1984). However, nitrogen regeneration in organic-poor
sediments may be inadequate for eelgrass growth (Short 1983a).

ECOLOGICAL FUNCTION
Effect on water motion and sediment dynamics
Several papers have discussed the role of eelgrass in slowing current and promoting sedimentation and
sediment stability (Petersen 1918; Marshall and Lukas 1970; Orth 1977; Kenworthy et al. 1982; Fonseca et
al. 1982b, 1983). In the laboratory, water currents cause the eelgrass leaf canopy to bend into a compact
layer over the substrate, and current speed is reduced in the leaf canopy due to drag forces (Fonseca et al.
1982b; Gambi et al. 1990; Ackerman and Okubo 1993). The reduced velocity near the bottom should
enhance the deposition of fine particles and reduce scour (Fonseca et al. 1982b). The network of roots and
rhizomes helps to bind sediment, reduce erosion, and preserve the sediment microfauna.
The patterns of water motion and sediment dynamics in an eelgrass meadow are dependent on a number of
variables including the size of the meadow, the density of the eelgrass, current velocity, wave energy, the
interaction between currents and waves, water depth, fetch, and proximity to inlets, channels, and sediment
sources (Thayer et al. 1984). Eelgrass can tolerate current velocities of 120-150 cm/sec, but the
configuration of the meadow and percent of fine particles and organic matter in the surface sediment varies
with current velocity (Fonseca et al. 1983). Meadows in high current areas (>90cm/sec) may be sources
and those in low current areas (<50 cm/sec) may be sinks of eelgrass-based organic matter (Fonseca et al.
1983).
Waves, particularly shallow water waves (water depth:wave length < 1:2), can generate lift forces as a wave
trough passes that suspends sediment into the water column (Thayer et al. 1984). Seagrasses are as
effective as emergent marsh plants in damping out waves, providing the seagrass canopy extends to the
surface of the water. Under these conditions, both seagrasses and marsh plants dampen out wave energy

22
approximately 1 m into the vegetation (Wayne 1975; Knutson et al. 1982; Fonseca and Cahalan 1992) and
may prevent erosion.
Nutrient Cycling

Chemical elements in an ecosystem cycle back and forth between the living organisms and the physical
environment in a series of transfers termed biogeochemical cycles. The cycles involving carbon, nitrogen,
phosphorus, and the trace metals are extremely important because these elements are vital to the
maintenance of life. These elements make up the biomass of living organisms, and also exist in the
sediment and water column as inorganic compounds, particulate organic matter, and dissolved organic
matter.

Primary producers (autotrophs) alone absorb inorganic carbon, nitrogen, phosphorus, and trace metals, and
use the sun's energy to convert them into living organic matter (biomass). Important primary producers in
the Gulf of Maine include single-celled organisms (e.g. bacteria, diatoms, dinoflagellates) that drift in the
water column as plankton or are associated with the substrate, macroalgae (e.g. rockweed, kelp, Irish moss)
that usually are attached to a substrate, and flowering plants (e.g. salt marsh grasses and eelgrass) that are
rooted in the substrate and inundated to various degrees. Biomass synthesized by primary producers is the
ultimate source of nutrients and energy for heterotrophs (bacteria, other single-celled organisms, and multi-
celled animals). Heterotrophs in tum excrete waste products containing inorganic nutrients which are
absorbed and utilized by primary producers.
Eelgrass meadows are important in the cycling of nutrients because they 1) can produce large quantities of
organic matter, 2) tend to be depositional areas where organic matter is retained, 3) harbor a diverse
community of tightly-coupled heterotrophs that consume organic matter and release inorganic nutrients, and
4) are able to absorb nutrients from the water column or the sediment (Thayer et al. 1984). Research in
different geographical areas has shown that sediments in eelgrass meadows have a higher concentration of
organic matter when compared to sediments in adjacent unvegetated areas (Marshall and Lucas 1970;
Thayer et al. 1975; Orth 1977; Kenworthy et al. 1980).
In the sediment, dissolved and particulate organic matter are consumed by heterotrophs ranging from
anaerobic bacteria to large invertebrates that require oxygen. These organisms excrete inorganic waste
products (ammonium and phosphate). Compared to unvegetated areas, sediments in eelgrass meadows
have greater concentrations of ammonium and phosphate (Kenworthy et al. 1982; Billen 1978; Blackbum
1979; lizumi et al. 1982; McRoy et al. 1972). Other bacteria in the sediment convert nitrogen gas to usable
nitrates and nitrites. In addition, eelgrass beds, with their high rates of sedimentation, may act as sinks for
metallic elements, and the activity of anaerobic bacteria can make them available for absorption (Wolfe et
al. 1976).

In the water column, dissolved and particulate organic matter are consumed by aerobic heterotrophs. These
organisms then excrete inorganic waste products containing ammonium and phosphate into the water
column.

Eelgrass is able to absorb nitrogen (as ammonium or nitrate) and phosphorus (as phosphate) through its
roots, rhizomes, and leaves (Penhale and Thayer 1980; lizumi et al. 1982; lizumi and Hattori 1982; Thursby
and Harlin 1982; Short and McRoy 1984; Brix and Lyngby 1985). Recent evidence suggests that eelgrass
does not significantly contribute to cycling of phosphorus between the sediment and water column (Brix and
Lyngby 1985). Eelgrass also is able to absorb elemental metals (cadmium, zinc, iron, copper, manganese,
lead, vanadium, and nickel) through its leaves and root/rhizome complex (Faraday and Churchill 1979;
Driftmeyer 1980; Driftmeyer et al. 1980; Brinkhuis et al. 1980; Penello and Brinkhuis 1980; Lyngby and Brix
1982; Kurata et al. 1979; Lyngby et al. 1982; Brix et al. 1983).

23
Food Source

Grazing on live eelgrass

Few organisms feed directly on living eelgrass, perhaps because of the reduced availability of nitrogen
compounds, the indigestible structural components of the cell wall, or the presence of toxic or inhibitory
compounds (Thayer et al. 1984). Living plant material (eelgrass leaves, shoots, and seeds) reportedly is
eaten by approximately 80 species of annelid worms, molluscs, crustaceans, echinoderms, fishes, reptiles,
and birds (McRoy and Helifferich 1980; Thayer et al. 1984). American brant, which declined dramatically
following the decline of eelgrass in the 1930s used to feed exclusively on eelgrass. In Maine, eelgrass
meadows are important winter feeding areas for migratory populations of Canada geese and other waterfowl
(Corr pers. comm.).

Grazing on epiphytes
Eelgrass leaves support a large number of epiphytes, organisms that grow on the leaves (use them as a
substrate), and that may or may not derive nutrition from the eelgrass (Kikuchi and Peres 1977; Harlin
1980). Epiphytes on eelgrass include bacteria, fungi, microalgae, and macroalgae (Harlin 1980). Epiphytic
bacteria utilize organic carbon derived from actively photosynthesizing leaves (Kirchman et al. 1984).
Thayer et al. (1978) estimated that 50% of the carbon in the epiphytes may be derived from carbon released
by the eelgrass leaf.

Epiphytes are an important source of food for many species of invertebrates, fishes, and birds, some of
which are commercially important (Ewald 1969; Caine 1980; Howard 1982; Van Montfrans et al. 1982;
Robertson and Mann 1982; Thayer et al. 1984; Van Montfrans et al. 1984). The biomass of the epiphytes
may be comparable to the biomass of the eelgrass leaves themselves (Penhale 1977) and their productivity
may approach 20% that of eelgrass leaves. In addition to obtaining nutrition, grazing herbivores may
enhance the productivity of the eelgrass by cleaning the leaves; excessive epiphytes can reduce eelgrass
photosynthesis by shading and perhaps by decreasing diffusion (Sand-Jensen 1977; Van Montfrans et al.
1984; see THREATS).

Consumption of eelgrass-derived detritus

Detritus refers to dissolved organic matter, particulate organic matter, and the assemblage of decomposing
bacteria and fungi that inhabits the organic matter (Zieman 1982; Thayer et al. 1984). Dissolved organic
matter includes soluble carbohydrates and proteins (Zieman 1982); particulate organic matter includes dead
Primary producers or autotroph debris, dead heterotroph or heterotroph debris of any size, fecal pellets, and
combinations of these items (Cousins 1980).

As part of the natural life cycle, old eelgrass leaves break away from the base of the shoot and fall to the
bottom or are carried away by water currents (Phillips 1984). Whole leaves probably become fragmented
by the shredding action of herbivorous invertebrates and by physical factors such as desiccation and ice
grinding (Robertson and Mann 1980). In the laboratory, intact green leaves are consumed and shredded by
several species of crustaceans, resulting in the production of small detrital particles that are released in the
faeces (Robertson and Mann 1980). The intact leaves and leaf fragments become colonized by bacteria
and fungi which begin to decompose the eelgrass.
The processes of particle size reduction and microbial colonization and decomposition are essential to the
detrital food web,and have been described in detail by Zieman (1982). Particle size reduction is important
because some organisms are able to ingest only fine particles, and the greater surface to volume ratio of
fine particles allows greater colonization by microbes. The microbes enzymatically break down plant
material that most organisms cannot digest, and convert it to microbial biomass which can be assimilated.
Microbes also enrich the detritus by taking up nitrogen and phosphorus from the surrounding medium and
incorporating it into microbe biomass. Organisms that consume eelgrass-derived detritus obtain most of
their nutrition, especially nitrogenous compounds, from the microbes (Zieman 1982). During growth and
decomposition, seagrasses release substantial amounts of dissolved organic carbon which is quickly

24
assimilated by microorganisms. Dissolved organic carbon becomes available to animals only after
conversion to microbial biomass (Zieman 1982).

Eelgrass roots and rhizomes are a substantial source of organic matter into the sediment (Thayer et al.
1984; Kenworthy and Thayer 1984). The production and decay of roots and rhizomes contribute large
quantities of dissolved and particulate organic matter into the sediments which decays more slowly than leaf
material (Thayer et al. 1984). Unlike decomposition in the water column, decomposition in the sediment is
largely anaerobic.

Habitat

An eelgrass meadow adds structural complexity to the substrate and water column and offers a variety of
habitats for flora and fauna of all sizes. Leaves of varying age and size extend into the water column, and
change their orientation with passing waves and as current changes. Epiphytes on the leaves add additional
complexity. The dense intervwven roots and rhizomes that penetrate the substrate provide additional
unique structure.

Researchers working in different places have reported a greater diversity and abundance of species in
eelgrass meadows than adjacent unvegetated areas (Kikuchi 1966; Orth 1973; Thayer et al. 1975; Orth and
Heck 1980; Weinstein and Brooks 1983; Stoner 1983; Summerson and Peterson 1984; Bell and Pollard
1989; Heck et al. 1989; Thayer and Chester 1989; Heck and Thoman 1984). However, few comparisons
have been made between eelgrass and other vegetated areas (e.g. Sogard and Able 1991).
A functional classification of the flora and fauna of Japanese eelgrass beds (Kikuchi 1966, 1980; Kukuchi
and Peres 1977) probably is valid for Maine and includes: 1) epiphytic organisms on eelgrass blades
(macroalgae, microalgae and their associated microfauna and, meiofauna; sessile fauna attached to the
blades; mobile fauna crawling on the leaves; swimming fauna that rest on the leaves; 2) biota that attach to
blade stem and rhizomes, 3) mobile fauna that swim within and over the leaf canopy (includes diurnal or
seasonal transients or permanent residents), and 4) epibenthic and infaunal invertebrates that dwell on or
within the sediments. Harlin (1980) has compiled a list of 124 species of animals that have been reported
on eelgrass blades.

The benthic and epibenthic fauna associated with eelgrass varies with latitude because there is a wide
variation in temperature, other physical factors, and biological factors over the range of eelgrass (Thayer et
al. 1984). Cape Cod and Cape Hatteras, divide the eastern United States into three zones that differ in
climate, physiography, and hydrography. The species associated with eelgrass in each area also differs
(Thayer et al. 1984). Species composition may even change within the same estuary or within a particular
eelgrass meadow depending on salinity, sediment grain size, hydraulic regime, degree and type of physical
disturbance (Thayer et al. 1984).

Many researchers have demonstrated that eelgrass meadows are nursery areas for commercially or
recreationally important species or for species which are important as food for fish and birds (Thayer et al.
1984). Research conducted in Chesapeake Bay, North Carolina and other temperate areas showed, in
general, that the density and diversity of the fish and decapod assemblage associated with eelgrass
meadows was greater than that of unvegetated areas, and that the assembly displayed diel, tidal, and
seasonal changes in abundance and composition (reviewed in Thayer et al. 1984). However, species of
fishes that are common in eelgrass beds also are characteristic of other vegetated habitats (Weinstein and
Brooks 1983). The large mobile animals found in seagrasses can be placed into four functional groups:
permanent residents, seasonal residents, transients, and casual species (Kikuchi 1980).
In the southem Gulf of Maine, 40 species of fishes and 9 species of invertebrates have been collected from
eelgrass beds (Table 2). These collections have been made in Hingham Harbor, Nahant Harbor and
Gloucester Harbor, Massachusetts (Colarusso pers comm.); Great Bay, New Hampshire (Short 1992); the
Nauset Harbor area, Massachusetts (Heck et al. 1989); and Jordan River, Maine (Newell et al. 1991). Some
of the species are commercially important (e.g. Atlantic cod and American lobster), some are important
forage species (Atlantic silversides and sand shrimp), and others are top-level predators (wolffish).

25
Stomachs of many of the larger fish (striped bass, cod, pollack, hake, tautog, and flounders) contained
invertebrates which are more abundant in eelgrass beds than in adjacent unvegetated areas (Colarusso
pers. comm.).
Eelgrass beds at the mouths of some estuaries in Maine are the sites of extensive pdmary settling by
drifting blue mussel (Mytilus edulis) larvae (Hidu et al. 1988; Newell et al. 1991). The baffling effect caused
by the undulation of eelgrass in response to rapid tidal currents (the "honami" effect) may serve to trap the
larvae (Thayer et al. 1984; Hoven et al. 1991).
Several researchers have demonstrated the use of eelgrass beds by juvenile American lobsters (Homarus
americanus). Barshaw and Bryant-Rich (1988) conducted an 8-month study on the behavior, growth, and
survival of eady juvenile American lobsters in three habitats: mud, rock with algae, and eelgrass. Rock and
eelgrass were better habitats for the creation and maintenance of burrows, but lobster in eelgrass had a
lower mortality rate than lobsters in rock or mud. In a Cape Cod estuary, the abundance of juvenile lobsters
was greater in eelgrass beds than in adjacent unvegeated areas (Heck et al. 1989). Short et al. (in press)
found that juvenile lobsters (average carapace length of 50 mm) were residents of eelgrass beds in the
Piscataqua River, New Hampshire and Maine.


CONCLUSION
Extensive studies of eelgrass have been conducted along the east coast of the United States (specifically
North Carolina, Chesapeake Bay, and the Gulf of Maine), the west coast of the United States, in Europe and
in Japan since the 1970s. These studies collectively have shown that eelgrass 1) helps stabilize sediments
and control erosion, 2) contributes to nutrient cycling, 3) directly provides food to grazers and indirectly
provides food to detritus feeders and epiphyte grazers, 4) helps maintain water quality, and 5) provides
habitat for numerous species including commercially and recreationally important species.


THREATS
The major verified threats to eelgrass include reductions of light levels due to eutrophication, degradation of
water quality, and physical structures; episodic outbreaks of the wasting disease; physical disturbance; and
overharvesting. In addition, a number of chemical substances may affect eelgrass, but cannot be
considered as verified threats at this time. For example, the literature is ambiguous concerning the effects
of petroleum hydrocarbons on eelgrass, but this ambiguity could be explained by eelgrass being
differentially sensitive to the various types of hydrocarbons (Fred Short pers. comm.). Little data are
available regarding the effects of pesticides, herbicides, chlorine, CCA (chromated copper arsenate)
pressure-treated wood, and heavy metals have on eelgrass, especially in areas with strong tidal mixing.
Future research may reveal these chemicals to be a threat to eelgrass.
Reduction of light levels due to degradation of water quality by nutrient loading is one of the greatest threats
to eelgrass (Dennison et al. 1993). Nutrient loading can result from point sources such as wastewater
treatment facilities and aquaculture sites and from non-point sources such as runoff from agriculture, animal
wastes, and fertilized lawns (Short 1992). Excessive nutrients promote the rapid growth of phytoplankton in
the water column, epiphytes on eelgrass leaves, and entangling macroalgae, all of which can reduce
amount of light reaching the eelgrass, and thus destroy the eelgrass beds (Orth and Moore 1983b; Kemp et
al. 1983; Twilley et al. 1985; Dennison 1987; Dennison et al. 1989).
Reduction of light levels due to sediment loading also is one of the greatest threats to eelgrass. Sediment
loading can result from point. and nonpoint sources of upland runoff and from resuspension of sediment by
tidal currents, wave action, dredging, filling, boating activity, and fishing activity (Short 1992; Colarusso
pers. comm.). Sediment suspended in the water column reduces the amount of light that penetrates the
water, thus reducing the light that reaches eelgrass rooted on the bottom (Dennison 1987). The impact of
suspended sediment in the vicinity of eelgrass may vary with hydrographic conditions, sediment grain size,
time of year, distance from the sediment, health of the rockweeds, and a host of other variables. As

26
Table 2. Species of fishes and invertebrates collected from eelgrass beds in the southemn Gulf of Maine.
Species/Reference            Common name                 Season of use             Frequency of
occurrence
Fishes
Alosa aestivalis
Alosa pseudoharengus 1,2
Alosa sapidissima2
Amrnmodytes americanus3
Anarhichas lupus 1
Anguilla rostrata 3
Apeltes quadricus 23
Brevoortia tyrannus
Centropristis striata 3
Clupea harengus harengus3
Cyclopterus lumpus 1,2,3
Fundulus heteroclitus 2,3
Fundulus majalis
Gadus morhua 1
Gasterosteus aculeatus 12,3
Macrozoares americanus t
Menidia menidia 1,23
Microgadus tomcod 1,2,3
Morone saxitilis 1
Myxocephalus aenaeus 1"3
Myxocephalus scorp io 1
Odontaspis taurus
Osmerus mordax 1,2
Paralichthys dentatus '
Pholis gunnelus 1
Pomatomus saltatrix 1
Pollachius virens "3
Prionotus carolinis 1
Pleuronectes americanus 1,2,3
Raja erinacea
Raja ocellata
Scomber scombrus
Pungitius pungitius 2
Scopthalmus aquosus 1
Syngnathidae fuscus ,23
Stenotomus chrysops 3
Stichaeus punctatus 3
Tautogolabrus adspersus 23
Tautoga onitus 1,3
Urophycis chuss 3
Urophycis tenuis
Invertebrates
Crangon septemspinosa 3
Carcinus maenas3
Cancer irroratus 3
Hippolyte zostericola 3
Homarus amenricanus 3
Libinia dubia 3
Mytilus edulis (larvae) 4
Pagarus acadianus 3
Palaemonetes vulgaris 3
Blueback herring
Alewife
American shad
Sandlance
Wolffish
American eel
Four-spine stickleback
Menhaden
Rock sea bass
Herring
Lumpfish adults; juveniles
Mummichog
Killifish
Atlantic cod
Three-spine stickleback
Ocean pout
Atlantic silverside
Atlantic tomod
Striped bass
Grubby
thorthom sculpin
Sand tiger shark
Rainbow smelt
Summer flounder
Rock gunnel
Bluefish
Pollock
Searobin
Winter flounder
Little skate
Winter skate
Atlantic mackerel
Nine-spine stickleback
Windowpane flounder
Northern pipefish
Scup
Arctic shanny
Cunner
Tautog
Red hake
White hake

Sand shrimp
Green crab
Rock crab

American lobster

Blue mussel

Grass shrimp
spring
spring

spring

spring, summer, fall

Spring; summer, fall

summer, fall
spring, summer, fall
spdng.
spring, summer, fall
summer, fall
spring, summer, fall
spring, summer, fall
spring, summer, fall
summer
occasional
occasional
occasional
occasional
spring, summer
spring
spring, summer, fall
spring, summer, fall
spring, summer, fall
spring
spring
spring, summer, fall

spring, summer, fall
spring; summer
spring, summer
common
common

rare

common

occasional; common

common
common
common
common
common
common
common
common
rare
spring
spring
summer
summer
common
rare
common
common
common
occasional
occasional
common

common
occasional; rare
occasional
1 = Colarusso pers comm.; 2 = Short 1992; 3 = Heck et al. 1989; 4 = Newell et al. 1991.

27
eelgrass beds decline or are eliminated, their capacity for stabilizing the sediment is reduced or eliminated.
The sediment may be more easily resuspended, making regrowth of eelgrass problematic. Unstable
sediments result in communities with reduced abundance and reduced diversity. Reduction of light levels
also occurs when physical structures such as docks and piers are placed over or in the vicinity of eelgrass
beds.

Presently there are no treatments that can prevent the wasting disease or relieve its deleterious effects. For
example, between 1981 and 1989, most of the eelgrass in the Piscataqua River and approximately 80% of
the eelgrass in Great Bay, New Hampshire, was lost due to the wasting disease (Short et al. 1993). In
addition, the loss of eelgrass from the wasting disease is exacerbated by reduction of light levels and
physical disturbances (Short et al. 1991). If these other stresses are reduced then eelgrass may be better
able to survive or recover from an outbreak of wasting disease.

Direct physical disturbances include the burying and uprooting of eelgrass. Indirect physical disturbances
include alteration of the current velocity, wave action, substrate or depth where eelgrass plants are growing.
Filling projects can cover eelgrass with large amounts of sediment. Dredging, fishing and boating activity
can uproot eelgrass plants and resuspend sediments. Dredging and filling projects which alter the physical
environments by changing current velocity, wave activity, substrate composition, and depth to the bottom
can prevent eelgrass from maintaining itself through vegetative and sexual reproduction.

Overharvesting is a potential threat to eelgrass. In Canada, eelgrass currently is harvested and made into a
natural, non-toxic insulation (Fred Short pers. comm.). If this new industry is profitable, harvesting of
eelgrass could easily arise along the coast of Maine.

28
REFERENCES
Ackermrnan, J.D. and A. Okubo. 1993. Reduced mixing in a marine macrophyte canopy. Funct. Ecol. 7: 305-
309.
Backman, T.W. and D.C. Barilotti. .1976. Irradiance reduction: Effects on standing crops of the eelgrass
Zostera marina in a coastal lagoon. Mar. Biol. 34: 33-40.
Bagnold, R.A. 1963. Mechanics of marine sedimentation. Pages 507-528. In: M.N. Hill (ed.). The Sea, Vol.
3. Interscience Publishers, New York.
Barshaw, D.E. and D.R. Bryant-Rich. 1988. A long-term study on the behavior and survival of early juvenile
American lobster, Homarus americanus, in three naturalistic substrates: Eelgrass, mud and rocks.
Fish. Bull. 86: 789-796.
Bell, J.D. and D.A. Pollard. 1989. Ecology of fish assemblages and fisheries associated with seagrasses,
chap 17:565-609. In: Biology of Seagrasses: A treatise on the biology of seagrasses with special
reference to the Australian region. A.W.D. Larkum, A.J. Comb and S.A. Shepherd (eds.). Elsevier,
NY.
Billen, G. 1978. A budget of nitrogen cycling in North Sea sediments off the Belgian coast. Estuarine
Coastal Mar. Sci. 7: 127-146.
Blackburn, T.H. 1979. Method for measuring rates of NH4 tumover in anoxic marine sediments using a 15H-
N44 dilution technique. Appl. Environ. Microbiol. 37: 760-765.
Brinkhuis, B.H., W.F. Penello, and A.C. Churchill. 1980. Cadmium and manganese flux in eelgrass, Zostera
marina. II. Metal uptake by leaf and root-rhizome tissues. Mar. Biol. 58: 187-196.

Brix, H. and J.E. Lyngby. 1985. Uptake and translocation of phosphorus in eelgrass (Zostera marina). Mar.
Biol. 90: 111-116.
Brix, H., J.E. Lyngby, and H.H. Schierup. 1983. Eelgrass (Zostera marina L.) as an indicator organism of
trace metals in the Limfjord, Denmark. Mar. Environ. Res. 8: 165-181.
Brown, B. 1993. A classification system of marine and estuarine habitats in Maine: an ecosystem approach
to habitats. Part h: Benthic habitats. Maine Natural Areas Program, Department of Economic and
Community Development. 51pp.
Burdick, D.M. and F.S. Short. The effects of boat docks on eelgrass beds in Massachusetts coastal waters.
Burdick, D.M., F.T. Short, and J. Wolf. 1993. An indes to assess and monitor the progression of wasting
disease in eelgrass Zostera marina. Mar. Ecol. Prog. Ser. 94: 83-90.
Burkholder, P.R. and T.E. Doheny. 1968. The biology of eelgrass with special reference to Hempstead and
South Oyster Bays, Nassau County, Long Island, New York. Lamont Geol. Observatory, Contrib.
1227. 120 pp.
Caine, E.A. 1980. Ecology of two littoral species of caprellid amphipods (Crustacea) from Wasington, USA.
Mar. Biol. 56: 327-335.
Churchill, A.C. 1983. Field studies on seed germination and seedling development in Zostera marina L.
Aquat. Bot. 16: 21-29.
Churchill, A.C. and M.l. Riner. 1978. Anthesis and seed production in Zostera marina L. from Great South
Bay, NewYork, USA. Aquat. Bot. 4: 83-93.

29
Conover, J.T. 1964. Environmental relationships of benthos in salt ponds (plant relationships). R.l. Grad.
Sch. Oceanog. Tech. Rep. 3. 2 vols.
Conover, J.T. 1968. The importance of natural diffusion gradients and transport of substances related to
benthic plant metabolism. Bot. Mar. 11: 1-9.
Cottam, C. 1934. Past periods of eelgrass scarcity. Rhodora 36: 261-264.
Cottam, C. and D.A. Monroe. 1954. Eelgrass status and environmental relations. J. Wildl. Manage. 18: 449-
460.
Cottam, C., J.F. Lynch, and A.L. Nelson. 1944. Food habits and management of American sea brant. J.
Wildl. Manage. 82: 36-56.
Cousins, S.H. 1980. A trophic continuum derived from plant structure, animal size, and a detritus cascade.
J. Theor. Biol. 82: 607-618.
DeCock, A.W.A.M. 1980. Flowering, pollination and fruiting in Zostera marina L. Aquat. Bot. 9: 201-220.
den Hartog, C. 1970. The seagrasses of the world. North-Holland Publishing Co., Amsterdam. 275 pp.
den Hartog, C. 1977. Structure, function and classification in seagrass communities. Pages 89-122 In: C.P.
McRoy and C. Helfferich (eds.). Seagrass ecosystems, a scientific perspective, Marcel Dekker, New
York.
Dennison, W.C. 1987. Effects of light on seagrass photosynthesis, growth and depth distribution. Aquat. Bot.
27: 15-26.
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., G.J. Marshall and C. Wigand. 1989. Effect of "brown tide" shading on eelgrass (Zostera
marina L. ) distributions. Lecture Notes Marine Estuarine Studies.
Dexter, R.W. 1985. Changes in the standing crop of eelgrass, Zostera marina L., at Cape Ann,
Massachusetts, since the epidemic of 1932. Rhodora 87: 357-366.
Driftmeyer, J.E. 1980. Uptake of 65Zn by eelgrass, Zostera marina L. Sci. Total Environ. 16: 263-266.
Driftmeyer, J.E., G.W. Thayer, F.A. Cross, and J.C. Zieman. 1980. Cycling of Mn, Fe, Cu, and Zn by
eelgrass, Zostera marina L. Am. J. Bot. 67: 1089-1096.
Ewald, J.J. 1969. Observations on the biology of Tozeuma carolinense (Decapoda, Hippolylidae) from
Florida, with special reference to larval development. Bull. Mar. Sci. 19: 510-549.
Faraday, W.E. and A.C. Churchill. 1979. Uptake of cadmium by the eelgrass Zosfera marina. Mar. Biol. 53:
293-298.
Fonseca, M.S. and J.A. Cahalan. 1992. A preliminary evaluation of wave attenuation by four species of
seagrass. Estuar. Coast. Shelf Sci. 35: 565-576.
Fonseca, M.S., W.J. Kenworthy, G.W. Thayer, and D.Y. Heller. 1982a. Transplanting of the seagrasses
Zostera marina and Halodule wrightii for the stabilization of subtidal dredge material. Annual Report of
the National Marine Fisheries Service, Southeast Fisheries Center, Beaufort Laboratory, Division of

30
Estuarine and Coastal Ecology, Beaufort, N.C. to the US Army corps of Engineers, Coastal
Engineering Research Center, Fort Belvoir, Va. 61 pp.

Fonseca, M.S., J.S. Fisher, J.C. Zieman, and G.W. Thayer. 1982b. Influence of the seagrass, Zostera
manrina L., on current flow. Estuaine Coastal Shelf Sci. 15: 351-364.

Fonrseca, 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 Shelf Sci. 17: 367-380

Fonseca, M.S. and W.J. Kenworthy 1987. Effects of current on photosynthesis and distribution of
seagrasses. Aquat. Bot. 27: 59-78.

Gagnon, P.S., R.L. Vadas, D.B. Burdick, and B.P. May. 1980. Genetic identity of annual and perennial
forms of Zostera marina L. Aquat. Bot. 8: 157-162.

Gambi, M.C., A.R.M. Nowell, and P.A. Jumars. 1990. Flume observations on flow dynamics in Zostera
marina (eelgrass) beds. Mar. Ecol. Prog. Ser. 61: 159-169.

Gutsell, J.S. 1930. Natural history of the bay scallop. Bull. US Bur. Fish. 45: 569-632.

Harlin, M.M. 1980. Seagrass epiphytes. Pages 117-151. In: R.C. Phillips and C.P. McRoy (eds.). Handbook
of seagrass biology, an ecosystem perspective. Garland STPM Press, NewYork

Heck, K.L. and T.A. Thoman. 1984. The nursery role of seagrass meadows in the upper and lower reaches
of the Chesapeake Bay. Estuaries 7: 70-92.

Heck, K.L.,Jr., K.W. Able, M.P.Fahay and C.T. Roman. 1989. Fishes and decapod crustaceans of Cape
Cod eelgrass meadows: Species composition, seasonal abundance patterns and comparison with
unvegetated substrates. Estuaries 12: 59-65.

Hidu, H., B. McAlice, L. Kindlom, C. Newell. and G. Podniesinski. 1988. Recruitment and commercial seed
procurement of the blue mussel Mytilus edulis L. in Maine. J. Shellfish. Res. 7: 568-569.

Hoven, H., R. Grizzle, F. Short, and L. Kindblom. 1991. Eelgrass as a larval trap: The "Honami" effect. J.
Shellfish Res. 10: 275-276.

Howard, R.K. 1982. Impact of feeding activities of epibenthic amphipods on surface-fouling of eelgrass
leaves. Aquat Bot. 14: 91-97.

lizumi, H., A. Hattori, and C.P. McRoy. 1982. Ammonium regeneration and assimilation in eelgrass (Zostera
marina) beds. Mar. Biol. 66: 59-65.

lizumi, H. and A. Hattori. 1982. Growth and organic production of eelgrass (Zostera marina L.) in temperate
waters of the Pacific coast of Japan. IlIl. The kinetics of nitrogen uptake. Aquat. Bot. 12: 245-256.
Kemp, W.M., W.R. Boynton, R.R. Twilley, J.C. Stevenson, and J.C. Means. 1983. The decline of
submerged vascular plants in Upper Chesapeake Bay: Summary of results concerning possible
causes. Mar. Soc. Tech. J. 17: 78-89.

Kenworthy, W.J., M.S. Fonseca, J. Homziak, and G.W. Thayer. 1980. Development of a transplanted
seagrass (Zostera marina L.) meadow in Back Sound, Carteret County, North Carolina. Pages 175-
193. In: D.P. Cole (ed.). Proceedings of the Seventh Annual Conference on the Restoration and
Creation of Wetlands. Hillsborough Community College, Tampa. Fla.

31
Kenworthy, W.J. and G.W. Thayer. 1984. Production and decomposition of the roots and rhizomes of
seagrasses, Zostera marina and Thalassia testudinum in termperate and subtropical marine
ecosystems. Bull. Mar. Sci. 35: 364-379.
Kenworthy, W.J., J.C. Zieman, and G.W. Thayer. 1982. Evidence for the influence of seagrass on the
benthic nitrogen cycle in a coastal plain estuary near Beaufort, North Carolina (USA). Oecologia 54:
152-158.
Kikuchi, T. 1966. An ecological study on animal communities of the Zostera marina belt in Tomioka Bay,
Amakusa, Kyushu. Publ. Amakusa Mar. Biol. Lab. Kyushu Univ. 1: 1-106.
Kikuchi, T. 1980. Faunal relationships in temperate seagrass beds. Pages 153-172. In: R.C. Phillips and
C.P. McRoy (eds.). Handbook of seagrass biology, an ecosystem perspective. Garland STPM Press,
NewYork
Kikuchi, T. and J.M. Peres. 1977. Consumer ecology of seagrass beds. Pages 147-193. In: C.P. McRoy and
C. Helfferich (eds.). Seagrass ecosystems, a scientific perspective, Marcel Dekker, NewYork.

Kirby-Smith, W.W. 1972. Grovwth of the bay scallop, the influence of experimental water current. J. Exp.
Mar. Biol. Ecol. 8: 7-18.
Kirby-Smith, W.W. and R.T. Barber. 1974. Suspension-feeding aquaculture systems: effects of
phytoplankton concentration and temperature on the growth of the bay scallop. Aquaculture 3: 135-
145.
Kirchman, D.L, L. Mazella. R.S. Alberte, and R. Mitchell. 1984. Epiphytic bacterial production on Zostera
marina. Mar. Ecol. Prog. Ser. 15: 117-123.
Knutson, P.L., W.N. Seelig, and M.R. Ilnskeep. 1982. Wave dampening in Spartina altemiflora marshes.
Wetlands 2: 87-104.
Kurata, A., Y. Yoshida, and F. Taguchi. 1979. Accumulation of Ni from the environmental sea water and
sediments by various marine organisms. Umi-Mer 17:11-17.
Kurland, J. M. 1991. Habitat mitigation efforts in the Gulf of Maine: Stemming the tide of environmental
degradation. Habitat and Protected Resources Division, Northeast region, National Marine Fisheries
Service, NOAA. 67 pp. Prepared for the Gulf of Maine Council on the Marine Environment.

Lamounette, R. 1977. A study of the germination and viability of Zostera marina L seed. M.S. Thesis.
Adelphi Univ., Garden City, N.Y. 41 pp.

Lough, R.G., P.C. Valentine, D.C. Potter, P.J. Auditore, G.R. Boltz, J.D. Neilson, and R.I. Perry. 1989.
Ecology and distribution of juvenile cod and haddock in relation to sediment type and bottom currents
on eastem Georges Bank. Mar. Ecol. Prog. Ser. 56: 1-12.
Lyngby, J.E. and H. Brix. 1982. Seasonal and environmental variation in cadmium, copper, lead and zinc
concentrations in eelgrass (Zostera marina L) in the Limfjord, Denmark. Aquat. Bot. 14: 59-74.

Lyngby, J.E., H. Brix, and H.H. Schierup. 1982. Adsorption and translocation of zinc in eelgrass (Zostera
marina L.). J. Exp. Mar. Biol. Ecol. 58: 259-270.
Marshall, N. and K. Lukas. 1970. Preliminary observations on the properties of bottom sediments with and
without eelgrass, Zostera marina cover. Proc. Natl. Shellfish Assoc. 60: 107-112.
McRoy, C.P. and R.J. Barsdate. 1970. Phosphate adsorption in eelgrass. Limnol. Oceanogr. 15: 6-13.

32


McRoy, C.P. and J.J. Goering. 1974. Nutrient transfer between the seagrass Zostera marina and its
epiphytes. Nature 248: 173-174.
McRoy, C.P. and C. Helfferich. 1980. Applied aspects of seagrasses. Pages 297-343. In: R.C. Phillips and
C.P. McRoy (eds.). Handbook of seagrass biology, an ecosystem perspective. Garland STPM Press,
New York.

McRoy, C.P., R.J. Barsdate, and M. Nebert. 1972. Phosphorus cycling in an eelgrass (Zostera marina L.)
ecosystem. Limnol. Oceanogr. 17: 58-67.

Muehlstein, L.K. 1989. Perspectives of the wasting disease of eelgrass Zostera marina. Dis. Aquat. Org. 7:
211-221.

Muehistein, L.K., D. Porter, and F.T. Short. 1988. Labyrinthula sp., a marine slime mold producing the
symptoms of wasting disease in eelgrass Zostera marina. Mar. Biol. 99: 465-472.

Muehistein, L.K., D. Porter, and F.T. Short. 1991. Labyrinthula zosterae sp. nov., the causative agent of
wasting disease of eelgrass, Zostera marina. Mycologia 83: 180-191.

Newell, C.R., H. Hidu, B.J. McAlice, G. Podniesinski, F. Short, and L. Kindblom. 1991. Recruitment and
commercial seed procurement of the blue mussel Mytilus edulis in Maine. J. World Aquacult. Soc. 22:
134-152.

Orth, R.J. 1973. Benthic infauna of eelgrass, Zostera marina beds. Chesapeake Sci. 14: 258-269.

Orth, R.J. 1977. The importance of sediment stability in seagrass communities. Pages 281-300 In: B.C.
Coull (ed.). Ecology of marine benthos. University of South Carolina Press, Columbia.

Orth, R.J. and K.A. Heck, Jr. 1980. Structural components of eelgrass (Zostera manha) meadows in the
lower Chesapeake Bay - Fishes. Estuaries 3: 278-288.

Orth, R.J. and K.A. Moore. 1981. Submerged aquatic vegetation of the Chesapeake Bay: past, present and
future. Trans. N. Am. Wildl. Nat. Resour. Conf. 46: 271-283.

Orth, R.J. and K.A. Moore. 1982. Distribution and abundance of submerged aquatic vegetation in
Chesapeake Bay: a scientific summary. Pages 381-427. In: E.G. Macalaster, D.A. Barber, and M.
Kasper (eds.). Chesapeake Bay Program technical studies: a synthesis. US Environmental Protection
Agency, Washington, DC.

Orth, R.J. and K.A. Moore. 1983a. Seed germination and seedling growth of Zostera marina L (eelgrass) in
the Chesapeake Bay. Aquat. Bot. 15: 117-131.

Orth, R.J. and K.A. Moore. 1983b. Chesapeake Bay: An unprecedented decline in submerged aquatic
vegetation. Science 22: 51-52.

Ostenfeld, C.H. 1908. On the ecology and distribution of grass wrack (Zostera marina) in Danish waters.
Rep. Dan. Biol. Stn. 16: 1-62.

Penello, W.F. and B.H. Brinkhuis. 1980. Cadmium and manganese flux in eelgrass Zostera marina. 1.
Modelling dynamics of metal release from labelled tissues. Mar. Biol. 58: 181-186.

Penhale, P.A. 1977. Macrophyte-epiphyte biomass and productivity in an eelgrass (Zostera marina L.)
community. J. Exp. Mar. Biol. Ecol. 26: 211-224.

33
Penhale, P.A. and G.W. Thayer. 1980. Uptake and transfer of carbon and phosphorus by eelgrass (Zostera
manrina L.) and its epiphytes. J. Exp. Mar. Biol. Ecol. 42: 113-123.
Petersen, C.G.J. 1918. The sea bottom and its production of fish food: a summary of the work done in
connection with valuation of Danish waters from 1883 to 1917. Rep. Dan. Biol. Stn. 25: 1-82.
Phillips, R.C. 1972. Ecological life history of Zostera marina L. (eelgrass) in Puget Sound, Washington.
Ph.D. Diss. Univ. of Washington, Seattle. 154 pp.
Phillips, R.C., W.S. Grant and C.P. McRoy. 1983a. Reproductive strategies of eelgrass (Zostera marina L.).
Aquat. Bot. 16: 1-20.
Phillips, R.C., C. McMilian, and K.W. Bridges. 1983b. Phenology of eelgrass, Zostera marina L., along
latitudinal gradients in North America. Aquat. Bot. 15: 145-156.
Phillips, R.C. 1984. The ecology of eelgrass meadows in the Pacific Northwest: a community profile. US
Fish & Wildl. Serv. FWSIOBS-84/24 85pp.
Rasmussen, E. 1973. Systematics and ecology of'the Isef Fjord marine fauna (Denmark) with a survey of
the eelgrass (Zostera) vegetation and its communities. Ophelia 11(2-3): 1-507.
Rasmussen, E. 1977. The wasting disease of eelgrass (Zostera marina) and its effects on environmental
factors and fauna. Pages 1-52. In: C.P. McCoy and C. Helfferich (eds.). Seagrass Ecosystems, a
scientific perspective. Marcel Dekker, New York.

Riggs, S.A. and R.A. Fralick. 1975. Zostera marina L., its growth and distribution in the Great Bay Estuary,
New Hampshire. Rhodora 77: 456-466.
Robertson, A.l. and K.H. Mann. 1980. The role of isopods and amphipods in the initial fragmentation of
eelgrass detritus in Nova Scotia, Canada. Mar. Biol. 59: 63-69.
Robertson, A.l. and K.H. Mann. 1982. Population dynamics and life history adaptations of Littorina neglecta
Bean in an eelgrass meadow (Zostera marina L.) in Nova Scotia. J. Exp. Mar. Biol. Ecol. 63: 151-171.
Robertson, A.l. and K.H. Mann. 1984. Disturbance by ice and life-history adaptations of the seagrass
Zostera marina. Mar. Biol. 80: 131-141.
Sand-Jensen, K 1977. Effect of epiphytes on eelgrass photosynthesis. Aquat. Bot. 3: 55-63.
Sculthorpe, C.D. 1967. The biology of aquatic vascular plants. Amold, London, 618 pp.
Setchell, W.A. 1929. Morphological and phenological notes on Zostera marina L. Univ. Calif. Publ. Bot. 14:
389-452.
Short, F.T. 1975. Eelgrass production in Chariestown Pond: an ecological analysis and simulation model.
M.S. Thesis. University of Rhode Island, Kingston, RI. 180 pp.
Short, F.T. 1981. Nitrogen resource analysis and modelling of an eelgrass (Zostera marina L.) meadow in
Izembek Lagoon, Alaska. Ph.D. Diss. Univ. of Alaska, Fairbanks. 173 pp.
Short, F.T. 1983a. The response of interstitial ammonium in eelgrass (Zostera marina L.) beds to
environmental perturbations. J. Exp. Mar. Biol. Ecol. 68: 195-208.
Short, F.T. 1983b. The seagrass, Zostera marina L.: plant morphology and bed structure in relation to
sediment ammonium in Izembek Lagoon, Alaska. Aquat. Bot. 16: 149-161.

34
Short, F.T. 1991. (ed.). An Estuarine Profile of Great Bay, New Hampshire. Report to NOAA National
Estuarine Research Reserve Program and New Hampshire Fish and Game Department. 223pp.

Short, F.T. (ed.). 1992. The Ecology of the Great Bay Estuary, New Hampshire and Maine: An estuarine
profile and bibliography. NOAA - Coastal Ocean Program Publ. 222 pp.

Short, F.T. and A.C. Mathieson. 1985. Wasting disease: Eelgrass decline in the Great Bay Estuary.
Estuaries 8: 49A.

Short, F.T. and C.P. McRoy. 1984. Nitrogen uptake by leaves and roots of the eelgrass Zostera marina L.
Bot. Mar. 27: 547-555.

Short, F.T., B.W. Ibelings and C. den Hartog. 1988. Comparison of a current eelgrass disease to the wasting
disease in the 1930s. Aquat. Bot. 30: 295-304.

Short, F.T, A.C. Mathieson, and J.J. Nelson. 1986. Recurrence of the eelgrass wasting disease at the
border of New Hampshire and Maine, USA. Mar. Ecol. Prog. Ser. 29: 89-92.
Short, F.T., L.K. Muehlstein and D. Porter. 1987. Eelgrass wasting disease: cause and recurrence of a
marine epidemic. Biol. Bull. 173: 557-562.
Short, F.T., D.M. Burdick, J. Wolf, and G.E. Jones. 1993. Eelgrass in estuarine research reserves along the
East Coast, USA. Part I: Declines from pollution and disease and Part II: Mangement of eelgrass
meadows. NOAA-Coastal Ocean Program publication. 107pp.

Silberhom, G.M., R.J. Orth, and K.A. Moore. 1983. Anthesis and seed production in Zostera marina L.
(eelgrass) from the Chesapeake Bay. Aquat. Bot. 15: 133-144.

Sogard, S.M. and K.W. Able. 1991. A comparison of eelgrass, sea lettuce macroalgae, and marsh creeks as
habitats for epibenthic fishes and decapods. Estuar. Coast. Shelf. Sci. 33: 501-519.

Stauffer, R.C. 1937. Changes in the invertebrate community of a lagoon after disappearance of the
eelgrass. Ecology 18: 427-431.

Stoner, A.W. 1983. Distribution of fishes in seagrass meadows: role of macrophyte biomass and species
composition. Fish. Bull. US 81: 837-846.

Summerson, H.C. and C.H. Peterson. 1984. The role of predation in organizing benthic communtities in and
around a temperate zone seagrass bed. Mar. Ecol. Prog. Ser. 15: 63-77.
Thayer, G.W. and A.J. Chester. 1989. Distribution and abundance of fishes among basin and channel
habitats in Florida Bay. Bull. Mar. Sci. 44: 200-219.
Thayer, G.W. and H.H. Stuart. 1974. The bay scallop makes its bed of eelgrass. US Natl. Mar. Fish. Serv.
Mar. Fish. Rev. 36(7): 27-39.
Thayer, G.W., S.M. Adams, and M.W. La Croix. 1975. Structural and functional aspects of a recently
established Zostera marina community. Pages 517-540. In: L.E. Cronin (ed.). Estuarine research, Vol.
1. Academic Press, NewYork.

Thayer, G.W., W.J. Kenmorthy, and M.S. Fonseca. 1984. The ecology of eelgrass meadows of the Atlantic
coast: a community profile. US Fish & Wilfl. Serv. FWS/OBS-84/24. 85 pp.

Thayer, G.W., P.L. Parker, M.W. LaCroix, and B. Fry. 1978. The stable carbon isotope ratio of some
components of an eelgrass, Zostera marina, bed. Oecologia 35: 1-12.

35
Thayer, G.W., H.H. Stuart, W.J. Kenworthy, J.F. Ustach, and A.B. Hall. 1979. Habitat values of salt
marshes, mangroves, and seagrasses for aquatic organisms. pages 235-247 In: P.E. Greeson, J.R.
Clark, and J.E. Clark (eds.). Wetland functions and values: the state of our understanding. American
Water Resources Association, Minneapolis.
Thursby, G.B. and M.M. Harlin. 1982. Leaf-root interaction in the uptake of ammonia by Zostera marina.
Mar. Biol. 72:109-112.
Tomlinson, P.B. 1980. Leaf morphology and anatomy in seagrasses. Pages 7-28 In: R.C. Phillips and C.P.
McRoy (eds.). Handbook of seagrass biology, an ecosystem perspective. Garland STPM Press, New
York.

Tutin, T.G. 1942. Zostera. J. Ecol. 30: 217-226.
Twilley, R.R., W.M. Kemp, K.W. Staver, J.C. Stevenson and W.R. Boynton. 1985. Nutrient enrichment of
estuarine submerged vascular plant communities. I. Algal growth and effects on production of plants
and associated communities. Mar. Ecol. Prog. Ser. 23: 179-191.

Van Montfrans, J., R.J. Orth, and S.A. Vay. 1982. Preliminary studies of grazing by Bittium varium on
eelgrass periphyton. Aquat. Bot. 14: 75-89.

Van Montfrans, J., R.L. Wetzel, and R.J. Orth. 1984. Epiphytic-grazer relationships in seagrass meadows:
Consequences for seagrass growth and production. Estuaries 7: 289-309.

Wahle, R.A. and R.S. Steneck 1991. Recruitment habitats and nursery grounds of the American lobster
(Homarus americanus Milne Edwards): a demographic bottneck? Mar. Ecol. Prog. Ser. 69: 231-243.

Wahle, R.A. and R.S. Steneck. 1992. Habitat restrictions in early benthic life: experiments on habitat
selection and in situ predation with the American lobster. J. Exp. Mar. Biol. ecol. 157: 91-114.

Wayne, C.J. 1975. Sea and marsh grasses: their effect on wave energy and nearshore transport. M.S.
Thesis. Florida State University, Tallahasee. 135 pp.
Weinstein, M.P. and H.A. Brooks. 1983. Comparative ecology of nekton residing in a tidal creek and
adjacent seagrass meadow community composition andstructure. Mar. Ecol. Prog. Ser. 12: 15-27.
Wolfe, D.A., G.W. Thayer, and S.M. Adams. 1976. Manganese, iron, copper and zinc in an eelgrass
(Zostera marina) community. Pages 256-270. In: C.E. Cushing, Jr (ed.). Radioecology and energy
resources. Dowden, Hutchinson and Ross, Stroudsburg, Pa.

Zieman, J.C. 1982. The ecology of the seagrasses of South Florida: a community profile. US Fish Wildl.
Serv. Biol. Serv. Program FWSIOBS-82/25. 124. 26 pp.

36
ROCKWEEDS

IDENTIFICATION AND DESCRIPTION
Five species of brown fucoid algae, Ascophyllum nodosum (and its unattached scorpiodes form), Fucus
vesiculosus, F. evanescens (formerly F. distichus subsp evanescens and subsp edentatus), F. distichus, and
F. spiralis are found along the Maine coast. In terms of biomass and percent cover of substrate, A.
nodosum, F. vesiculosus, and F. evanescens are the most important species (Vadas, pers. comm.). F.
distichus is described as a small tidepool species, however, a non-tidepool member of the complex has
been reported in southern Maine (Sideman and Mathieson 1983) and Nova Scotia (C. Bird pers. comm). F.
spiralis has a limited distribution in the high intertidal zone.
Algae are relatively simple in structure. They lack well-developed tissue systems, and have no true roots,
stems or leaves. The five species of fucoid algae described here, commonly referred to as rockweeds,
range in color from dark brown to olive green due to the presence of fucoxanthin, a brown xanthophyll
accessory pigment which tends to obscure the greenish chlorophyll (Luning 1990). Rockweeds attach to the
substrate by an expanded holdfast. The thalli or fronds of the algae have dichotomous or pinnate strap-
shaped branches, usually with buoyant air bladders (Taylor 1957). The branches of the four Fucus species
have a thickened midrib, whereas the branches of A. nodosumrn do not (Taylor 1957).

DISTRIBUTION
The five species of rockweeds occur primarily in the Cold Temperate North Atlantic marine biogeographic
region, limited by the 10ï¿½C summer/0OC winter isotherms in the north and by the 15oC summer/10ï¿½C winter
isotherms in the south, and to a lesser extent in the Arctic region (Lining 1990). A. nodosum and F.
vesicufosus are found on both sides of the Atlantic Ocean from Greenland to Virginia and from the Barents
Sea to Portugal, i.e., from approximately 71ï¿½N to 38ï¿½N (Luning 1990). Scattered populations of F.
vesiculosus occur father north along Baffin Island and south on the northwest coast of Africa (Luning 1990).
A small, introduced population of A. nodosum also exists in San Francisco Bay (P. Silva pers. comm.). F.
spiralis occurs from Newfoundland to Delaware (50ï¿½N to 38ï¿½N), from Finland to northwest Africa (70ï¿½N to
20ï¿½N), and sporadically along the coast of Washington, British Columbia, and Alaska (Lulning 1990). F.
evanescens and F. distichus range from Greenland and Iceland (80ï¿½N) to Delaware (38ï¿½N) and the English
Channel (50ï¿½N) (Luning 1990). Members of the F. distichus Powell complex also occur on the Pacific coast
(Luning 1990).
In the Gulf of Maine, the rockweeds typically occur in the intertidal zone at a particular height, and these
striking pattems of vertical zonation have received a great deal of attention (Norton 1986). F. spiralis is
found in the upper intertidal zone, F. vesiculosus and A. nodosum in the mid-intertidal zone, F. evanescens
in the low-intertidal zone, and F. distichus in tidepools in the intertidal zone (Taylor 1957; Mathieson and
Guo 1992). Desiccation and overheating (see Ecology) and biological interactions have been identified as
critical factors controlling the zonation of intertidal fucoid algae. The experimental removal of algal
competitors and animal grazers allowed some species of rockweeds to expand upshore and downshore of
their normal range (Burrows and Lodge 1951; Menge 1975; Schoenbeck and Norton 1980; Hawkins and
Hartnoll 1985).
The rockweeds can grow in the subtidal zone (e.g., Lubchenco 1980; Serrao et al. 1996), although some
appear to be better suited to the intertidal zone. In Rhode Island, plant density was greater in an intertidal
population of A. nodosum (91 plants/m2) than a subtidal population (50 plants/m2), and the intertidal plants
had a higher capacity for photosynthesis and nitrate uptake (Peckol et al. 1988).
In many areas, the distribution and abundance of the species of rockweeds varies with exposure to waves
and currents. Menge (1976) examined the effect of exposure on the rocky intertidal community at six sites
in Maine and Massachusetts. He reported that at the most exposed site (Pemaquid Point), F. evanescens

37
.covered less than 10% of the substrate; at the third most exposed site (Chamberlain), F. evanescens
covered 50 to 90% of the substrate; at the fifth most exposed site, F. vesiculosus covered 50 to 90% of the
substrate; at the most protected site (in Massachusetts), A. nodosum covered 50 to 90% of the substrate.
However, in the Bay of Fundy, A. nodosum dominated the intertidal zone with high biomass under all
conditions of exposure, and F. vesiculosus was common, but its distribution was unrelated to exposure
(Thomas 1994).
A. nodosum is the most abundant intertidal alga in the Gulf of Maine, and F. vesiculosus ranks second in
abundance (Keser et al. 1981). On moderately exposed and sheltered shores in the Damariscotta-
Pemaquid region of Maine, A. nodosum accounts for 50 to 90% of the substrate cover with average
densities ranging from 390 to 521 plants/mr2 (Keser et al. 1981). Estimates of A. nodosum biomass in this
region were 3 to 6 kg dry weight/m2  depending on the exposure (Vadas 1972; Vadas et al. 1976). In Cedar
Point, New Hampshire, A. nodosum (plus its unattached form) and F. vesiculosus accounted for an average
of 98.3% of the intertidal algal biomass each month of a 15-month study, and maximum summer biomass
was 1.36 to 1.68 kg dry weight/m2 (Chock and Mathieson 1983).


LIFE CYCLE

The generalized life cycle of the five species of fucoid algae is rather simple. The gametes are released into
the water column where fertilization occurs. The resulting embryos quickly sink to the substratum, and
attach within several hours after fertilization as adhesive cell wall materials accumulate (Evans et al. 1982).
Embryos of summer and winter reproducing species may take four hours to two days to attach, respectively,
because the accumulation of adhesive material is temperature dependent (Bird and McLachlan 1974). The
microscopic embryos begin to grow, becoming visible at a length of approximately one millimeter. Growth
involves an increase in thallus length and the number of branches. Typically, each A. nodosum shoot forms
an air bladder or vesicle each year, a growth pattem that has been used to estimate age and productivity
(Cousens 1984). During the year, portions of the thallus may be tom off as a result of wave action or ice
scouring, and vegetative propagation may occur from the holdfast. Eventually the adult plant develops
reproductive structures vwere gametes are produced.
The time of reproduction varies with each species. A. nodosum reproduces in late spring and early summer,
F. vesiculosus principally in autumn but also in early winter and spring to early summer in Maine, F.
evanescens in autumn and spring, F. spiralis in summer and F. distichus principally in winter (Mathieson and
Guo 1992; Brawley and Johnson 1992). The release of gametes in A. nodosum is correlated with specific
temperatures (Bacon and Vadas 1991), and has a lunar or possibly a semilunar periodicity in estuaries
(Lambert and Brawley 1993). Fucoid algae do not release gametes when hydrodynamic conditions are
turbulent, even when other factors are favorable (Serr5o et al. 1996a)

The tiny zygotes are passively dispersed and probably come into contact with substrate as a result of
downward turbulence (Norton 1986). Survival "depends on settlement and attachment in an appropriate
place because the zygotes are not equipped for dormancy like seeds," they are probably not able to actively
select an appropriate substrate, and they cannot move if conditions later become unfavorable (Norton
1986). Most zygotes settle within a few meters of the parent plant, but Burrows and Lodge (1950) reported
Fucus plants arising more than 60 m away from the nearest group of parent plants.
Mortality is high for the microscopic, early post-settlement stages (zygotes and embryos). Vadas et al.
(1992) reviewed 23 factors and/or processes that have been reported to affect mortality in these stages.
These factors and processes include grazing by herbivores, canopy effects (e.g. cover, sweeping and
reduced light), turf effects (barrier to settlement), desiccation, water motion, competition by faster growing
ephemeral algae, sediment/silt, and scour. Research has established that 90 to 100% of F. vesiculosus
zygotes die within 17 days of settlement (Bray and Norton unpublished data) and that the mean life
expectancy of newly settled A. nodosum zygotes is 10 days (Miller and Vadas 1984).

38
For some species (e.g., F. gardneri on the west coast of Canada), new recruits to the population arise only
from zygotes (Ang 1991), but for others, populations may be maintained by vegetative propagation from the
holdfast. Vadas and Wright (1986) observed no sexual recruitment in over 20 years in a population of A.
nodosum on exposed shores in Maine; all the new shoots arose from vegetative sprouting. This lack of
recruitment of A. nodosum in exposed areas is probably due to its slow growth and inability to settle when
exposed to energetic wave action and high flow rates. Vegetative regrowth and canopy regeneration may
be reduced if the holdfast is damaged. McCook and Chapman (1992) found only 20 to 30% of
experimentally damaged holdfasts of F. vesiculosus and F. evanescens sprouted large numbers of shoots.

Growth rates vary with species, developmental stage, location, and season. Young A. nodosum growvery
slowly compared to the four species of Fucus, and may reach only I to 2 mm at one year of age (Chapman
and Johnson 1990). Growth rates of adult A. nodosum plants in Canada range from 5 to 20 cm/year, and
are temperature dependent (CAFSAC 1992). The growth of species of Fucus was highly variable within and
among sites in Connecticut (Keser and Larsen 1984), but the general pattern consisted of slowto moderate
growth during winter and early spring and rapid growth through summer and autumn (Brinkhuis 1977;
Brnkhuis and Jones 1976; Chock and Mathieson 1976; Mathieson et al. 1976). In Norway, the apical
growth of F. evanescens (F. distichus spp edentatus) was greatest in July, but growth in March at 4 to 4.2ï¿½C
was 45% of growth in July (Str6mgren 1985).

The five species of fucoid algae are perennial plants with very different life spans. Keser et al. (1981) and
Vadas (unpublished data) counted the number of vesicles (air bladders) on the longest unbroken shoot of
plants in mid-coast Maine and estimated that shoots of A. nodosum plants may reach 16 to 18 years of age.
The holdfasts may be a great deal older (Vadas pers. comm.). In Monsweag Bay and at Pemaquid Point,
98 to 99% of the A. nodosum shoots were 0 to 6 years old, while in the Damarscotta estuary 98% of the
shoots were 0 to 13 years old (Keser et al. 1981). In contrast, the life span of F. vesiculosus in the region
was 2 to 4 years (Keser et al. 1981).
There are a number of productivity estimates for adult A. nodosum. Cousens (1984) estimated the net
productivity in Canada to range from 0.61 to 1.9 kg. dry weight/m2n/year. Chock and Mathieson (1979)
estimated productivity in New England at 1.5 kg/m /year by measuring standing crop on two dates. Keser
(1978) estimated the productivity of harvested areas in Maine to be 1.86 kg/m2.

POPULATION TRENDS
No overall trends have been reported for the rockweeds, but there are some regional trends in Europe. A.
nodosum and F. vesiculosus disappeared from the inner part of the Oslofjord (Bokn and Lein 1978), but
recently populations of these species have reappeared (Bokn et al. 1992). The depth penetration of F.
vesiculosus decreased in the Baltic Sea from 11.5 m in 1943-44 to 8.5 m in 1984 at 10 of 11 stations
(Kautsky et al. 1986; Kautsky 1991). Near Kiel, Germany, Fucus currently occurs no deeper than two
meters, and there has been a 95% decline in biomass since the 1950s (Vogt and Schramm 1991). In the
Archipelago Sea in Finland, there has been a decline in biomass of F. vesiculosus (Roennberg et al. 1985;
Kangas et al. 1982; Haahtela 1984). In each case, the changes have be attributed, in part, to increased
nutrient levels and eutrophication, which have changed the competitive balance between fucoid algae,
ephemeral green algae, and grazers.


ECOLOGICAL REQUIREMENTS

Light

Sunlight is a fundamental requirement for the five species of rockweeds because light energizes the
reactions of photosynthesis. However, the amount and spectral characteristics of available light varies with
latititude, season, water type, and depth (reviewed in LOning 1990). As light passes through water and the
substances dissolved and suspended in it, different wavelengths of light are selectively absorbed, causing

39
the total amount of available light and the range of available wavelengths to decrease with depth. Dring
(1987) developed a model of the light environment in the intertidal zone to overcome the difficulties of
making intertidal measurements. The model incorporated surface integrated light values, constantly
changing water levels, water clarity, and the interactions of tidal and solar cycles.
The relationship between photosynthesis and light intensity (irradiance) is described by a parabolic curve
(Lijning 1990; Chapman 1995b). The compensation irradiance (Ic) is the minimum amount of light needed
for the plant to survive, i.e., at this light level the rate of photosynthesis equals the rate of respiration. As
light intensity increases above the compensation irradiance, the photosynthetic rate first increases in a linear
fashion, then plateaus as photsynthetic enzymes become light-saturated, and finally may decrease due to
photoinhibition (LCining 1990; Chapman 1995b). For a particular plant, the photosynthesis-versus-light
relationship is the result of genetic adaptation and acclimation (Ramus 1981). Saturation irradiances for
growth were estimated to be 90 to 100 W/m2 for A. nodosum and 60 to 70 WI/2 for F. vesiculosus and F.
spiralis from the west coast of Sweden in a natural daylight regime in September (Str6mgren and Nielsen
1986). Str6mgren and Nielsen (1986) also reported that maximum growth rate occurred early in the
morning and declined throughout the remainder of the day. However, at this site, tides are almost
nonexistent, and growth patterns wmuld differ in the intertidal zone.
Insufficient irradiance in turbid water has been posited as controlling the lower limits of intertidal fucoid
algae (Ding 1984). Ramus et al. (1977) found decreased growth in excised pieces of A. nodosum and F.
vesiculosus grown at various depths in the turbid waters of Long Island Sound when compared with pieces
grown at the surface, despite an increase in light-harvesting pigments at depth. In addition, the maximum
photosynthetic capacity of F. vesiculosus was about twice that of A. nodosum (Ramus et al. 1977).

Fucoid algae may be more sensitive to light quality than to light quantitity. They can survive in the dark for
long periods, and they are protected against damage at high light levels (summarized in Chapman 1995b).
Young fucoid plants die after exposure to yellow-green or orange light at photon fluence rates of
100pzmol/m/s for 4 to 6 weeks (McLachlan and Bidwell 1983). These are the colors of light that penetrate
best in turbid coastal waters. Under normal conditions, however, young fucoid plants have plenty of full
spectrum light at and near low tide, even in turbid waters.
Light also may be used as an environmental signal to synchronize growth and reproduction to the seasonal
cycle. For A. nodosum the onset of reproduction is under the control of short daylengths (Terry and Moss
1980).
Water
The rockweeds that occur in the intertidal zone can grow and thrive under conditions of continual
submergence in water, however, they are physiologically adapted to periodic emersion in air during low tide
and the subsequent stress of desiccation (Chapman 1995b). Schonbeck and Norton (1979a) showed that F.
spiralis grew best when submerged, and grew rapidly only if illuminated while submerged. Hydrated F.
evanescens and A. nodosum photosynthesized faster in air than in water, perhaps because of better
illumination or the higher diffusion rate of CO2 (inorganic carbon is required for photosynthesis) in air than
water (Johnson et al. 1974; Johnston and Raven 1986), but photosynthesis began to shut down as the plants
became dehydrated (Dring and Brown 1982). Photosynthesis in air was not greater for F. vesiculosus
(Bidwell and McKachlan 1985).
The rockweeds have no mechanisms to prevent or slow water loss when exposed to air, they simply tolerate
desiccation (Norton 1986). The species differ in their ability to tolerate desiccation and then resume
photosynthesis, and these differences in tolerance are related to their vertical position in the intertidal zone
(Schonbeck and Norton 1978). For example, F. spiralis survived four times longer than F. vesiculosus and
A. nodosum under similar drying conditions (Schonbeck and Norton 1978). The rate and extent of recovery
of phosphate uptake following severe desiccation and resubmergence increased with height on the shore
(Hurd and Ddng 1991). For five species of brown algae from different tidal heights along the shores of
Ireland and Helgoland, photosynthesis increased as up to 25% of tissue water was lost, then photosynthetic

40
rate decreased with further desiccation (Dring and Brown 1982). Recovery from severe desiccation took
about two hours in all species, but the extent of recovery from a given degree of desiccation was greatest in
upper shore species. F. spiralis showed complete recovery from 80 to 90% water loss and F. vesiculosus
from 70%.
Desiccation tolerance varies as a result of previous experience (hardening) and season, being maximal in
summer and declining in winter (Schonbeck and Norton 1979b). It is likely that the smallest stages in the
life history are more susceptible to desiccation than large plants. Brawiey and Johnson (1993) showed that
mortality of zygotes and embryos was highly correlated with desiccation as measured by agarose bead
hydration, a technique that should facilitate study of these microscopic stages.
Temperature
Temperature tolerance is considered a major determinant of seaweed distribution (van den Hoek 1975;
LOning 1990). The rockweeds tolerate a wide range of water temperatures from 0 to 28ï¿½C (LQning 1990),
although high summer temperatures in combination with desiccation represent potential lethal stress to
intertidal rockweeds (Norton 1986). In mid-coastal Maine, A. nodosum and F. vesiculosus grew where the
water temperature ranged from 0ï¿½ to 40C in winter and 12ï¿½ to 22ï¿½C in summer (Keser et al. 1981). In
northern Europe, A. nodosum survived from Oï¿½ to 25ï¿½C when collected in winter, but survived to 28ï¿½C when
collected in summer; F. spiralis and F. vesiculosus survived from 0O to 28ï¿½C (LOning 1984). When several
species of fucoid algae were grown at temperatures from 2.50C to 35ï¿½C, a temperature optimum below
17.5ï¿½C was indicated, and high temperatures of 30ï¿½ to 35ï¿½C were lethal to all species (Str6mgren 1977). In
other experiments, summer plants were more tolerant of high temperature and had a higher temperature
optimum than winter plants, but winter plants had a higher net rate of photosynthesis at 51 to 30 C than
summer plants (Niemeck and Mathieson 1978). Thermal discharges from the Maine Yankee Atomic Power
Plant in Monsweag Bay raised water temperatures 7ï¿½ to 150C, and resulted in decreased biomass and
percent cover of A. nodosum and F. vesiculosus near the discharge channel (Vadas et al. 1976).
Low winter temperatures also represent potential lethal stress to intertidal rockweeds (Davison et al. 1989;
Pearson and Davison 1993). Experiments show that the species do not differ in the freezing point of their
tissue during slow cooling, but they do differ in their freezing tolerance (Pearson and Davison 1993).
Overall, laboratory experiments on fucoids from Maine shows that freezing tolerance is often correlated with
vertical position in the intertidal zone (Davison et al. 1989; Pearson and Davison 1993; reviewed in
Chapman 1995b). The freezing points of nine species of red and brown algae from the coast of Maine were
similar, however, the recovery of photosynthetic capacity following freezing and thawing was related to tidal
height, and species lower on the shore recovered slower (Pearson and Davison 1993). However, the
importance of stress itolerance n determining the community structure of intertidal algae in nature is not well
understood (reviewed in Davison and Pearson 1996).
In Nova Scotia, most species of Fucus reprbduce in winter (Chapman 1995b) and low temperatures may
negatively impact gametes, zygotes, and embryos. Bird and McLachlan (1974) tested the cold hardiness of
these life history stages of Fucus and found reduced mortality for zygotes at and below -10 C. Older
embryos were more resistant to cold, and full cold hardiness was achieved by F. evanescens after 2 days of
age.
Salinity
It is generally believed that salinity is a major determinant of algal distribution because it can place osmotic
stress on the cells (Back et al. 1992 a,b; Chapman 1995b), but there has been little experimental work to
substantiate the idea. High salinity can potentially result in water loss, and low salinity can cause cells to
swell and/or burst. In the rockweeds, intracellular inorganic (K, CI) and organic (mannitol) substances are
used to regulate osmotic potentials of cells. In Maine, adult rockweed plants have been studied in
Monsweag Bay and the Damadscotta River Estuary where salinities ranged from 10 to 33 ppt (Keser et al.
1981). It has been suggested that salinity stress may be especially inhibiting to successful fertilization
(Brawley 1992) and embryonic growth (Wright and Reed 1990). Polyspermy, multiple fertilizations which

41
are lethal, increases when sodium is limiting (Brawley 1991, 1992). However, sperm lack negative
phototaxis under brackish conditions (SerrSo et al. 1996b) which may reduce the sperm:egg concentration in
the water column. Although this may reduce the susceptibility to polyspermy, it should decrease fertilization
success. Hence, A. nodosum populations in estuaries are probably restricted by spring reproduction when
run-off is high and salinity is low (S. Brawley pers. comm.).

Water motion

The intertidal rockweeds generally are found in areas that are protected or partially protected from wave
action and tidal currents. Studies of several habitats in New Hampshire indicated that A. nodosum and F.
spiralis were absent or stunted in mid-channel areas where the maximum current was 5.5 knots (Mathieson
et al. 1983).

Several investigators have tested the hypothesis that differences in adult plant distributions may be
explained by differences in hydrodynamic tolerance of early post-settlement stages. In a series of field and
laboratory experiments, Vadas et al. (1990, 1992) demonstrated that wave action and currents are a major
source of mortality to recently settled zygotes. Pottery substrates were seeded with zygotes and exposed in
the field to 100 waves or quiet water for an equivalent period of time; the number of zygotes remaining was
0.4 to 0.5% when exposed to waves and 68 to 83% for quiet water (Vadas et al. 1990). The survival of A.
nodosum and F. evanescens zygotes.was compared when exposed to currents between 0 and 25 cm/sec in
a flow tank (Vadas et al. 1992). Survival of F. evanescens was 30 to 100 times greater than survival of A.
nodosum at speeds above 7.5 cm/s. Chapman (1995a) made pairwise comparisons to determine the
resistance of 1-day zygotes to hydrodyamic forces. He found the survival of A. nodosuin to be much
reduced compared to F. vesiculosus at the highest water pressure tested, but no difference in survivorship
of the different Fucus species.
Substrate
Along the Maine coast, the five species of rockweeds have been found attached to rock (>3m diameter),
boulder (10 inches to 3 m diameter), and cobble (2.5 to 10 inches diameter) substrate in the intertidal and
shallowsubtidal zone (Brown 1993).
Nutrients

Uptake of nutrient ions vades with the ion, algal species, individual tested, life history stage, temperature,
and light (Chapman 1995b). McLachlan (1977) made a general survey of minerals required for the growth
of F. evanescens embryos. Deficiencies in the media of nitrogen, phosphorus or iron retarded growth rates
of the embryos, and an increase in iron together with nitrogen and phosphorus stimulated growth; boron and
bromine additions stimulated growth but vitamins had no effect. A requirement for zinc, copper and
manganese was not demonstrated, although these essential metals were probably sufficiently available as
contaminants.

Hurd and Ddng (1990) reported that the rate of phosphorus uptake by three species of rockweeds was
related to their normal zonation with uptake rate in F. spiralis > F. vesiculosus > A. nodosum. However,
plants lower on the shore have a longer period to take up phosphorus, thus the total amount of phosphorus
taken up was about the same for the three species when uptake rate and time for uptake were considered.
The differences in uptake within a species may be related to development of hairs such as those found in
Fucus species in Europe (Hurd et al. 1993).
A. nodosum and F. vesiculosus collected from Rhode Island were able to store nitrate (Asare and Harlin
1983), but Topinka and Robbins (1976) showed that fucoids are commonly nitrogen limited dudring the
summer and that F. spiralis plants receiving supplements of either nitrate or ammonium showed a great
increase in growth. In winter, supplements of nitrogen produce no growth enhancement (Rosenberg et al.
1984).

42
ECOLOGICAL FUNCTION
Nutrient cycling
Chemical elements in an ecosystem cycle back and forth between the living organisms and the physical
environment in a series of transfers termed biogeochemical cycles. The cycles involving carbon, nitrogen,
phosphorus, and the trace metals are extremely important because these elements are vital to the
maintenance of life. These elements make up the biomass of living organisms, and also exist in the
sediment and water column as inorganic compounds, particulate organic matter, and dissolved organic
matter,

Autotrophs alone absorb inorganic carbon, nitrogen, phosphorus, and trace metals, and use the sun's energy
to convert them into living organic matter (biomass). Important autotrophs in the Gulf of Maine include
single-celled organisms (e.g. bacteria, diatoms, dinoflagellates) that drift in the water column as plankton or
are associated with the substrate, macroalgae (e.g. rockweeds, kelps, Irish moss) that usually are attached
to a substrate, and flowering plants (e.g. salt marsh grasses and eelgrass) that are rooted in the substrate
and inundated to various degrees. Autotrophic biomass is the ultimate source of nutrients and energy for
'heterotrophs (bacteria, other single-celled organisms, and multi-celled animals). Heterotrophs in turn
excrete waste products containing inorganic nutrients which are absorbed and utilized by autotrophs.
Rockweeds are important in the cycling of nutrients because they 1) can produce and release large
quantities of organic matter, 2) harbor a diverse community of heterotrophs that consume organic matter
and release inorganic nutrients, and 3) are able to absorb nutrients from the water column. Following
experimental dehydration and rehydration, A. nodosum released 2 to 10 mg C/100 g dry weight into the
water, and F. vesiculosus released 10 to 50 times more (Moebus et al. 1974). Aerobic heterotrophs that are
attached to rockweed shoots or the substrate or that are in the Water column consume dissolved and
particulate organic matter and excrete inorganic waste products (ammonium and phosphorus) into the water
column. Rockweed plants are able to absorb macronutrients like nitrogen (as ammonia, ammonium, or
nitrate) and phosphorus, and micronutrients like elemental metals (cadmium, zinc, iron, copper, manganese
lead, vanadium, and nickel) through the thallus. Rockweeds, with their high rates of accumulation, may act
as a sink for metallic elements.

Uptake of pollutants
The tissues of rockweeds tend to accumulate many metal ions. F. vesiculosus accumulates silver, (Ag),
arsenic (As), cadmium (Cd), cobalt (Co) copper (Cu), lead (Pb), manganese (Mn) and zinc (Zn) (Floc'h
1982; Bryan 1983; Luoma et al. 1982; Munda and Hudnick 1986; Holan and Volesky 1994) and A. nodosum
accumulates Mn, Pb, and Zn (Floc'h 1982; Holan and Volesky 1994). Accumulation of Cd, Co, Mn, and Zn
in F. vesiculosus is greatest in the vegetative shoots and is temperature dependent (Munda 1986).

Accumulation of metal ions may be a passive process. Holan et al. (1993) demonstrated the biosorption of
cadmium by nonliving dried brown algae. Polyphenols extracted from A. nodosurn and F. vesiculosus
chelated a number of heavy metals in weakly acidic solution in the laboratory (Ragan et al. 1979). Ragan et
al. (1980) noted that exudates from brown algae may contribute to the natural chelating capacity of
nearshore Water, important because some phytoplankton are very sensitive to heavy metals.

Bryan (1983) suggested that F. vesiculosus is a good indicator of bioavailable forms of Ag, Cd, Cu. Pb, and
Zn. Luoma et al. (1982) reported that concentrations of Ag, As, Cu, Pb, and Zn in the tissue of F.
vesiculosus correlated significantly with concentrations in the sediment. Within 57 days of being
transplanted from an unpolluted to a polluted site, the concentration of Zn and Cu in the tissue of A.
nodosum and F. vesiculosus had increased by at least a factor of 10 (Ho 1984).
Some metals inhibit the growth of rockweeds. Munda and Hudnik (1986) studied the effects of heavy
metals (singly or in dual combinations at total concentrations of 2.5 and 5 ppm) on the growth of F.

43
vesiculosus. The inhibitory sequence was Cu > Ni > Cd > Zn > Co > Mn (with Cu lethal and Mn favorable
for growth), and was not consistent with the sequence of accumulation. Some metals reduced the inhibitory
effects of other metals or decreased their accumulation. Studies by Stromgren (1980b) demonstrated
decreased growth of F. spiralis, F. vesiculosus, and A. nodosum when exposed to 810g/I Pb, 450gA Cd and
more than 10 g/l Hg. According to Stromgren (1980a), resistance to copper was greatest in A. nodosum,
midrange in F. vesiculosus and least in F. spiralis, however, Smith et al. (1985) stated that Fucus species
have a relatively high tolerance to high environmental levels of dissolved copper.

Food source

Grazing on live rockweeds

Grazing activity by herbivores appears to be the greatest source of mortality to the early post-settlement
phases of many species of algae, although the effect depends on the species involved, developmental
stage, height on the shore, interactions with other biological factors, and the experimental design (Menge
and Farrell 1989; Vadas et al. 1992; McCook and Chapman 1993; Chapman 1995b). Gastropod snails
(Littorina littorea, L. obtusata, and Lacuna vincta) and gammarid amphipods are the most common grazers
in the intertidal zone. Lubchenco (1978, 1983) hypothesized that grazers remove ephemeral algae, which
are able to outcompete the fucoid algae for space, thus allowing fucoid algae to establish iteself. In two-way
choice experiments using adult plants, L. Iittorea prefered ephemeral green algae (Ulva and Enteromorpha)
over Ascophyllum and Fucus (Lubchenco 1978; Watson and Norton 1985) while L. obtusata preferred
Ascophyllum and Fucus over ephemeral green algae (Watson and Norton 1987). When recently-settled
plants (germlings) were used in two-way choice experiments, L. Ittorea also preferred Ulva over
Ascophyllum and Fucus (Watson and Norton 1985). However, when presented with germlings of a single
species of Fucus, L. Ifftorea consumed 70 to 100% and L. obtusata consumed 20 to 50% of the indivduals
(Barker and Chapman 1990). Other multiple choice experiments demonstrated the preference of gammarid
amphipods in Nova Scotia for F. vesiculosus (Denton and Chapman 1991). Under natural conditions in New
Brunswick, Canada, F. evanescens was heavily grazed by the gastropod Lacuna vincta (Thomas and Page
1983). When the mean maximum number of snails was 280/m, grazing removed 79% of the net
production of F. evanescens, which averaged 61 g dry weightlm /day.

Differences in the concentration of phlorotannins (Denton et al. 1990) and polyphenols (Geiselman and
McConnell 1981) may inhibit the herbivores. F. gardneri responded to grazing damage by increasing its
polyphenol content (Van Alstyn 1988) and A. nodosum responded by increasing the toughness of its tissue
(Lowell et al. 1991).

Grazing on epiphytes

There is no comprehensive study of rockweed epiphytes in the Gulf of Maine. One of the most important
epiphytes on A. nodosum, on both sides of the Atlantic Ocean, is the red alga Polysiphonia lanosa (Pearson
and Evans 1990) which is probably resistant to grazers. In the Baltic Sea, F. vesiculosus was found to have
26 species of epiphytic algae that were distributed over the entire length of the thallus, but decreased toward
the holdfast and at the growing tips (Colina 1981). In the Baltic Sea and in northern NewZealand, diatoms
epiphytic on F. vesiculosus were found to contribute 1 to 10 mg Clm2/host/h (Booth 1987).

Consumption of rockweed-derived detritus

Detritus refers to dissolved organic matter, particulate organic matter, and the assemblage of decomposing
bacteria and fungi that inhabits the organic matter (Zieman 1982; Thayer et al. 1984; Mann 1988).
Dissolved organic matter includes soluble carbohydrates and proteins (Zieman 1982; Mann 1988);
particulate organic matter includes dead autotrophs or autotroph debris, dead heterotroph or heterotroph
debris of any size, fecal pellets, and combinations of these items (Cousins 1980; Mann 1988).

As part of the natural life cycle, rockweed blades are eroded at the end by wave action or ice scouring and
fall to the bottom or are carried away by water currents. Whole fronds probably become fragmented by the

44
shredding action of herbivorous invertebrates and by physical factors such as desiccation and ice grinding.
The intact blades or fragments become colonized by bacteria and fungi, which begin to decompose the
algae.
The processes of particle size reduction and microbial colonization and decomposition are essential to the
detrital food web,and have been summarized in detail for,algae by Mann (1988). Particle size reduction is
important because some organisms are able to ingest only fine particles, and the greater surface to volume
ratio of fine particles allows greater colonization by microbes. The microbes enzymatically break down plant
material that most organisms cannot digest, and convert it to microbial biomass which can be assimilated.
Microbes also ennch the detritus by taking up nitrogen and phosphorus from the surrounding medium and
incorporating it into microbe biomass.

Macroalgae like the rockweeds make an important contribution to the detdtal cycle. Josselyn and Mathieson
(1980) conducted an 18-month study in Great Bay Estuary system and the adjacent open coast to determine
the seasonal contribution and detrital processing of plant litter that is locally derived (Spartina altemiflora,
Zostera marina, A. nodosum, and F. vesiculosus). Rockweed lifter accounted for 35 to 85% of the total
strand line throughout the year, and contained I to 3 times as much detrital material as the vascular plants
within the estuary and 50 times as much on the open coast. Seaweeds decomposed 3 to 10 times faster
than the vascular plant litter under similar conditions. Decomposition rates and changes in nutdrient content
of litter were dependent on surrounding environmental conditions, and were maximal in spring and summer.

Smith and Foreman (1984) assessed litter and detrital decomposition and nitrogen composition of
decomposing litter for 10 important species of seaweeds in British Columbia, Canada. The decomposition
rate for F. gardneri was about 70 days. As decomposition proceeded there was an accelerating increase in
the nitrogen:dry weight ratio of the remnant litter.

Habitat
The rockweeds provide habitat for a variety of species by adding structural complexity to the water column
when the plants are submerged in water and by ameliorating the extremes of temperature and desiccation
when they are emersed in air (Vadas and Elner 1992). Using a functional approach, the flora and fauna
associated with intertidal rockweeds include: 1) epiphytes on the thallus), 2) biota that attach to the holdfast
and thallus, 3) mobile fauna that move on or swim over the thallus and 4) biota that dwell on or in the
underlying substrate.
Researchers working in the Baltic Sea, where rockweeds are the dominant subtidal vegetation on hard
bottoms, have reported a greater biomass of macroscopic animals in the F. vesiculosus zone than in
shallow areas with soft bottoms (reviewed in Lehtinen et al. 1988). In addition to being important source of
food and shelter for fish and invertebrates, the rockweed beds in the Baltic Sea are important as a spawning
area for some species of pelagic fishes, especially herring, Clupea harengus (Aneer 1989).

During a 13-month study of epifauna on A. nodosum and F. vesiculosus in Nova Scotia, the total density of
epifauna declined by two orders of magnitude between summer and winter (Johnson and Scheibling 1987).
Harpacticoid copepods and their nauplii, nematodes and halacarid mites were numerically the most
abundant species during the study, but considerable numbers of young annelid worm and gastropods also
were present (Johnson and Scheibling 1987). Density of most of the epifaunal taxa was positively
correlated with epiphytic algal biomass (Johnson and Scheibling 1987).

At a site in Scotland, experimental evidence showed that the species diversity and abundance of individuals
settling on the substrate was greater under a canopy of A. nodosum than outside of the canopy (Hruby and
Norton 1979). The authors attributed this to the greater humidity found under the A. nodosum when the tide
was out. However, shelter from water motion and predation may contribute to the increase of settling
species.

45
Table 1 lists the species of fishes and invertebrates that have been collected from the rockweed zone in
southwestern Nova Scotia (Black and Miller 1991; Rangeley 1994), Passamaquoddy Bay (Thomas et al.
1983; Rangeley and Kramer 1995a,b), and the Sheepscot River Estuary (Tort 1993). Black and Miller
(1991) obtained evidence that the fishes were feeding on amphipods, isopods, Littorina sp., green crabs,
and young fish, all of which associate with A. nodosum in the intertidal zone.

Commercial uses
Salts of alginic acid (alginates) occur in all brown algae as structural components of the cell walls. Alginates
are commercially extracted from Macrocystis pyrifera, A. nodosum, and Laminaria sp., and are used for
their thickening, stabilizing, film forming and gel producing properties (Whyte 1988). The products are used
extensively in the baking industry and as a stabilizer in the manufacture of ice cream. Extracts of A.
nodosum have been used alone or with other ingredients as a biostimulant for field grown bedding plants
(Pdincelot 1994) and hydroponically grown barley (Steveni et al. 1992). In Maine, rockweeds are collected
for use as packing matedal for live lobsters (L. Mercer pers. comm.). In some parts of Europe, F.
vesiculosus is sold as a health food (S. Brawley pers. comm.).

Small quantities of A. nodosum are harvested in Maine, but large-scale commercial harvesting in Nova
Scotia, Canada, has been active for 35 years (Sharp et al. 1994). Harvesting in Canada is traditionally done
by raking from a boat or by cutting at lowtide with a sickle (Sharp et al. 1994). However, suction cutters, if
used properly, cause the least damage to the holdfast (R. Vadas pers. comm.).

CONCLUSION
Studies of the rockweeds have been conducted on both sides of the northern Atlantic Ocean. Many of these
studies have concentrated on the causes of the striking patterns of vertical zonation that occur in the
intertidal zone. Few studies have examined the ecological functions of these species, but collectively they
have shown that the five species of rockweeds 1) contribute to nutrient cycling, 2) provide food to grazers
and detritus feeders, 3) help maintain water quality, and 4) provide habitat for numerous species including
commercially and recreationally important species.


THREATS
The major threats to rockweeds indude reductions of light levels due to degradation of water quality or
physical structures; reproductive inhibition due to petroleum hydrocarbons; local overharvesting; chlorate
effluent from pulp. mills; competition from nonnative species; and thermal discharges from power plants.

Reduction of light levels due to degradation of water quality by nutdent loading is a threat to the rockweeds.
Nutrient loading can result from point sources such as wastewater treatment facilities and aquaculture sites
(Roennberg et al. 1992) and non-point sources such as runoff of animal wastes and fertilizers from lawns
(Short 1992). Excessive nutrients promote the rapid growth of phytoplankton in the water column and
epiphytes on the rockweed thallus, both of which can reduce amount of light reaching the rockweeds. In the
Baltic Sea, unlike the Gulf of Maine, F. vesiculosus grows subtidally and is the dominant coastal vegetation
(Lehtinen et al. 1988). Increased inshore nutrient levels in the Baltic Sea (both nitrate, nitrite and
phosphate) have been accompanied by increased blooms of plankton and filamentous algae, decreased
water transparency, decreased depth penetration of F. vesiculosus, increased deposition of organic matter
on the bottom, decreased oxygen in bottom waters, and reduced bottom fauna (Lainianen et al. 1989;
reviewed by Cederwall and Elmgren 1990; Kautsky 1991). The disappearance of A. nodosum, P.
vesiculosus, and F. evanescens from the inner part of Oslofjord was attributed to the increased discharge of
sewage (Bokn and Lein 1978). Surveys conducted in 1988 to 1990 indicated that fucoid algae have
recovered in many parts of Oslofjord, and this recovery corresponds to improved sewage treatment and
discharge practices (Bokn et al. 1992).

46
Table 1. Species of fishes and invertebrates collected from intertidal fucoid algae and fucoid algae habitat in
the Gulf of Maine.

Species                                   Common name
Fishes
Alosa pseudoharengus 1 2 3	Alewife
Ammodytes americanus 2	Sandlance
Anguilla rostrata '	American eel
Clupea harengus 2	Herring
Cyclopterus lumpus 2	Lumpfish
Fundulus heteroclitus' 3	Mummichog
Fundulus majalis 2	Killifish
Gadus morhua 2	Atlantic cod
Hemitripterus americanus	Atlantic sea raven
Menidia menidia 2 3	Atlantic silverside
Microgadus tomcod 1' 2	Atlantic tomcod
Myxocephalus aenaeus 1	Grubby
Myxocephalus octodecemspinosus	Longhorn sculpin
Myxocephalus scorpius1	Shorthom sculpin
Osmerus mordax 1'2,3	Rainbow smelt
Pholis gunnellus 2	Rock gunnel
Pollachius virens ' 4	Pollock
Pleuronectes americanus 1 2 3	Winter flounder
Scomber scombrus 1	Atlantic mackerel
Pungitius pungitius 3	Nine-spine stickleback
Tautogolabrus adspersus	Cunner
Tautoga onitus 1	Tautog
Urophycis tenuis ' 2	White hake
Invertebrates
Carcinus maenas	Green crab
Flustrellida hispida 5	Bristly bryozoan
Gammarus oceanicus5	Amphipod
Littorina obtusata 5	Smooth periwinkle
Nucella lapillus 5	Dogwinkle
Potamilla neglecta 5	Fan worm
Sertularia pumila 5	Hydroid
1 = Black and Miller 1991; 2 = Rangeley 1994; 3 =Tort 1993; 4 = Rangeley and Kramer 1995a, b;
5 = Thomas et al. 1983; 6 = Maria Tort pers comm.

47
Reduction of light levels can also be caused by sediment loading. Sediment suspended in the water column
reduces the amount of light that penetrates the water, thus reducing the light that reaches rockweeds on the
bottom. Suspended sediments result from point and nonpoint sources of upland runoff and from
resuspension of sediment by tidal currents, wvave action, dredging, filling, boating activity, and fishing
activity (Short 1992; Colarusso pers. comm.). The impact of suspended sediment in the vicinity of
rockweeds may vary with hydrographic conditions, sediment grain size, time of year, distance from the
sediment, health of the rockweeds, and a host of other variables.
The reproductive stages of rockweeds are especially sensitive to oil pollution. In laboratory expenments,
minute quantities of crude oil, No. 2 fuel oil, and two jet fuels killed Fucus sperm being released from
receptacles, and fuel oil and jet fuels at a concentration of 20 ppm completely dilled Fucus zygotes (Steele
1977, 1978). A shoreline survey conducted after No. 2 fuel oil was spilled in Narragansett Bay revealed
little impact on adult Fucus, but sexual reproduction was inhibited due to mortality of gametes and/or
embryos for approximately three weeks (Thursby et al. 1990). In A. nodosum, as well as other species of
brown algae, unsaturated hydrocarbons serve as pheromones, attracting male to female gametes (MOller
1981). It has been suggested, but not demonstrated, that petroleum hydrocarbons could interfere with
normal functioning of these pheromones (Maier and MIller 1986).
Overharvesting is a potential threat to fucoid algae, particularly to the slow growing A. nodosum. Large
amounts of rockweeds are harvested in Canada, but the harvest is partially subsidized by the government
(Inka Milewski pers. comm.). The harvest in Maine is small by comparison, but if the price for rockweeds
increases sufficiently, harvesting would probably increase.
A study conducted in mid-coastal Maine to determine the regrowth of A. nodosum and F. vesiculosus
following harvest demonstrated that regrowth varied with site, exposure, and harvest method (Keser et al.
1981). When A. nodosum was harvested to the holdfast, recovery to average pre-cut biomass required 2 to
3 years at protected sites and more than 3 years at moderately exposed sites. When A. nodosum was
harvested by cuts at heights of 15 or 25 cm, recovery was faster at sheltered sites than exposed sites, and
was faster than when the algae was harvested to the holdfast.
Few studies have addressed the ecological impact of rockweed harvesting. Only one study, conducted in
Nova Scotia where harvests exceeded 20, 000 tons in 1989, has investigated the use by fishes of intertidal
habitats where A. nodosum was dominant or had been harvested (Black and Miller 1991). However,
sampling bias (Rangeley 1994) and the use of nets with a large mesh, which would not have captured
juvenile fishes, make the results of this study inconclusive.
In the Baltic Sea, the abundance and biomass of fucoid algae declined regionally after cellulose pulp mills
started releasing chlorate (CLO3-) in their effluent (reviewed in Lehtinen et al. 1988 and Kautsky 1992). The
chlorate is formed as a by-product in the bleaching of chemical pulp (Lehtinen et al. 1988). Laboratory tests
indicated that chlorate, effective at 10 to 20 gg/L, was the main causal agent for the decline of F.
vesiculosus (Rosemarin et al. 1985). Other species of brown algae were highly sensitive to chlorate, while
cyanobacteria, green algae, red algae, and Zostera had much lower sensitivities (Rosemarin et al. 1985;
1994). Transplantation experiments, quantitative surveys along transects, and mesocosm experiments
were used to investigate the effect of chlorate. Briefly, these experiments revealed that the abundance,
biomass and depth distribution of F. vesiculosus and the abundance, biomass and diversity of
macroinvertebrates increased with increased distance from the chlorate source (Lehtinen et al. 1988;
Kautsky 1992). Kautsky (1992) noted that reduction of chlorate effluent has resulted in the partial recovery
of F. vesiculosus in the effluent receiving area.
The native species of rockweeds may be outcompeted by introduced or nonnative species. F. serratus, a
species of fucoid algae that occurs in Europe, has been introduced into Nova Scotia. Its spread has been
rapid in some areas but slow in others (Dale 1982). A study of the phytosociological structure of the
seaweed community in Nova Scotia has found evidence that the spread of this species into new areas is not
slowed by competition from Chondrus crispus but competition by F. evanescens is an important factor (Dale
1982).

48
REFERENCES

Aneer, G. 1989. Herring (Clupea harengus L.) spawning and spawning ground characteristics in the Baltic
Sea. Fish. Res. 8: 189-195.
Ang, P.O. 1991. Natural dynamics of a Fucus distichus (Phaeophyceae, Fucales) population. Mar. Ecol.
Prog. Ser. 78: 71-85.
Asare, S.O. and M.M. Harlin. 1983. Seasonal fluctuations in tissue nitrogen for five species of perennial
macroalgae in Rhode Island Sound. J. Phycol. 19: 254-257.
Bacon, L.C. and R.L. Vadas. 1991. A model for gamete release in Ascophyllum nodosum (Phaeophyta). J.
Phycol. 27: 166-173.
Back, S., J.C. Collins, and G. Russell. 1992a. Comparative ecophysiology of Baltic and Atlantic Fucus
vesiculosus. Mar. Ecol. Prog. Ser. 84: 71-82.
Back, S., J.C. Collins, and G. Russell. 1992b. Effects of salinity on growth of Baltic and Atlantic Fucus
vesiculosus. Br. Phycol. J. 27: 39-47.
Barker, K.M. and A.R.O. Chapman. 1990. Feeding preferences of periwinkles among four species of Fucus
. Mar. Biol. 106: 113-118.
Bidwell, R.G.S. and J. McLachlan. 1985. Carbon nutrition of seaweeds: photosynthesis, photorespiration,
and respiration. J. Exp. Mar. Biol. Ecol. 86: 15-46.
Bird, C.J. and J. McLachlan. 1974. Cold-hardiness of zygotes and embryos of Fucus (Phaeophyceae:
Fucales). Phycologia 13: 215-225.
Black, R., and R.J. Miller. 1991. Use of the intertidal zone by fish in Nova Scotia. Env. Biol Fish. 31: 109-
121.
Bokn, T. and T.E. Lein. 1978. Long-term changes in fucoid association of the inner Oslofjord, Norway.
Norway J. Bot. 25: 9-14.
Bokn, T. L., S.N. Murray, F.E. Moy, and J.B. Magnusson. 1992. Changes in fucoid distributions and
abundances in the inner Oslofjord, Norway: 1974-80 versus 1988-90. Acta Phytogeogr. Suec. 78:
Booth, W.E. 1987. Contribution by diatoms to marine algal host epiphyte photosynthesis. Bot. Mar. 30: 129-
140.
Brawley, S.H. 1992. Fertilization in natural populations of the dioecious brown alga Fucus ceranoides and
the importance of the polyspermy block. Mar. Biol. 113: 145-157.
Brawley, S.H. and L.E. Johnson. 1992. Gametogenesis, gametes and zygotes: an ecological perspective on
sexual reproduction in the algae. Br. Phycol. J. 27: 233-252.
Brawley, S.H. and L.E. Johnson. 1993. Predicting desiccation stress in microscopic organisms: the use of
agarose beads to determine evaporation within and between intertidal microhabitats. J. Phycol. 29:
528-535.
Brinkhuis, B.H. 1977. Seasonal variations in salt-marsh algae photosynthesis. II. Fucus vesiculosus and
Ulva lactuca. Mar. Biol. 44: 177-186.
Brinkhuis, B.H. and R.F. Jones. 1976. The ecology of temperate salt-marsh fucoids. II. In situ growth of
transplanted Ascophyllum nodosum ecads. Mar. Biol. 34: 339-348.
Brown, B. 1993. A classification system of marine and estuarine habitats in Maine: an ecosystem approach
to habitats. Part I. Benthic habitats. State of Maine Department of Economic and Community
Development, Maine Natural Areas Program. 61 pages.
Bryan, G.W. 1983. Brown seaweed, Fucus vesiculosus, and the gastropod, Littorfina littorea, as indicators of
trace-metal availability in estuaries. Sci. Total Environ. 28: 91-104.

49

Burrows, E.M. and S.M. Lodge. 1950. A note on the inter-relationships of Patella, Balanus and Fucus on a
semi-exposed coast. Rep. Mar. Biol. Stn. Port Erin 62: 30-34.
Burrows, E.M. and S.M. Lodge. 1951. Autecology and the species problem in Fucus. J. Mar. Biol. Ass. U.K.
30:161-176.
Canadian Atlantic Fishedes Scientific Advisory Committee. 1992. Rockweed in southwestern New
Brunswick CAFSAC Advisory Document 92113.
Cederwall, H. and R. Elmgren. 1990. Biological effects of eutrophication in the Baltic Sea, particularly the
coastal zone. Ambio 19: 109-112.
Chapman, A.R.O. 1995a. Ecology of fucoid recruits: effects of grazing and hydrodynamic forces. In:
Grassle, J.P., A. Kelsey, E. Oates, and P.V. Snelgrove (eds.). Twenty Third Benthic Ecology Meeting.
Rutgers the State University, New Brunswick, New Jersey. Inst. Marine Coastal Sciences 1995.
Chapman, A.R.O. 1995b. Functional ecology of fucoid algae: twenty-three years of progress. Phycologia
34:1-32.
Chapman, A.R.O. and C.R. Johnson. 1990. Disturbance and organization of macroalgal assemblages in the
Northwest Atlantic. Hydrobiologia 192: 77-121.
Chock, J.S. and A.C. Mathieson. 1976. Ecological studies of the salt marsh ecad scorpioides (Homemann)
Hauck of Ascophyllum nodosum (L.) Le Jolis. J. Exp. Mar. Biol. Ecol. 23: 171-190.
Chock, J.S. and A.C. Mathieson. 1979. Physiological ecology of Ascophyllum nodosum and its detached
ecad scorpioides (Fucales, Phaeophyta). Bot. Mar. 22: 21-26.
Chock, J.S. and A.C. Mathieson. 1983. Variations of New England estuarine seaweed biomass. Bot. Mar.
26: 87-97.
Colina, A. 1981. Investigation on the structure, composition and productivity of the epiphyte communities on
Fucus vesiculosus L. in the western Baltic. Diss. (Dr. rer. math. nat.). 117 pages.
Cousens, R. 1984. Estimation of annual production by the intertidal brown alga Ascophyllum nodosum (L.)
Le Jolis. Bot. Mar. 27: 217-227.
Cousins, S.H. 1980. A trophic continuum derived from plant structure, animal size, and a detritus cascade.
J. Theor. Biol. 82: 607-618.
Dale, M. 1982. Phytosociological structure of seaweed communities and the invasion of Fucus serratus in
Nova Scotia. Can. J. Bot. 60: 2652-2658.
Davison, l.R., S.R. Dudgeon and H.-M. Ruan. 1989. The effect of freezing on seaweed photosynthesis. Mar.
Ecol. Prog. Ser. 58: 123-131.
Davison, l.R. and G.A. Pearson. 1996. Stress tolerance in intertidal seaweeds. J. Phycol. 32:197-211.
Denton, A.B., A.R.O. Chapman, and J. Markham. 1990. Size-specific concentrations of phlorotannins (anti-
herbivore compounds) in three species of Fucus. Mar. Ecol. Prog. Ser. 65: 103-104.
Denton, A.B. and A.R.O. Chapman. 1991. Feeding preferences of gammarid amphipods among four
species of Fucus . Mar. Biol. 109: 503-506.
Dring, M.J. 1984. Mean irradiance levels at different heights in the intertidal and upper subtidal zones: a
general model and its application to shores in the Bristol Channel. Br. Phycol. J. 19: 192.
Dring, M.J. 1987. Light climate in intertidal and subtidal zones in relation to photosynthesis and growth of
benthic algae: a theoretical model. In: Crawford, R.M.M.(ed.). Plant life in aquatic and amphibious
habitats. Pages 23-34: Blackwells, Oxford.
Dring, M.J. and F.A. Brown. 1982. Photosynthesis of intertidal brown algae during and after periods of
emersion: a renewed search for physiological causes of zonation. Mar. Ecol. Prog. Ser. 8: 301- 308.

50
Evans, L.V., J.A. Callow, M.E. Callow. 1982. The biology and biochemistry of reproduction and early
development in Fucus. In: Round,  and A.R.O. Chapman (eds.). Progress in Phycological Research,
Vol. 1. Elsevier Biomedical Press. Pages 68-98.
Floc'h, J.Y. 1982. Uptake of inorganric ions and their long distance transport in Fucales and Laminariales. In:
Srivastava, L.M. (ed.). Synthetic and degrative processes in marine macrophytes. Pages 139-166.
Walter de Gruyter, Berlin.
Geiselman, J.A. and O.J. McConnell. 1981. Polyphenols in brown algae Fucus vesiculosus and
Ascophy#lum nodosum: chemical defenses against the marine herbivorous snail, Littorina littorea. J.
Chem. Ecol. 7:1115-1133.
Haahtela, I. 1984. A hypothesis of the decline of the bladder wrack (Fucus vesiculosus L.) in SW Finland in
1975-1981. Limnologica 15: 345-350.
Hawkins, S.J. and R.G. Hartnoll. 1985. Factors determining the upper limits of intertidal canopy-forming
algae. Mar. Ecol. Prog. Ser. 20: 265-271.
Ho, Y.B. 1984. Zn and Cu concentrations in Ascophyllum nodosum and Fucus vesiculosus (Phaeophyta,
Fucales) after transplantation to an estuary contaminated with mine wastes. Conserv. Recydcling 7:
329-337.
Holan, Z.R. and B. Volesky. 1994. Biosorption of lead and nickel by biomass of marine algae. BiotechnoL
Bioeng. 43: 1001-1009.
Holan, Z.R., B. Volesky, and I. Prasetyo. 1993. Biosorption of cadmium by biomass of marine algae.
Biotechnol. Bioeng. 41: 819-825.
Hruby,T. and T.A. Norton. 1979. Algal colonization on rocky shores in the Firth of Clyde. J. Ecol. 67: 65-77.
Hurd, C.L. and M.J. Dring. 1990. Phosphate uptake by intertidal algae in relation to zonation and season.
Mar. Biol. 107: 281-289.
Hurd, C.L. and M.J. Dring. 1991. Desiccation and phosphate uptake by intertidal fucoid algae in relation to
zonation. Br. Phycol. J. 26: 327-333.
Hurd, C.L., R. Galvin, T.A. Norton, and M.J. Dring. 1993. Production of hyaline hairs by intertidal species of
Fucus (Phaeophyceae) and their role in phosphate uptake. J. Phycol. 29: 160-165.
Johnson, S.C. and R.E. Scheibling. 1987. Structure and dynamics of epifaunal assemblages on intertidal
macroalgae Ascophyllum nodosum and Fucus vesiculosus in Nova Scotia, Canada. Mar. Ecol. Prog.
Ser. 37: 209-227.
Johnson, W.S., A. Gigon, S.L. Gulman and H.A. Mooney. 1974. Comparative photosynthetic capacities of
intertidal algae under exposed and submerged conditions. Ecology 55: 450-453.
Johnston, A.M. and J.A. Raven. 1986. The analysis of photosynthesis in air and water of Ascophyllum
nodosum (L.) Le Jol. Oecologia 69: 288-295.
Josselyn, M.N. and A.C. Mathieson. 1980. Seasonal influx and decomposition of autochthonous macrophyte
litter in a north temperate estuary. Hydrobiologia 71: 197-208.
Kangas, P., H. Autio, G. Haellfors, H. Luther, A. Niemi, and H. Salemaa. 1982. A general model of the
decline of Fucus vesiculosus at Tvaerminne, south coast of Finland in 1977-81. Acta Bot. Fenn. 118:
1-27.
Kautsky, H. 1991. Influence of eutrophication on the distribution of phytobenthic plant and animal
communities. Int. Rev. Gesamt. Hydrobiol. 76: 423-432.
Kautsky, H. 1992. The impact of pulp-mill effluents on phytobenthic communities in the Baltic Sea. Ambio
21: 308-313.

51
Kautsky, N., H. Kautsky, U. Kautsky, and M. Waem. 1986. Decreased depth penetration of Fucus
vesiculosus (L.) since the 1940's indicates eutrophication of the Baltic Sea. Mar. Ecol. Prog. Ser. 28:
1-8.
Keser, M. 1978. Ecological effects of harvesting on the growth of Ascophyllum nodosum and the growth
dynamics of Fucus. Ph.D. thesis, University of Maine at Orono, 137 pp.
Keser, M., R.L. Vadas, and B.R. Larson. 1981. Regrowth of Ascophyl/um nodosum and Fucus vesiculosus
under various harvesting regimes in Maine, USA. Bot. Mar. 24: 29-38.
Keser, M. and B.R. Larson. 1984. Colonization and growth dynamics of three species of Fucus . Mar. Ecol.
Prog. Ser. 15: 125-134.
Lambert, A.J. and S.H. Brawley. 1993. Fertilization success and polyspermy in estuarine Ascophyllum
nodosum. Northeast Algal Symposium, Woods Hole. Page 15.
Launiainen, J., J. Vainio, A. Voipio, J. Pokki, and J. Nieminaa. 1989. Long-term changes in the secchi depth
in the northern Baltic Sea. XIV Geofysiikan Paivat Helsingissa 3.4.5.1989:117-121.
Lehtinen, K.J., M. Notini, J. Mattsson, and L. Landner. 1988. Disappearance of bladder wrack (Fucus
vesiculosus L.) in the Baltic Sea: relation to pulp-mill chlorate. Ambio 17: 387-393.
Lowell, R.B., J.H. Markham, and K.H. Mann. 1991. Herbivore-like damage induces increased strength and
toughness in a seaweed. Proc. R. Soc. B. 243: 31-38.
Lubohenco, J. 1978. Plant species diversity in a marine intertidal community: importance of herbivore food
preference and algal competitive abilities. Am. Nat. 112: 23-39.
Lubchenco, J. 1980. Algal zonation in a New England rocky intertidal community: an experimental analysis.
Ecology 61: 333-344.
Lubchenco, J. 1983. Littorina and Fucus: effects of herbivores, substratum heterogeneity, and plant escapes
during succession. Ecology 64: 1116-1123.
L0ning, K. 1984. Temperature tolerance and biogeography of seaweeds: the marine algal flora of Helgoland
(North Sea) as an example. Helgolander Meeresunters. 38: 305-317.
Luning, K. 1990. Seaweeds: their environment, biogeography and ecophysiology. Wiley, NewYork. 527
pages.
Luoma, S.N., G.W. Bryan, and W.J. Langston. 1982. Scavenging of heavy metals from particulates by
brown seaweed. Mar. Pollut. Bull. 13: 394-396.
Maier, I. and D.G. Muller. 1986. Sexual pheromones in algae. Biol. Bull. 170: 145-175.
Mann, K.H. 1988. Production and use of detritus in various freshwater, estuarine, and coastal marine
ecosystems. Limnol. Oceanogr. 33: 910-930.
Mathieson, A.C. and Z. Guo. 1992. Patterns of fucoid reproductive biomass allocation. Br. Phycol. J. 27:
271-292.
Mathieson, A.C., J.W. Shipman, R.J. O'Shea, and R.C. Hasevlat. 1976. Seasonal growth and reproduction
of estuarine fucoid algae in New England. J. Exp. Mar. Biol. Ecol. 25: 273-284.
Mathieson, A.C., C.D. Neefus, and C.E. Penniman. 1983. Benthic ecology in an estuarine tidal rapid. Bot.
Mar. 26: 213-230.
McCook, L.J. and A.R.O. Chapman. 1992. Vegetative regeneration of Fucus rockweed canopy as a
mechanism of secondary succession on an exposed rocky shore. Bot. Mar. 35: 35-46.
McCook, L.J. and A.R.O. Chapman. 1993. Community succession following massive ice scour on an
exposed rocky shore: recruitment, competition, and predation during early, primary succession. Mar.
Biol. 115: 565-575.

52
 McLachlan, J. 1977. Effects of nutrients on growth and development of embryos of Fucus edentatus Pyl.
(Phaeophyceae, Fucales) Phycologia 16: 329-338.
McLaughlan, J. and R.G.S. Bidwell. 1983. Effects of colored light on the growth and metabolism of Fucus
embryos and apices in culture. Can. J. Bot. 61: 1993-2003.
Menge, B.A. 1976. Organization of the New England rocky intertidal community: role of predation,
competition, and environmental heterogeneity. Ecol. Monogr. 46: 355-393.
Menge, B.A. and T.M. Farrell. 1989. Community structure and interaction webs in shallow marine hard-
bottom communities: tests of an environmental stress model. Adv. Ecol. Res. 19: 189-262.
Menge, J. 1975. Effect of herbivores on community structure of the New England rocky intertidal region:
distribution, abundance, and diversity of algae. Ph.D. thesis, Harvard University, Cambridge, MA, 164
PP.
Miller, S.L. and R.L. Vadas. 1984. The population biology of Ascophyllum nodosum: biological and physical
factors affecting survivorship of germlings. Br. Phycol. J. 19: 198.
Moebus, L., K.M. Johnson and J.McN. Sieburth. 1974. Rehydration of desiccated intertidal brown algae:
release of dissolved organic carbon and water uptake. Mar. Biol. 26: 127-134.
Muller, D.G. 1981. Sexuality and sex attraction. In: Lobban, C.S. and Wynne, M.J. (eds.). The biology of
seaweeds. Pages 661-674. Blackwell, Oxford.
Munda, l.M. 1986. Differences in heavy metal accumulation between vegetative parts of the thalli and
receptacles in Fucus spiralis L. Bot. Mar. 29: 341-349.
Munda, l.M. and V. Hudnik. 1986. Growth response of Fucus vesiculosus to heavy metals, singly and in dual
combinations, as related to accumulation. Bot. Mar. 29: 401-412.
Niemeck, R.A. and A.C. Mathieson. 1978. Physiological studies of intertidal fucoid algae. Bot. Mar. 21: 221-
227.
Norton, T.A. 1986. The zonation of seaweeds on rocky shores. In: Moore, P.J. and R. Seed (eds). The
ecology of rocky coasts. Columbia University Press, NewYork. Pages 7-21.
Pearson, G.A. and l.R. Davison. 1993. Freezing rate and duration determine the physiological response of
intertidal fucoids to freezing. Mar. Biol. 115: 353-362.
Pearson, G.A. and L.V. Evans. 1990. Settlement and survival of Polysiphonia lanosa (Ceramiales) spores
on Ascophyllum nodosum and Fucus vesiculosus (Fucales). J. Phycol. 26: 597-603.
Peckol, P., M.M. Harlin, and P. Krumscheid. 1988. Physiological and population ecology of intertidal and
subtidal Ascophyllum nodosum (Phaeophyta). J. Phycol. 24: 192-198.
Poincelot, R.P. 1994. The use of a commercial organic biostimulant for field grown bedding plants. J. Home
Consum. Hortic. 1: 95-110.
Ragan, M.A., C.M. Ragan, and A. Jensen. 1980. Natural chelators in sea water: detoxification of Zn+2 by
brown algae polyphenols. J. Exp. Mar. Biol. Ecol. 44: 216-267.
Ragan, M.A., O. Smidsrod, and B. Larsen. 1979. Chelation of divalent metal ions by brown algal
polyphenols. Mar. Chem. 7: 265-271.
Ramus, J., F. Lemons, F. and C. Zimmerman. 1977. Adaptation of light-harvesting pigments to downwelling
light and the consequent photosynthetic performance of the eulittoral rockweeds Ascophyllum
nodosum and Fucus vesiculosus. Mar. Biol. 42: 293-303.
Ramus, J. 1981. The capture and transduction of light energy. In: Lobban, C.S. and M.J. Wynne (eds.). The
biology of seaweeds. Pages 458-492. Blackwell, Oxford.
Rangeley, R.W. 1994. The effects of seaweed harvesting on fishes: a critique. Env. Biol. Fish. 39: 319-323.

53
Rangeley, R.W. and D.L. Kramer. 1995a. Tidal effects on habitat selection and aggregation by juvenile
pollack Pollachius virens in the rocky intertidal zone. Mar. Ecol. Prog. Ser. 126: 19-29.
Rangeley, R.W. and D.L. Kramer. 1995b. Use of rocky intertidal habitats by juvenile pollock Pollachius
virens. Mar. Ecol. Prog. Ser. 126: 9-17.
Roennberg, O., J. Lehto, and I. Haahtela. 1985. Recent changes in the occurrence of Fucus vesiculosus in
the Archipelago Sea, SW Finland. Ann. Bot. Fenn. 22: 231-244.
Roennberg, O., K. Aadjers, C. Ruokolahti, and M. Bondestam. 1992. Effects of fish farming on growth,
epiphytes and nutrient content of Fucus vesiculosus L. in the Aaland Archipelago, northern Baltic
Sea. Aquat. Bot. 42: 109-120.
Rosenberg, G., T.A. Probyn and K.H. Mann. 1984. Nutrient uptake and growth kinetics in brown seaweeds:
response to continuous and single additions of ammonium. J. Exp. Mar. Biol. Ecol. 80:125-146.
Rosemarin, A., K.J. Lehtinen, M. Notini, B. Axelsson, and J. Mattsson. 1985. Effects of pulp mill chlorate on
algae. Swedish Environmental research group (SERG). Report No. K5015:1.
Rosemarin, A., K.J. Lehtinen, M. Notini, and J. Mattsson. 1994. Effects of pulp mill chlorate on Baltic Sea
algae. Environ. Pollut. 85: 3-13.
Schonbeck, M.W. and T.A. Norton. 1978. Factors controlling the upper limits of fucoid algae on the shore. J.
Exp. Mar. Biol. Ecol. 31: 303-313.
Schonbeck, M.W. and T.A. Norton. 1979a. The effects of brief periodic submergence on intertidal fucoid
algae. Estuar. Coast. Mar. Sci. 8: 205-211.
Schonbeck, M.W., and T.A. Norton. 1979b. Drought-hardening in the upper-shore seaweeds Fucus spiralis
and Pelvetia canaliculata. J. Ecol. 67: 687-696.
Schonbeck, M.W. and T.A. Norton. 1980. Factors controlling the lower limits of fucoid algae on the shore. J.
Exp. Mar. Biol. Ecol. 43: 131-150.
Serrao, E.A., G. Pearson, L. Kautsky, and S.H. Brawley. 1996a. Successful external fertilization in.turbulent
environments. Proc. Natl. Acad. Sci. 00:00-00.
Serrao, E.A., L. Kautsky, and S.H. Brawley. 1996b. Distributional success of the marine seaweed Fucus
vesiculosus L. in the brackish Baltic Sea correlates with osmotic capabilites of Baltic gametes.
Oecologia 00:00-00.
Sharp, G., P. Ang, Jr., and D. MacKinnon. 1994. Rockweed (Ascophyllum nodosum (L.) Le Jolis) harvesting
in Nova Scotia, Canada: its socioeconomic and biological implications for coastal zone management.
In: Wells, P.G. and P.J. Ricketts (eds.). Coastal Zone Canada' 94, Cooperation In The Coastal Zone.
Conference Proceedings, Volume 4. Dartmouth, NS Canada Coastal Zone Canada Assoc. Pages
1632-1644.
Short, F.T. (ed.). 1992. The Ecology of the Great Bay Estuary, New Hampshire and Maine: An estuarine
profile and bibliography. NOAA - Coastal Ocean Program Publ. 222 pp.
Sideman, E.J. and A.C. Mathieson..1983. Ecological and genecological distinctions of a high intertidal dwarf
form of Fucus distichus (L.) Powell in New England. J. Exp. Mar. Biol. Ecol. 72: 171-188.
Smith, B.D., and F.E. Foreman. 1984. An assessment of seaweed decomposition within a southern Strait of
Georgia seaweed community. Mar. Biol. 84: 197-205.
Smith, K.L., G.W. Bryan, and J.L. Harwood. 1985. Changes in endogenous fatty acids and lipid synthesis
associated with copper pollution in Fucus spp. J. Exp. Bot. 36: 663-669.
Steele, R.L. 1977. Effects of certain petroleum products on reproduction and growth of zygotes and juvenile
stages of the alga Fucus edentatus de La Pyl (Phaeophycear: Fucales). In: Wolfe, D.A. (ed.). Fate
And Effects Of Petroleum Hydrocarbons In Marine Ecosystems And Organisms, Proc. Symp.
November 10-12, 1976, Seattle, Washington. Pergamon Press, NewYork. Pages 138-142.

54
Steele, R.L. 1978. Sensitivity of some brown algal reproductive stages to oil pollution. In: Jensen, A. and
J.R. Stein (eds.). Proceedings Of The Ninth International Seaweed Symposium, Santa Barbara,
California, 20-27 August 1977. Science Press, Princeton, New Jersey. Pages 181-190.
Steveni, C.M., J. Norrington Davies, and S.D. Hankins. 1992. Effect of seaweed concentrate on
hydroponically grown spring barley. J. Appl. Phycol. 4: 173-180.
Str6mgren, T. 1977. Short-term effects of temperature upon the grovwth of intertidal Fucales. J. Exp. Mar.
Biol. Ecol. 29: 181-195.
Str6mgren,T. 1980a. The effect of dissolved copper on the increase in length of four species of intertidal
fucoid algae. Mar. Environ. Res. 3: 5-13.
Str6mgren,T. 1980b. The effect of lead, cadmium, and mercury on the increase in length of five intertidal
Fucales. J. Exp. Mar. Biol. Ecol. 43:107-119.
Str6mgren, T. 1985. Effect of light on apical length growth of Fucus distichus ssp. edentatus (Phaeophyta).
Mar. Biol. 86: 263-269.
Str6mgren, T. and M.V. Nielsen. 1986. Effect of diurnal variations in natural irradiance on the apical length
growth and light saturation of growvth in five species of benthic macroalgae. Mar. Biol. 90: 467- 472.
Taylor, W.R. 1957. Marine algae of the northeastemrn coast of North America. University of Michigan Press,
Ann Arbor, Michigan. 509 pages.
Terry, L.A. and B.L. Moss. 1980. The effect of photoperiod on receptacle initiation in Ascophyllum nodosum
(L.) Le Jol. Br. Phycol. J. 15: 291-301.
Thayer, G.W., W.J. Kenworthy, and M.S. Fonseca. 1984. The ecology of eelgrass meadows of the Atlantic
coast: a community profile. US Fish & Wilfl. Serv. FWSIOBS-84124. 85 pp.
Thomas, M.L.H. 1994. Littoral communities and zonation on rocky shores in the Bay of Fundy, Canada: an
area of high tidal range. Biol. J. Linn. Soc. 51: 149-168.
Thomas, M.L.H. and F.H. Page. 1983. Grazing by the gastropod, Lacuna vincta, in the lower intertidal area
at Musquash Head, New Brunswick, Canada. J. Mar. Biol. Assoc. U.K. 63: 725-736.
Thomas, M.L.H., D.C. Amold, and A.R.A. Taylor. 1983. Rocky intertidal communities. In: Thomas, M.L. H.
(ed.). Marine and Coastal Systems of the Quoddy Region, New Brunswick. Canadian Special
Publications of Fisheries and Aquatic Sciences, No. 64. Pages 35-73.
Thursby, G., R. Steele, M. Tagliabue, and C. Sheehan. 1990. Sexual reproduction in species of brown
seaweed, Fucus, to assess damage and recovery form the World Prodigy oil spill. In: Spaulding, M.L.
and M. Reed (eds.). Oil spills: Management and Legislative Implications. Proc. Conf., Newport, R.I.,
May 15-18, 1990. Pages 291-301.
Topinka, J.A. and J.V. Robbins. 1976. Effects of nitrate and ammonium enrichment on growth and nitrogen
physiology of Fucus spiralis. Limnol. Oceanog. 21: 659-664.
Tort, M.J. 1993. Environmental variable influencing species composition and abundance of intertidal fish
assemblages in a north-temperate estuary. M.S. thesis. University of Maine, Orono, Maine.
Vadas, R.L. 1972. Marine algae. Third annual report, Maine Yankee Atomic Power Company, pages 250-
310.
Vadas, R.L., Sr. and R.W. Elner. 1992. Plant-animal interactions in the north-west Atlantic. In: John, D.M.,
S.J. Hawkins and J.H. Price (eds.). Plant-Animal Interactions in the Marine Benthos. Systematics
Association Special Volume No. 46, Clarendon Press, Oxford. Pages 33-60.
Vadas, R.L., Sr., S. Johnson, and T.A. Norton. 1992. Recruitment and mortality of early post-settlement
stages of benthic algae. Br. Phycol. J. 27: 331-351.

55
Vadas, R.L, M. Keser, and P.C. Rusanowski. 1976. Influence of thermal loading on the ecology of intertidal
algae. In: Esch, G. and R. MacFarlane (eds.). Thermal ecology. II. AEC Symposium Series (Conf-
750425). Augusta, GA. Pages 202-212.
Vadas, R.L and W.A. Wright. 1986. Recruitment, growth, and management of Ascophyllum nodosum. In:
Westermeier, R. (ed.). Actas Segundo Congreso Nacional Sobre Algas Marinas Chilensas.
Universidad Austral de Chile, Valdivia. Pages 101-113.
Vadas, R.L, W.A. Wright, and S.L. Miller. 1990. Recruitment of Ascophyllum nodosum: wave action as a
source of mortality. Mar. Ecol. Prog. Ser. 61: 263-272.
Van Alstyn, K.L. 1988. Herbivore grazing increases polyphenolic defenses in the intertidal brown alga Fucus
distichus. Ecology 69: 655-663.
van den Hoek, C. 1975. Phytogeographical provinces along the coasts of the northern Atlantic Ocean.
Phycologia 14: 317-330.
Vogt and Schramm. 1991. Conspicuous decline of Fucus in Kiel Bay (westem Baltic): what are the causes?
Mar. Ecol. prog. Ser. 69:189-194.
Watson, D.C. and T.A. Norton. 1985. Dietary preferences of the common periwinkle, Littorina littorea (L.). J.
Exp. Mar. Biol. Ecol. 88: 193-211.
Watson, D.C. and T.A. Norton. 1987. The habitat and feeding preferences of Littoina obtusata (L.) and L.
maniae Sacchi et Rastelli. J. Exp. Mar. Biol. Ecol. 112: 61-72.
Whyte, J.N.C. 1988. Extraction of alginic acid from a brown seaweed. In: Lobban, C.S., D.J. Chapman, and
B.P. Kremer (eds.) Experimental Phycology: a Laboratory Manual. Pages 168-173.
Wright, P.J. and R.H. Reed. 1990. Effects of osmotic stress on gamete size, rhizoid initiation and germling
growth in fucoid algae. Br. Phycol. J. 25:149-155.
Zieman, J.C. 1982. The ecology of the seagrasses of South Florida: a community profile. US Fish Wildl.
Serv. Biol. Serv. Program FWS/OBS-82/25. 124. 26 pp.

56
KELPS

IDENTIFICATION AND DESCRIPTION

Seven species of brown laminarian algae (Agarum cribrosum, Alania esculenta, Chorda filum, Laminaria
digitata, L. Iongicruris, L. saccharina, and Saccorhiza dermatodea) are found in the Gulf of Maine, and are
the dominant fleshy macrophytes (Vadas and Elner 1992). A. cribrosum, L. digitata, and L. Iongicruris are
the most important species in terms of biomass and percent cover of substrate. An aerial survey and
supporting groundtruth surveys of kelps in southwestern Nova Scotia revealed an average biomass of 42.2
t/ha that was 60% L. Iongicruris and 40% L. digitata (Sharp and Carter 1986). In St. Margaret's Bay in Nova
Scotia, A. crTibrosum, L. digitata, and L. Iongicruris accounted for at least 72% of the total biomass of
intertidal and subtidal seaweeds (Mann 1972a).

Algae are relatively simple in structure. They lack well-developed tissue systems, and have no true roots,
stems or leaves. The seven species of laminarian algae described here, commonly referred to as kelps, are
dark brown due to the presence of fucoxanthin, a brown xanthophyll accessory pigment which tends to
obscure the greenish chlorophyll (LOning 1990).

Six of the species of kelps are plants of moderate to large size with a holdfast (which attaches to the
substrate), stalk, and greatly expanded flattened blade of various forms (Taylor 1957). A. esculenta has a
midrib, lateral bladelets belowthe main blade, and a costate (ribbed) main blade; A. cribrosum has a midrib
and a perforated blade; L. digitata has no midrib and a wide blade that is split into 5 to 30 straplike fingers;
L. Iongicruris has no midrib, a stalk which is distally hollow and expanded, and a solid blade; L. saccharina
has no midrib, a solid stalk and a solid blade; S. dermatodea has a cuplike holdfast, no midrib, and a solid
blade with tufts of hairs on its surface (Taylor 1957). C. filum, a slender whiplike plant without branches, is
not discussed further.

DISTRIBUTION

The six species of kelps occur primarily in the Arctic and cold temperate marine biogeography regions;
detailed maps of their distribution, the basis for this section, have been compiled by LCining (1990). A.
cribrosum is found along the western Atlantic Ocean from Greenland (80 N) to Cape Cod (42 N) and in the
Pacific Ocean from British Columbia to Korea. A. esculenta occurs on both sides of the Atlantic Ocean from
Greenland (78 N) to Cape Cod (42 N) and from the Barents Sea to the English Channel; it also is found
sporadically in the Pacific Ocean. L. digitata is restricted to both sides of the Atlantic Ocean from Greenland
(78 N) to Cape Cod (42 N) and from the Barents Sea to France. L. Iongicruris is endemic to the eastem
coast of North America. L. saccharina occurs in the Atlantic Ocean from north of 80 N to Long Island
Sound and Portugal (40 N) and in the Pacific Ocean from British Columbia to Korea. S. dermatodea is
restricted to the Atlantic Ocean and occurs from Greenland (70 N) to Cape Cod (42 N), along the coast of
Norway, and sporadically in the Barents Sea.

Kelp beds flourish in the Bay of Fundy and offshore island and ledges in the Gulf of Maine (Vadas and
Steneck 1988; Vadas and Elner 1992). Nearshore refuge areas that are free of seas urchins contain
numerous patches of kelps (Steneck 1986).

Recent review articles have reported clear vertical zonation of communities on subtidal rocky substrate
around the British Isles and in the Gulf of Maine, with a zone of kelps beginning in the subtidal fringe (level
of lowest spring tide to 1 m above) and extending to varying depths (Hiscock 1986; Sebens 1986). A.
esculenta often forms a band from the subtidal fringe to just belowthe low-water (Hiscock 1986; Edelstein et
al. 1969; Mann 1972a; LOning 1990). The slow-growing long-lived L. digitata and fast-growing short-lived L.
longicruris are found at intermediate depths (Luning 1990). L. sacchaina grows farther south and in more
protected areas (LOning 1990). A. cnribrosum is found at the greatest depths (Edelstein et al. 1969; Mann
1972a; Vadas and Steneck 1988; LOning 1990). Boden's (1979) quantitative survey off Appledore Island,

57
Maine, confirmed this zonation pattern. L. digitata dominated from 0 to 1 m, L. saccharina from 2 to 8 m,
and A. cribrosum below 11 m. However, in other studies, the vertical distribution of the six species of kelps
often showed considerable overlap wthen found at the same site (Edelstein et al. 1969; Mathieson 1979;
Mann 1972a; Sebens 1986; Vadas and Steneck 1988).
The depth distribution and species mix of kelps in the Gulf of Maine depends, in part, on available light,
available substrate, exposure, and the presence of sea urchins (Mathieson 1979). In the southem Gulf of
Maine (Massachusetts), a zone of L. saccharina, L. digitata, A. esculenta, and A. cnbrosum extended from
the sublittoral fringe to a depth of I to 3 m where sea urchins were present, but extended to at least 15 m
where urchins were scarce (Sebens 1986). At exposed sites in the Gulf of Maine, L. digitata and L.
saccharina dominated at 4 to 8 m, and removal experiments indicated they were limited to these depths by
sea urchins (VVitman 1987). Near Halifax, Canada, a zone of mixed L. saccharina, L. Iongicruris, L. digitata,
A. esculenta, and S. dermatodea extended to 15 m only where the substrate was stable and consisted of
rock and large boulders (Edelstein et al. 1969). In St. Margaret's Bay, Nova Scotia, Mann (1972a)
recognized four subtidal zones that were characterized by kelps. At exposed sites there was a mix of L.
digitata and L. Iongicruris from 3 to 13 m; at sheltered sites there was a bed of L. Iongicruris from 4 to 18 m;
a mix of L. Iongicruris and A. cribrosum occurred from 8 to 18 m, and A. cribrosurn was found at 15 to 20 m
(Mann 1972a).


LIFE CYCLE
The life cycles of the six species of kelps are similar, and altemate between a large asexual (sporophyte)
stage and a microscopic sexual (gametophyte) stage. The sporophyte cells are diploid, i.e., they contain
two of each type of chromosome characteristic of the species. At maturity, the sporophyte produces and
releases haploid spores, which contain one of each type of characteristic chromosome. After being released
into the water, each spore quickly attaches to the substrate and germinates into a microscopic gametophyte.
The gametophyte produces either eggs or sperm. Sperm are released into the water, and are attracted to
egg-containing gametophytes. In L. digitata, the release of male gametes is triggered by pheromones
secreted by the female gametes (Maier 1982). Fertilization of the egg results in an embryonic sporophyte.
Growth of the sporophyte occurs at the base of the blade where it joins the stipe, and involves an increase
in thallus length, width, and thickness (Mann 1972b). During the year, portions of the thallus may be tom off
as a result of wave action or ice scouring.
A single study of kelp reproduction, survival of early stages, and longevity has been conducted in the Gulf of
Maine in southem Nova Scotia (Chapman 1984). The study showed that the average density of L. digitata
(3.22/m2) was 2.5 times greater than that of L. longicruris (1.241m2), and yearly spore production of L.
digitata (20.02 X 10l) was 2.3 times greater than thatof L. Iongicruris (8.9 X 10). Spore production
occurred over a protracted period, but was greatest in October for both species. Recruitment of microscopic
plants onto a substrate occurred year round, but was greatest from February to April. Yearly recruitment
success of microscopic plants was 10 times greater for L. Iongicruris (8.89 X 106) than for L. digitata (0.98 X
10I), and recruitment of visible plants was enhanced for both species by removal of red algal turf species.
Mortality rate of macroscopic plants, although constant through time for each species, was greater in L.
Iongicruris. When a mixed assemblage of plants was examined, the maximum ages of L. longicruris and L.
digitata were approximately 2 and 4 years old, respectively (Chapman 1984). In a more recent study, in
which a single cohort was followed through time, some L. Iongicruris lived to 4 years of age (Chapman
1986).
Kelp reproduction pattems in Long Island Sound, the southem limit of kelp distribution on the east coast of
the United States, may or may not be similar to pattems in the Gulf of Maine. For instance, L. longicruris
reproduced in the late fall with the highest percentage of fertile plants occurring from October to December
(van Patten and Yadsh 1993), similar to the pattern described for Nova Scotia. However, L. sacchanina
plants appeared to be primarily annuals, reproducing in early spring with maximum spore release in May
(Lee and Brinkhuis 1986) and disintegrating between August and September (Lee 1984).

58
-Seasonal growth patterns of kelps have been shown to vary and to be affected by external factors such as
light, nutrients, and to a lesser extent, temperature. Recent experiments with L. cligitata from the North Sea
demonstrated that seasonal growth pattems are controlled by an endogenous circannual clock that is
synchronized by daylength cycles (Schaffelke and Laning 1994).
A characteristic growth pattern of kelps in the North Atlantic is rapid growth in late winter and early spring
followed by a quiescent period in late summer and early fall (Chapman and Lindley 1980). A. cribrosum, L
digitata, and L. Iongicruris in St. Margaret's Bay, Nova Scotia, demonstrated this characteristic pattern of
increasing growth rate from October to May and decreasing growth rate from May to October (Mann 1972b).
The same pattern was reported for L. Iongicruris in Long Island Sound (Egan and Yarish 1990), L.
saccharfina in Rhode Island (Brady-Campbell et al. 1984), and rope-cultured L. saccharina in Long Island
(Bnnkhuis et al. 1983). Field observations and laboratory experiments have shown that summer growth of
L. Iongicruris in Nova Scotia is severely limited by the low nitrate availability (Chapman and Craigie 1977,
1978).

Irradiance and nitrogen have been proposed as factors limiting kelp growth at different times during the year
(Brinkhuis et al. 1984). The growth of L. digitata and L. saccharina collected from the North Sea off
Scotland and cultured in the laboratory in nutdent enriched seawater in April to May (151LM nitrate; 3 pM
phosphate), June to July (7.5 ItM nitrate; 1.8 pM phosphate) and September (7.5 j.M nitrate; 3 ptM
phosphate) was about twice that of plants cultured without enriched seawater or growing in their natural
habitat (Conolly and Drew 1985b). However, in January the growth rate of L. saccharina cultured under
increased daylength (17.5 versus 7.5 h) was 3 to 4 times higher than plants cultured under conditions of
higher temperature or nutrient enrichment (Conolly and Drew 1985b). The growth rate of a different
population of L. digitata in the North Sea was nitrogen-limited between June and October, and was light-
limited with a possible temperature effect for the remainder of the year (Davison et al. 1984).
A different growth pattern may occur in areas where nutrient levels are generally high. In the Gulf of St.
Lawrence, growth of L. longicruris was maximal in June and minimal from December to February when ice
cover limited light levels (Anderson et al. 1981). In Cobscook Bay, where nutrients levels are high all year
and tidal energy prevents ice cover, growth of kelp continues year round, but is greatest in summer (Vadas
unpublished data). Similarly, blade and stipe growth of L. Iongicruris at twvo depths in Quebec did not appear
to be nitrogen limited, and showed an annual cycle characterized by a June peak (Gendron 1989). The
growth rate of mature L. saccharina at Appledore Island in Maine, transplanted to 7 depths between I and
21 m, was maximal at 9 m in July (Boden 1979). The growth of transplants at 1 and 3 m and in undisturbed
plants on shore at 0 m was 40% of the growth at 9 m and was attributed to low levels of nitrate and high
temperatures near the surface and attenuated light at depth (Boden 1979).
In some studies, growth rate is apparently unrelated to a single environmental factor. Machalek and
Davison (1992) studied a population of L. digitata growing in tidepools at Schoodic Point, Maine, and found
no apparent correlation between photosynthesis irradiance parameters and growth rate. However, nutrients
were not measured. In the North Sea, L. digitata and L. saccharina transplanted to three depths reduced or
ceased growth at different times of the year under otherwise identical conditions of temperature, nutrients,
light intensity or light quality suggesting that none of these factors actually trigger their seasonal growth
behavior (LOning 1979).
Numerous studies in St. Margaret's Bay, Nova Scotia, have produced an understanding of the internal
storage capacity of L. Iongicruris, resulting in an ability to extend its growth period when light energy or
dissolved nitrogen are in short supply (Hatcher et al. 1977; Chapman and Craigie 1978; Gagne and Mann
1981; Gagne et al. 1982). From late winter to eady spring when ambient nitrate levels were high, L.
Iongicruris grew rapidly and created an internal store of nitrogen. In the beginning of this growth phase, the
carbon produced in newtissue exceeded that fixed by photosynthesis, and carbon reserves (mannitol and
laminadin) were used for carbon and energy, but from January to May, growth was sustained by
photosynthesis. During and after the spring plankton bloom when ambient nitrate levels were low, kelp grew
at a decelerating rate and used up internal reserves of nitrate. During the summer when nitrate was not
available and photosynthetic rates were high, carbon reserves were built up and stored in the blade tissue.

59
Of the annual net assimilation of carbon, approximately 45% appeared as new blade tissue, 8% as new
stipe tissue, 12% as storage products (mannitol or laminarin) and 35% was assumed to be lost as dissolved
organic carbon (Hatcher et al. 1977).


POPULATION TRENDS
Two alternate community states exist in the rocky subtidal zone of the Gulf of Maine and the Atlantic coast
of Nova Scotia: highly productive kelp beds dominated by L. Iongicruris (Johnson and Mann 1988) and an
unproductive sea urchin-coralline alga community. Initially, the kelp beds are heavily grazed by populations
of sea urchins (Strongy(ocentrotus droebachiensis) that can increase dramatically in southwestern Nova
Scotia (Johnson and Mann 1988), but that increase more gradually elsewhere in the Gulf of Maine (Robert
Steneck pers. comm.). The sea urchin-coralline alga community then becomes dominant. Eventually,
disease induced mortality of green sea urchins (Johnson and Mann 1988) or urchin harvesting (Steneck et
al. 1995; Vadas and Steneck 1995) decreases the urchin population, reduces grazing pressure, and allows
regrowth of the kelp community.

Urchins destructively grazed kelps along the coast of Nova Scotia during the 1970s (reviewed in Vadas and
Elner 1992), however, A. cribrosum, a defended algae, was able to coexist with urchins in deeper zones
(Vadas and Elner 1992). In St. Margaret's Bay, Nova Scotia, natural and experimentally induced
regeneration of kelps Was destroyed by sea urchins under normal grazing pressure within 10 months
(Chapman 1981). Apparently, benthic microalgae with a mean biomass of 2.2 g C/im2 and productivity of 15
g C/m2/y supported a population of sea urchins large enough to suppress kelp regeneration, and re-
establishment of kelp beds seemed unlikely (Chapman 1981). Removal of sea urchins from a heavily
grazed site in Conception Bay, Newfoundland, initially resulted in an increase in the biomass of macroalgal,
but urchins grazed away all algae except A. cribrosum when urchin removal ceased (Gordon and Sand-
Jensen 1990).
Ultimately, an urchin die-off in the eady 1980s provided an opportunity to study the ability of L. Iongicruris to
recover its former dominant status and to compete with other seaweeds and wthen perturbed by storms and
grazers other than urchins. Recovery of seaweeds following overgrazing by urchins Was studied extensively
in St. Margaret's Bay, Nova Scotia. Rates of recolonization of L. Iongicruris depended on the proximity to a
refugial source of spores (Johnson and Mann 1988), resulting in the rapid re-establishment of a canopy at
some sites but not at others. Recovery by understory species also Was rapid (Johnson and Mann 1993), and
the recolonized communities resembled those documented at the same sites prior to the destructive grazing
(Edelstein et al. 1969; Johnson and Mann 1993).
Miller (1985) and Elner and Vadas (1990) reviewed the hypotheses relating to the destruction of kelp beds
by sea urchins that emerged, often with very little supporting data, during the 1970s and 1980s. Elner and
Vadas (1990) summarized four scenarios that were proposed during this time period. Briefly, two scenarios
predicted irreversible degradative spirals, ultimately resulting in a decline of system productivity and
predators. Two other scenarios predicted cycles which involved dramatic growth of the urchin population,
decimation of kelp beds by urchin grazing, decline of the urchin population, and recovery of the kelp beds.
In retrospect, a cyclic scenario was correct. Kelp populations were repressed for nearly 15 years, but
recovered quickly when urchin populations declined due to disease, changing environmental conditions, or a
combination of both factors.
Kelps also are affected by a "stipe blotch" disease, caused by Phycomelaina laminariae. This disease was
studied in infected L. saccharina population off Gloucester, Massachusetts. Susceptibility to infection
appeared to be part of the aging process, as infections occurred primarily in the larger (older) size groups
(Schatz 1984).

60
ECOLOGICAL REQUIREMENTS

Light
Sunlight is a fundamental requirement for the six species of kelps because light energizes the reactions of
photosynthesis. However, the amount and spectral characteristics of available light varies with latitude,
season, water type, and depth (reviewed in LOning 1990). As light passes through water and the substances
dissolved and suspended in it, different wavelengths of light are selectively absorbed, causing the total
amount of available light and the range of available wavelengths to decrease with depth. LOning and Dring
(1979) made continuous measurements of underwater light at different depths for over a year in the North
Sea, and related the available light to the lower limits of algal zones.
The relationship between photosynthesis and light intensity (irradiance) is described by a parabolic curve
(LOning 1990; Chapman 1995). The compensation irradiance (1,) is the minimum amount of light needed for
the plant to survive, i.e. at this light level the rate of photosynthesis equals the rate of respiration. As light
intensity increases above the compensation irradiance, the photosynthetic rate first increases in a linear
fashion, then plateaus as photosynthetic enzymes become light-saturated, and finally may decrease due to
photoinhibition (LOning 1990; Chapman 1995). For a particular plant, the photosynthesis-versus-light
relationship is the result of genetic adaptation and acclimation (Ramus 1981). The saturation irradiance for
photosynthesis was estimated to be 150 limol photon/m2/s for L. saccharina and L. digitata transplanted to
three depth in the North Sea (LOning 1979). Fortes and LOning (1980) estimated the saturation irradiance
for growth of L. saccharina cultured in the laboratory to be 70 limol photonlm2/s, and noted reduced growth
at higher light levels. Growth rates of small L. digitata plants in the laboratory increased with increasing
irradiance up to 50 pmol photon/ml2/s, at which point they became light saturated (Davison 1988). Th&se
values are typical for shade plants (LOning 1990). For young L. digitata, the light saturated rate of
photosynthesis was 50 to 100% higher in blue light than in red light, but at low irradiances the photosynthetic
rates were slightly higher in red than in blue light (Dring 1989).
Experiments have confirmed ecotypic differences in light-related traits. L. saccharina plants from turbid
Long Island Sound grewfaster under light-limiting and saturating daily irradiances, had higher
photosynthetic capacity and efficiency, low respiration rates, and higher surface area:weight ratios than
plants from shallow and deep habitats along the Atlantic coast of Maine (Gerard 1990).
Temperature

Experiments on young L. Iongicruris under crossed gradients of light (5 to 70 tmol photon/m2/s) and
temperature (50 to 25ï¿½C) indicated that temperature was the primary factor regulating reproduction and
vegetative growth and that the plants demonstrated seasonal acclimation responses (Egan et al. 1989).
Germination of spores at 5ï¿½C was greater in January than in July, but at 25ï¿½C was greater in July. Optimal
growth of gametophytes occurred at 100 to 15ï¿½C in March and 150 to 20ï¿½C in July. Growth of sporophytes
was optimal at 100 to 150C in all months (Egan et al. 1989). Other experiments on juvenile L. longicruris
under crossed gradients of temperature (5 0 and 15ï¿½C) and nutdent concentrations demonstrated
significantly greater growth at 15ï¿½C than 5ï¿½C at all nutrient concentration (Penniman et al. 1988).

L. saccharina on rope culturing systems in Long Island Sound grew slowly in winter when temperature
ranged from -1ï¿½ to 4ï¿½C (Brinkhuis et al. 1983). In one study, Davison (1987) demonstrated almost constant
photosynthetic rate in L. saccharina grown at temperatures between 0ï¿½ and 200C, and in a later study
reported that L. saccharina grown at 15 ï¿½C contained significantly more chlorophyll and had increased
photosynthetic efficiencies than similar plants grown at 5ï¿½C (Davison et al. 1991).

Gerard (1988) reported that genetic variation among populations of L. saccharina exists in relation to
several environmental factors, including light and temperature, and as a result of this variation, plants from
different populations exhibit differential rates of carbon assimilation, growth, and survivorship under
common garden conditions.

61
Salinity
It is generally believed that salinity is a major determinant of algal distribution because it can place osmotic
stress on the cells (Chapman 1995), but there has been little experimental work to substantiate the idea.
High salinity can potentially result in water loss, and low salinity can cause cells to swell or burst. In L.
digitata, intracellular inorganic (Na, K, CL, NO3) and organic (mannitol) substances are used to regulate
osmotic potentials of cells (Davison and Reed 1985).

Water motion
Mechanical stress imposed by currents and waves may result in hydrodynamic streamlining of blade
morphology without effecting productivity. Laboratory grown L. saccharina subjected to mechanical stress in
the form of constant longitudinal tension had significantly narrower blades and higher rates of blade
elongation than plants grown with no stress, however, the rates of biomass production were similar for the
two groups (Gerard 1987)
Blade elongation was compared in two populations of L. Iongicruris in Nova Scotia; one population was
exposed to and the other was sheltered from high intensity water movement. The maximum and minimum
elongation rates were similar for the two populations, but the sheltered plants grew more rapidly during 8
months of the year. Low concentrations of dissolved nutrients in summer and low levels of illumination in
fall and winter were more limiting to growth at the exposed site than the sheltered site (Gerard and Mann
1979).

Substrate

Along the Maine coast, the six species of kelps have been found attached to rock (>3m diameter), boulder
(10 inches to 3 m diameter), and cobble (2.5 to 10 inches diameter) substrate in the subtidal zone (Brown
1993).

Nutrients

Kelps take up nutrients through the mature blade because they do not have a root system. The nutrients
are transported to the meristem and used for growth or are stored in an internal pool for later use. In a
population of L. digitata growing in Scotland, approximately 70% of the nitrogen demand of the meristem
was supplied by nitrogen taken up by the blade and transported to the meristem (Davison and Stewart
1983). L. longicruris absorbed exogenous nitrate under both summer and winter conditions (Hardin and
Craigie 1978). Translocation of iodine occurs in L. saccharina and follows a 'source-to-sink" pattern i.e. it is
translocated from the blade to the meristem, or if located at the meristem it stays there (Amat and
Srivastava 1985).
The rate of uptake of nitrate (NO3) and phosphate (PO4) by L. digitata and L. saccharina collected from the
North Sea off Scotland and maintained in laboratory increased with increasing nutrient concentration up to
about 20 4M NO3 and 10 VM P04 (Conolly and Drew 1985b). At these concentrations, uptake rates were
approximately 24 jigN/g dry weight/h and 8 4gP/g dry weight/h (Conolly and Drew 1985b). Nitrite,
ammonium and urea were also taken up, independently of each other, and supported growth (Conolly and
Drew 1985b). Other laboratory experiments conducted on young L. saccharina from the North Sea
confirmed the apparent saturation between 10 and 20 VM NO3 and the development of internal reserves of
nitrate at extemal concentrations above 20 VM N03' (Chapman et al. 1978).
Studies of L. longicruris from nitrogen-poor (St. Margaret's Bay) and nitrogen-rich (Bay of Fundy) regions of
Nova Scotia indicated genetic differences in the populations. Plants from the nitrogen-poor site had lower
half-saturation values for growth, higher maximum specific growth rates, higher maximum uptake rates, and
a greater capacity to accumulate nitrogen in an internal tissue pool than plants from the nitrogen-rich site
(Espinoza and Chapman 1983).

62
ECOLOGICAL FUNCTION

Effect on water motion

In shallow water, dense canopies of macrophytes like kelps greatly reduce water movements throughout the
water column and increase sedimentation (Neushul 1972).

Nutrient cycling

Chemical elements in an ecosystem cycle back and forth between the living organisms and the physical
environment in a series of transfers termed biogeochemical cycles. The cycles involving carbon, nitrogen,
phosphorus, and the trace metals are extremely important because these elements are vital to the
maintenance of life. These elements make up the biomass of living organisms, and also exist in the
sediment and water column as inorganic compounds, particulate organic matter, and dissolved organic
matter.
Autotrophs alone absorb inorganic carbon, nitrogen, phosphorus, and trace metals, and use the sun's energy
to convert them into living organic matter (biomass). Important autotrophs in the Gulf of Maine include
single-celled organisms (e.g. bacteria, diatoms, dinoflagellates) that drift in the water column as plankton or
are associated with the substrate, macroalgae (e.g. rockweed, kelp, Irish moss) that usually are attached to
a substrate, and flowering plants (e.g. salt marsh grasses and eelgrass) that are rooted in the substrate and
inundated to various degrees. Autotrophic biomass is the ultimate source of nutrients and energy for
heterotrophs (bacteria, other single-celled organisms, and multi-celled animals). Heterotrophs in tum
excrete waste products containing inorganic nutrients which are absorbed and utilized by autotrophs.
Kelp beds are important in the cycling of nutrients because they 1) can produce large quantities of organic
matter, 2) harbor a diverse community of heterotrophs that consume organic matter and release inorganic
nutrients, and 3) are able to absorb nutrients from the water column. Aerobic heterotrophs that are attached
to kelp blades or the substrate or that are in the water column consume dissolved and particulate organic
matter and excrete inorganic waste products (ammonium and phosphorus) into the water column. . Kelp
plants are able to absorb macronutrients including nitrogen (as ammonia, ammonium, or nitrate) and
phosphorus, micronutrients (e.g., elemental metals), and other elements such as bromine and iodine
(Pedersen and Romans 1983) through the blade. Kelp beds, with their high rates of accumulation, may act
as sinks for nitrogen.

Kelps may be useful in reducing nutrient loading because of its high nitrogen requirements for growth.
Subander et al. (1993) reported that kelp grown in the effluent of salmon aquaculture pens removed 26 to
40% of the incoming dissolved inorganic nitrogen during light periods. Uptake was greatest at the highest
flow rates and at the lowest kelp density where no shading occurred. Conolly and Drew (1985a) studied the
effect of a coastal eutrophication gradient (dissolved nitrate, nitrite, ammonium and phosphate on the
growth and tissue composition of L. digitata and L. saccharina in St. Andrews, Scotland. The growth rates
were considerably enhanced at the eutrophic stations in spring and summer.

Toxic accumulations
There is little information available regarding accumulation of toxics by kelps. All the tissues of L.
fongicruris collected from Long Island Sound had consistently lower concentrations of cadmium than copper
for all months of the year-long study, and new blade tissue had significantly higher trace metal
concentrations than other tissues for the months of June and October (Shimshock et al. 1992).
Food source
Several investigators have made estimates of kelp productivity. An annual net carbon budget, determined
for a typical mature L. Iongicruris plant in Canada, revealed annual net assimilation was 6 to 8 mg C/cmr of
blade surface or 71 cal/cm2 (Hatcher et al. 1977). A seasonal carbon budget for a L. saccharina population
in Scottish showed net annual production in excess of 120 g C/m2/y (Johnston et al. 1977). Miller et al.

63
(1971) estimated that the production of seaweeds in St. Margaret's Bay was greater than 7000 kcallm2lyr,
and that it far exceeded the ingestion rate by herbivores. In Long Island Sound, L. longicruris production
was 10.6 kg fresh weight/m2 in 1986, and was estimated to have been 46 to 50 kg/m2 in 1987 (Egan and
Yarish 1990).
Grazing on live kelp
The most important and destructive grazer on living kelp in the Gulf of Maine is the green sea urchin, S.
droebachiensis. In the early 1970s, population increases of sea urchins resulted in intensive grazing of kelp
and the conversion of kelp forests into coralline barrens along the Atlantic coast of Nova Scotia (Mann
1977). In 1968, the rocky floor of St. Margaret's Bay was covered with a dense bed L. Iongicruris, L.
digitata, and A. cribrosum to a depth of approximately 20 m (Mann 1982). The average urchin biomass in
the kelp was 150 gIm2, but there were barren patches where kelp had been completely removed and urchin
biomass averaged 1200 g/m2 (Mann 1982). These barrens expanded due to urchin grazing, and by 1978
only 10% of the original cover of kelp remained (Mann 1982).
After sea urchin populations declined in the rocky subtidal area of the Gulf of Maine and dense beds of
canopy-forming L. Iongicruris grew back, the major grazers on the kelps were snails (Lacuna vincta)
(Johnson and Mann 1986) and perhaps amphi pods (Robert Steneck pers. comm.). Johnson and Mann
(1986) also reported that limpets and chitons grazed on kelps, but Steneck (1982) has shown that limpets
graze on coralline algae. The presence of anti-herbivore chemicals in the tissue and variations in toughness
and nutritional quality of tissues may have reduced grazing by snails (Johnson and Mann 1986).

Studies of carbon isotope ratios can be used to determine the importance of marine autotrophs in coastal
food webs. A study of animals occurring in eelgrass and kelp beds in St. Mary's Bay, Nova Scotia, indicated
that most herbivores, the filter feeders, and their predators derived their carbon from kelps, and only the
herbivorous gastropod L. vincta utilized eelgrass more (Stephenson et al. 1984; Stephenson et al 1986).

Consumption of kelp-derived detritus
Detritus refers to dissolved organic matter, particulate organic matter, and the assemblage of decomposing
bacteria and fungi that inhabits the organic matter (Zieman 1982; Thayer et al. 1984; Mann 1988).
Dissolved organic matter includes soluble carbohydrates and proteins (Zieman 1982; Mann 1988);
particulate organic matter includes dead autotrophs or autotroph debris, dead heterotroph or heterotroph
debris of any size, fecal pellets, and combinations of these items (Cousins 1980).
As part of the natural life cycle, kelp blades are eroded at the end by wave action and fall to the bottom or
are carried away by water currents (Mann 1972b). Whole blades probably become fragmented by the
shredding action of herbivorous invertebrates and by physical factors such as ice grinding. The intact
blades or fragments become colonized by bacteria and fungi which begin to decompose the algae.
The processes of particle size reduction and microbial colonization and decomposition are essential to the
detrital food web, and have been described in detail for algae by Mann (1988). Particle size reduction is
important because some organisms are able to ingest only fine particles, and the greater surface to volume
ratio of fine particles allows greater colonization by microbes. The microbes enzymatically break down plant
material that most organisms cannot digest, and convert it to microbial biomass which can be assimilated.
Microbes also enrich the detritus by taking up nitrogen and phosphorus from the surrounding medium and
incorporating it into microbe biomass.
Carbon budgets have been developed for various populations of kelps, and permit estimates of the amount
of biomass that enters the detrital cycle. Hatcher et al. (1977) estimated that 35% of the annual net carbon
of L. Iongicruris in Nova Scotia and was lost as dissolved organic carbon. Johnston et al. (1977) estimated
that over 13% of gross carbon production was lost as extracellular secretions, and 40 to 45% was lost by
distal decay and entered detdtal food chains. Miller et al. (1971) estimated the production of seaweeds in
St. Margaret's Bay and the ingestion rate of herbivores, and concluded that most of the seaweed production
was exported out of the coastal system in the form of suspended particulate matter.

64
Habitat
All types of macroalgae, including kelps, add structural complexity to the substrate and water column and
offer a variety of habitats for flora and fauna of all sizes (Hacker and Steneck 1990). Blades of varying age
and size extend into the water column, and change their orientation with passing waves and as current
changes. Epiphytes on the blades add additional complexity. Using a functional approach, the flora and
fauna associated with intertidal rockweeds include: 1) epiphytes on the thallus), 2) biota that attach to the
holdfast and thallus, 3) mobile fauna that move on or swim over the thallus and 4) biota that dwell on or in
the underlying substrate.
There is very little data regarding the use of A. esculenta, A. cribrosum, L. digitata, L. longicruris, L.
saccharina, and S. dermatodea habitat by invertebrates and fishes on the east coast of the United States.
However, on the west coast of the United States the giant kelp (Macrocystis pyrifera) forms dense forests,
and use of this habitat has been documented. Research has shown that many species of fishes use giant
kelp habitat (Feder et al. 1974). The abundance and diversity of fish species declined after experimental
removal of the plants (Bodkin 1988), and the density of many fishes, especially early life stages, were
significantly and positively related to plant density (De Martini and Roberts 1990). The presence of giant
kelp enhanced the local abundance of some species of temperate reef fishes and reduced the abundance of
others (Carr 1989).
In the Gulf of Maine, 12 species of fishes and 51 species of invertebrates have been collected from kelp
beds or have been observed in kelp beds (Table 1). Sampling was conducted in Passamaquoddy Bay
(Logan et al. 1983) and near Pemaquid Point (Robert Steneck pers. comm). Logan et al. (1983) termed the
area they sampled a subtidal crustose-coralline algae community because the encrusting coralline algae
often covered 100% of the hard substrate. However, several species of kelps (Agarum, Alaria, and
Laminaria) were listed as abundant or common, and the natural community they described is a kelp
community.
In Maine, habitats with low disturbance potential (i.e., low herbivore grazing) and high productivity potential
(i.e., well mixed water and sufficient light) are dominated by kelps and have high algal biomass and
functional group diversity (Steneck 1994). Habitats with higher disturbance potential or lower productivity
potential are dominated by crustose algae and have low algal biomass and functional group diversity
(Steneck 1994).
Bologna and Steneck (1993) investigated the use of kelp beds by the American lobster (Homarus
amencanus) in mid-coast Maine. Lobster population density and biomass were significantly higher in plots
of real and artificial (plastic) kelp than in adjacent non-kelp control plots, and the change in lobster density
was apparent the day following the beginning of the experiment (Bologna and Steneck 1993). Early benthic
stage lobsters are also found in kelp beds (Wahle and Steneck 1991).
One clear habitat-related use of macroalgae is as a substrate for numerous epifauna (Himmelman and
Lavergne 1985; Himmelman et al. 1983). Some fish, such as herring (Clupea harengus), use macroalgae
as egg-deposition sites (Cooper et al. 1975; Aneer 1989). Finally, juvenile Atlantic cod (Gadus morhua)
may be more abundant among kelp beds (Keats et al. 1987).
Commercial uses
Salts of alginic acid (alginates) occur in all brown algae as structural components of the cell walls. Alginates
are commercially extracted from Macrocystis pyrifera, A. nodosum, and Laminaria sp., and are used for
their thickening, stabilizing, film forming and gel producing properties (Whyte 1988). The products are used
extensively in the baking industry and as a stabilizer in the manufacture of ice cream. In Maine, kelp may
dragged for sea urchin feed (L. Mercer pers. comm.).

65
Table 1. Species of fishes and invertebrates collected from kelp beds and kelp habitat in northem and mid-
coastal Gulf of Maine.
Species
Fishes
Anarhichas lupus'
Cyclopterus lumPus
Gadus morhua
Hemitnripterus amencanus 3
Myxocephalus aenaeus
Myxocephalus scorpio1
Myxocephalus
octodecemspinosus
Pholis gunnellus
Pollachius virens
Prionotus carolinis
Stichaeus punctatus 1
Ulvaria subbifurcata '
Invertebrates
Acanthodoris pilosa 3
Acmaea testudinalis 3' 4
Alcyonium digitatum 3
Amphitritejohnstoni3
Asterias sp.3
Balanus sp. 3
Boltenia sp.3
Buccinum undatum 3
Cancersp. 3
Caprella sp. 3
Carcinus maenas 3
Chlamys islandicus 3
Colus stimpsoni 3
Coryomorpha pendula
Coryphella sp. 3

Cucumaria frondosa 3
Dendronotus sp. 3
Gersemia rubiformis 3
Haliclona sp.3
Halichondria sp. 3
Species
Invertebrates
Halocynthia pyrfornnis 3
Harmothoe sp. 3
Henricia sanguinolenta 3
Hiatella arctica 3
Homarus americanus 3
Hyas sp. 3
lophon pattersoni3
Ischnochiton Spp. 3
Lacuna vincta
Lepidonotus sp. 3
Littorina littorea 4
Obelia sp 3.
Onchidorus sp. 3
Margarites sp. 3
Metrindium senile 3
Modiolus modiolus 3
Molgula sp. 3
Mytilus edulis 3
Myxicola infundibulum 3
Neptunea decemcostata 3
Pagurus sp. 3
Pandalus montagui 3
Potamilla sp. 3
Psolus fabricii 3
Solaster endeca 3
Spirontocanrs spinus 3
Spirorbis sp. 3
Strongylocentrotus
droebachiensis 3, 4
Tealia felina 3
Tonicella rubra 4
Tubularia sp. 3
Common name
Common name
Wolffish
Lumpfish
Atlantic cod
Sea Raven
Grubby
Shorthom sculpin
Longhom sculpin

Rock gunnel
Pollock
Searobin
Arctic shanny
Radiated shanny

Nudibranch
Limpet
Octocoral
Terebellid wvrm
Blood star
Barnacle
Sea squirt
Whelk
Rock crab
Skeleton shrimp
Green crab
Iceland scallop
Whelk
Hydroid
Nudibranch

Sea cucumber
Nudibranch
Soft coral
Sponge
Sponge
Sea squirt
Scale vworm
Sea star
Boring clam
Lobster
Toad crab
Sponge

Chiton
Chink shell
Scale worm
Periwinkle
Hydroid
Nudibranch
Gastropod
Sea anenome
Horse mussel
Sea squirt
Blue mussel
Fanworm
Whelk
Hermit crab
Boreal red shrimp
Fanworm
Sea cucumber
Sea star
Caridean shrimp
Tubeworm
Sea urchin

Sea anenome
Chiton
Hydroid
Bryozoans3
1 = Robert Steneck pers. comm.; 2 = Keats et al.
1990; 5 = Johnson and Mann 1986.
1987; 3 = Logan et al. 1983; 4 = Schleibling and Raymond

66
CONCLUSION

Studies of the kelps have been conducted on both sides of the North Atlantic Ocean, and a great deal of
work has been done in Nova Scotia, particularly St. Margaret's Bay. Few studies have examined the
ecological functions of these species, but collectively they have shown that the five species of rockweed 1)
contribute to nutrient cycling, 2) provide food to grazers and detritus feeders, 3) help maintain water quality,
and 4) provide habitat for numerous species including commercially and recreationally important species.


THREATS

The major threats to kelps are destructive overgrazing by sea urchins, reproductive inhibition due to
petroleum hydrocarbons, competition from nonnative species, chlorate effluent from pulp mills, local
overharvesting, and toxic effects of heavy metals.

Destructive overgrazing of kelps by sea urchins appears to be sporadic problem that can last for at least 15
years. In Maine, sea urchins are harvested commercially, and in the past 10 years urchins have become
the state's second most valuable marine harvest. However, in the past two years the catch has declined
steadily, an indication of overfishing. A sustainable sea urchin harvest would not only provide jobs for
fishermen, but would be a means of protecting kelp habitat.

The reproductive stages of Laminaria are especially sensitive to oil pollution. In laboratory experiments,
more than 20 ppb of No. 2 fuel oil prevented Laminaria spores from germinating, and growth of male
gametophytes was decreased by 64% in the presence of 2 ppb (Steele 1978). Petroleum products had less
effect on female gametophytes and gametophytes more than 21 days old (Steele 1978). There was no
evidence that adult L. saccharina and L. digitata were detrimentally affected by the release of No. 2 fuel oil
following the grounding of the tanker World Prodigy on Brenton Reef (Peckol et al. 1990). However, in situ
primary production of L. saccharina was significantly inhibited by all types and concentrations of oil tested in
Canada including crude oils and the oil dispersant Corexit (Hsiao et al. 1978).

Nowthat populations of kelps in the Gulf of Maine have recovered from ove.rgrazing by sea urchins, they
are threatened by a non-native species of lacy crust bryozoan (Membranipora). The bryozoan colonizes and
can completely cover a kelp blade during the growing season. The encrusting bryozoan makes the kelp
blade inflexible and more likely to break when water motion increases (Robert Steneck pers. comm.).

In the Baltic Sea, the abundance and biomass of brown (fucoid) algae declined regionally after cellulose
pulp mills started releasing chlorate (CLO3) in their effluent (Lehtinen et al. 1988; Kautsky 1992). The
chlorate is formed as a by-product in the bleaching of chemical pulp (Lehtinen et al. 1988). Laboratory tests
indicated that chlorate, effective at 10 to 20 jig/L, was the main causal agent for the decline of the algae
(Rosemarin et al. 1985). All species of brown algae tested were highly sensitive to chlorate, while
cyanobacteria, green algae, red algae, and Zostera had much lower sensitivities (Rosemarin et al. 1985,
1994).

Overharvesting is a potential threat to kelps and their habitat. Current harvesting in Maine is minimal, but if
the price for kelps increases sufficiently, harvest would probably increase. In Canada, harvesting of
macroalgae, including Laminaria, by the dragrake method resulted in a by-catch of associated biota and
alteration of the abiotic structure of the habitat (Pringle and Sharp 1980). Unfortunately, there is little data
regarding the impact of harvesting on kelp recruitment, kelp regrowth, and on the assemblage of organisms
that use kelp habitat. Population variables and the understory community composition were monitored
before and after an experimental total harvest of L. Iongicruris and L. digifata in Nova Scotia. Both plots
were characterized by high kelp standing crop that had not been affected by recent sea urchin grazing. In
the deeper plot (3 to 4 m below MSL), L. Iongicruris attained maximum observed abundance within one
year, and both species required two years to mature to pre-harvest population characteristics (Smith 1985).

67

Copper, a heavy metal, appears to have an effect on kelps. L. saccharina plants were cultured in various
concentrations of copper. The release of spores was reduced at concentrations greater than or equal to 50
pg/I, development of early stages was delayed in concentrations greater than or equal to 5 pg/I, and growth
of adult plants was inhibited at concentrations greater than 10 pg/1 (Chung and Brinkhuis 1986)

68
REFERENCES
Amat, M.A. and L.M. Srivastava. 1985. Translocation of iodine in Laminaria saccharina (Phaeophyta). J.
Phycol. 21: 330-333.
Anderson, M.R., A. Cardinal, and J. Larochelle. 1981. An altemnate growth pattern for Laminaria Jongicruris.
J. Phycol. 17: 405-411.
Aneer, G. 1989. Herring (Clupea harengus L.) spawning and spawning ground characteristics in the Baltic
Sea. Fish. Res. 8: 189-195.
Boden,G.T. 1979. The effect of depth on summer growth of Laminaria saccharina (Phaeophyta,
Laminariales). Phycologia 18: 405-408.
Bodkin, J.L. 1988. Effects of kelp forest removal on associated fish assemblages in central Califomia. J.
Exp. Mar. Biol. Ecol. 117: 227-238.
Bologna, P.A.X and R.S. Steneck. 1993. Kelp beds as habitat for American lobster Homerus americanus
Mar. Ecol. Prog. Ser. 100: 127-134.
Brady-Campbell, M.M., D.B. Campbell, and M.M. Hariin. 1984. Productivity of kelp (Laminaria spp.) near
the southern limit in the northvwestem Atlantic Ocean. Mar. Ecol. Prog. Ser. 18: 79-88
Brinkhuis, B.H., H.G. Levine, and S.L. Williams. 1984. Alternative alternates: factors affecting growth of
Laminaria saccharina (Phaeophyceae) at its southern limit of distribution. J. Phycol. 20 suppl. 34.
Brinkhuis, B.H., V.A. Breda, S. Tobin, S., and B.A. Macler. 1983. NewYork marine biomass program-
culture of Laminaria saccharina. J. World Maricult. Soc. 14: 360-379.
Brown, B. 1993. A classification system of marine and estuarine habitats in Maine: an ecosystem approach
to habitats. Part I. Benthic habitats. State of Maine Department of Economic and Community
Development, Maine Natural Areas Program. 61 pages.
Carr, M.H. 1989. Effects of macroalgae assemblages on the recruitment of temperate zone reef fishes. J.
Exp. Mar. Biol. Ecol. 126: 59-76.
Chapman, A.R.O. 1981. Stability of sea urchin dominated barren grounds following destructive grazing of
kelp in St. Margaret's Bay, eastern Canada. Mar. Biol. 62: 307-311.
Chapman, A.R.O. 1984. Reproduction, recruitment and mortality in two species of Laminaria in southwest
Nova Scotia. J. Exp. Mar. Biol. Ecol. 78: 99-109.
Chapman, A.R.O. 1986. Age versus stage: an analysis of age- and size-specific mortality and reproduction
in a population of Laminaria Iongicruris Pyl. J. Exp. Mar. Biol. Ecol. 97: 113-122.
Chapman, A.R.O. 1995. Functional ecology of fucoid algae: twenty three years of progress. Phycologia 34:
1-32.
Chapman, A.R.O. and J.S. Craigie. 1977. Seasonal growth in Laminaria Iongicruris: relations with dissolved
inorganic nutrients and internal reserves of nitrogen. Mar. Biol. 40: 197-205.
Chapman, A.R.O. and J.S. Craigie. 1978. Seasonal growth in Laminaria longicruris: relations with reserve
carbohydrate storage and production. Mar. Biol. 46: 209-213.
Chapman, A.R.O. and J.E. Lindley. 1980. Seasonal growth of Laminaria solidungula in the Canadian High
Arctic in relation to irradiance and dissolved nutrient concentrations. Mar. Biol. 57: 1-5.
Chapman, A.R.O., J.W. Markham, and K. Luning. 1978. Effects of nitrate concentration on the growth and
physiology of Laminaria saccharina (Phaeophyta) in culture. J. Phycol. 14: 195-198.
Chung, Ik Kyo and B.K. Brinkhuis. i986. Copper effects in early stages of the kelp, Laminaria saccharina
Mar. Pollut. Bull. 17: 213-218.

69


Conolly, N.J. and E.A. Drew. 1985a. Physiology of Laminaria. III. Effect of a coastal eutrophication gradient
on seasonal patterns of growth and tissue composition in L. digitata Lamour and L. saccharina (L.)
Lamour. Mar. Ecol. 6: 181-195.
Conolly, N.J. and E.A. Drew. 1985b. Physiology of Laminaria. IV. Nutrient supply and daylength, major
factors affecting growth of L. digitata and L. saccharina. Mar. Ecol. 6: 299-320.
Cooper, R.A., J.R. Uzmann, R.A. Clifford, and K.J. Pecci. 1975. Direct observations of herning (Clupea
harengus harengus) egg beds on Jeffreys Ledge, Gulf of Maine in 1974. International Commission for
the Northwest Atlantic Fisheries Research Document 75/93.
Cousins, S.H. 1980. A trophic continuum derived from plant structure, animal size, and a detritus cascade.
J. Theor. Biol. 82: 607-618.
Davison, l.R. 1987. Adaptation of photosynthesis in Laminaria saccharina (Phaeophyta) to changes in
growth temperature. J. Phycol. 23: 273-283.
Davison, I.R. 1988. The effect of irradiance on carbon and nitrogen metabolism in the brown alga Laminaria
digitata. J. Phycol. 24 Suppl: 17.
Davison, l.R., M.R. Greene, and E.J. Podolak. 1991. Temperature acclimation of respiration and
photosynthesis in the brown alga Laminaria saccharina. Mar. Biol. 110: 449-454.
Davison, l.R. and R.H. Reed. 1985. Osmotic adjustment in -Laminaria digitata (Phaeophyta) with particular
reference to seasonal changes in intemal solute concentrations. J. Phycol. 21: 41-50.
Davison, l.R. and W.D.P. Stewart. 1983. Occurrence and significance of nitrogen transport in the brown alga
Laminaria digitata. Mar. Biol. 77: 107-112.
Davison, l.R., M. Andrews, and W.D.P. Stewart. 1984. Regulation of growth in Laminaria digitata: use of in-
vivo nitrate reductase activities as an indicator of nitrogen limitation in field populations of Laminaria
ssp. Mar. Biol. 84: 207-217.
DeMartini, E.E. and D.A. Roberts. 1990. Effects of giant kelp (Macrocystis) on the density and abundance of
fishes in a cobble-bottom kelp forest. Bull. Mar. Sci. 46: 287-300.
Dring, M.J. 1989. Stimulation of light saturated photosynthesis in Laminaria (Phaeophyta) by blue light. J.
Phycol. 25: 254-258.
Edelstein, T., J.S. Craigie, and J. McLachlan. 1969. Preliminary survey of the sublittoral flora of Halifax
county. J. Fish. Res. Bd. Can. 26: 2703-2713.
Egan, B. and C. Yarish. 1990. Productivity and life history of Laminaria Iongicruris at its southern limit in the
western Atlantic Ocean. Mar. Ecol. Prog. Ser. 67: 263-273.
Egan, B., A. Vlasto, and C. Yarish. 1989. Seasonal acclimation to temperature and light in Laminaria
Iongicruris de la Pyl. (Phaeophyta). J. Exp. Mar. Biol. Ecol. 129: 1-16.
Elner, R.W. and R.L. Vadas. 1990. Inference in ecology: the sea urchin phenomenon in the North-west
Atlantic. Am. Nat. 136: 108-125.
Espinoza, J. and A.R.O. Chapman. 1983. Ecotypic differentiation of Laminaria longicruris in relation to
seawater nitrate concentration. Mar. Biol. 74: 213-218.
Feder, H.M., C.H. Turner, and C. Limbaugh. 1974. Observations on fishes associated with kelp beds in
southern Califomnia. Calif. Dep. Fish Game, Fish Bull. 160. 144 pages.
Fortes, M.D. and K. Luning. 1980. Growth rates of North Sea macroalgae in relation to temperature,
irradiance and photoperiod. Helgolander Meeresunters. 34: 15-29.
Gagne, J.A. and K.H. Mann. 1981. Comparison of growth strategy in Laminaria populations living under
differing seasonal patterns of nutrient availability. In: Levring, T. (ed.). 10th international seaweed
symposium proceedings. Pages 297-302.

70
Gagne, J.A., K.H. Mann, and A.R.O. Chapman. 1982. Seasonal pattems of growth and storage in Laminaria
Iongicruris in relation to differing patterns of availability of nitrogen in the water. Mar. Biol. 69: 91-101.
Gendron, L. 1989. Seasonal growth of the kelp Laminaria Iongicruris in Baie des Chaleurs, Quebec, in
relation to nutrient and light availability. Bot. Mar. 32: 345-354.
Gerard, V.A. 1987. Hydrodynamic streamlining of Laminaria saccharina Lamour. in response to mechanical
stress. J. Exp. Mar. Biol. Ecol. 107: 237-244.
Gerard, V.A. 1988. Ecotypic differentiation in light-related traits of the kelp Lamrninaria saccharina. Mar. Biol.
97: 25-36.
Gerard, V.A. 1990. Ecotypic differentiation in the kelp Laminaria saccharina: phase-specific adaptation in a
complex life cyde. Mar. Biol. 107: 519-528.
Gerard,V.A. and K.H. Mann. 1979. Growth and production of Laminaria Iongicruris (Phaeophyta) populations
exposed to different intensities of water movement. J. Phycol. 15: 33-41.
Gordon, D.M. and K. Sand-Jensen. 1990. Effects of 02, pH and DIC on photosynthetic net 02 evolution by
marine macroalgae. Mar. Biol. 106: 445-451.
Hacker, S.D. and R.S. Steneck. 1990. Habitat architecture and the abundance and body-size-dependent
habitat selection of a phytal amphipod. Ecology 71: 2269-2285.
Harlin, M.M. and J.S. Craigie. 1978. Nitrate uptake by Laminaria longicruris (Phaeophyceae). J. Phycol. 14:
464-467.
Hatcher, B.G., A.R.O. Chapman, and K.H. Mann. 1977. An annual carbon budget for the kelp Laminaria
longicruris. Mar. Biol. 44: 85-96.
Himmelman, J.H. and Y. Lavergne. 1985. Organization of rocky subtidal communities on the St. Lawrence
Estuary. Le Naturaliste canadien 112: 143-154.
Himmelman, J.H., A. Cardinal, and E. Bourget. 1983. Community development following removal of
urchins, Strongylocentrotus droebachiensis, from the rocky subtidal zone of the St. Lawrence Estuary,
eastern Canada. Oecologia (Berlin) 59: 27-39.
Hiscock, K. 1986. Aspects of the ecology of rocky sublittoral areas. In: Moore, P.J. and R. Seed (eds). The
ecology of rocky coasts. Columbia University Press, NewYork. Pages 290-328.
Hsiao, S.l.C., D.W. Kittle, and M.G. Foy. 1978. Effects of crude oils and the oil dispersant Corexit on
primary production of arctic marine phytoplankton and seaweed. Environ. Pollut. 15: 209-221.
Johnson, C.R. and K.H. Mann. 1986. The importance of plant defense abilities to the structure of subtidal
seaweed communities: the kelp Laminaria Iongicruris de la Pylaie survives grazing by the snail Lacuna
vincta (Montagu) at high population densities. J. Exp. Mar. Biol. Ecol. 97: 231-267.
Johnson, C.R. and K.H. Mann. 1988. Diversity, patterns of adaptation, and stability of Nova Scotian kelp
beds. Ecol. Monogr. 58: 129-154.
Johnson, C.R. and K.H. Mann. 1993. Rapid succession in subtidal understory seaweeds during recovery
from overgrazing by sea urchins in eastern Canada. Bot. Mar. 36: 63-77.
Johnston, C.S., R.G. Jones, and R.D. Hunt. 1977. A seasonal carbon budget for a laminarian population in a
Scottish sea-loch. Helgolander Meeresunters. 30: 527-545.
Kautsky, H. 1992. The impact of pulp-mill effluents on phytobenthic communities in the Baltic Sea. Ambio
21: 308-313.
Keats, D.W., D.H. Steele, and G.R. South. 1987. The role of fleshy macroalgae in the ecology of juvenile
cod (Gadus morhua L.) in inshore waters off eastern Newfoundland. Can. J. Zool. 65: 49-53.
Lee, J.A. 1984. Seasonal reproduction patterns of Laminaria saccharina (L.) Lamour. (Phaeophyta) at its
southern limit of distribution in the western Atlantic Ocean. J. Phycol. 20 suppl: 3.

71
Lee, J. A. and B.H. Brinkhuis. 1986. Reproductive phenology of Laminaria saccharina (L.) Lamour.
(Phaeophyta) at the southern limit of its distribution in the northwestern Atiantic Ocean. J. Phycol. 22:
276-285.
Lehtinen, K.J., M. Notini, J. Mattsson, and L. Landner. 1988. Disappearance of bladder wrack (Fucus
vesiculosus L.) in the Baltic Sea: relation to pulp-mill chlorate. Ambio 17: 387-393.
Logan, A., A.W. McKay, and J.P.A. Noble. 1983. Sublittoral hard substrates. In: Thomas, M.L.H. (ed.).
Marine and Coastal Systems of the Quoddy Region, New Brunswick. Canadian Special Publications
of Fisheries and Aquatic Sciences, No. 64. Pages 119-139.
Luning, K. 1979. Growth strategies of three Laminaria species (Phaeophyceae) inhabiting different depth
zones in the sublittoral region of Helgoland (North Sea). Mar. Ecol. Prog. Ser. 1: 195-207.
Luning, K. 1990. Seaweeds: their environment, biogeography and ecophysiology. Wiley, NewYork. 527
pages.
Luning, K. and M.J. Dring. 1979. Continuous underwater light measurement near Helgoland (North Sea)
and its significance for characteristic light limits in the sublittoral region. Helgolander Meeresunters. 32:
403-424.
Machalek, K.M. and l.R. Davison. 1992. Seasonal changes in photosynthesis in the brown alga, Laminaria
digitata. J. Phycol. 28 Suppl: 16.
Maier, I. 1982. New aspects of pheromone-triggered spermatozoid release in Laminaria digitata
(Phaeophyta). Protoplasma 113: 137-143.
Mann, K.H. 1972a. Ecological energetics of the seaweed zone in a marine bay on the Atlantic coast of
Canada. I. Zonation and biomass of seaweeds. Mar. Biol. 12: 1-10.
Mann, K.H. 1972b. Ecological energetics of the sea-weed zone in a marine bay on the Atlantic coast of
Canada. II. Productivity of the seaweeds. Mar. Biol. 14: 199-209.
Mann, K.H. 1977. Destruction of kelp-beds by sea urchins: a cyclical phenomenon or irreversible
degradation? Helgolinder Meeresunters. 30: 455-467.
Mann, K.H. 1982. Kelp, sea urchins, and predators: a review of strong interactions in rocky subtidal systems
of eastem Canada, 1970-1980. Neth. J. Sea. Res. 16: 414-423.
Mann, K.H. 1988. Production and use of detritus in various freshwater, estuarine, and coastal marine
ecosystems. Limnol. Oceanogr. 33: 910-930.
Mathieson, A.C. 1979. Vertical distribution and longevity of subtidal seaweeds in northern New England,
U.S.A. Bot. Mar. 30: 511-520.
Miller, R.J. 1985. Seaweeds, sea urchins, and lobsters: a reappraisal. Can. J. Fish. Aquat. Sci. 42: 2061-
2072.
Miller, R.J., K.H. Mann, and D.J. Scarratt. 1971. Production potential of a seaweed-lobster community in
eastem Canada. J. Fish. Res. Bd. Can. 28: 1733-1738.
Neushal, M. 1972. Functional interpretation of benthic marine algal morphology. In: Abbot, J.A. and M.
Kurogi (eds.). Contributions to the Systematics of Benthic Marine Algae of the North Pacific. Japanese
Society of Phycology, Kobe. Pages 47-73.
Peckol, P., S.C. Levings, and S.D. Garrity. 1990. Kelp response following the World Prodigy oil spill. Mar.
Pollut. Bull. 21: 473-476.
Pedersen, M. and G.M. Romans. 1983. Ultrastructural localization of bromine and iodine in the stipes of
Laminaria digitata (Huds.) Lamour., Laminaria saccharina (L.) Lamour. and Laminaria hyperborea
(Gunn.) Fosl. Bot. Mar. 26:113-118.

72
Penniman, C., B. Egan, and C. Yarish. 1988. Comparative growth and tissue chemistry responses of
juvenile Laminaria Iongicruris sporophytes cultured in temperature and nitrogen concentration gradients.
J. Phycol. 24 Suppl: 27.
Pringle, J.D. and G.J. Sharp. 1980. Multispecies resource management of economically important marine
plant communities of eastern Canada. Helgolander Meeresunters. 33: 711-720.
Ramus, J. 1981. The capture and transduction of light energy. In: Lobban, C.S. and M.J. Wynne (eds.). The
biology of seaweeds. Blackwell, Oxford. Pages 458-492.
Rosemarin, A., K.J. Lehtinen, M. Notini, B. Axelsson, and J. Mattsson. 1985. Effects of pulp mill chlorate on
algae. Swedish Environmental research group (SERG). Report No. K5015:1.
Rosemarin, A., K.J. Lehtinen, M. Notini, and J. Mattsson. 1994. Effects of pulp mill chlorate on Baltic Sea
algae. Environ. Pollut. 85: 3-13.
Schaffelke, B. and K. LOning. 1994. A circannual rhythm controls seasonal growth in the kelps Laminaria
hyperborea and L. digitata from Helgoland (North Sea) Eur. J. Phycol. 29: 49-56.
Schatz, S. 1984. The Laminaria-Phycomelaina host-parasite association: seasonal patterns of infection,
grovwth and carbon and nitrogen storage in the host. Helgolander Meeresunters. 37: 1-4.
Scheibling, R.E. and B.G. Raymond. 1990. Community dynamics on a subtidal cobble bed following mass
mortalities of sea urchins. Mar. Ecol. Prog. Ser. 63: 127-145.
Sebens, K.P. 1986. Community ecology of vertical rock walls in the Gulf of Maine, USA: small-scale
processes and altemative community states. In: Moore, P.J. and R. Seed (eds). The ecology of rocky
coasts. Columbia University Press, New York. Pages 346-371.
Sharp, G.J. and J.A. Carter. 1986. Biomass and population structure of kelp (Laminaria spp.) in
southwestern Nova Scotia. Can. Manuscr. Rep. Fish. Aquat. Sci. No. 1907. 46 pages.
Shimshock, N, G. Sennefelder, M. Dueker, F. Thurberg, and C. Yarish. 1992. Patterns of metal
accumulation in Laminaria Iongicruris from Long Island Sound (Connecticut). Arch. Environ. Contam.
Toxicol. 22: 305-312.
Smith, B.D. 1985. Recovery following experimental harvesting of Laminaria Iongicruris and L. digitata in
southwestern Nova Scotia. Helgol. Meeresunters. 39: 83-101.
Steele, R.L. 1978. Sensitivity of some brown algal reproductive stages to oil pollution. In: Jensen, A. and
J.R. Stein (eds.). Proceedings Of The Ninth International Seaweed Symposium, Santa Barbara,
Califomia, 20-27 August 1977. Science Press, Princeton, NewJersey. Pages 181-190.
Steneck, R.S. 1978. Factors influencing the distribution of crustose coralline algae (Rhodophyta,
Corallinaceae) in the Damariscotta River, Maine. M.S. Thesis. University of Maine.
Steneck, R.S. 1982. A limpet-coralline algal association: adaptations and defenses between a selective
herbivore and its prey. Ecology 63: 507-522.
Steneck, R.S. 1986. The ecology of coralline algal crusts: convergence pattems and adaptive strategies.
Ann. Rev. Ecol. System. 17: 273-303.
Steneck, R.S. and M.N. Dethier. 1994. The structure of algal-dominated communities: a functional group
approach. Oikos 69: 476-498.
Steneck, R.S., D. McNaught and S. Zimsen. 1995. Spatial and temporal patterns in sea urchin populations,
herbivory and algal community structure in the Gulf of Maine. 1994 Workshop on the Management and
Biology of the Green Sea Urchin (Strongylocentrotus droebachiensis). Pages 34-73.
Steneck, R.S. and R.S. Vadas. 1995. Overfishing and inferences in kelp-sea urchin interactions. In: Skjodal,
H.R., C. Hopkins, K.j E. Erikstad, and H.P. Leinaas (eds.). Ecology of Fjords and Coastal Waters.
Elsevier Science B.V. Pages 509-524.
Stephenson, R.l., F.C. Tan, and K.H. Mann, K.H. 1984. Stable carbon isotope variability in marine
macrophytes and its implications for food web studies. Mar. Biol. 81: 223-230.

73
Stephenson, R.L., F.C. Tan, and K.H. Mann. 1986. Use of stable carbon isotope ratios to compare plant
material and potential consumers in a seagrass bed and a kelp bed in Nova Scotia, Canada. Mar. Ecol.
Prog. Ser. 30: 1-7.
Subandar, A., R.J. Petrell, and P.J. Harrison. 1993. Laminaria culture for reduction of dissolved inorganic
nitrogen in salmon farm effluent J. Appl. Phycol. 5: 455-463.
Taylor, W.R. 1957. Marine algae of the northeastern coast of North America. University of Michigan Press,
Ann Arbor, Michigan. 509 pages.
Thayer, G.W., W.J. Kenworthy, and M.S. Fonseca. 1984. The ecology of eelgrass meadows of the Atlantic
coast: a community profile. US Fish & Wilfl. Serv. FWSIOBS-84/24. 85 pp.
Vadas, R.L. and R.S. Steneck 1988. Zonation of deep water benthic algae in the Gulf of Maine. J. Phycol.
24: 338-346.
Vadas, R.L., Sr. and R.W. Elner. 1992. Plant-animal interactions in the north-west Atlantic. In: John, D.M.,
S.J. Hawkins and J.H. Price (eds.). Plant-Animal Interactions in the Marine Benthos. Systematics
Association Special Volume No. 46, Clarendon Press, Oxford. Pages 33-60.
van Patten, M.S. van and C. Yarish. 1993. Allocation of blade surface to reproduction in Laminaria
longicruris of Long Island Sound (USA). Hydrobiologia 260-261: 173-181.
Wahle, R.A. and R.S. Steneck. 1991. Recruitment habitats and nursery grounds of the American lobster
(Homarus americanus Milne Edwards): A demographic bottleneck? Mar. Ecol. Prog. Ser. 69: 231-243.
Whyte, J.N.C. 1988. Extraction of alginic acid from a brown seaweed. In: Lobban, C.S., D.J. Chapman, and
B.P. Kremer (eds.). Experimental Phycology: A Laboratory Manual. Pages 168-173.
Witman, J.D. 1987. Subtidal coexistence: storms, grazing, mutualism, and the zonation of kelps and
mussels. Ecol. Monogr. 57:167-187.
Zieman, J.C. 1982. The ecology of the seagrasses of South Florida: a community profile. US Fish Wildl.
Serv. Biol. Serv. Program FWS/OBS-82125. 124. 26 pp.