SEAGRASS RESTORATION:
Success, Failure, and the Costs of Both
Selected Papers
presented at a workshop
Mote Marine Laboratory
Sarasota, FL
March 11–12, 2003
S.F. Treat & R.R. Lewis III, editors
Published by
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June 2006
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CONTENTS
PREFACE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . iii
PERSPECTIVES ON TWO DECADES OF EELGRASS (ZOSTERA MARINA L.) . . . . . . . . . . . . . . . 1
RESTORATION USING ADULT PLANTS AND SEEDS IN CHESAPEAKE BAY
AND THE VIRGINIA COASTAL BAYS, USA
R.J. Orth, J. Bieri, J.R. Fishman, M.C. Harwell, S.R. Marion,
K.A. Moore, J.F. Nowak & J. vanMontfrans
EVALUATION OF THE SUCCESS OF SEAGRASS MITIGATION AT PORT MANATEE, . . . . . . . 19
TAMPA BAY, FLORIDA
R.R. Lewis III, M.J. Marshall, S.A. Bloom, A.B. Hodgson & L.L. Flynn
A SHALLOW WATER TECHNIQUE FOR THE SUCCESSFUL RELOCATION AND/OR . . . . . . . . 41
TRANSPLANTATION OF LARGE AREAS OF SHOALGRASS (HALODULE WRIGHTII)
G.J. Montin & R.F. Dennis III
SPECIES SELECTION, SUCCESS, AND COSTS OF MULTI-YEAR, MULTI-SPECIES . . . . . . . . . 49
SUBMERGED AQUATIC VEGETATION (SAV) PLANTING IN SHALLOW CREEK,
PATAPSCO RIVER, MARYLAND
P. Bergstrom
USING TERFS AND SITE SELECTION FOR IMPROVED EELGRASS . . . . . . . . . . . . . . . . . . . . 59
RESTORATION SUCCESS
F.T. Short, R.C. Davis, B.S. Kopp, J.L. Gaeckle & D.M. Burdick
SEAGRASS SCARRING IN TAMPA BAY: IMPACT ANALYSIS . . . . . . . . . . . . . . . . . . . . . . . . . . . 69
AND MANAGEMENT OPTIONS
J.F. Stowers, E. Fehrmann & A. Squires
ASSESSMENT OF A CONSTRUCTION-RELATED EELGRASS RESTORATION . . . . . . . . . . . . . 79
IN NEW JERSEY
P.A X. Bologna & M.S. Sinnema
EXPERIMENTAL HALODULE WRIGHTII AND SYRINGODIUM FILIFORME . . . . . . . . . . . . . . 91
TRANSPLANTING IN HILLSBOROUGH BAY, FLORIDA
W. Avery & R. Johansson
COSTS AND SUCCESS OF LARGE-SCALE EELGRASS (ZOSTERA MARINA L.) . . . . . . . . . . . . 103
PLANTINGS IN NEW ENGLAND (NEW HAMPSHIRE AND MAINE)
R.C. Davis, J.T. Reel, F.T. Short & D. Montoya
BISCAYNE BAY SEAGRASSES AND RECENT RESTORATION EFFORTS . . . . . . . . . . . . . . . . . 115
G.R. Milano & D.R. Deis
TOPOGRAPHIC RESTORATION OF BOAT GROUNDING DAMAGE AT . . . . . . . . . . . . . . . . . 131
THE LIGNUMVITAE SUBMERGED LAND MANAGEMENT AREA
P.L. McNeese, C.R. Kruer, W.J. Kenworthy, A.C. Schwarzschild,
P. Wells & J. Hobbs
CULTIVATION STUDIES OF THE HALOPHILA SEAGRASSES H. JOHNSONII . . . . . . . . . . . . 147
AND H. DECIPIENS
B. Baca, G. Stone & A. Sanchez-Gomez
SUITABILITY OF ALTERNATIVE SAV MEASUREMENTS AS AN INDICATOR . . . . . . . . . . . . . 155
OF WATER QUALITY EFFECTS, LOWER ST. JOHNS RIVER, FLORIDA
A.M. Steinmetz, D. Dobberfuhl & N. Trahan
IMPLEMENTATION, REGULATORY AND RESEARCH PRIORITIES . . . . . . . . . . . . . . . . . . . . . 165
FOR SEAGRASS RESTORATION: RESULTS FROM THE SEAGRASS
RESTORATION WORKSHOP, MARCH 11–12, 2003
H. Greening
WRAP-UP OF SEAGRASS RESTORATION: SUCCESS, FAILURE . . . . . . . . . . . . . . . . . . . . . . . . 169
AND LESSONS ABOUT THE COSTS OF BOTH
M. Fonseca
ii
PREFACE
The three day symposium, Submerged Aquatic Habitat Restoration in Estuaries: Issues,
Options and Priorities, was held at the Mote Marine Laboratory on March 11–13, 2003,
and was attended by over 200 scientists and managers.
I had the pleasure of helping organize and chair the all-day workshop on March 11
devoted to the topic of Seagrass Restoration: Success, Failure and the Costs of Both. This
volume contains the submitted, accepted and peer-reviewed manuscripts from that
session.
During my introductory remarks I noted that determining the actual cost of restoration
projects is not something biologists are prone to do. This is a task often left to engineers
and scientists. Restoration biologists, however, need to tackle this issue if they are to see
that limited restoration funding is well spent on documented successful projects, and if
they expect to see future restoration funding for their projects. There is a perception
among the so-called stake-holder groups, including ordinary citizens, fisherman and
politicians, that restoration does not work. This is one of the reasons that artificial reefs
have received millions of dollars from the collection of salt water fishing licenses in
Florida for “fishery enhancement” while seagrass restoration has received only a pittance.
Artificial reefs obviously “work,” or at least the media assures us they do, and many
believe they do. It is not so with most natural habitat restoration.
We all must participate in establishing a sound financial as well as ecological basis for
seagrass restoration, and that requires accounting for all costs. I noted in my presentation
that most cost accounting for restoration projects is of limited scope, and may only look
at the cost of plant materials and their planting, ignoring important cost items like project
design, permitting, construction and monitoring. All of these latter costs must be
included, as they are essential to carrying out the project.
We were fortunate to have with us at the meeting experienced seagrass restoration
scientists including Dr. Mark Fonseca, Dr. Bob Orth and Dr. Fred Short who have dealt
with these cost issues and welcomed the opportunity to share their experience with the
audience.
Thank you to all who attended and made presentations, and I hope the documentation
of these presentations will enable future restoration scientists to more cost effectively
conduct their projects.
Robin Lewis
Salt Springs, Florida
May 2006
iii
˜
A REVIEW OF TECHNIQUES USING ADULT PLANTS AND SEEDS TO
TRANSPLANT EELGRASS (ZOSTERA MARINA L.) IN CHESAPEAKE BAY
AND THE VIRGINIA COASTAL BAYS
R.J. Orth, J. Bieri, J.R. Fishman, M.C. Harwell,
S.R. Marion, K.A. Moore, J.F. Nowak & J. van Montfrans
ABSTRACT
In many areas of the Chesapeake Bay region, including the coastal bays of the Delmarva
Peninsula, eelgrass (Zostera marina L.) is much less widespread today than in the past due to the
eelgrass wasting disease in the 1930s and the more general seagrass population decline in the
1970s in Chesapeake Bay resulting from increasing nutrients and sediments entering the bays’
watershed. In 1978, an experimental eelgrass restoration program was initiated in lower
Chesapeake Bay as part of a larger research effort on the biology and ecology of eelgrass beds. In
this paper we provide an overview of both manual and mechanized techniques we have used in
efforts to restore eelgrass at a number of different locations using either adult plants or seeds,
highlighting the importance of the timing of transplanting, use of fertilizer, labor requirements,
and initial success. Much of the earliest transplant work was conducted in a variety of locations
with different vegetation histories and water quality characteristics to facilitate addressing
questions related to growth and habitat requirements.
We found that planting eelgrass in fall rather than spring was optimal because plants had a longer
growing period to become established. Addition of fertilizer to transplants increased plant density
but did not enhance the long-term survival. Techniques utilizing adult plants (e.g., mesh mats
with bare rooted shoots, sods and cores of seagrass and sediment, bundles of bare root shoots
with anchors, single shoots without anchors) were generally successful, with the manually
planted single shoot method being both successful and requiring the least time. Mechanized
planting with a planting boat had lower initial planting unit survivorship and did not result in
significant savings of time. Techniques using seeds (e.g., peat pots, seed broadcasting, burlap bags
to protect seeds) rather than adult plants had varying degrees of success with highest seedling
establishment noted where seeds were protected in burlap bags. Current issues with seeds deal
primarily with the low survival rate of seeds (generally between 5 and 10% of seeds establishing
as seedlings in field experiments). However, broadcast of seeds is one of the least labor-intensive
techniques used to date in our program and is currently proving successful in restoring eelgrass
to Virginia’s seaside coastal bays that have been unvegetated since the 1930s pandemic wasting
disease.
INTRODUCTION
Eelgrass (Zostera marina L.) populations in the Chesapeake Bay region are significantly
different today than in the recent past due to the pandemic eelgrass wasting disease of the
1930s (Cottam, 1934; Orth, 1976), changes due to the passage of Tropical Storm Agnes in
June 1972 (Orth and Moore, 1983a; 1984), and continued poor water quality due to high
anthropogenic nutrient and sediment inputs. Although eelgrass re-populated some regions
following both perturbations, many areas have either not recovered or remain sparsely
populated (Orth et al., 2002; Orth et al. 2006).
During the last 25 years, we have transplanted eelgrass using several techniques, with either
adult plants or seeds, to a number of sites around Chesapeake Bay (Fig. 1; Table 1). We have
used transplant ‘gardens’ to test various hypotheses regarding the influence of
environmental parameters (light, nutrients, suspended solids) and ecological processes (e.g.,
predation, seed interactions, habitat utilization) on survival and growth of eelgrass
1
Orth et al.
(Dennison et al., 1993; Orth et al., 1994, 2003; Moore et al., 1996; 1997; Lombana, 1996;
Fishman and Orth, 1996; Williams and Orth, 1998; Marion, 2002; Harwell and Orth,
2002a). Transplant ‘gardens’ have also been used successfully in developing efficient and
effective techniques for transplanting eelgrass (Orth et al., 1999, Fishman et al., 2004).
Figure 1. Locations of sites in Chesapeake Bay and the Virginia coastal bays where
transplant studies with adult plants and seeds were conducted.
2
Table 1. Summary of transplant projects conducted in Chesapeake Bay and the coastal bays by VIMS scientists, 1979–2003, using seeds or
adult plants with various techniques.
Eelgrass Transplant Techniques in Chesapeake Bay
3
Orth et al.
In this paper we provide an overview of those techniques we have used in efforts to restore
eelgrass, both manually and mechanized, highlighting the importance of the timing of
transplanting, use of fertilizer, labor requirements, and initial success of various transplant
techniques.
EELGRASS LIFE HISTORY CHARACTERISTICS
IN THE CHESAPEAKE BAY REGION
Eelgrass in Chesapeake Bay is near its southern range of distribution along the Atlantic coast
of the United States (Thayer et al., 1984). Chesapeake Bay populations, which grow in
water depths from just below mean low water (MLW) to -2.0 m (MLW) (Orth and Moore,
1988), are perennial and exhibit a bimodal growth period, with maximum growth and peak
biomass occurring in late May to early June. Beginning in June, temperatures above 25o C
and high light attenuation by the water column result in minimal growth and widespread
leaf defoliation (Moore et al., 1996; 1997). A second period of growth follows in midSeptember when water temperatures drop below 25o C. Shoot density increases rapidly, but
shoot biomass is less than the spring biomass peak. In winter, with temperatures below 10o
C, growth is at its minimum (Marsh, 1973; Orth and Moore, 1986; Moore et al., 1996).
Flowering begins in January, with anthesis occurring from late March to April, followed by
seed release from mid-May to early June (Silberhorn et al., 1983). Once released, seeds fall
rapidly to the sediment surface and remain within meters where they settle (Orth et al.,
1994). However, detached reproductive shoots with seeds can float many kilometers from
beds of origin providing a mechanism for long distance dispersal (Harwell and Orth,
2002a). Seeds remain dormant through summer. Germination occurs from mid-November
to December (Orth and Moore, 1983b; Moore et al., 1993). Seed banks are transient, lasting
no longer than six months (Harwell and Orth, 2002b).
METHODS AND RESULTS
Selection of Transplant Sites
Areas that either historically supported eelgrass beds or where eelgrass has exhibited some
recovery were used for transplant ‘garden’ experiments (Fig. 1), a key recommendation of
Fonseca et al. (1998) and Short et al., (2002). Transplanting was generally not conducted at
sites that had existing eelgrass. Much of the earliest transplant work was conducted in the
York River because this region provided a variety of locations with different vegetation
histories and facilitated addressing questions related to growth and habitat requirements
(Batiuk et al., 1993; Moore et al., 1996, 1997), while simultaneously minimizing logistical
constraints. Areas selected included those which had: (1) some eelgrass loss in the 1970s but
recovered almost completely to pre-1970s levels; (2) complete eelgrass loss in the 1970s
with moderate levels of recovery; and (3) complete eelgrass loss in the 1970s without
recovery, despite transplanting efforts (Moore et al., 1996).
Success of Transplant Techniques
Success of the different transplant techniques using adult plants or seeds was evaluated at
several time intervals within the first year of transplanting. In the first interval we evaluated
initial success of the transplant process. This usually entailed assessing percent survivorship
4
Eelgrass Transplant Techniques in Chesapeake Bay
of the planting units (PUs) after one to two months, allowing time for the planting unit to
become physically established prior to significant new growth. In the second evaluation
period, five to eight months after transplanting, we evaluated structural equivalency
(depending on project goals) of transplant plots with natural beds. Structural equivalency
was defined and evaluated through measures of bottom cover and shoot density. Success
in seed experiments was evaluated five to nine months after seeds were placed in the field,
allowing necessary time for germination and growth of seedlings to detectable size. Finally,
in the third evaluation period, nine to 12 months after transplanting, we continued to
quantify only structural equivalency with natural beds. Evaluation nine to 12 months after
planting ensures that the transplants go through one complete cycle of growth and
reproduction. While there are other measures for success (e.g. faunal colonization, sediment
changes, etc.) we concentrated only on primary plant characteristics over short temporal
scales (one year or less). While a much longer time frame would be necessary to evaluate
whether a transplant effort had succeeded in approximating the ecological function of a
natural seagrass bed, our concern here was solely to evaluate the utility of transplant
methods for establishing plants.
Transplanting With Adult Plants
Our initial work was conducted with adult plants (shoots with well-developed leaves and
attached rhizomes and roots were considered adult plants). We focused on evaluating
transplanting techniques, planting season, and fertilizer application on transplant success
primarily within the first year after transplanting.
Effects of Season on Transplant Success
The distinct annual growth cycle of eelgrass in Chesapeake Bay led us to hypothesize a
seasonal component to transplant success. Our initial experiments conducted in the York
River tested this hypothesis by planting 10 cm diameter cores of sediment and eelgrass in
small test plots (2 x 2 m on 0.5 m centers) in the spring, summer, and fall of 1979 and 1980
(see below for description of method). Monitoring growth and survival over a period of one
year (Fig. 2) revealed that transplanting in the spring (April) and fall (October) yielded
better survival for this time period, especially under marginal or poor water quality
conditions (Orth and Moore, 1982a; Moore et al., 1996; 1997). Further, our data suggested
that planting in the fall rather than spring was optimal because plants would have eight
months (versus two or less) of suitable environmental conditions to become established and
grow prior to experiencing elevated water temperatures and diminished light levels of late
spring and summer (Moore et al., 1996; 1997). Based on these data we determined that the
optimal transplant period for adult plants in Chesapeake Bay is during the fall, between
mid-September and mid-November, when water temperatures range from 25o to 10o C and
when light attenuation levels (Kd) are less than 1.5 (Dennison et al. 1993).
Fertilizer Additions
We conducted a series of fertilizer addition experiments concurrently with the seasonal
transplant experiments discussed above. Results showed that plant density significantly
increased with addition of fertilizer (Orth and Moore, 1982b). However, fertilizer did not
5
Orth et al.
enhance survival of transplants when other habitat conditions (e.g. light) were limiting (Fig.
2).
Figure 2. Influence of season and fertilizer effects on survival of transplanted eelgrass at three sites along a
gradient of good to poor water quality in the York River estuary. The Guinea Marsh site was considered a high
quality seagrass site as this area experienced some eelgrass loss in the 1970s but recovered almost completely
to pre-1970s levels; the Gloucester Point site was considered a marginal quality seagrass site complete eelgrass
loss in the 1970s with moderate levels of recovery; The Mumfort Island site was considered a poor quality site
as this areas experienced complete eelgrass loss in the 1970s without recovery, despite transplanting efforts.
Planting Techniques
Several techniques for planting adult plants were attempted with each method having
advantages and disadvantages. As cost per unit effort can vary as a function of species used
(Fonseca et al., 1994), geographical location (e.g., environmental conditions, distance from
planting resources), and time of year (e.g., warm summer versus cold winter), we report
effort as time per planting unit. Time required per planting unit is the sum of collection,
preparation, and planting time divided by total number of planting units (Table 2). It does
not include transport time (either in the collection or planting process, which can be
considerable) or organizational time. As time required to plant 1 m2 depends on the choice
of spacing, we standardized our technique comparisons on a per PU basis. In Table 2 all
transplant work and results are reported per PU, unless otherwise noted, regardless of the
number of shoots used. For each technique, the preparation time, physical labor, and
susceptibility of a PU to washing out were rated qualitatively as high, medium, or low.
6
Table 2. Summary of eelgrass transplanting effort and survivorship in Chesapeake Bay utilizing different techniques for planting adult plants and seeds.
Survivorship of planting units is reported for different time scales reflecting criteria used to define success (see text). Preparation time refers to time needed
to collect and prepare a planting unit for transplanting. Physical labor refers to effort required to handle a planting unit (i.e., a more labor intensive technique
scores higher). Susceptibility to washing out is a qualitative assessment of whether a planting unit could be dislodged by typical hydrodynamic forces, such as
currents and waves.
Eelgrass Transplant Techniques in Chesapeake Bay
7
Orth et al.
Preparation time refers to time needed to collect and prepare a PU for transplanting.
Physical labor reflects effort required to handle a PU (i.e., a more labor-intensive technique
scores higher). Susceptibility to washing-out is a qualitative assessment of whether a PU
could be dislodged by currents and waves typical of Chesapeake Bay.
Bare-Root Plants in Mesh Fabric
In 1979, plants were dug with shovels from donor sites and sediments immediately sieved
from the roots and rhizomes. Rhizomes with attached shoots were woven into “HOLDGRO”® mesh mats. An entire mat (the PU) was planted with metal anchors. In this way
shoot densities were standardized (15 shoots PU-1); however, mat preparation required the
most time (30.0 minutes PU-1) of all methods (Table 2). Even though anchored, mats were
susceptible to rapidly being washed out, especially in rough weather, resulting in a 95% loss
of shoots and mats within one month (Table 2).
Cores
In 1979 and 1980, cores (10 cm diameter) of eelgrass and sediment were collected,
transported, and planted intact into pre-dug holes in unvegetated areas. This method was
moderately labor intensive (primarily due to digging and moving large quantities of
sediment), but required less time than the woven mats (5.7 minutes PU-1). However,
standardizing the number of shoots per core was difficult (Table 2). Cores with sediment
provided adequate anchoring, and 100% of the planting units survived the initial one to two
month period, with an average of 57% survival up to one year (Table 2).
Sods
In 1983 and 1984, sods of eelgrass and sediment were collected using a 0.1 m2 U-shaped
aluminum sod-cutter (10 cm sides, 25 cm bottom) by pushing it horizontally below the
rhizome layer. The sod was removed from the cutter, wrapped in a “HOLD-GRO”® mesh
fabric (to maintain integrity of the sod during transport), and transplanted into pre-dug
holes in unvegetated areas. Collection, preparation, and planting of sods required more time
than that with cores (6.4 minutes PU-1) because transport of large numbers of sods was
logistically difficult due to sediment weight (Table 2). Like cores, there was difficulty in
standardizing the number of shoots per sod. Similar to cores, however, the weight of
sediment provided an anchoring mechanism, and 94% of planting units survived the initial
one to two month period with an average of 77% survival up to five to six months (Table
2). Low survivorship after nine months at many transplant sites (Table 2) was attributed to
their locations in regions of the Bay and rivers where water quality was marginal for seagrass
growth (Dennison et al., 1993).
Bundles of Bare-Root Shoots With Anchors
From 1982 to 1996, we used the bundle method for most of our transplanting, a method
proven effective in transplanting programs elsewhere (see Fonseca, 1994; Fonseca et al.,
1998, for details of this method). Plants were dug up using shovels, and sediments separated
from roots and rhizomes by sieving. Ten to 15 individual shoots, each with at least a 1 to
2 cm rhizome fragment, were fastened by a twist-tie (a thin wire wrapped in paper) to a
8
Eelgrass Transplant Techniques in Chesapeake Bay
metal anchor. Each bundle was planted by pushing the anchor into a hole dug with fingers
or trowel so the roots and rhizomes were buried. Although this method required less time
than cores and sods, significant time (4.9 minutes PU-1) was required for bundle preparation
(Table 2). Survival of planting units in the initial one to two month period was 100%,
decreasing to 81% after six months, and 67% at more than the nine months.
Bare-Root Shoots Inserted into Sediments Without Anchors
In 1995, we began using single shoots with attached rhizomes and roots with no anchor.
Shoots with attached roots and rhizomes were collected by shovel, and sediments were
sieved out immediately. An individual shoot with an intact section of rhizome and root was
gently inserted by hand at a slight angle (usually with a single finger) into the sediment to
a depth of between 25 and 50 mm (see Orth et al., 1999, for details of this technique). Since
sediment remains unconsolidated for some time after planting, the rhizome was inserted
under a more compact area of the sediment, which assisted in anchoring the plant. This
technique was a modification of the ‘horizontal rhizome method’ (Davis and Short, 1997)
that used two shoots with rhizomes planted on opposite sides of a bamboo skewer anchor.
We used a single shoot with no anchor thereby minimizing preparation and planting time
substantially (0.35 minutes PU-1) when compared to all previous methods. Although the
initial one to two month success rate (74%) was lower than the core, sod,, or bundled shoot
techniques (Table 2), it was the most time-efficient of adult plant methodologies and had
a high overall success rate (Fig. 3). Though each shoot was initially vulnerable to biological
and physical perturbation within a few days of transplanting, the technique was robust
considering ease of planting and labor requirements (Orth et al., 1999). The plant spacing
we chose for the restoration effort (15 cm centers), and the rapid growth rates of individual
shoots, precluded distinguishing individual planting units beyond the second evaluation
period. We therefore relied on percent cover and shoot density as measures of overall
success (Orth et al., 1999) for a two-year period.
Comparison of Mechanized versus Manual Transplanting
In 2001, we compared PU success of our manual, single shoot method with a mechanized
planting boat (Seagrass Recovery, Inc). The boat was a 7.3 m pontoon boat with two
aluminum wheels. The wheels, approximately 0.91 m apart, were attached to a winch and
counterweight system that allows the wheels to roll freely along the bottom as the boat
moves forward. Grass bundles are placed in plastic clips mounted on the wheels (clips are
approximately 0.91 m apart) as the wheel rotates. Each clip pushed a hole in the sediment,
and friction between the grass and the sediment releases the bundles from the clip. The
machine can be set to deliver a pulse of pressurized air through the clip to aid in the release
of the bundle in the sediment. As the boat moves forward, it leaves two lines of transplants
spaced approximately 0.91 m apart. While the machine was able to plant planting units
(PUs) faster than the manual method (2.2 seconds PU-1 vs. 5.8 seconds PU-1, respectively),
this speed was offset by poorer overall success in the proportion of attempted transplants
that were successfully transplanted into the sediment (8.9 to 65% of attempted PUs for
machine planted PUs compared with 100% success for manual transplanting (Fig. 4). This
resulted in a much greater total labor investment for each machine planted PU that
persisted to 24 weeks than for each similarly persisting manually planted PU (40.6 person
9
Orth et al.
seconds/PU and 22.4 person seconds/PU, respectively, averaged across all sites). (Fishman
et al., 2004). The high rate of loss at the time of planting from the mechanized planter was
due to PUs either being planted upside down, or simply remaining on the wheel and not
getting planted into the bottom (see Fishman et al., 2004, for details of this study).
James River
York River
Figure 3. Low-level oblique aerial view of two transplant sites (one in the James River
and one in the York River) taken eight months after planting in October 1996. Each
of the small squares is a 2 × 2 m patch of eelgrass where, initially, 70 single,
unanchored adult shoots were inserted into sediment at 15 cm intervals along five
rows spaced 0.5 m apart. Medium and large patches had 13 and 50, 2 × 2 m plots,
respectively, arranged in a checkerboard pattern with alternating vegetated and
unvegetated plots.
Transplanting With Seeds
In the late-1980s, we began an examination of the potential use of eelgrass seeds in
restoration. We believed that working with seeds could potentially be more labor and time
efficient.
Harvesting of seeds was accomplished by hand collection of mature reproductive shoots
from established beds when seeds were being released from the flowering shoots (normally
10
Eelgrass Transplant Techniques in Chesapeake Bay
mid-to the end of May into early June). Harvested shoots were placed in nylon mesh bags,
returned to the laboratory, and placed in flow-through circular, 3.8 m3 outdoor tanks that
were shaded (approx. 50%) and aerated. Shoots were maintained in the tanks for up to 12
weeks until mid- to late July to allow for decomposition of the shoots and release of seeds.
Remaining stem and leaf material was removed by sieving. Harvested seeds were then kept
in aerated flow-through tanks (also see Granger et al., 2002). Seed experiments were
initiated between mid-August, and mid-October, prior to the initiation of seed germination
(Orth and Moore, 1983b; Moore et al., 1993). The seed collection method is fairly effective
and has resulted in collections of up to 2.5 to 6.6 million viable seeds per year requiring
between 246 to 204 total collecting hours, respectively (actual number of seeds used in
transplant tests). During the holding period we noted some proportion of seeds suffered
mortality as observed by seed husks floating from the tank. We did not measure this
mortality.
Figure 4. Percent of attempted planting units (PUs) observed by divers to
be present in the sediment immediately after planting at two sites in
Chesapeake Bay that were either machine or manually planted (from
Fishman et al., 2004, reprinted with permission of Restoration Ecology).
Several techniques for planting seeds were attempted. Seed germination is usually complete
by the end of November (Moore et al., 1993), but seedlings are generally difficult to detect
from direct field observations until February or March (Orth and Moore, 1983b). We
measured initial success in the spring once seedlings were large enough to be observed.
Thus, initial success was usually evaluated five to eight months after the beginning of an
experiment. These techniques were also rated for preparation time, physical labor, and, in
experiments where seed containers were used, susceptibility to washing out.
11
Orth et al.
Seeds Broadcast by Hand
Initially starting in 1987 we simply broadcasted seeds by hand either from a boat or by
walking in shallow water (Orth et al., 1994, 2003). This technique was effective because
eelgrass seeds are rapidly incorporated into the sediment and generally do not move far from
where they settle under various hydrodynamic regimes (Orth et al., 1994). Topographic
complexity of the bottom due to biological and physical processes appears to be responsible
for seed retention (Luckenbach and Orth, 1999). A major limitation of the broadcast
method was the low number of seedlings present after six months (overall average of 5-15%
of seeds broadcast; Table 2, Fig. 5). Causes for low success may be attributable to wash-out
(Orth et al., 1994; Harwell and Orth, 1999), germination failure (Moore et al., 1993), or
predation (Fishman and Orth, 1996). However, because of the simplicity of the method we
have continued with this method through 2003.
Figure 5. Mean (±1 SD) number of eelgrass seedlings in treatments
where 10, 100, 1000 and 5000 seeds were broadcast into 4 m2 plots at
4 sites in the Chesapeake Bay region. Horizontal lines indicate sites that
were not significantly different (SNK post-hoc multiple comparison
test) (reprinted with permission from Marine Ecology Progress Series).
12
Eelgrass Transplant Techniques in Chesapeake Bay
Seeds Planted in Peat Pots
In 1994, we attempted to protect seeds in peat pots (5 x 5 x 5 cm) to minimize seed loss.
Peat pots, each containing ten seeds buried 15 to 20 mm in natural sediments, were held in
greenhouse tanks until after seed germination in the fall, and then planted in the field in
spring. This method yielded a low level of survival, which was similar to that of the
broadcast method (14.5% survival; Table 2). A major problem with peat pots was their
susceptibility to being washed-out during periods of high wave action.
Seeds Broadcast and Then Covered with Burlap
In an effort to protect seeds from predation, we attempted to cover seeds in the field with
burlap secured by an anchored wire mesh in 1994. However, the burlap and wire mesh were
extremely vulnerable to wave action, which quickly scoured the plots, resulting in the
lowest success of all transplant methods (0.7%; Table 2).
Seed Placed in Burlap Bags and Buried in the Field
In 1995, we planted seeds in 1 mm mesh burlap bags (5 x 5 cm) to protect them from
potential predation (Fishman and Orth, 1996) and to minimize burial and/or lateral
transport (Harwell and Orth, 1999). Burlap bags increased individual seedling success four
fold-over simple broadcast methods, and success was similar to that in control greenhouse
tanks (see Harwell and Orth, 1999, for details of this experiment). This method had the
highest planting unit success (95%) of all methods using seeds but was only evaluated at one
time interval only (Table 2). Because the objective of the experiment was to assess initial
seedling success, all bags were destructively sampled at the first sampling interval,
precluding subsequent sampling to assess longer-term success.
DISCUSSION
A variety of transplant techniques have been used in the Chesapeake Bay region since 1978
using adult plants and seeds in both basic experiments on the biology and ecology of
eelgrass populations, as well as applied aspects of transplant technology, mirroring
concurrent transplant efforts elsewhere (Fonseca et al., 1998). Below are some of the lessons
we have learned from our transplanting efforts.
Optimal Transplant Season
Optimal season for transplanting eelgrass adult plants in Chesapeake Bay is the fall between
mid-September and mid-November (temperature range during this period is 20o to 10oC).
Planting adult plants in the fall allows the transplanted PU the longest period to establish
and grow before exposure to summertime stresses. Because of the wide latitudinal range of
this species (Thayer et al., 1984) optimal seasons should vary. Fonseca et al. (1998)
recommend planting later in the season for sites at the southern end of eelgrass distribution
versus early in the year for the northern parts of the range. Davis and Short (1997) noted
success in spring and summer eelgrass plantings in New Hampshire, USA, especially in
sub-tidal areas not impacted by winter ice. Christensen et al. (1995) found spring to be
optimal for eelgrass transplanting in Danish waters because it allowed the longest interval
for transplant growth. Summer plantings experienced an initial growth lag and subsequent
13
Orth et al.
high mortality, while autumn plantings did not develop roots and rhizomes adequate to
provide anchoring prior to onset of winter storms.
Use of seeds is optimal between June and October. Seeds are generally fully released from
flowering shoots by mid-June and do not germinate until mid-November (Moore et al.,
1993). We have broadcast seeds through the end of October and have noted viable seedlings
the following spring (unpublished data). Seed availability, as well as germination periodicity,
is likely to vary latitudinally (Silberhorn et al., 1983).
Addition of Fertilizer
Addition of fertilizer to transplants increased shoot density and spread of the transplant unit
in our studies (Orth and Moore, 1982b) and others (Fonseca et al., 1994; Christensen et al.,
1995; Sheridan et al., 1998), although results are equivocal due to varying release rates of
nutrients by the fertilizer (Kenworthy and Fonseca, 1992). Sheridan et al. (1998) found
addition of fertilizer to transplanted shoalgrass (Halodule wrightii) enhanced propagation
(coverage and shoot densities) but did not influence survival. They recommended fertilizer
additions be integrated into seagrass restoration. Long-term benefit to survivorship of a
transplant unit may be negligible especially in areas where water quality parameters are
marginal for survival as was noted at one of our test locations (Fig. 2). Thus, the benefit of
fertilizer use in seagrass restoration projects must be balanced by the cost of fertilizer and
time needed to prepare and deploy a fertilizer supplement.
Transplant Techniques — Adult Plants
Each of the transplant techniques described above for adult plants was generally considered
successful in the short term except for woven mats (Table 2). Time required per PU for
three of the four techniques (cores, sods, and bundles) was similar. Although preparation
time was lower for cores and sods, the physical energy required by field personnel and time
necessary for moving planting units from donor to transplant sites and that required for
planting was excessive when compared to other approaches. Single, unanchored shoots were
equally successful, but required an order-of-magnitude less time per PU than the other
methods. The single shoot technique might require anchors in areas with currents greater
than the 25 cm sec-1 maximum typical of our transplant sites (Orth et al. 1994). Anchored,
bundled shoots have been shown to withstand tidal velocities of up to approximately 50 cm
sec-1 (Fonseca, 1994).
Our time estimates for transplanting adult plants using the various techniques (except the
single, unanchored shoots) are higher than other reported efforts. Fonseca’s (1994)
techniques required 3.5 minutes and 2.0 minutes PU-1 for cores and bundling, respectively,
compared to our estimates of 5.7 minutes and 4.7 minutes PU-1, respectively, for these same
techniques. Elements of the transplant process such as collection effort, environmental
conditions (e.g., temperature, water depth), availability of donor plants, and manual labor
for either field or laboratory work, could account for such differences. These times are
much greater than the time required for planting single, unanchored shoots (21 seconds
PU-1).
14
Eelgrass Transplant Techniques in Chesapeake Bay
Mechanized planting increased the rate of planting adult plants but this speed was offset by
poorer planting success. Because PU success was so low, mechanized transplanting did not
result in a significant time savings (Fishman et al., 2004). Potential does exist for
improvement in mechanical planting but subsequent testing must be conducted and results
published to ensure validity of the improved technology.
Transplant Techniques — Seeds
Transplant techniques using seeds were highly variable in their success. Burlap with wire
covering seeds, and peat pots with buried seeds, were very susceptible to being washed out.
Seedlings, if washed out, are unlikely to become re-established naturally. Broadcast seeds
have yielded generally low rates of initial seedling establishment (avg. 5-15%, Orth et al.,
2003). Protected seeds in burlap bags had a higher survival rate than other seed methods
(Table 2). Utility in larger scale restoration efforts will require further assessment at time
scales greater than in the experiment that assessed this method.
Improved methodologies are needed to increase seedling success in the field. Recent
research with an underwater planter has shown promise in Rhode Island, where
broadcasting seeds has resulted in extremely low germination rates, by elevating germinating
rates to approximately 10% (S. Granger, personal communication). The potential for use
of seeds in restoration efforts remains open for continued discussion but has shown promise
in our attempts to restore eelgrass in the Delmarva coastal bays where we have established
approximately 25 hectares of eelgrass by broadcasting seeds from 1999 through 2002 in one
acre plots (Orth et al. 2006; Table 1).
CONCLUSIONS
The increasing interest in seagrass restoration worldwide has resulted in efforts to develop
new and improved methodologies in transplanting seagrass, both manual and mechanized
(Fonseca et al., 1998; Paling et al. 2001a, b; Fishman et al., 2004). Results of the various
techniques we have used here in Chesapeake Bay parallel those efforts from other regions
(Fonseca et al., 1998). Most techniques we attempted were successful in the short term but
longer term success in Chesapeake Bay is more influenced by water quality than any other
factor (Dennison et al., 1993). While the development of more effective techniques using
adult plants and seeds, either manually or mechanically, will likely continue, stronger efforts
will be needed to provide appropriate water and sediment quality conditions for healthy
seagrass to persist and spread.
ACKNOWLEDGMENTS
Work described in this review paper was funded in part by the Virginia Saltwater Recreational Fishing License
Fund, the Commonwealth of Virginia, and private grants from the Allied-Signal Foundation and Norfolk
Southern. We thank F. Short and J. Kenworthy for helpful comments on the manuscript. Finally, we
acknowledge the numerous individuals who participated in the transplant effort during the twenty-five years
we have been conducting eelgrass research in Chesapeake Bay. This is contribution number 2594 from the
Virginia Institute of Marine Science, College of William and Mary.
15
Orth et al.
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Christensen PB, Sortkjaer O, Glathery KJ. 1995. Transplantation of eelgrass. National Environmental
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Cottam C. 1934. Past periods of eelgrass scarcity. Rhodora 36:261–264.
Davis RC, Short FT. 1997. Restoring eelgrass, Zostera marina L., habitat using a new transplanting technique:
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Fishman JR, Orth RJ. 1996. Effects of predation on Zostera marina L. seed abundance. Journal of Experimental
Marine Biology and Ecology 198: 11–26.
Fishman JR, Orth RJ, Marion S, Bieri J. 2004. A comparative test of mechanized and manual transplanting of
eelgrass, Zostera marina, Chesapeake Bay. Restoration Ecology 12:214–219.
Fonseca MS. 1994. A guide to planting seagrasses in the Gulf of Mexico. Sea Grant Publ. TAMU-SG-94-601.
Texas A&M University, College Station, Texas.
Fonseca MS, Kenworthy WJ, Courtney FX, Hall MO. 1994. Seagrass planting in the Southeastern United
States: Methods for accelerating habitat development. Restoration Ecology 2:198–212.
Fonseca MS, Kenworthy WJ, Thayer GW. 1998. Guidelines for the conservation and restoration of seagrasses
in the United States and adjacent waters. NOAA Coastal Ocean Program Decision Analysis Series No.
12. NOAA Coastal Ocean Office, Silver Spring, MD. 222 pp.
Granger S, Traber MS, Nixon SW, Keyes R. 2002. A practical guide for the use of seeds in eelgrass (Zostera
marina L.) restoration. Part 1. Collection, processing and storage. M. Schwartz (ed.), Rhode Island Sea
Grant, Narragansett, R. I. 20pp.
Harwell MC, Orth RJ. 1999. Eelgrass (Zostera marina L.) seed protection for field experiments and implications
for large-scale restoration. Aquatic Botany 64:51–61.
Harwell MC, Orth RJ. 2002a. Long distance dispersal potential in a marine macrophyte. Ecology
83:3319–3330.
Harwell MC, Orth RJ. 2002b. Seed bank patterns in Chesapeake Bay eelgrass (Zostera marina L.): A baywide
perspective. Estuaries 25:1196–1204.
Kenworthy WJ, Fonseca MS. 1992. The use of fertilizer to enhance growth of transplanted seagrasses Zostera
marina L. and Halodule wrightii Aschers. Journal of Experimental Marine Biology and Ecology 163:141–161.
Lombana A. 1999. Habitat fragmentation in transplanted eelgrass (Zostera marina L.) beds: Effects on decapods
and fish. M. A. Thesis, College of William and Mary.
Marion S. 2002. Effects of habitat fragmentation on the utilization of eelgrass (Zostera marina) habitat by mobile
epifauna and macrofauna. M. A. Thesis, College of William and Mary.
Marsh GA. 1973. The Zostera epifaunal community in the York River, Virginia. Chesapeake Science 14:87-97.
Moore KA, Neckles HA, Orth RJ. 1996. Zostera marina L. (eelgrass) growth and survival along a gradient of
nutrients and turbidity in the lower Chesapeake Bay. Marine Ecology Progress Series 142:247–259.
Moore,KA, Orth RJ, Nowak JF. 1993. Environmental regulation of seed germination in Zostera marina L.
(eelgrass) in Chesapeake Bay: Effects of light, oxygen, and sediment burial depth. Aquatic Botany
45:79–9l.
Moore KA, Wetzel RL, Orth RJ. 1997. Seasonal pulses of turbidity and their relations to eelgrass (Zostera
marina L.) survival in an estuary. Journal of Experimental Marine Biology and Ecology 215:115–134.
Orth RJ. 1976. The demise and recovery of eelgrass, Zostera marina, in the Chesapeake Bay, Virginia. Aquatic
Botany 2:l4l–l59.
Orth, RJ, Harwell MC, Fishman JR. 1999. A rapid and simple method for transplanting eelgrass using single,
unanchored shoots. Aquatic Botany 64:77–85.
Orth RJ, Luckenbach ML, Moore KA. 1994. Seed dispersal in a marine macrophyte: Implications for
colonization and restoration. Ecology 75:1927–1939.
Orth RJ, Luckenbach ML, Marion SR, Moore KA, Wilcox DJ. 2006. Recovery of the seagrass Zostera marina
(eelgrass) in the Delmarva coastal bays, USA. Aquatic Botany 84:26–76.
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Eelgrass Transplant Techniques in Chesapeake Bay
Orth RJ, Moore KA. 1982a. The biology and propagation of Zostera marina, eelgrass, in the Chesapeake Bay,
Virginia. Final Report U.S. EPA. Grant No. R805953, Washington, DC. 187 pp.
Orth RJ, Moore KA. 1982b. The effect of fertilizers on transplanted eelgrass, Zostera marina L. in the
Chesapeake Bay. pp. l04-l3l. In: Proceedings Ninth Annual Conference on Wetlands Restoration and
Creation, ed. Webb F. J., Hillsborough Community College, Tampa, FL, May 20–21.
Orth RJ, Moore KA. 1983a. Chesapeake Bay: An unprecedented decline in submerged aquatic vegetation
Science 222:5l–53.
Orth RJ, Moore KA. 1983b. Seed germination and seedling growth of Zostera marina L. (eelgrass) in the
Chesapeake Bay. Aquatic Botany l5:ll7–l3l.
Orth RJ, Moore KA. 1984. Distribution and abundance of submerged aquatic vegetation in Chesapeake Bay:
An historical perspective. Estuaries 7:531–540.
Orth RJ, Moore KA. 1986. Seasonal and year-to-year variations in the growth of Zostera marina L. (eelgrass)
in the lower Chesapeake Bay. Aquatic Botany 24:335–341.
Orth RJ, Moore KA. 1988. Distribution of Zostera marina L. and Ruppia maritima L. sensu lato along depth
gradients in the lower Chesapeake Bay. Aquatic Botany 32:291–305.
Orth RJ, Wilcox DJ, Whiting JR, Nagey LS, Tillman A. 2002. Distribution and abundance of submerged
aquatic vegetation in the Chesapeake Bay and tributaries and Chincoteague Bay–2001. U.S.E.P.A.
Chesapeake Bay Program Final Report. Annapolis, MD. http://www.vims.edu/sav/bio/sav01.
Orth RJ, Fishman JR, Harwell MC, Marion SR. 2003. Seed density effects on germination and initial seedling
establishment in eelgrass, Zostera marina, in the Chesapeake Bay region. Marine Ecology Progress Series
250:71–79.
Paling EI, van Keulen M, Wheeler K, Phillips J, Dyhrberg R. 2001a. Mechanical seagrass transplantation in
western Australia. Ecological Engineering 16:331–339.
Paling EI, van Keulen M, Wheeler K, Phillips J, Dyhrberg R, Lord DA. 2001b. Improving mechanical seagrass
transplantation. Ecological Engineering. 18:107–113.
Sheridan P, McMahan G, Hammerstrom K, Puich W. 1998. Factors affecting restoration of Halodule wrightii
to Galveston Bay, Texas. Restoration Ecology 6:144–158.
Silberhorn G, Orth RJ, Moore KA. 1983. Anthesis and seed production in Zostera marina L. (eelgrass) from the
Chesapeake Bay. Aquatic Botany l5:l33–l44.
Short FT, Davis RC, Kopp BS, Short CA, Burdick DM. 2002. Site-selection model for optimal transplantation
of eelgrass Zostera marina in the northeastern US. Marine Ecology Progress Series 227:253–267.
Thayer GW, Kenworthy WJ, Fonseca MS. 1984. The ecology of eelgrass meadows of the Atlantic Coast: A
community profile. U. S. Fish Wildlife Service, FWS\OBS-84\02, Washington, D.C., 147p.
RJO, JRF, SRM, KAM, JFN, JvM (Virginia Institute of Marine Science, School of Marine Science, College
of William and Mary, Gloucester Point, VA 23062: jjorth@vims.edu); JB (Chesapeake Bay Foundation, 142
West York Street, Norfolk, VA 23510); MCH (A. R. M. Loxahatchee National Wildlife Refuge, United States
Fish and Wildlife Service, Boynton Beach, FL 33437)
17
❖
EVALUATION OF THE SUCCESS OF SEAGRASS MITIGATION
AT PORT MANATEE, TAMPA BAY, FLORIDA
R.R. Lewis III, M.J. Marshall, S.A. Bloom, A.B. Hodgson & L.L. Flynn
ABSTRACT
Restoration methods for three species of subtropical seagrass (shoal grass, Halodule wrightii,
manatee grass, Syringodium filiforme, and turtle grass, Thalassia testudinum) were evaluated over two
growing seasons in the Tampa Bay estuary on the Gulf Coast of southwest central Florida, USA.
Seven methods were tested: (1) hand-planted field-collected floating shoots with an attached
viable rhizome, (2) hand-planted bundles of viable shoots with attached rhizomes, (3) handplanted 5 cm plugs of viable shoots and rhizomes embedded in sediment within fibrous peat
pots, (4) mechanized transplanting of shoal grass only, using a modified pontoon boat with a
hydraulically driven rotary wheel, (5) mechanized transplanting of primarily turtle grass with 1.21
m x 1.51 m mechanized jaws that extracted and planted units of seagrass, (6) a modified manual
shovel method of transplanting shoalgrass into a former dredged material disposal area excavated
prior to transplanting and (7) passive recovery of damaged seagrass within an area closed to boats
with combustion engines. Only the latter two methods proved successful. The failure of the
other five hand planting, hand transplanting and mechanical transplanting techniques appeared
to be due to a failure to appropriately transplant turtle grass, the more robust species. A portion
of this failure appears to have been due to forceful tidal currents that uprooted any planted shoal
grass, and uncontrollable bioturbation by large schools of southern stingrays (Dasayatis sabina)
foraging within the seagrass restoration areas. Method 6, which involved transplanting 0.78 ha
of shoalgrass to 2.77 ha of excavated dredged spoil and adjacent unvegetated areas, resulted in
1.09 ha of seagrass cover. Method 7, establishing and enforcing a restriction on motorized boat
traffic within a seagrass protection zone, achieved rapidly observable increases in seagrass cover
and distribution and was the least expensive method. As a result of passive protection, 6.23 ha of
mixed beds of shoalgrass and turtle grass containing 1.34 ha of bare sand due to prop scarring
showed a reduction to 0.57 ha of bare sand that resulted in 0.77 ha of seagrass recovery in
eighteen months. These combined efforts have successfully established 1.86 ha of primarily shoal
grass to offset impacts to 2.94 ha of both turtle grass and shoalgrass at an estimated cost of
approximately USD$6.3 million as of December 2002. These figures indicate a cost for successful
seagrass mitigation, at this stage of the project, of US$3,387,097 per ha.
INTRODUCTION
Several researchers have examined the feasibility of restoring seagrass meadows by planting
or transplanting various seagrass species during the past three decades, primarily in the
United States and Australia (Lewis et al.1987, Lewis et al 1994, Fonseca et al. 1994, Fonseca
et al. 1996, Fonseca et al. 1998, Short et al. 2000, Paling et al. 2001a, Paling et al. 2001b,
Fonseca et al. 2002, Orth et al. 2002). Generally, the adverse effects of impaired water
quality, bioturbation caused by marine organisms, elevated tidal and wind driven current
energy, and other factors limit the successful reestablishment of seagrasses in most areas.
This paper describes an attempt to apply lessons learned from these decades of efforts in
conducting seagrass mitigation to a major port expansion project in Tampa Bay, Florida,
USA.
Seagrass salvage was first described by Lewis (1987) who noted that Van Breedveld (1975)
first reported successfully transplanting intact turtle grass, manatee grass and shoal grass
plugs (e.g., plugs with seagrass leaves and rhizomes embedded in sediment). Experimental
transplants at Craig Key during 1979–1981 in the Florida Keys (Monroe County, Florida)
for the Florida Keys Bridge Replacement Project showed that in experiments with
19
Lewis, Marshall, Bloom, Hodgson & Flynn
transplanted turtle grass, with careful handling, over 90% of the plugs remained in place and
began to coalesce after just 18 months. Manatee grass plugs placed on 1 m centers showed
similar success; however, on 2 m centers, none survived. Inter-plot success was variable;
30% of the shoalgrass plugs survived in one plot, none in two other plots. For the larger
bridge replacement project 54.3 ha of seagrass were restored, or documented to recover
naturally from damage (Lewis 1987, Lewis et al. 1994). The original damage to seagrass beds
totaled 37.8 ha, so that 16.5 ha more than the impact area were restored.
Fonseca et al. (1994) reported on experimental transplanting of shoal grass and manatee
grass in Florida (Tampa Bay) and shoal grass and eelgrass (Zostera marina) in North
Carolina. Methods of transplanting, addition of fertilizers and caging to prevent
bioturbation by rays were tested. They tested three methods of transplanting, cores, staples,
and peat pots, and concluded that all could provide persistent seagrass plantings but that not
all may be applicable to every situation. Bioturbation by rays was reported as a significant
negative factor and caging of plantings in Florida significantly increased survival even after
the cages were removed at 90 days post-installation. Townsend and Fonseca (1998)
reviewed the literature on bioturbation in seagrass beds and concluded from experiments
in North Carolina that bioturbation may play an important role in the maintenance of
seagrass landscape patterns. Fonseca et al. (1996) reported on eleven experimental plantings
of shoal grass and manatee grass at sites in Tampa Bay in July 1987 and May 1988. Six sites
showed enough survival to produce persistent beds through the end of sampling after three
years. Five of these were shoal grass plantings, and only one was a manatee grass plot. Again,
bioturbation by rays was implicated in the loss of much of the planted material.
Paling et al. (2001a, b) describe a very large-scale program in Western Australia where
Posidonia spp. and Amphibolis griffithii were mechanically transplanted with success ranging
from 44.3% to 76.8% after two years. At the end of three years, total survival was
approximately 70% and further modifications were being made to the process (Paling et al.
2001b).
Project Permit Requirements
The Manatee County Port Authority (MCPA) received a Florida Department of
Environmental Protection (FDEP) Conceptual Environmental Resource Permit (ERP)
(ERP # 0129291-001-EC) on December 10, 1999 for dredging and mitigation activities
associated with a port expansion plan covering 57.67 ha of submerged lands around Port
Manatee, Manatee County, Florida (Figures 1 and 2). Seagrass habitat impacts were
estimated to be 5.14 ha at the time of permit application submittal. The areal coverage of
seagrass beds within the proposed dredge area was re-mapped as required under Specific
Condition 3 of the consolidated final permit (ERP # 0129291-002-EI) issued for the project
on August 29, 2000, and totaled 2.16 ha (Figure 3). The U.S. Army Corps of Engineers
permit issued for the project is designated # 199801210(IP-MN), which provided federal
authorization for the port expansion plan.
Dredging and filling for the port expansion project was calculated to result in permanent
loss of the following communities: 2.16 ha of seagrass habitat (1.21 ha of primarily turtle
20
Seagrass Mitigation at Port Manatee
grass and 0.95 ha of primarily shoalgrass), 14.92 ha of shallow unvegetated bay bottom
habitat (occurring in depths less than –2 m MLW), and 0.75 ha of intertidal mangrove fringe
and saltmarsh habitat. Temporary impacts to 17.83 ha of deeper unvegetated bay bottom
habitat (occurring in depths greater than –2 m MLW) and 0.78 ha of shoal grass associated
with a mangrove mitigation area located south of the main port facility were also calculated
to occur. Thus, impacts to seagrass totaled 2.94 ha.
Figure 1. Location of Port Manatee, Palmetto, Florida.
The MCPA proposed to offset the project impacts by performing six mitigation and public
interest activities: transplanting the seagrass salvaged from the impact areas (2.94 ha) to
unvegetated areas, planting additional seagrass collected from other local undisturbed
seagrass beds and floating drift and wrack material in up to 57.67 ha of unvegetated
submerged areas in order to restore and enhance seagrass habitat in Tampa Bay (note that
21
Lewis, Marshall, Bloom, Hodgson & Flynn
Figure 2. True color vertical aerial photograph of the project area taken on May 3, 1998. (Source: Gandy Aerial
Photography, Inc.)
22
Seagrass Mitigation at Port Manatee
mitigation work was required to be performed prior to construction); dredging a total of
6.18 ha of wetlands to remove silt and open intertidal creeks within degraded mangrove and
saltmarsh habitat; removing invasive vegetation and grading 19.97 ha of a 26.71 ha spoil
island to enhance mangrove habitat and create seabird nesting habitat; creating a 194.26 ha
motorized vessel exclusion zone for the protection of manatees (Trichechus manatus) and
seagrass beds; and donating two parcels totaling 5.26 ha of privately-owned submerged lands
in exchange for a 0.23 ha area filled at one of the ship berths. The resource protection,
restoration, enhancement, and management activities proposed on the 26.7 ha dredged
material island and within the 194.26 ha manatee and seagrass protection motorized vessel
restriction zone would be included, also, in a long-term management agreement (the
‘seagrass mitigation plan’) to be implemented by the port.
Figure 3. Seagrass distribution prior to disturbances related to the Port Manatee
construction project, 1999.
23
Lewis, Marshall, Bloom, Hodgson & Flynn
The Seagrass Mitigation Plan
The seagrass mitigation plan was based on salvaging all seagrass within the proposed
dredging areas prior to dredging. The seagrass mitigation plan (Figures 4–7, Table 1)
included transplanting 2.16 ha of the 2.94 ha seagrass present in two of the three donor sites
(Figure 4, A and B) into 15 discrete areas (Figure 5), within 4.71 ha of seagrass beds that had
been damaged extensively by boat traffic as documented by Sargent et al. (1995, Figure B-7)
and by direct observation by the senior author over a 20-year period. These areas were
designated sites 1, 2 and 3 for purposes of calculating mitigation credits (Table 1). Sites 1–3
and adjacent areas were further protected through the legal exclusion of motorized vessels.
The remaining 0.78 ha of seagrass from donor site C (Figure 4) was to be transplanted to
2.78 ha of excavated dredged material and 1.87 ha of partially diked natural mudflats that
were further protected through the installation of protective breakwaters (Figures 6 and 7).
These areas were designated sites 4, 5 and 6 (Table 1). Additionally, two passive methods
were proposed: natural recovery to enhance up to 43.30 ha of seagrass beds partially
damaged by propeller scars and grounding sites within those beds (Figure 8), and protecting
194.26 ha of shallow water habitat (including the mitigation areas listed above) through
implementation of a motorized vessel restriction zone (Figure 4, dotted line). These
methods were incorporated into a sovereign submerged land management agreement. The
credits given for mitigation as described above were to be available when mitigation was
deemed successful (Table 1).
Figure 4. Overall summary plan showing the three seagrass salvage areas (A, B and C), and the Manatee/
Seagrass protection area (within dotted lines).
24
Seagrass Mitigation at Port Manatee
Figure 5. Approved seagrass mitigation areas within Areas 1, 2 and 3. Vertical aerial
photograph taken October 3, 2001.
Mitigation “success” was defined two ways for permitting purposes. The first method was
by comparing measured percent cover of all species of seagrass to the mean value of all
reference sites (see the discussion of cover computation in Methods). Partial success was
achieved if the combined percent cover of all seagrass species was less than that of the
combined reference sites. Thus, if one acre (0.4 ha) of mitigation achieved 74% cover, and
the mean reference cover was 74%, one credit of successful mitigation was awarded;
however, if the same mitigation area achieved only 37% cover 0.5 credit was awarded. The
second method of defining “success” was the measurement (using remote sensing
techniques agreed to in advance as part of the permitting) of recovery of visible prop scars
caused by boats within specific defined sites (e.g., Site 8, Table 1) where boats operating
with internal combustion engines were banned and regular enforcement occurred. After
application of the mitigation ratios (Table 1), mitigation credits were awarded. A total of
12.7 mitigation credits were required to achieve full success of the seagrass mitigation
program with a maximum potential availability of 20.37 credits from all sites (Table 1).
In this paper we present summary sample parameters and statistical analyses of percent
cover estimates obtained during quantitative monitoring at nineteen sites at Port Manatee
for all periods from July 1, 2000 through September 10, 2002, or 26 months after the
initiation of the seagrass mitigation program, demarcation of the Manatee/Seagrass
Protection Area, and initiation of patrolling.
25
Lewis, Marshall, Bloom, Hodgson & Flynn
Figure 6. Vertical aerial photographs of Piney Point restoration areas showing conditions before and after
restoration was complete on September 23, 2001. Arrows show the two 300'-long breakwaters.
METHODS
Study Location
Port Manatee, a multi-berth shipping facility for oceangoing marine traffic and holiday
cruise ships, is located along the southeastern shoreline of Hillsborough Bay within the
Tampa Bay estuary, on subtidal and upland land areas, in Manatee County, FL, USA
(82º35’ N, 27º37.5’W) (Figures 1 and 2).
Seagrass Mapping
True color vertical aerial photography of the seagrass donor and mitigation sites was flown
on September 30, 1999, September 27, 2000, October 10, 2001, and October 3, 2002.
Thirty-nine numbered concrete monuments were placed within the seagrass beds as
permanent reference points, surveyed by a marine surveyor, and included as reference
points on all seagrass maps. Each concrete marker was approximately 0.5 m2, weighed 10
kg and was stabilized by a 2 m long coated iron rebar inserted through the center into the
substrate. A series of 72 styrofoam panel targets were anchored in the seagrass mitigation
sites, and geo-referenced to local surveyed reference points using sub-meter accuracy
Trimble® GPS during the September 2000 photography acquisition. The photographic
sequences were scanned at 1200 dpi, then mosaiced and spatially rectified in ERDAS
Imagine™ (Redlands, CA) using the panel ground control points and additional points
synthesized from existing surveyed features. Divers swam transects within the mitigation
26
Seagrass Mitigation at Port Manatee
sites to identify seagrass bed species composition. Four seagrass species—shoalgrass, turtle
grass, manatee grass, and wigeongrass (Ruppia maritima)—were identified within the study
site. The distribution of each seagrass species, based on an inventory of seagrass cover before
disturbances related to the Port Manatee construction project, was mapped by Lewis
Environmental Services Inc. in 1999 (Figure 3). Seagrass community composition was
digitized as ArcView™ 3.2a (ESRI, St. Louis, MO) shape files overlaid on the rectified aerial
photography, and attributed within polygons. Mapping was updated on each annual
photographic series.
Figure 7. Seagrass planting areas completed by EAC, Inc.,
August–September 2001.
Overview of Seagrass Restoration and Protection Methods
The seagrass mitigation project began on April 1, 2000, as the Early Start Program (ESP),
necessitated since permits to disturb seagrasses in the proposed dredging areas had not yet
been issued, in which floating seagrass rhizomes were planted into several sites. The
Manatee/Seagrass Protection Area was also marked and patrolling began at this same time
(Figure 4). After permits were issued to allow salvage of seagrass in areas of proposed
dredging in August 2000, final ecological engineering plans were prepared for that portion
of the project and implementation began in June of 2001. A total of seven methods of
27
Lewis, Marshall, Bloom, Hodgson & Flynn
seagrass restoration were tested: (1) hand-transplanting individual shoots with an attached
viable rhizome; (2) hand-planted bundles of viable shoots with attached rhizomes; (3) hand
planted shoots and rhizomes in fibrous peat pots; (4) mechanized transplanting using a
modified pontoon boat with a hydraulically driven rotary wheel; (5) mechanized
transplanting of mega-units; (6) a modified manual shovel method and; (7) passive recovery
of damaged seagrass within an area closed to boats with combustion engines. Detailed
methodological descriptions are presented below. Table 2 identifies the individual planting
areas and respective restoration methods.
Table 1. This table is modified from page 7, consolidated ERP #0129291-002-EI (FDEP 2000), which shows
mitigation credits to be approved for each seagrass mitigation site. The FDEP permit states that mitigation
credits will be approved after the mitigation work is completed and a site is determined to be successful; 12.7
credits are required for final success of the seagrass mitigation program.
MITIGATION
SITES
MITIGATION
METHODS
AREA (acres)
MAXIMUM
MITIGATION
CREDIT
1–3
Plant and transplant
salvaged seagrass
11.64
2.33
4B, 6B1,
6B2
Install breakwater
and plant
3.04
0.76
5
Remove sand spit
and plant
1.98
0.50
4A, 6A
Scrape down
and plant
6.47
3.24
7
Scrape down
and plant
12.82
6.41
8
Repair prop scars
107.00
7.13
TOTAL
142.95
20.37*
*potential maximum credits
Method 1— Installation of Bare-root Floating Seagrass
Vigorous, floating, bare-root shoots of turtle grass, manatee grass and shoalgrass, with
attached rhizomes approximately varying from approximately 2 to 60 cm in length, were
collected, hand planted individually and anchored in place with 15 cm wire landscaping
staples. Planting unit leaves were trimmed as the units were planted to reduce possible
uprooting from tidal current drag in the planting areas. LES technicians planted
approximately 8,000 planting units in mitigation sites 3B, 3F, and 3G during April–June
2000 (Figure 4).
Method 2—Hand Planting of Floating Seagrass Bundles
LES technicians and trustee prisoners from the Manatee County correctional facility
collected floating plants, primarily, with visible roots and rhizomes of all three seagrass
28
Seagrass Mitigation at Port Manatee
species from the water surface near Port Manatee when westerly winds drove the seagrass
towards the port, or from mid-Tampa Bay when wind blew the floating seagrass fragments
away from the port. Manatee grass was the most commonly collected species. Shoalgrass
and turtle grass were collected but were not abundant. Floating plants were bound together
into small bundles, with paper-covered wire (‘twist ties’). The seagrass shoots were held
before transplanting in floating PVC pens anchored nearshore, then transported to the
planting sites in floating plastic baskets or the floating PVC pens to the planting sites. The
bundles were hand-planted approximately 10 cm into the substrate to bury the rhizome
mass on approximately one meter spacing in sites 2 B, 3E, 3F, and 3G (Figure 5). The
planted seagrass bundles were anchored in place with 15 cm wire landscaping staples. The
seagrass bundles held in the pens were planted within a few days of their collection to
ensure that they were vigorous.
Figure 8. Computer interpretation (ERDAS Imagine) of prop scar recovery in the Manatee/Seagrass Recovery
Area.
Method 3—Shoalgrass Plugs in Peat Pots
Excavated plugs of shoalgrass approximately 5 cm in diameter were hand bedded in an
additional small amount of sediment, inserted into 3 inch peat pots and hand installed using
a plastic garden trowel. A total of 4,582 peat pot units were installed in planting sites 2B, 3B,
3F, and 3G (Figure 5) between June 1 and August 24, 2000. In August 2001 an additional
29
Lewis, Marshall, Bloom, Hodgson & Flynn
2000 peat pots were planted in linear plots approximately 20 m long by 2 m wide in planting
site 2D (Figure 5). The planted rectangular plots were covered with galvanized chain link
fence that was anchored in place over the substrate with landscape staples in order to test
survival of planting units protected from bioturbation by rays.
Table 2. Designation, planting methods, and size of the experimental mitigation sites.
SITE
DESIGNATION
1A
1B
2A1
2A2
2B
3A1
3A2
3A3
3B
3C
3D
3E
3F
3G
3H
METHODS
AREA (hectares)
none
none
machine & hand
machine & hand
machine & hand
machine & hand
machine & hand
machine & hand
machine & hand
machine & hand
machine & hand
machine & hand
machine & hand
machine & hand
machine
0.52
0.28
0.38
0.31
0.23
0.26
0.43
0.46
0.23
0.37
0.30
0.36
0.15
0.26
0.18
Subtotal, sites 1–3
4.72
4A and 6A*
remove dredged material
and transplant with shovel
2.62
5
remove dredged material
0.80
install breakwater and
transplant with shovel
1.23
4B, 6B1, 6B2*
Subtotal, sites 4–6
4.65
7
removed dredged material
and transplant with shovel
5.12
8
close to motorized boat access
and monitor natural recovery
43.30
TOTAL
57.79
*4A nd 4B are Piney Point S (PPS) experimental sites; 6A, 6B1 and 6B2 are Piney Point North (PPN)
experimental sites.
Method 4—Machine Planting of Shoalgrass Bare-root Units
Sites 1A, 1B, 2A1, 2A2, 3A1, 3A2, and 3A3 (Figure 5) were planted by a rotary planting
machine in 2000. This method is further described and analyzed by Fishman et al. (2004).
Plugs of shoal grass were collected from seagrass beds in Tampa Bay pursuant to a collection
permit issued by the Florida Department of Environmental Regulation to the Port’s
30
Seagrass Mitigation at Port Manatee
landscape contractor Seagrass Restoration, Inc. (SRI). These were divided by hand into
individual bare root planting units, and planted using a modified pontoon boat operated by
SRI. The rotary wheel was approximately 2 m in diameter with a 0.5 m wide rotating face
positioned in the center of the boat, rotated by an 8 hp hydraulic drive. On the rotating
surface of the wheel were a series of alternating planting nozzles, each approximately 10 cm
long. As the wheel turned, a technician placed a plug of seagrass on each nozzle. As the
wheel made contact with the substrate the plug was then spiked into the substrate and
‘planted.’ Using this boat, 44,321 planting units were installed in mitigation sites 2A1, 2A2,
3A1, 3A2, 3A3, and 3C between April 1 and September 7, 2000, as certified by Gee and
Jenson Consulting Engineers, Inc., consultants to the MCPA.
Method 5—Seagrass Sods Planted by Hydraulic Extraction
Beginning in June 2001 and continuing through December 2001, a second machine
planting technique, “hydraulic extraction and transplanting of large sods,” was used to
transplant 1.21 m × 1.51 m (1.83 m2) sods of primarily turtle grass with some shoalgrass by
SRI. The donor units were excavated from the proposed dredging areas (donor sites A and
B, Figure 4). Approximately 1.21 ha of turtle grass and 0.94 ha of shoal grass were
transplanted. The planting units were extracted using a modified pontoon boat with three
sets of jaws operated by an eight hp hydraulic pump. The jaws were lowered to excavate
blocks of seagrass from the bottom, removing plants, rhizomes, and sediment in a large sod.
The units were raised and held suspended above the water in the jaws, then the boat moved
to the transplant area, and dropped the units to the substrate on approximately 2 m centers.
A total of 11,609 of these units were transplanted into 13 planting sites totaling 3.91 ha
within the same 4.71 ha boat damage area that had initially been planted by hand, then
subsequently by the rotary planting machine, in 2000 (Figure 5). The designed and
permitted plan was to install the units flush with the sediment surface and clustered tightly
to minimize sediment loss between the large sods.
Method 6—Modified Shovel Method
Within the circulation cuts in seagrass mitigation site 7 (Area C, Figure 4) 0.78 ha of
shoalgrass (13,000 planting units, each 0.06 m²), was moved by hand and shovel into 2.78
ha of excavated dredged spoil in seagrass mitigation sites 6A and 4A (Figure 6), and
surrounding areas protected by installed breakwaters (Figure 7) (G. Montin and R. Dennis,
pers. comm.).
Method 7—Seagrass Protection and Natural Recovery
In addition to physically transplanting seagrass from the donor sites, a 194.26 ha West Indian
manatee and seagrass protection area (site 8) was legally established in June 1999 within
which the use of combustion motors (propeller-driven boat motors) was legally excluded
(Figure 3, dotted line). Port Manatee security vessels and the Florida Marine Patrol
patrolled the protective zone intermittently. Patrols were run seven days a week during
daylight hours except during inclement weather. Boaters were initially given brochures as
they approached the protected area explaining the reason for the closures and providing a
map. Enforcement through the issuance of tickets eventually was necessary for repeated
31
Lewis, Marshall, Bloom, Hodgson & Flynn
violators of the closure. Motorized boat traffic has essentially ceased within the closed area,
but patrolling and monitoring continue.
MONITORING
Baseline monitoring to determine statistically valid sampling parameters was conducted
March 25, 2000. A survey to establish the density of seagrasses to be impacted by dredging
was conducted June 2000 (discussed above). Transplanted seagrass unit survival was
sampled by determining percent cover at periodic intervals after the initial planting was
completed. The first (‘time zero’), monitoring sampling was conducted July 1 2000, two
weeks after initial planting, and represented the initial monitoring event to track changes
during the duration of the mitigation project.
Each experimental unit (e.g., a mitigation site [Table 2]) was located in the field by LES
staff. The sampling area was chosen by the haphazard placement of a 400 m2(20 m × 20 m)
within the experimental unit. To form the sampling area, a right triangle with 20 m sides
and a 28.3 m hypotenuse was staked by driving ½-inch PVC pipe into the sediment at the
corners of the triangle, and stretching polypropylene cord attached to the poles across the
surface of the water to delineate the triangle. The polypropylene cord was then reflected
across the diagonal corners of the experimental unit to define the other side of the triangle
and a center pole was emplaced in the middle. Each measurement consisted of 50 replicate
0.25 m² quadrats, subdivided by string into 100 units (25 cm² each), with each quadrat
representing a data point for visual percent cover by species present made at random, preplotted points distributed in a 20m × 20m area. Sample points were re-randomized within
each experimental unit for each sampling event.
Percent cover data were collected at fourteen sites on July 1, 2000 (time zero). Five
measurements were obtained in reference beds determined by the senior author to represent
the best undisturbed examples within the general area and nine were obtained in
experimental sites. On October 21 and 22, 2000 (time zero plus three months), 19 sites
were measured using the same methodology. The additional five measurements were made
in experimental sites not sampled in July. On April 7, 2001, measurements were repeated
in the same 19 sites and two additional seagrass reference beds of manatee grass. On June
29 and 30, 2001 measurements were repeated in all 21 sites. With the excavation of the
dredged spoil areas in the summer of 2001, and transplanting of shoal grass to them, two
additional reference sites (Piney Point North reference, PPNR, and Piney Point South
reference, PPSR), and two experimental sites where the transplanted shoal grass were placed
after construction (PPNE and PPSE) were added to the sampling matrix. One additional
turtle grass/shoal grass transplant site (3H) was also added. This increased the total sampling
sites to 26. Subsequently, two of the experimental sites (1A and 1B) were dropped since no
mitigation work took place in them. This reduced the number of sampling sites to 24 for
the three remaining sampling periods. Sampling concluded on September 10, 2002.
Statistical Analysis
An experimental unit was defined as one planted mitigation site, or reference bed area, a
sampling unit was defined as one 400 m2 area within a planted or reference bed, a sub32
Seagrass Mitigation at Port Manatee
sampling unit was equivalent to one 0.25 m2 quadrat (composed of 100 cells) within a
sampling unit, and a sub-sub-sampling unit was one cell within one quadrat. The quadrat
data were compiled using customized Pascal programming and Excel 2000 software
(Microsoft, Inc., Redlands, WA, USA). Statistical analysis was conducted using customized
Pascal programs and SPSS (SPSS Inc., Chicago, IL). The sample statistics (mean, standard
deviation, and number of observations) are summarized in Table 3. In quadrats with more
than one grass species present, the numbers of cells occupied by a single grass species, as
well as cells occupied by combination of species, were separately recorded, although those
data are not separated in the summary statistics presented here. In some cases, the 20 m2 area
overlapped natural beds. Where a sampling site fell in the natural bed (hereinafter referred
to as ‘edge’ measurements), the measurement was recorded separately from the
measurements of the actual experimental areas. The edge measurements are also not
included within the statistics. The percent cover data were determined to be non-normal
due to bimodality using a Shapiro-Wilk normality test (α = 0.05) across all samples
(reference and experimental sites at all sampling times) so that parametric statistical analyses
on the untransformed data were not conducted (Shapiro and Wilk 1965). The remotely
sensed areal change in seagrass cover was the criterion for success in site 8. The amount of
seagrass cover within an area of interest (AOI) was analyzed for sample years 1999 and 2002
using unsupervised maximum likelihood classifier change detection analysis in Erdas
Imagine® (Redding, CA).
Table 3. Summary of combined seagrass cover data for all reference and experimental sites, July 1,
2000–September 10, 2002.
DATE
7/1/00
10/22/00
4/7/01
6/30/01
9/22/01
4/20/02
6/24/02
9/9–10/02
MONITORING
EVENT
Time Zero
Time Zero + 3
Time Zero + 9
Time Zero + 12
Time Zero + 15
Time Zero + 21
Time Zero + 24
Time Zero + 26
REFERENCE SITES
Mean S.D.
n
94.57 16.18
250
90.55 16.02
251
81.56 28.50
335
92.31 19.11
348
92.12 19.63
450
96.95 14.29
450
99.34
4.89
450
91.07 23.84
450
TOTAL SAMPLES
2984
EXPERIMENTAL SITES
Mean S.D.
n
0.28
1.10
400
0.39
1.45
673
0.26
1.75
651
0.58
3.50
669
3.09
9.77
817
2.34
8.41
737
5.46
6.96
745
10.17
3.19
738
5430
RESULTS
Several conclusions can be reached by examining the data in Table 3 (Figures 9, 10).
Reference sites exhibited 94.57% in July 2000 and showed persistent cover of 91.07% in
September 2002. Among the experimental sites, there was a net increase in mean cover from
0.28% in July 2000 to 10.33% in September 2002; however, the distribution of cover within
experimental sites was not uniform. Mean cover of the experimental sites did not increase
from Time Zero except in sites 4A (PPS experimental) and 6A (PPN experimental) (Figures
6 and 7). These two sites, which had no seagrass cover prior to their excavation, initially
showed a mean cover of 21.48% immediately after 13,000 shoalgrass planting units were
33
Lewis, Marshall, Bloom, Hodgson & Flynn
installed in September 2001; however, they declined to 11.65% in April of 2002, then
increased to 72.8% in September of 2002, when sampling ended twelve months later.
100
]
]
station
]
]
]
]
]
EX
RB
]
Error Bars show 95.0% Cl of Mean
Mean Cover (%)
75
Bars show Means
50
25
]
0
]
]
]
]
]
]
]
01-JUL-2000
07-APR-2001
22-SEP-2001
24-JUN-2002
22-OCT-2000
30-JUN-2001
20-APR-2002
10-SEP-2002
Date
Figure 9. Mean percent cover of reference and experimental sites on all sample dates.
Mean cover in the 13 other planting sites (excluding 1A, 1B, PPSE and PPNE) ranged from
0.00% in Site 2A1 to 1.86% in Site 3C, with a mean of 0.54% in September 2002, which is
not significantly different (p=0.953) from the mean for the same 13 sites of 0.53% in
September 2001, after all sites had been planted two or three times using various methods.
There was no discernable difference in mean cover among the machine-planted, handplanted, or combination thereof throughout sample dates (p=0.000). With two minor
exceptions, planted or transplanted shoalgrass, turtle grass or manatee grass in these areas
did not survive transplanting or, where they remained rooted, the plants did not expand.
These exceptions were not monitored quantitatively and, due to their small sizes, have not
contributed to the cover in the planted sites.
The first exception area includes approximately 40 m² of shoalgrass plugs in peat pots over
which a chain link fence was laid horizontally in August 2001 in site 2D (Figure 5). The site
is clearly visible both when snorkeling the site and from aerial photographs. It has not
expanded as hoped, however, outside of the boundaries of the fencing. The second similarly
sized area in site 3C received approximately 250 bare root turtle grass rhizomes. Using bare
root turtle grass rhizomes is a modification of a method described by Tomasko et al. (1991),
who collected material by fanning the sediment away from the edges of a turtle grass
34
Seagrass Mitigation at Port Manatee
meadow in a high energy location, and removing selected rhizomes. This method shows
great promise, as very large numbers (hundreds at a time) of bare root, and apparently
viable, turtle grass rhizomes with attached leaves wash up on shorelines in Tampa Bay
during and after winter storms. The method differs from Tomasko et al. (1991), but
produces a similar type of plant material without disturbing existing seagrass beds. These
shoreline stranded plants eventually die and, if they were collected and planted, we believe,
based upon anecdotal observations of the plantings in site 3C, could be a significant source
of non-destructive plant material sources for restoration efforts in Tampa Bay. The same
cannot be said of shoalgrass or manatee grass rhizomes, as the brief exposure to sunlight
apparently caused irreversible mortality in all we attempted to either collect or plant. Ron
Phillips (pers. comm.) confirmed this belief from his experience. Wigeongrass occurred
sparsely in areas of fluctuating salinity within the PPS lagoon and was not sampled.
Location
2A1
2A2
2B
80
]
70
]
3A1
3A2
3A3
3B
3C
3D
3E
3F
60
3G
3H
Mean Cover (%)
50
PPN
PPS
40
]
Error Bars show 95.0% Cl of Mean
]
30
Bars show Means
]
]
20
]
10
0
]
]] ]]] ]
]
]]
]
]
]]]]]]] ]
]
]
]]
]
]
]] ]]]]]] ]]
]] ]]]]]
07-APR-2001
30-JUN-2001
]] ]]]]]]]]]
]]
]
]
]]
]]]]
]
] ] ]
] ] ]
] ]
] ]]]
]]
]
]]]
]
]
]]
]]]]]]] ]]]
-10
01-JUL-2000
22-OCT-2000
22-SEP-2001
20-APR-2002
24-JUN-2002
10-SEP-2002
Date
Figure 10. Mean percent cover of the 17 experimental sites on all sample dates. Note that Piney Point
transplant site is the only site doing well.
Field observations during the sampling over the last two years demonstrate clearly that most
planting units were unable to survive long enough to become securely anchored through
expansion of their root systems. Throughout the area, whether in natural seagrass beds,
planted areas, or bare sand areas, there have been obvious signs of significant bioturbation
notable as depressions ranging up to a 1 m in diameter and 10-25 cm in depth. In many of
the depressions, areas of disturbed sediment, as evidenced by gray redox discolorations on
the sediment surface, dislodged Diopatra spp. tubes, broken crab shells and shattered broken
35
Lewis, Marshall, Bloom, Hodgson & Flynn
clam shells (primarily hard shell clams, Mercenaria mercenaria) were evident, and a large and
active population of southern stingrays (Dasyatis sabina) was continuously observed within
the planting sites. Once planting units are loosened by bioturbation, the strong diurnal tidal
current system within the channelized planting sites quickly removed the units.
Loss of planted seagrass from bioturbation was not unexpected, as previous work in Tampa
Bay attempting to transplant both shoalgrass and manatee grass showed similar “…heavy
bioturbation by rays,” and loss of essentially all planting units in most of the planting sites
(Fonseca et al. 1996). Orth (1975) described similar bioturbation by the cownose ray,
Rhinoptera bonasus in eelgrass seagrass meadows in Chesapeake Bay as it preys on a common
bivalve, Maya arenaria. Sharks are known to prey on stingrays. For example, Dasyatis spp. has
been reported from stomach content analyses of lemon sharks, Negaprion brevirostris, in
Florida (Schmidt 1986) and Baum and Myers (2004) have recently estimated that
commercially harvested sharks in the Gulf of Mexico have declined by 90–99% since the
1950s. The regional population of stingrays appears to be increasing based only on anecdotal
observations by the senior author over the last 37 years. We hypothesize that the magnitude
of decline in sharks may have reduced predation on stingrays and could contribute to their
apparent population expansion, and result in increased damage to seagrass beds.
Other organisms also cause patchiness in seagrass meadows. In tropical areas, seagrass cover
is negatively correlated with burrow density of the ghost shrimp, Callianassa spp. (Suchanek
1983, Suchanek et al. 1986). Clearly, bioturbation can have a major impact on attempts to
restore seagrasses in Tampa Bay (Fonseca et al. 1994, Fonseca et al. 1996). The success of
the shoalgrass plantings at Piney Point, which have similar common excavations by the
southern stingray, may be due to their shallower position, reduced current velocities, and
wave action behind a protective breakwater.
Mechanical Planting
Using the mechanical rotary wheel planting boat method, 44,321 bare root-planting units
were installed in mitigation sites 2A1, 2A2, 3A1, 3A2, 3A3, and 3C between April 1 and
September 7, 2000. On several occasions, it was apparent that the planting contractor replanted sites that had been designated as ‘completed,’ presumably after observing that there
were no live seagrass planting units in the site. The existence of re-planted units was
obvious to the monitoring scientists, although the exact number of re-planted units could
not be accurately determined. At least one site (3C) was planted three times unsuccessfully.
Based on observations of these planting units, it was apparent that a high proportion (>90%
based upon weekly counts) of any units planted by this method were either washed out
during the first few tidal cycles or died soon thereafter as a result of being buried in the
shifting sediments.
Salvage and planting of turtle grass plugs using the mechanical excavation and planting
machine was initially designed and permitted to be implemented meticulously to insure that
the apical meristems are preserved intact, and the edges of the transplant units were smooth
and flush with the existing sediment surface. This would allow for the maximum rate of
expansion and coverage from this relatively slow growing species, and help prevent exposing
36
Seagrass Mitigation at Port Manatee
the edges of the planting units to currents and bioturbation that could result in the removal
of all or portions of the transplanted plugs. This method, using hand labor as reported in
Lewis (1987) and Lewis et al. (1994) in the Florida Keys, worked well, and resulted in the
establishment of a viable turtle grass bed. However, the majority of the mechanically
excavated planting units in the Port Manatee mitigation area appeared to have been
inadequately placed to a sufficient depth in the sediment to withstand the tidal drag, and
most appeared to have been inverted completely or partially at placement so that the
rhizomes were exposed on the surface and the ‘green side’ was buried in the sediments.
Visual examination of many of the ‘transplants’ revealed rhizomes protruding into the water
column with only buried blades preventing the grass from floating away immediately. The
monitoring contractor reported that approximately 50% of the 11,609 excavated units had
disappeared within two weeks of installation. Overall, our counts of surviving units indicate
less than 1% remained after several months. We believe, however, based upon the previous
work of Lewis (1985) and Lewis et al. (1994), that had the units been correctly handled as
originally designed and permitted, a much greater chance of success would have been
possible, as this approach is basically the same method reported to be successful in Western
Australia (Paling et al. 2001a, b).
Future attempts at using these types of units should always take into account, however, the
wave energy and current velocities at a proposed planting site. Wave energy determinations
are being made in Tampa Bay using a relative exposure index (REI) which predicts whether
a given site appears to be too exposed to support any seagrasses without wave protection
(Fonseca et al. 2002a). Wave protection was added to the mitigation plan for sites 4, 5 and
6 as a result of review and comment on the draft mitigation plan by Mark Fonseca, Center
for Coastal Fisheries, NOAA. A current velocity of less than 15 cm/sec is considered
appropriate for seagrass plantings. Velocities between 15–50 cm/sec need to be evaluated
carefully as within this range these velocities can impact seagrass cover. Any site with
velocities over 50 cm/sec should be rejected as a planting site (Fonseca et al. 1998).
Prop Scar Protection Area
Unlike most of the planting sites, the prop-scarred inshore waters within seagrass mitigation
site 8 showed an obvious response to protection. Change detection analysis was conducted
on an AOI encompassing 6.23 ha within seagrass mitigation site 8. As a result of protection,
6.23 ha of mixed beds of shoalgrass and turtle grass containing 1.34 ha of bare sand due to
prop scarring and boat groundings showed a reduction to 0.57 ha of bare sand, or 0.77 ha
of seagrass recovery in eighteen months (Figure 8). Continued protection of these sites
should allow additional recovery within this site and adjacent areas. As a result, both the
FDEP and the Corps have awarded one mitigation credit towards the needed 12.7.
Costs of Restoration
Total costs through December of 2002 were approximately US$6.3 million apportioned as
shown in Table 4. Based upon the successful establishment of 1.86 ha of primarily shoal
grass to offset impacts to 2.94 ha of both turtle grass and shoalgrass, these figures indicate
a cost for successful seagrass mitigation, at this stage of the project, of US$3,387,097 ha-1.
37
Lewis, Marshall, Bloom, Hodgson & Flynn
In contrast, Fonseca et al. (2002b) reported a legally accepted cost estimate of US$ 940,000
ha-1 of restoration for a case of seagrass damage in the Florida Keys.
These costs will continue to increase, as the mitigation project is still underway, with less
than two of the 12.3 credits necessary for final success having been achieved. As certain sites
perhaps expand in areas of successful seagrass transplantation (i.e., the Piney Point dredged
material excavation and planting site) the fixed costs may be distributed among, and
decrease, on a per successful hectare basis. Within other sites, particularly all the sites
planted using the mechanical planting method, a total of US$1,266,837 was spent; however,
it is our opinion that there is no chance for successful results, and the public funds spent
testing the demonstrably unsuccessful methods have been spent without realizing any
benefits to the MCPA.
Table 4. Costs (US$) by category for the Port Manatee seagrass mitigation project through December 2002
(source: Manatee County Port Authority).
CATEGORY
COST
Consultant Services (design, permitting, construction
and planting supervision, monitoring)
Land Purchase and Survey
Construction
Enforcement of No-motor Zone*
Manual Planting
Mechanical Planting
$2,482,844
TOTAL COST
$6,319,449
550.000
1,189,768
300,000
530,000
1,266,837
*Costs incurred during the first three years, including closed area buoys and maintenance.
The total cost of passively protecting seagrass beds by the measures implemented in this area
was approximately US$100,000 per year, which included purchasing and installing buoys
to mark the motorized exclusion zone, the manpower, operation of boats, including fuel,
to patrol the site after the buoys were installed, and routine maintenance of the buoys. On
a per hectare basis, this was the cheapest method, which generated 0.77 ha of seagrass
recovery, and 1.0 mitigation credit.
CONCLUSIONS
Seagrass communities at Port Manatee were observed during 1999–2002. Our observations
confirmed that attempts to use seven methods of seagrass restoration as mitigation for
potential dredging impacts to seagrasses in Tampa Bay have largely failed at a cost to date
of over US$6 million. Naturally occurring bioturbation and strong tidal currents, combined
with a lack of quality control on mechanical efforts to salvage and transplant seagrass prior
to dredging impacts, have resulted in a net loss of seagrass as of the Time Zero + 26 months
monitoring in September 2002. Only continued monitoring will reveal if no net loss, and
the net gain of seagrass in Tampa Bay required by the permit, can be achieved.
Although the state and federal permitting agencies have issued permission for dredging of
the entire project based upon the success to date, final berthing of vessels within the
38
Seagrass Mitigation at Port Manatee
excavated portions of the project cannot occur until all 12.7 mitigation credits have been
certified as having been earned. As of January 2003, less than two credits have been awarded.
LITERATURE CITED
Baum JK, Myers RA. 2004. Shifting baselines and the decline of pelagic sharks in the Gulf of Mexico. Ecology
Letters 7(2):135-145.
Fishman JR, Orth RJ, Marion S, Bieri J. 2004. A comparative test of mechanized and manual transplating of
eelgrass, Zostera marina, in Chesapeake Bay. Rest. Ecol. 12(2): 214-219.
Fonseca MS, Kenworthy WJ, Courtney FX, Hall MO. 1994. Seagrass planting in the Southeastern United
States: methods for accelerating habitat development. Rest. Ecol. 2(3): 198-212.
Fonseca MS, Kenworthy WJ, Courtney FX. 1996. Development of planted seagrass beds in Tampa Bay,
Florida, USA. I. Plant components. Mar. Ecol. Prog. Ser. 132:127-139.
Fonseca MS, Kenworthy WJ, Thayer GW. 1998. Guidelines for the conservation and restoration of seagrasses
in the United States and adjacent waters. NOAA Coastal Ocean Program, Decision Analysis Series No.
12. NOAA, Beaufort, NC.
Fonseca MS, Robbins BD, Whitfield PE, Wood I, Clinton P. 2002a. Evaluating the effect of offshore sandbars
on seagrass recovery and restoration in Tampa Bay through ecological forecasting and hindcasting of
exposure to waves. Tampa Bay Estuary Program, St. Petersburg, Florida.
Fonseca MS, Kenworthy WJ, Julius BE, Shutler S, Fluke S. 2002b. Seagrasses. Pages 149-170 in Perrow MR,
Davy AJ (eds.), Handbook of Ecological Restoration, Volume 2, Restoration in Practice. Cambridge
University Press, Cambridge, UK.
Greig-Smith, P. 1964. Quantitative plant ecology (2nd. Edition). Butterworth, London.
Lewis RR. 1987. The restoration and creation of seagrass meadows in the southeastern United States. Pages
153-173 in M. J. Durako, R. C. Phillips and R. R. Lewis (eds.), Proceedings of the Symposium on
Subtropical Seagrasses of the Southeastem United States. Fla. Dept. of Natural Resources Mar. Res. Pub.
No.42. St. Petersburg, Florida. 209 pp.
Lewis RR, Kruer CR, Treat SF, Morris SM. 1994. Wetland mitigation evaluation report – Florida Keys Bridge
Replacement. Report No. FL-ER-55-94. Florida Department of Transportation, Tallahassee, Florida.
Milner C, Hughes RE. 1968. Methods for the measurement of the primary productivity of grasslands. IBP
Handbook No. 6. Blackwell, Oxford.
Orth RJ. 1975. Destruction of eelgrass, Zoster marina, by the cownose ray, Rhinoptera bonasus, in the Chesapeake
Bay. Chesapeake Sci. 16:205-208.
Orth RJ, Batuik RA, Bergstrom PW, Moore KA. 2002. A perspective on two decades of policies and regulations
influencing the protection and restoration of submerged aquatic vegetation in Chesapeake Bay, USA. Bull.
Mar. Sci. 7(3):1391-1403.
Paling EI, van Keulen M, Wheeler KD, Phillips J, Dyhrberg R, Lord DA. 2001a. Mechanical seagrass
transplantation in Western Australia. Ecological Engineering 16: 331-339.
Paling EI, van Keulen M, Wheeler KD, Phillips J, Dyhrberg R, Lord DA. 2001b. Improving mechanical
seagrass transplantation. Ecological Engineering 18:107-113.
Sargent FJ, Leary TJ, Crewz DW, Kruer CR. 1995. Scarring of Florida seagrasses: Assessment and
management options. Florida Department of Environmental Protection. FMRI Technical Report TR-1.
St. Petersburg, Florida.
Schmidt TW. 1986. Food of young juvenile lemon sharks, Negaprion brevirostris (Poey), near Sand Key, western
Florida Bay. Fla. Sci. 49(1):7-10.
Shapiro SS, Wilk MB. 1965. An analysis of variance test for normality (complete samples). Biometrika 52 (3
and 4):591-611.
Short, FT, Burdick DM, Short CA, Davis RC, Moore PA. 2000. Developing success criteria for restored
eelgrass, salt marsh and mud flat habitats. Ecol. Eng. 15:239-252.
Suchanek TH. 1983. Control of seagrass communities and sediment distribution by Callianassa (Crustacean,
Thallassinidea) bioturbation. J. Mar. Research 41: 281-298.
Suchanek TH, Colin PL, McMarty GM, Suchanek CS. 1986. Bioturbation and redistribution of sediment
radionucleides in Enewetak Atoll Lagoon by callianassid shrimp: biological aspects. Bull. Mar. Sci. 38(1):
144-154.
39
Lewis, Marshall, Bloom, Hodgson & Flynn
Tomasko DA,Dawes CJ, Hall MO. 1991. Effects of the number of short shoots and presence of the rhizome
apical meristems on the survival and growth of transplanted seagrass, Thalassia testudinum. Contrib. Mar.
Sci. 32:41-48.
Townsend EC, Fonseca MS. 1998. Bioturbation as a potential mechanism influencing spatial heterogeneity
of North Carolina seagrass beds. Mar. Ecol. Prog. Ser. 169:123-132.
Van Breedveld J. 1975. Transplanting of seagrasses with emphasis on the importance of substrate. Florida
Department of Natural Resources, St. Petersburg, Fl. Fla. Mar. Res. Publ. 17.
RRL (Lewis Environmental Services, Inc., P. O. Box 5430, Salt Springs, FL 32134-5430); MJM (Coastal Seas
Consortium, Inc., P. O. Box 20818, Braden River, FL 34204-0818); SAB (Ecological Data Consultants Inc.,
P. O. Box 760, Archer, FL, 33618-0760); ABH (Resource Designs, Inc., 911 Silver Palm Way, Apollo Beach,
FL 33572-2005); LLF (Lewis Environmental Services, Inc., 2824 Falling Leaves Drive, Valrico, FL 33594)
40
A SHALLOW WATER TECHNIQUE FOR THE SUCCESSFUL
RELOCATION OR TRANSPLANTATION OF LARGE AREAS OF
SHOALGRASS (HALODULE WRIGHTII)
G.J. Montin & R.F. Dennis III
ABSTRACT
In August 2001, Environmental Affairs Consultants, Inc. (EAC) was contracted to transplant 0.78
ha (1.92 ac) of existing shoalgrass (Halodule wrightii) at Port Manatee, in Tampa Bay, Florida. This
seagrass was excavated from within 3 shallow areas, designated as flushing channels, and
replanted within approximately 2.78 ha (6.86 ac) of two remote mitigation sites. The previously
prepared mitigation areas, sites 4 and 6, are located immediately north of Port Manatee near the
western end of Piney Point Road. Given the inadequate results of formerly documented hand
planting methods to reliably and economically transplant large areas of seagrass, EAC devised and
implemented a “modified shovel method.” Using this method, personnel collected shoalgrass
with a modified, sharpened, flat shovel facilitating the removal of seagrass planting units (PUs)
that are 0.06 m2 (0.69 ft 2) by 7.62 cm (3 in) thick. The units were transported and installed flush
to the existing substrate and stapled at densities averaging 0.61 m (2.0 ft) centers within the
prepared mitigation sites. Transplantation was completed in 24 working days at an average rate
of 324 square meters (3,485 sq ft) per day. One year following planting (September 10, 2002),
the documented mean percent coverage of seagrass was 72.8% as determined from the two
quantitative monitoring areas located within the Piney Point North and Piney Point South
mitigation sites. Based on the observed rate of expansion of the installed PUs, coalescence of the
seagrass in the transplanted areas is expected within 3 years.
INTRODUCTION
In August 2001, Environmental Affairs Consultants, Inc. (EAC) was contracted to transplant
0.78 ha (1.92 ac) of existing shoalgrass (Halodule wrightii) at Port Manatee, in Tampa Bay,
Florida. This transplantation effort was one of the major seagrass mitigation projects
undertaken by the Manatee County Port Authority (MCPA). Pursuant to the federal and
state permit conditions for this project, the compensatory mitigation was required to
demonstrate success prior to initiating the maintenance dredging and port expansion
activities. This requirement was due to the inherent difficulties and risk associated with the
reestablishment of seagrasses (Fonseca 1994). Please refer to Figure 1 for the project
overview depicting the locations of the donor areas, recipient sites, and overland transport
route.
Development and implementation of a rigorous organizational plan was necessary to
accomplish the successful transplantation of this quantity of seagrass. Aspects of each
transplantation element were stringently evaluated, which eliminated bottlenecks and
produced an overall efficient coordinated methodology necessary to complete this massive
transplantation project.
Although different techniques of hand planting seagrass have been employed, such as the
plug method, staple method, and peat pot method, establishment of seagrass habitat using
these methods has historically been poor. Given the inadequate results of formerly
documented hand planting methods to reliably and economically transplant large areas of
seagrass (Fonseca et al. 1998), EAC devised and implemented a “modified shovel method”
to improve the process and ultimate success of transplanting seagrass. This method was
developed after evaluating the physical conditions of both the recipient and donor sites and
implementing a procedure which recognized the growth habits, life histories, and the
physical limitations of the species to be transplanted.
41
Montin & Dennis
Figure 1. Aerial photograph of project site, showing donor areas, recipient sites, and transport routes.
42
Transplantation Technique for Shoalgrass
METHODS
Subsequent to the approval of the modified shovel method by the Florida Department of
Environmental Protection (DEP), EAC obtained and fabricated the necessary equipment
to transplant seagrass including: 25 predrilled and sharpened flat 22.86 cm (9 in) x 27.94 cm
(11 in) shovels (Figure 2); 60 wooden framed 10.16 cm (4 in) thick bead board floats
measuring 1.22 m (4 ft) x 1.22 m (4 ft) as shown in Figure 3; 1,000 heavy duty 27.31 cm
(10.75 in) x 53.67 cm (21.13 in) plastic trays; and 125,000 sod staples 15.24 cm (6 in) long.
Figure 2. Shovel modified for transplantation of Halodule wrightii.
Figure 3. Bead board (styrofoam) floats with trays of seagrass.
43
Montin & Dennis
The source material for transplantation was collected from within 3 shallow areas,
designated as the future flushing channels for site 7, and replanted within the excavated
spoil and adjacent unvegetated areas totaling approximately 2.78 ha (6.86 ac) of two
designated remote mitigation sites (4 and 6, totaling 3.63 ha (8.96 ac). The two mitigation
sites, located immediately north of Port Manatee near the western end of Piney Point Road,
were prepared by removing excess sand in the former spoil deposition areas to elevations
suitable for shoalgrass propagation. Once the spoil material was removed, breakwaters were
installed to stabilize and protect the near shore areas. This created a quiescent location
facilitating the establishment of the seagrass to be transplanted within the areas and
maintenance of the existing seagrass (Lewis 2002a).
Prior to project mobilization, the MCPA constructed temporary accesses to each of the 3
donor sites and demarcated the previously surveyed limits of the seagrass collection areas.
The donor sites were further subdivided by EAC into separate parallel ~3 meter (10 ft)
wide harvest lanes in order to track the seagrass collection progress and clearly identify work
areas during the transplantation process. The limits of the recipient mitigation areas were
also marked with PVC pipe by the MCPA to demarcate the planting area. Both the donor
and recipient sites were systematically inspected to identify the water depths during daily
tidal ranges, locate existing seagrass, and evaluate substrate consistency to determine labor
limitations, equipment placement, and site accessibility. Staff gauges were installed at both
the donor and recipient sites to provide water depth information necessary to shift donor
removal based upon water depth. Due to the instability of the substrate at points of entry
along the shorelines of both the donor and recipient sites, metal runway matting was laid
to stabilize the substrate and improve accessibility for repetitive vehicular and pedestrian
utilization. Separate equipment storage, sheltered rest/first aid areas, potable water, and
restroom facilities were centrally located at both the recipient and donor sites. The haul
route, approximately 3.2 km (2 miles), for the transportation process was also established.
A unique organizational infrastructure was developed to simultaneously coordinate seagrass
collecting, transporting, and replanting. The three different but simultaneous operations
were staged with separate labor crews whose members were assigned specialized tasks. Each
operation also included crew leaders in constant communication with at least 1 of the 2
onsite supervisors to troubleshoot and to maintain quality control. Daily progress meetings
were held during the transplantation process to discuss safety issues and/or observed
problems, thereby increasing the efficiency and safety of each individual operation. Prior
to the actual commencement of transplantation a final on-site organizational meeting was
held where the two supervisors selected and trained the crew leaders for each of the
individual operations as shown below. The labor crew, totaling 30 personnel, was then
subdivided, assigned, and trained.
Seagrass Collection
9 Harvesters
1 Crew Leader/Harvester
3 Loaders/Runners
1 Supervisor
Seagrass Transport
3 Drivers/Loaders
1 Crew Leader/Driver
Seagrass Installation
9 Planters
1 Crew Leader/Planter
3 Loaders/Runners
1 Supervisor
Utilizing the modified shovel, the harvesters collected the 0.06 m2 (0.69 ft 2) planting units
(PUs) by vertically inserting the shovel into the substrate to a depth of approximately 10 cm
(3 in). The handle of the shovel is then pushed downward toward the surface of the
44
Transplantation Technique for Shoalgrass
substrate, slowly manipulating the sharpened front edge of the shovel until it is parallel to
the bottom, while simultaneously pushing it forward until the carrying capacity of the
shovel is achieved. Finally, the shovel is lifted to the surface completing the collection of the
planting unit. The multiple 0.64 cm (0.25 in) holes predrilled into the shovel prevents
suction between the saturated substrate and the shovel, allowing the harvesters to remove
the PUs carefully by sliding them into the heavy duty plastic sod trays. The shovels’
sharpened front and side edges cut the rhizomes cleanly and reduced tearing and/or
displacement during collection. The associated substrate collected with the seagrass at a
depth of ~10 cm, coupled with the support provided by the trays, minimized the potential
for stress during transport attributed to desiccation and/or further physical rhizome damage.
The seagrass planting units were placed 2 per tray at rate of up to 2 per minute per harvester.
The trays, 8 in total, were conveniently carried on the wooden framed bead board floats
during the collection, transport, and installation processes.
The installation process utilized the same sharpened shovel to create a hole in the substrate
allowing the seagrass to be inserted flush with the adjacent elevation of the bottom,
preventing the potential loss of material and/or erosion. Following the excavation of the
hole, the material was cast aside and the shovel was then used to lift the PU’s from the
plastic trays and install the seagrass. Sod staples measuring 3.81cm x 15.24 cm or 1.5 in x 6
in were inserted into the PU’s, 1 per PU, for the purpose of anchoring the unit in place.
The PU’s were installed at various densities averaging 0.61 m (2.0 ft) centers within the
prepared mitigation sites in accordance with the needs of the MCPA.
RESULTS AND DISCUSSION
The total transplantation of approximately 0.78 ha (1.92 ac) of shoalgrass into 2.78 hectares
(6.86 ac) of the two mitigation areas was completed utilizing 30 laborers and 2 on-site
supervisors in 24 working days (9 hours/day) at an average production rate of 324 square
meters (3,485 sq ft) per day. The cost associated with this project was calculated at $3.44 per
0.06 m2 (0.69 ft 2) planting unit transplanted including all mobilization, labor, equipment,
supplies, and demobilization expenditures. However, it is important to consider the factors
that influenced both the production rates and the related cost such as: the water depths of
the donor and recipient sites, the quantity and planting spacing of the transplanted seagrass,
and the mode of transportation.
The water depths, at both the donor and recipient sites, in conjunction with the daily tidal
ranges, determined the available work windows for optimal production. Although the
modified shovel method has recently been adapted to deeper water through the use of
multiple diver surface supplied air systems, the appropriate water depths for a maximized
rate of production throughout the collection and planting phases were between 10 cm (4
in) and 1 m (~3 ft).
The quantity of material transplanted coupled with the planting spacing also influenced the
overall cost of the project. As per the contract at Port Manatee, all of the seagrass within the
3 shallow areas designated as flushing channels had to be transplanted within the mitigation
sites.
45
Montin & Dennis
Because direct land access was convenient at the donor and recipient sites and they were
linked by a haul route designated and prepared by the MCPA, the seagrass was transported
by means of multiple customized submersible trailers. Due to the distance between the sites
and their proximity to shallow water, transportation by boat and/or barge was projected to
be less efficient and more expensive.
The modified shovel method was designed and implemented with the goal of successfully
transplanting seagrass in a manner that left the rhizomes and associated substrate intact.
This “sod” included a protective buffer against physical damage attributed to handling
and/or desiccation during the transplantation process. The surface area of the PU was large
enough to ensure the presence of growing tips or rhizome apical meristems and maximize
the number of short shoots on long shoots without being too large or difficult to efficiently
handle. In addition, the labor crews were subdivided, assigned, and trained for a specialized
task in order to simultaneously accomplish seagrass collection, transport, and planting. As
a result, the seagrass was typically replanted within 25 minutes of collection further
reducing the potential for damage attributed to desiccation.
In accordance with the state and federal permit requirements, the Manatee County Port
Authority (MCPA) conducted the Seagrass Mitigation Success Monitoring one year
following the completion of the transplantation process (September 2002). Following the
completion of the monitoring event, the documented mean percent coverage of seagrass was
72.8 percent as determined from the two quantitative, previously unvegetated monitoring
areas (400 m2 or 4,306 ft2 quads) located within the Piney Point North and Piney Point
South mitigation sites., (Lewis 2002b). As both the amount of seagrass that was transplanted
as well as the total area planted within the mitigation sites were determined by the
aforementioned regulatory permits, optimum spacing and PU expansion rates using the
modified shovel method were not evaluated. However, based on qualitative observations
of the expansion of the installed planting units and encroachment by existing adjacent
seagrass communities, complete coalescence of the transplanted areas is expected within
three years. Further analysis of the optimal spacing between planting units and the
associated expansion rates is recommended for future transplantation projects.
Due to its versatility and effectiveness, the modified shovel method could potentially be
adapted to supplement the modified compressed succession principle where a faster
growing, opportunistic seagrass such as a shoalgrass is used to stabilize propeller scars,
blowholes, and berms serving as temporary substitute for the climax species, turtle grass
(Thalassia testudinum) (Kenworthy et al. 2000). Further adaptations could be made to utilize
donor beds immediately adjacent to potential recipient sites in order to minimize transport
distance and associated costs. However, Fonseca et al (1994) have documented that existing
natural beds have recovered from excavations as large as 0.5 m2 (~5 ft2) within one year.
Therefore, the substantially smaller seagrass harvest units used with this method can
systematically be collected in appropriate sites without causing adverse, long term damage.
ACKNOWLEDGMENTS
Environmental Affairs Consultants, Inc. (EAC) would like to thank the Manatee County Port Authority for
the opportunity to develop and implement the modified shovel method and to be a part of this successful
seagrass mitigation project at Port Manatee in Tampa Bay, Florida.
46
Transplantation Technique for Shoalgrass
LITERATURE CITED
Fonseca MS. 1994. A Guide to Planting Seagrasses in the Gulf of Mexico. Texas A & M University Sea Grant
College Program, TAMU-SG-94-601. 26p.
Fonseca MS, Kenworthy WJ, Paling E. 1998. Restoring Seagrass Ecosystems in High Disturbance
Environments. Ocean Community Conference, Nov. 16-19, 1998, Baltimore, MD.
Fonseca MS, Kenworthy WJ, Courtney FX, Hall MO. 1994. Seagrass Planting in the Southeastern United
States: Methods for Accelerating Habitat Development. Restoration Ecology Vol. 2 No. 3, pp. 198-212.
Kenworthy WJ, Fonseca MS, Whitfield PE, Hammerstrom K, Schwarzshild AC. 2000. A Comparison of Two
Methods for Enhancing the Recovery of Seagrasses into Propeller Scars: Mechanical Injection of a
Nutrient and Growth Hormone Solution vs. Defecation by Roosting Seabirds. NOAA Damage
Assessment Center, NOAA Marine Sanctuaries Division. http://shrimp.bea.nmfs.gov/Fonseca/Reports.
40p
Lewis RR. 2002a. The Potential Importance of the Longshore Bar System to the Persistence and Restoration
of Tampa Bay Seagrass Meadows. Proceedings of the conference on “Seagrass Management: It’s Not Just
Nutrients.” August 22–24, 2000. St. Petersburg, Florida.
Lewis RR. 2002b. Annual Progress and Mitigation Success Report. Port Manatee Seagrass Mitigation Project,
Manatee County Port Authority October 2001–November 2002. Pursuant to FDEP Permit No. 0129291002-EI and ACOE Permit No. 199801210(IP-MN).
GJM, RFD (Environmental Affairs Consultants, Inc., 429 10th Avenue West, Palmetto, FL 34220)
47
❖
SPECIES SELECTION, SUCCESS, AND COSTS OF MULTI-YEAR, MULTISPECIES SUBMERGED AQUATIC VEGETATION (SAV) PLANTING IN
SHALLOW CREEK, PATAPSCO RIVER, MARYLAND
P. Bergstrom
ABSTRACT
The US Army Corps of Engineers Baltimore District (USACE) funded the US Fish & Wildlife
Service (USFWS) Chesapeake Bay Field Office to do small-scale submerged aquatic vegetation
(SAV) planting in Shallow Creek, at the mouth of the Patapsco River, in 1999, 2000 and 2001.
This was done to compensate for the loss of small beds of SAV that were growing in a channel
in Shallow Creek that was dredged by the USACE in 1998 to maintain navigation for commercial
watermen. Only small-scale SAV planting was done because there were not enough plants,
funding, or suitable shallow areas available to do larger scale planting in this creek. Costs of the
planting effort were calculated for comparison to other small-scale planting efforts. Water clarity
(Secchi depth) in the creek was marginal for SAV planting, so the plantings were quite shallow
(0.4–0.8 m at low tide). The SAV species planted were wild celery (Vallisneria americana), near the
high end of its salinity range, and two at the low end of their salinity range, redhead grass
(Potamogeton perfoliatus), and sago pondweed (Stuckenia pectinata). All were raised by
micropropagation, from seeds, or from cuttings. We planted a total of 2,000 shoots of the three
SAV species without exclosures in 1999, 4,000 shoots of the three species with exclosures in 2000,
and 600 shoots of redhead grass with exclosures in 2001. We found that exclosures greatly
increased plant survival, since survival from the 1999 planting (no exclosures) was less than 10%,
and survival was much higher from the 2000 and 2001 plantings with exclosures. From the 2000
planting, redhead grass had the highest survival of the three species over two years with
exclosures present, followed by wild celery. However, a year after the exclosures were removed
(in fall 2003), most of the redhead grass had disappeared, and most of the wild celery remained.
At the range of shoot densities used for large scale planting in Chesapeake Bay (7,500 to 15,000
shoots/acre), the range of estimated planting costs for the three years was $33,750–$99,000 per
planted acre. The costs per shoot were lowest in 2000 mainly because the planting was larger.
INTRODUCTION
Submerged aquatic vegetation (SAV) in tidal waters of Chesapeake Bay generally grows in
relatively shallow water (1 meter deep or less at low tide), limited in depth by relatively low
water clarity (Batiuk et al. 1992, Chapter III; USEPA 2003, Table IV-15, page 124). The
“shallow water designated use depths” in the latter document are based on the observed
maximum depths of SAV beds mapped in each segment in past SAV surveys (USEPA 2003).
As a result, it is rare for dredging for navigation to cause direct loss of SAV in lower salinity
waters of Chesapeake Bay, since the channels to be dredged are usually already too deep to
have SAV growing in them. However, in rare cases relatively shallow channels need
dredging, and sometimes these channels have SAV growing in them. This occurred when
channels used for commercial navigation in Shallow Creek were dredged by the US Army
Corps of Engineers Baltimore District (USACE). They funded the US Fish & Wildlife
Service (USFWS) Chesapeake Bay Field Office to do small scale Submerged Aquatic
Vegetation (SAV) planting to compensate for the loss of small beds of SAV in the area
dredged.
The plants lost in the dredging in Shallow Creek were Eurasian watermilfoil (Myriophyllum
spicatum), but we did not plant the same species of SAV. Since the planting was done for
compensation and not for mitigation, USFWS and USACE staff agreed that it was best to
49
Bergstrom
plant only native SAV species, even though they were not the same species that was
impacted. This was not considered a substitution because recent Chesapeake Bay Program
guidance considers all of the SAV in the Bay to be equivalent in ecological value
(Chesapeake Bay Program 1995).
However, Shallow Creek was not an area we would have chosen for planting SAV
otherwise, since Secchi depths there were well below the 1 m recommended minimum
water clarity for SAV growth in Chesapeake Bay (Batiuk et al. 1992). In addition, the salinity
was in a transitional range from lower to higher salinity SAV species and quite variable, and
we did not have a long-term record of water quality data. Because we were uncertain as to
which species would grow best, whether fencing was needed, and which were the best
planting sites and times, we planted three species at two different sites in two seasons, with
and without fencing. We refer to this planting strategy as “bet-hedging” to differentiate it
from the more usual “best species” approach. “Best species” is used more often in SAV
planting, although some strategies have used pioneer species to help establish more
persistent species (Fonseca et al. 1998, pp. 107–108). Bet-hedging is especially useful when
survival is hard to predict (Smart and Dick 1999; Stearns 2000), since it should raise the
chance that at least one of the species that are planted will grow well. The cost of this bethedging was that we had fewer units planted per species.
The high costs per acre that we found for this method of hand planting of individual shoots
were expected, since it is very labor intensive. Experiments in Chesapeake Bay are ongoing
to test a planting machine (Fishman et al. 2004) and to develop methods to plant eelgrass
directly from seed (Harwell and Orth 2002, Orth et al. 2003a). The goal of these projects
is to reduce the cost per acre of SAV planting by developing successful methods for largescale restoration, so we can do more of it.
STUDY SITE AND METHODS
Shallow Creek is a small tidal tributary at the north side of the mouth of the Patapsco River
on Chesapeake Bay, in Baltimore County, Maryland. The only SAV found there in 1998
were a few dense beds of Eurasian watermilfoil, Myriophyllum spicatum, and small, scattered
patches of four other SAV species: wild celery, common waterweed (Elodea canadensis),
muskgrass (Chara sp.), and horned pondweed (Zannichellia palustris) (Orth et al. 1999b).
We measured water quality during most visits to the creek. We measured Secchi depth with
an all- white Secchi disk, 20 cm in diameter, to the nearest 0.05 meter. We used the depth
at which the disk was just visible as a circle. If we took more than one measurement we used
the one closest to the planting site. We measured salinity with a refractometer that was
calibrated regularly with distilled water, to the nearest 1 ppt.
We planted three species of SAV in Shallow Creek in three plantings over three years.
Previous test plantings of the native redhead grass (Potamogeton perfoliatus or Ppf) in the
adjacent Old Road Bay had shown initial growth (followed by loss due to waterfowl
grazing), so we planted that species in Shallow Creek. We also planted two native species
that were found in Shallow Creek in small amounts, wild celery (Vallisneria americana or Va)
50
Species Selection, Success and Costs of SAV Planting
and sago pondweed (Stuckenia pectinata or Ppc) to see if either survived better than redhead
grass. All were grown by micropropagation, from seeds or from cuttings, at the
Environmental Center at Anne Arundel Community College (AACC) and the USDA
National Plant Materials Center (NPMC). Most were planted as bare root shoots and some
of the sago pondweed was grown and planted in small peat pots, but the survival of these
two types of propagules was not compared. There were almost always 2–3 stems per
“shoot” but the term “shoot” is used here for convenience, to mean the planting unit that
we got from the lab.
Details of the three plantings are given in Table 1. We planted more wild celery than the
other species in 1999 and 2000, because we expected it to grow better since it already had
healthy beds in the outer cove of Shallow Creek. However, since most of what survived
from the 1999 and 2000 plantings was redhead grass, we planted this species alone in 2001.
We changed the planting density from 25 shoots/m2 in 1999 to 64 shoots/m2 in 2000 and
2001. In 1999, the 25 shoots were spread over 5 rows of 5 plants each in a 1 m × 1 m
square, but we found that these squares were too large for one person to plant without
moving. Once the person planting moved, they tended to lose track of where they had
planted already, since the water was too murky to see the plants. In 2000 and 2001, we
planted 16 shoots in 4 rows of 4 plants each in a 0.5 m × 0.5 m square (0.25 m2), which
allowed the person planting to install a whole square without moving. We planted every
other square so the areal density (including the two unplanted squares) was 32 shoots/m2.
The advantage of planting every other square is that it makes it easier to do the planting
without trampling other plants, and it allows room for natural spread.
Table 1. Planting dates, number of shoots by species, and other details.
PLANTING
DATES
NUMBER
OF SHOOTS
by species
EXCLOSURES?
SHOOT
DENSITY
(per m2)
MONITORING
DATES
(months post-planting)
6/22–23/99
1200 Va
400 Ppf
400 Ppc
No
25
1, 12
6/27–28/00
2000 Va
1000 Ppf
1000 Ppc
Yes
64
3, 14, 25, 37
9/5/01
600 Ppf
Yes
64
11, 23
Planting was done by hand at low tide in shallow water, so that planters did not need
SCUBA or snorkels, and did not have to put their heads underwater. This limits how many
shoots can be planted per day, and it would make it very hard to plant if a wind-driven tide
made the low tide higher than usual. No anchors were used since there was little wave
energy. For the bare root shoots, the person planting excavated a hole with their fingers,
placed the roots in it, and pressed the sediment around it, similar to the single shoot method
described for eelgrass (Zostera marina) by Orth et al. (1999a). The peat pots were planted the
51
Bergstrom
same way except that sometimes a trowel was used to excavate the hole. The side of the pot
was split before it was planted to reduce the chance that the plant would become root bound
in the container. This method is similar to the peat pot method described for use with
eelgrass by Fonseca et al. (1998) except that the sago pondweed that we planted in peat pots
had been grown in the pots, while the eelgrass planted by this method was field harvested
as plugs and placed in the peat pots just before planting.
The rectangular exclosures used in 2000 and 2001 were placed with their long axis
perpendicular to the shore, so the planting depth at low tide ranged from 0.4 m at the
shallow end to 0.8 m at the deep end. Typical Secchi depths near the planting sites were
0.45 to 0.55 m, far less than the 1 m or more that is desirable to allow SAV to grow to 1 m
depth or more, so we planted in relatively shallow water. In Chesapeake Bay, SAV generally
grows as deep as the growing season median Secchi depth (Batiuk et al. 1992). Planting in
shallow water also made planting easier, because many people cannot comfortably plant
using this method in water more than about 0.6 m (2 ft) deep (unless they have long arms).
Exclosures were each about 4 m x 11 m and made out of 1.45 m (4 ft) high plastic fencing
with 5 cm (2 in) openings, attached to 3.8 cm (1.5 in) diameter PVC poles with black cable
ties. The exclosures were small enough to discourage waterfowl from landing inside (as they
have no top). We removed the fencing around the 2000 plants about 26 months after
planting, and plan to remove the fencing around the 2001 plants in 2004. We found that
some plastic fencing lasted much longer than others. We found that flat plastic “warning”
or “safety” fencing with oval openings did not last well (the top rows tore) and it also
became weighed down with barnacles on the wider areas of the plastic. Thicker plastic
warning fencing with diamond mesh did not get as many barnacles, but the top part became
very brittle where it was exposed to light, after as little as 8 months. Green plastic garden
fencing with square openings (made by Tenax Corp.) seemed to last longer than other
types, lasting about two years before it became brittle.
We planted in two areas within the creek in 1999 and 2000. One was on the main creek,
upstream from the former railroad bed that was built across the mouth of the creek. We did
most of the 1999 and 2000 planting here, and also all of the 2001 planting. It had firmer
sediments with lower organic matter than the other site. The second was in the Outer Cove,
which is outside the former railroad bed, where smaller numbers of plants were planted in
1999 and 2000. The sediments here were almost too soft to hold the plants, and there was
also broken glass present that made hand planting dangerous. We chose this cove because
it had very small natural beds of sago pondweed, and larger beds of wild celery, on the side
of the cove across from where we planted. Figure 1 shows the locations of the planting areas
and the natural beds.
Planting was done in June in 1999 and 2000, and in September 2001. The species we
planted are normally planted in the spring in Chesapeake Bay, although we know of no tests
done to compare the survival of spring and fall planting for the three species planted. The
smaller September 2001 planting was added to the original plans because there were some
funds left from 2000, plants were available from AACC, and it was apparent that most of
the surviving plants at that point were redhead grass.
52
Species Selection, Success and Costs of SAV Planting
USACE funding paid for plants and materials only. Planting was done by state and Federal
agency staff on work time and a few volunteers, and monitoring was done by USFWS and
AACC staff.
Figure 1
Dredged Channel
Main creek
(’99, ’00, ’01)
Natural Va beds
Outer Cove
(‘99, ‘00)
Figure 1. Map of Shallow Creek showing the two planting sites, and the location of natural wild celery (Va)
beds and the channel that was dredged.
Monitoring was done at 1 and 12 months after the June 1999 planting. Since coalescence
of plants did not occur, results are presented as percent survival of shoots.
Monitoring was done at 3, 14, 25, and 37 months after the June 2000 planting, and 11 and
23 months after the September 2001 planting (Table 1). When we monitored survival in
the same year in which we planted, we were able to track the survival of individual shoots.
However, due to coalescence of the plants that survived and spread after both plantings, this
was not possible in the following year and in subsequent years. On our visits one and two
years after planting we visually estimated the percent cover of the whole exclosure (about
4 m × 11 m) by species, using the method of Paine (1981) as modified by Orth et al.
(2003b). Since we did not estimate percent cover visually at the time of planting, we
estimated the starting percent cover based on the planting design. Because we planted every
other square in 2000 and 2001, we assumed the starting percent cover over all species in
those years was 50%, assuming there was 100% cover within each planted square. In 2000,
the starting percent cover that we used for wild celery was 25% (because they made up half
of the plants) and 12.5% each for redhead grass and sago pondweed (because they each
made up one- quarter of the plants). Percent cover after 3 months in 2000 was estimated
by multiplying the percent survival of shoots by the initial percent cover. We reported this
53
Bergstrom
visually estimated percent cover as our measure of plant survival over time for the 2000 and
2001 planting. Shoot length was measured only on one visit, in September 2000, so it could
not be compared over time, and is not reported here.
The estimated cost for the in-kind labor used the $550/day cost that USFWS used for
putting a biologist in the field, assuming that each person who did planting devoted a half
day including travel time. Most of those helping worked in the Baltimore or Annapolis area,
within an hour of the planting site by car. We calculated this cost per shoot (including
materials costs) and multiplied this cost by typical numbers of shots/acre for large scale
plantings, to estimate the cost to plant an acre by this method. We were unable to estimate
monitoring costs per acre because we did not keep detailed records on the time spent
monitoring, and we also do not know how long it would take to monitor an area that large.
RESULTS
Water Quality
Table 2 shows that surface salinity varied from 2.5 to 15 ppt, while Secchi depth was less
variable, from 0.45 to 0.85 m. This shows that SAV growing in Shallow Creek in these years
needed to be able to tolerate a wide range of salinity, from oligohaline (0.5–5 ppt) to
mesohaline (5–18 ppt). The relatively low Secchi depths (median 0.55 m) show why we had
to plant in relatively shallow water.
Table 2. Surface salinity and Secchi depth in Shallow Creek, 1998–2003 (nd–no data).
DATE
SALINITY (ppt)
SECCHI (m)
8/15/98
6/22/99
7/30/99
10/21/99
6/27/00
8/2/00
9/27/00
8/15/01
9/5/01
8/8/02
9/26/02
8/7/03
5
7
11
10
nd
5
10
8
nd
10
15
2.5
0.7
0.8
0.55
0.8
nd
0.45
0.85
0.45
0.55
0.45
0.55
0.45
Median
9.0
0.55
NOTES
10 months before first planting
first planting (also on 6/23)
one month after planting
all SAV had died back
second planting (also on 6/28)
3 months after second planting
third planting, 14 months after second planting
25 months after second planting
most SAV had died back
37 months after second planting
Survival of Planted SAV
June 1999 Planting
Plants in this first planting were unfenced (no exclosures). One month after planting,
between 42% and 73% of the shoots had survived (overall mean 52%), but there were signs
of grazing. There was also competition with Eurasian watermilfoil at the Outer Cove site.
We did not find any of these plants in 2000, and we put some of the new plants in 2000 in
the same areas where we had planted in 1999. We found about 10% of the 1999 redhead
grass plants in 2001, near the exclosures we had planted in 2000. We concluded that small
54
Species Selection, Success and Costs of SAV Planting
plantings in this creek had to be fenced, so we fenced the plantings that we did in 2000 and
2001.
June 2000 Planting
Figure 2 shows the percent cover over time of these plants, using mean percent cover over
the four exclosures. In late September 2000, three months after planting, the percent cover
of redhead grass shoots and wild celery shoots had fallen only slightly from the starting
values, but percent cover of sago pondweed shoots was about half of the starting value
(Figure 2). In August and September 2001, 14–15 months after planting, we found that
redhead grass had expanded to cover more of the bottom than we had planted (expanding
from 12.5% to 62% cover), while wild celery percent cover was about the same as the
original plant cover, and very little sago pondweed had survived (Figure 2). In general, the
deeper ends of each exclosure had more plants than the shallower ends, probably due to
grazing (see Discussion).
100
Figure 2
% cover over exclosure
90
80
70
Va
Ppc
Ppf
60
50
40
30
20
10
0
0
10
20
30
40
Time after June 2000 planting (mo)
Figure 2. Mean percent cover by species of SAV planted in Shallow Creek in June
2000, averaged over the four exclosures used.
In August 2002, 25 months after planting, redhead grass percent cover had fallen to 40%
averaged over all four exclosures, but all of the plants were in the three exclosures on the
main creek, where they had spread to small areas outside the fences. Redhead grass did not
survive as well in the fourth exclosure in the Outer Cove, covering only 1%. Wild celery had
an overall mean of 21% cover, slightly less than the starting cover, and a few of the plants
had formed seeds. No sago pondweed was found. As we found after 14 months, survival
was better at the deeper end of each exclosure after 25 months. In August 2001 some of the
fences had fallen down and there were cropped plants at the shallow ends of the exclosures.
55
Bergstrom
We repaired the fences on that visit and a month later there were new shoots in the shallows
of those exclosures.
Salinity in the creek reached 15 ppt (by refractometer) on 9/26/02, when most of the
redhead grass and all of the wild celery had died back. We removed the fencing on the
exclosures from 2000 on this visit, but left the poles in place.
In August 2003, 37 months after planting and after about a year without fencing, the
situation had changed. Almost all of the redhead grass was gone (0.5% cover) but the
percent cover of wild celery had increased slightly from the previous year (Figure 2).
September 2001 Planting
In August 2002, 11 months after planting, the whole exclosure was full of redhead grass, or
twice as much cover as was present at planting. There was also some spread of redhead grass
beyond the planted area. A month and a half later on 9/26/02, salinity in the creek reached
15 ppt by refractometer, and most of the redhead grass had died back. In August 2003, 23
months after planting, percent cover of redhead grass had fallen to 50%, or half of what it
was a year before, and about the same as it was at planting. The fences were left up and will
be removed in spring 2004.
Costs
The estimated costs by year for 1999–2001 suggest that larger plantings are more efficient,
since the largest planting had the lowest cost per unit planted. The estimated planting costs
by year were:
1999: Total cost $12,700 for 2,000 shoots ($6.35/shoot)
Plants & materials $2,800 (no fencing)
Labor $9,900 (in kind, 18 person-days @ $550)
2000: Total cost $18,100 for 4,000 shoots ($4.50/shoot)
Plants & materials $6,000 (added fencing)
Labor $12,100 (in kind, 22 person-days @ $550)
2001: Total cost $3,950 for 600 shoots ($6.60 /shoot)
Plants & materials $650
Labor $3,300 (in kind, 6 person-days @ $550)
Translating these costs per shoot into a cost per acre depends on how many shoots are
planted per acre. There is no consensus on how many shoots are required to consider an
acre has been “planted” with SAV. For example, the current Wilson Bridge mitigation
plantings in the Potomac River have an average density of 7,500 shoots/acre (2-foot
spacing), while tests of Jim Anderson’s SAV planting boat done in Maryland in summer
2003 planted single shoots of wild celery at twice that density, 15,000 shoots per acre.
56
Species Selection, Success and Costs of SAV Planting
Using that range of density, 7,500–15,000 per acre, the range of costs per planted shoot
documented in this project would result in costs of between $33,750 and $99,000 per
planted acre. Costs per hectare would be 2.47 times higher, or between $83,362 and
$244,530 per planted hectare.
DISCUSSION
The range of planting costs documented in this study included the cost per acre for
collection, preparation, and installation of seagrasses found by Fonseca et al. (2002), which
was $45,000 per acre. However, they found this was only 18% of the total project cost,
which they estimated was $245,000 per acre, with 59% of the total spent on monitoring
(Fonseca et al. 2002). Unfortunately, costs for larger scale SAV planting using a machine
in Chesapeake Bay are not yet available, since the one published study focused on survival
rather than cost (Fishman et al. 2004).
The high salinity (up to 15 ppt) in Shallow Creek during the drought in 2002 could have
limited the growth of wild celery that year. Redhead grass survived at 15 ppt in 2002 in the
Magothy River, the next tributary south of the Patapsco on the western shore of Chesapeake
Bay, but wild celery did not grow well. In 2002 the small beds of wild celery in the Magothy
on South Ferry Point were partly brown and smaller than in previous years (pers. obs.).
Grazing by waterfowl appeared to limit SAV growth, either when fencing on the exclosures
was partly down, or after it was removed. Mute swans, Canada geese, and mallards were all
seen in Shallow Creek on several visits (pers. obs.). We did not observe birds eating plants
but did find cropped plants in August 2001 that had new growth a month later after fencing
was repaired. The much higher survival of wild celery after the fencing was removed,
compared to redhead grass, could be due to a preference of the grazing waterfowl for
redhead grass over wild celery. The persistent natural wild celery beds in this creek, with
several species of waterfowl present, suggest that they may not graze it heavily. Direct
observations are needed on the feeding preferences of waterfowl on different SAV species,
studying the preferences of mute swans in particular.
Bet-hedging (by planting multiple species at the same place and time) was useful in this case
because we found that a species that did not grow naturally in the creek, redhead grass, grew
better over the first two years than two species that grew there naturally, wild celery and
sago pondweed. This difference in survival among the three species may have been partly
due to the high salinity in the planting years due to drought. Although the well-established
natural wild celery beds in the creek were apparently able to tolerate the elevated salinity,
it may be that the added stress of transplanting made it hard for the transplanted wild celery
to survive. Sago pondweed should be able to tolerate salinity higher than what was found
in Shallow Creek, but it generally has not survived well when transplanted in mesohaline
waters of Chesapeake Bay, so its poor survival was probably not related to the high salinity.
The decline of redhead grass and persistence of wild celery that we found three years after
planting, one year after the fencing was removed, may have been due to more grazing on
redhead grass, as discussed above.
57
Bergstrom
ACKNOWLEDGMENTS
Thanks to Mike Norman, AACC, for help with project planning, monitoring, and planting, USACE Baltimore
District for funding for plants and materials, federal & state Agency staff and citizens who did the planting each
year, USFWS and AACC for boats and staff support, and AACC (Mike Norman) and NPMC (Jen Kujawski)
for growing the plants. Thanks also to Steve Ailstock, Mike Norman, Jen Kujawski, Mark Mendelsohn, Mark
Fonseca and Jud Kenworthy for reviewing the manuscript.
LITERATURE CITED
Batiuk RA, Orth RJ, Moore KA, Dennison WC, Stevenson JC, Staver LW, Carter V, Rybicki NB, Hickman
RE, Kollar S, Bieber S, Heasly P. 1992. Chesapeake Bay Submerged Aquatic Vegetation Habitat
Requirements and Restoration Targets: A Technical Synthesis. Chesapeake Bay Program, Annapolis, MD.
CBP/TRS 83/92, 248 pp.
Chesapeake Bay Program. 1995. Guidance for Protecting Submerged Aquatic Vegetation in Chesapeake Bay
from Physical Disruption. Chesapeake Bay Program, Annapolis, MD. EPA 903-R-95-013,
CBP/TRS139/95. 29 pp. Available online: http://www.chesapeakebay.net/pubs/SAVguidance.pdf
Fishman JR, Orth RJ, Marion S, Bieri J. 2004. A Comparative Test of Mechanized and Manual Transplanting
of Eelgrass, Zostera marina, in Chesapeake Bay. Restoration Ecology 12: 214–219
Fonseca MS, Kenworthy WJ, Thayer GW. 1998. Guidelines for the conservation and restoration of seagrasses
in the United States and adjacent waters. NOAA Coastal Ocean Program Decision Analysis Series No.
12. NOAA Coastal Ocean Office, Silver Spring, MD.
Fonseca MS, Kenworthy WJ, Julius BE, Shutler S, Fluke S. 2002. Chapter 7: Seagrasses. Handbook of
Ecological Restoration, Vol. 2, M. R. Perrow & A. J. Davy, eds. Cambridge University Press, Cambridge.
Harwell MC, Orth RJ. 2002. Long-distance dispersal potential in a marine macrophyte. Ecology 83(12):
3319–3330.
Orth, RJ, Harwell MC, Fishman JR. 1999a. A rapid and simple method for transplanting eelgrass using single,
unanchored shoots. Aquatic Botany 64: 77–85.
Orth RJ, Nowak JF, Wilcox DJ, Whiting JR, Nagey LS. 1999b. Distribution of Submerged Aquatic Vegetation
in the Chesapeake Bay and Tributaries and the Coastal Bays–1998. VIMS Special Scientific Report
Number 140. Available online: http://www.vims.edu/bio/sav/sav98/index.html
Orth RJ, Fishman RJ, Harwell MC, Marion SR. 2003a. Seed-density effects on germination and initial
seedling establishment in eelgrass Zostera marina in the Chesapeake Bay region. Marine Ecology Progress
Series 250: 71–79.
Orth RJ, Wilcox DJ, Nagey LS, Owens AL, Whiting JR, Serio A. 2003b. 2002 Distribution of Submerged
Aquatic Vegetation in Chesapeake Bay and Coastal Bays . VIMS Special Scientific Report Number 139.
Available online: http://www.vims.edu/bio/sav/sav02/index.html
Paine DP. 1981. Aerial Photography and Image Interpretation for Resource Management. John Wiley & Sons, Inc., New
York City, NY. 571 pp.
Smart RM, Dick GO. 1999. Propagation and Establishment of Aquatic Plants: A Handbook for Ecosystem
Restoration Projects. Technical Report A-99-4, US Army Corps of Engineers, Waterways Experiment
Station. Accessed from <http://el.erdc.usace.army.mil/elpubs/pdf/tra99-4.pdf>
Stearns,SC. 2000. Daniel Bernoulli (1738): evolution and economics under risk. J. BioSci. 25: 221–228.
USEPA. 2003. Technical Support Document for Identification of Chesapeake Bay Designated Uses and
Attainability. USEPA Chesapeake Bay Program, Annapolis, MD. EPA 903-R-03-004. Online at:
http://www.chesapeakebay.net/uaasupport.htm
PB: NOAA Chesapeake Bay Office, 410 Severn Ave. Suite 107A, Annapolis, MD 21403. E-mail
peter.bergstrom@noaa.gov
58
USING TERFS AND SITE SELECTION FOR
IMPROVED EELGRASS RESTORATION SUCCESS
F.T. Short, R.C. Davis, B.S. Kopp, J.L. Gaeckle & D.M. Burdick
ABSTRACT
Techniques are being developed to try to offset losses of eelgrass, Zostera marina L. by
transplanting and seeding. For eelgrass restoration, both site selection and transplanting method
are critical to achieve success at the lowest possible cost. In the northeastern U.S., we have
developed a model for choosing sites that will most likely sustain eelgrass after restoration. The
site selection model is now being configured in a GIS format that will produce maps showing
the optimum restoration areas.
The high cost of restoring eelgrass beds in subtidal environments, and the difficulty in protecting
transplants from various bioturbating organisms, led us to develop a new method not requiring
SCUBA. “Transplanting Eelgrass Remotely with Frame Systems” (TERFS) is a modification of
bare-root transplanting methods. Eelgrass shoots are attached with biodegradable ties to weighted
wire frames that provide protection from uprooting and bioturbation. The TERFS are then
deployed from any small boat. After three to five weeks, the frames are retrieved and ready for
reuse, leaving behind dense patches of eelgrass. We tested TERFS first in the Great Bay Estuary,
NH and again in New Bedford Harbor, MA; in both cases the TERFS method was successful.
The ease and success of this technique provides a restoration approach that can involve citizen
volunteers. More importantly, it significantly reduces the cost of eelgrass restoration, $3.04 per
planting unit vs. $5.28 per planting unit for the earlier horizontal rhizome method. In addition
to costing less, the TERFS method provided a higher level of initial survival, and 75% of sites
transplanted with TERFS were successful after one year.
INTRODUCTION
Eelgrass (Zostera marina L.) meadows provide a wide array of ecological functions that are
important for maintaining healthy estuarine and coastal ecosystems. Eelgrass meadows form
a basis of primary production that supports ecologically and economically important species.
Over the last two decades, eelgrass populations have declined due to pollution associated
with increased human populations, as well as other human-induced and natural
disturbances. Because of the critical role eelgrass habitat plays in estuarine and coastal
systems, efforts are underway to prevent further losses and, more recently, to restore
eelgrass populations to historic distributions. However, once eelgrass cover is lost, physical
and biological site characteristics may change. Most notably, declines in water quality
commonly prevent success of restoration efforts. Other changes can prevent natural
recolonization of historic eelgrass sites even when water quality is adequate (Dennison et
al. 1993). Transplanting can establish eelgrass habitat decades before natural processes might
permit recolonization. Eelgrass transplanting has been used to restore habitat as well as to
mitigate for eelgrass loss or damage (Fonseca et al. 1998, Short et al. 2002a).
Selection of transplanting sites is probably the most critical step in any eelgrass restoration.
A site selection model was developed that synthesizes available historic data and literature,
data from reference sites, simple field measurements, and test transplants to identify and
prioritize locations for large-scale eelgrass transplanting (Short et al. 2002a, Short and
Burdick 2005). Development of our site selection model was based on the physical and
biological characteristics associated with the most successful transplants sites in a mitigation
project in New Hampshire and a restoration effort in Massachusetts. If resources are not
59
Short, Davis, Kopp, Gaeckle & Burdick
available for a complete site selection model analysis, an assessment of the depth range in
which eelgrass grows and a test transplanting study will provide some of the critical
information needed to identify successful transplant sites.
Many methods have been developed to restore degraded seagrass habitat or to mitigate
damage to seagrass beds. Although some work is being done to restore seagrasses from seed
(Harwell and Orth 1999, Granger et al. 2000), the most widely used and most consistently
effective methods involve transplanting seagrass shoots from healthy donor beds (Fonseca
et al. 1998, Short et al. 2002a). Because transplanting is labor intensive and often requires
heavy reliance on SCUBA, these methods are very costly. We have developed and tested a
less expensive eelgrass restoration method that simultaneously transplants and protects
eelgrass shoots without the need for SCUBA, using remotely deployed wire frames. We
named the method “Transplanting Eelgrass Remotely with Frame Systems” or TERFS1.
The TERFS method was used to meet the objectives of an eelgrass habitat restoration
project in New Bedford Harbor, Buzzards Bay, Massachusetts (USA). New Bedford
Harbor is an estuary contaminated with polychlorinated biphenyls (PCBs) and heavy metals
(particularly copper) from years of industrial discharge. The TERFS method allowed us to
transplant eelgrass while avoiding direct contact with these sediments. Additionally, the wire
frame of the TERFS provided structure that acted as caging of newly transplanted shoots,
protecting them from bioturbating organisms. The use of TERFS also allowed us to build
community support for the project by involving citizen volunteers in TERFS preparation
and deployment (Burdick-Whitney and Short 2002, Short et al. 2002c).
METHODS
The site selection model determines the best areas for eelgrass transplanting (Short et al.
2002a, Short and Burdick 2005). A Preliminary Transplant Suitability Index (PTSI) is first
calculated based on available data from the literature and project reports for the estuary; the
PTSI identifies areas with sufficient potential to merit test transplanting. In New Bedford
Harbor, twenty potential sites were test transplanted with TERFS in 1998. Then a final
Transplant Suitability Index (TSI) was calculated. The TSI uses a combination of the PTSI,
the success of the test transplants, and measured eelgrass growth and nitrogen content of
eelgrass leaf tissue to further refine the choice of transplanting sites. In New Bedford, we
narrowed the potential restoration areas from twenty to four sites with the TSI (Short et al.
2002b). Full scale transplanting of a total of 4 acres of eelgrass using TERFS at these four
sites in 1999 was the first major test of the site selection model.
TERFS is an eelgrass transplanting method that is a modification of the horizontal rhizome
method (HRM) developed by Davis and Short (1997). The HRM is a bare root
transplanting method using two overlapping, opposed shoots (a “planting unit” or PU)
stapled to the bottom with a bamboo skewer on 0.5m centers. A large restoration project in
Great Bay Estuary, NH was completed in 1995; 7 acres of eelgrass were transplanted using
1
TERFS: the acronym itself (not the method) is registered as a trademark of the University of New Hampshire. The
method may be used by anyone, with an acknowledgment of the trademark to UNH.
60
TERFS and Site Selection for Eelgrass
HRM. All of the planting required caging around the transplanted plots to reduce crab
damage and planting was done with SCUBA; while expensive, the project met its success
criteria (including 50% long-term survival; Short et al. 2000) established at the beginning
of the project, and much of the eelgrass persists today.
In the TERFS method, opposing pairs of eelgrass shoots are attached with biodegradable
ties to rubber-coated wire frames (Fig. 1). Twenty-five pairs of plants (or PUs) are tied to
each frame 5 cm apart (Short et al. 2002c). The frame, which is weighted with bricks and
deployed from a boat, presses the eelgrass rhizomes into the top centimeter of substrate
(Fig. 2). The weighted wire frame acts to hold the new transplants in place while they take
root, and also to protect the eelgrass shoots from bioturbating organisms. After a period of
about one month (depending on the season), eelgrass shoots have rooted, and the frame is
removed. The frame can then be used again. The technique creates a 0.25 m2 patch of
eelgrass at the relatively high shoot density of 200 m-2(Fig. 3). A detailed description of the
TERFS methodology is available on a CD, “Using TERFS for Community-based Eelgrass
Restoration” (Burdick-Whitney and Short 2002) which includes a downloadable manual,
A Manual for Community-Based Eelgrass Restoration (Short et al. 2002c), or the manual may be
obtained as a pdf from the first author.
Figure 1. A weighted wire mesh frame with eelgrass attached; the
Transplanting Eelgrass with Remote Frame Systems (TERFS) method.
Preliminary testing of the TERFS method was conducted in 1996 at a site in Broad Cove
in the Great Bay Estuary, NH. Donor shoots were harvested from an intertidal eelgrass bed
at Gerrish Island, ME within the same estuarine system. Eelgrass was previously
transplanted in Broad Cove using the HRM, but failed due to problems with bioturbation
by clam worms (Davis 1999). Bioturbation in this context is the disturbance of plants by
living creatures. One year after transplanting with TERFS, all four created eelgrass areas
persisted and had increased four times in shoot density.
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Short, Davis, Kopp, Gaeckle & Burdick
Figure 2. TERFS immediately after being placed on the sediment surface.
Figure 3. Eelgrass patches created by TERFS, one year after planting.
In the summer of 1998, the TERFS method was again tested at three sites in the Great Bay
Estuary. In addition, 20 sites were selected for test transplanting in the greater New Bedford
Harbor area. Donor beds for the New Bedford trials were located within the study region,
and shoots were transplanted within three days of their harvest. At each test site, four
TERFS were deployed at least 0.5 m apart. After a period of approximately one month, the
TERFS were retrieved and the initial success (percent shoot survival after one month)
determined. We compared TERFS with the HRM by planting 12 of the New Bedford test
sites using both methods. Similar to the TERFS, the HRM plots consisted of 25 PUs;
however, in the HRM, the PUs were on 0.5 m centers in 2 m x 2 m plots.
62
TERFS and Site Selection for Eelgrass
Costs were compared between HRM and TERFS. Direct cost comparison is difficult
because each method and each project has specific requirements. In both cases, the
monitoring costs represent a post-planting assessment of eelgrass survival one month after
transplanting plus measurements after one full year of canopy structure, biomass, and
habitat function (fish use, benthic infauna and epibenthic community). Administrative costs
were included and represent actual salaries and wages, including costs for design and
implementation of each project.
RESULTS
Our results showed the TERFS method to be highly effective in creating eelgrass patches,
even at sites where conventional transplanting had previously failed (Table 1). One month
survival of eelgrass planted using the TERFS method in the Great Bay Estuary in 1998
ranged from 53% to 86% at various sites (Table 1). Initial survival of TERFS in New
Bedford Harbor in 1998 ranged from 47% to 83%, with three of eight sites having at least
80% one month survival. Comparison of initial survival between the TERFS and HRM at
twelve sites in New Bedford Harbor showed that all twelve TERFS sites survived, while
only four of the twelve HRM sites had eelgrass survival. Eelgrass survival using the HRM
exceeded survival using TERFS at only one site. In the full transplanting in New Bedford
Harbor, 75% of sites using TERFS were successful after one year.
Table 1. Outcomes of eelgrass transplanting with TERFS in the Great Bay Estuary (NH) and New Bedford
Harbor (MA) study sites. Initial survival is a measure of shoots one month after transplanting. *Test
transplanting, four TERFS per site.
SITE
Great Bay Estuary 1998*
Broad Cove
INITIAL SUCCESS
OUTCOME
ONE YEAR SUCCESS
—
4x increase in density
Great Bay Estuary 1998*
Broad Cove
Bellamy River
Schiller
86%
53%
71%
—
—
—
New Bedford Harbor 1998*
Buzzards Bay
Site 20 New Bedford
Site 1 New Bedford
Site 5 New Bedford
Site 8 Dartmouth
Site 6 Dartmouth
Site 18 Fairhaven
Site 16 Fairhaven
Site 3 Fairhaven
80%
50%
47%
83%
47%
81%
48%
52%
—
—
—
—
—
—
—
—
—
75%
New Bedford Harbor 1999
4 acres transplanted
(full scale transplanting)
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Short, Davis, Kopp, Gaeckle & Burdick
We applied the site selection model (Short et al. 2002a) to the restoration sites in New
Bedford Harbor. The PTSI indicated ten sites for further analysis and test transplanting.
Of those ten sites, the TSI identified four as preferred locations for full-scale eelgrass
transplanting. Of those four sites, three were successful, producing our 75% success. The
one site that failed was heavily impacted by a bloom of Codium fragile that drifted over the
transplant area and smothered the eelgrass.
Costs were compared between the HRM and TERFS eelgrass transplanting methods on a
“per planting unit” basis (Table 2). A planting unit consists of two shoots of eelgrass.
Collecting plants from the donor site costs the same in both cases: $0.19 per planting unit.
Monitoring and administration costs were also the same for each method. Equipment was
more expensive for the HRM than for the TERFS method since SCUBA equipment is
required: $0.78 as opposed to $0.55. The HRM also had the cost of caging, at $0.42 per
planting unit. TERFS do not require caging, so that cost is avoided. The major difference
in cost between HRM and TERFS is in the actual transplanting. Transplanting with HRM
using SCUBA costs $2.39 per planting unit while transplanting without SCUBA with
TERFS costs only $0.80 per planting unit. In total, the cost per planting unit for HRM is
$5.28 while for TERFS it is $3.04. On a per acre basis, HRM costs $79,200 and TERFS
costs $45,600. All costs are expressed in 1998 dollars (US). The calculations for both
methods are based on transplanting 15,000 planting units per acre, which for TERFS means
deploying 600 frames per acre. We are conducting experiments to determine the best pattern
for deploying the TERFS on the bottom to maximize large scale restoration success.
Table 2. Costs of HRM and TERFS eelgrass transplanting methods on a “per planting unit” basis. A planting
unit consists of two shoots of eelgrass. All costs are expressed in 1998 dollars (US).
METHOD
HRM
TERFS
Collecting plants from the donor site
Monitoring
Administration
Equipment
Additional caging
Actual transplanting
$0.19
0.70
0.80
0.78
0.42
2.39
$0.19
0.70
0.80
0.50
0.00
0.80
Total cost per planting unit
$5.28
$3.04
$79,200
$45,600
On a per acre basis
DISCUSSION AND CONCLUSIONS
Choosing the site for eelgrass transplanting is the step that makes the difference between
success and failure of the restoration effort; site selection deserves detailed assessment and
analysis. For restoration as part of compensatory mitigation and most other transplanting
or seeding operations, a rigorous quantitative site selection model should be applied (Short
et al. 2002a). In the case of volunteer-based restoration activities, determining the depth
range of eelgrass growth and test transplanting can suffice. The use of test transplants is
64
TERFS and Site Selection for Eelgrass
recommended for any eelgrass restoration activity, because it provides the most absolute
method of evaluating all the water quality and other physical and biological conditions that
may limit restoration success at a particular site. If test transplants survive and spread, there
is a high likelihood that full-scale transplanting efforts will be successful. Ideally, test
transplant success should be judged a full year, or at least full growing season, after planting
to achieve the most reliable habitat evaluation.
The TERFS technique provides an ideal, quick and inexpensive methodology for test
transplanting. Because the TERFS are easy to deploy at any site and provide replicate
transplant shoots for evaluation, they can be used by any group interested in evaluating
restoration potential of an area. SCUBA is not required in using TERFS, which reduces the
cost and eliminates safety and liability issues arising in most other transplant methods.
The cost of seagrass transplanting can be greatly reduced using TERFS. Transplanting with
TERFS can be done at a lower cost per planting unit than other bare root methods. The one
acre costs of $79,200 for HRM and $45,600 for TERFS clearly shows the lower cost of
using the TERFS method. However, the relatively low cost of TERFS can be substantially
further reduced by using volunteer labor (Fig. 4). Much of the eelgrass collection and
sorting (see Davis and Short (1997) for collection method), the attaching of eelgrass shoots
to the planting frames, and the monitoring can be done by community volunteers (BurdickWhitney and Short, 2002), cutting the overall cost of a TERFS transplanting effort by almost
half.
Figure 4. Citizen volunteers attaching eelgrass shoots to TERFS
transplanting frames.
65
Short, Davis, Kopp, Gaeckle & Burdick
Additionally, the TERFS have the advantage of providing a self-caging structure, creating
protection from bioturbation. A wide variety of bioturbating organisms, from worms to
swans, have been identified in relation to seagrasses (Short et al. 2002a). Often areas that
have lost their eelgrass are difficult to restore due to bioturbation; new transplants are
disturbed by crabs, clam worms, lobsters, horseshoe crabs, rays, or snails. Bioturbation can
be a major challenge, and caging is often required to exclude these animals while the
eelgrass gets established. TERFS, with the wire frame that surrounds the transplanted
shoots, eliminates the need for additional caging and provides their own protection. By the
time the frame is removed, the eelgrass is established well enough to deter most organisms,
or at least survive most bioturbation problems.
No staples or other debris are left behind after the TERFS planting. The frames are
retrieved from the planting area, cleaned, dried and reused. Additionally, TERFS can be
used in contaminated areas or in conditions unsuitable for diving. Since the frames are
deployed and retrieved from a boat or by wading, SCUBA is not needed. Creating eelgrass
beds in these areas will accelerate the process of sediment accumulation and the burial of
contaminated bottom layers.
Restoration using TERFS and volunteer workers can be very effective because it provides
inexpensive labor, educates community groups, and creates advocates for the coastal
environment. In New Bedford Harbor (MA) and Portsmouth Harbor (NH), we found
community volunteers from a wide age span to be very capable, under the direction of a
scientist or graduate student, of participating fully in TERFS transplanting efforts. In
Narragansett Bay (RI), Save the Bay is successfully conducting volunteer-based eelgrass
restoration efforts using TERFS. Real savings are achieved when volunteer labor is used.
Of course, supervision is necessary and food, drink, shade and training must be provided.
There is the “front end investment” of phone calls and work with community groups to
organize volunteer participation. Overall, community-based restoration efforts are a good
way to create a group of knowledgeable advocates for the coastal environment while making
eelgrass restoration efforts affordable and more practicable. Beyond this, the involvement
of citizen volunteers in an eelgrass restoration effort often attracts local media coverage,
spreading the word about coastal habitat issues.
The TERFS method creates high density (200 shoots m-2) patches of eelgrass. We have
found that eelgrass transplanted in patches is more successful than planting units spaced on
the 0.5m centers typical of HRM and other bare root methods. The patches spread because
of the eelgrass rhizome growth habit and are more resistant to environmental and biological
challenges in the early stages of expansion. Additionally, patches of eelgrass as created by
TERFS appear to be more attractive to fish and invertebrates than sparsely spaced planting
units (unpublished data).
66
TERFS and Site Selection for Eelgrass
In conclusion, all restoration of subtidal habitat is costly. Using the site selection model in
conjunction with TERFS is a high-value investment. The site selection model eliminates
large scale expenditures at sites that are unlikely to support eelgrass. Once suitable sites are
selected, TERFS are a good transplanting method, since they yield greater transplant success
than other methods, invite community participation, and cost less.
REFERENCES
Burdick-Whitney CL, Short FT. 2002. Using TERFS for Community-based Eelgrass Restoration, CD-ROM,
Report to the NOAA Restoration Center. Jackson Estuarine Laboratory, University of New Hampshire,
Durham, NH.
Davis RC. Short FT. 1997. Restoring eelgrass, Zostera marina L., habitat using a new transplanting technique:
the horizontal rhizome method. Aquatic Botany 59:1-15.
Davis RC. 1999. The effect of physical and biological site characteristics on the survival and expansion of
transplanted eelgrass (Zostera marina L.). Ph.D. Thesis, University of New Hampshire, Durham, NH.
Fonseca,MS, Kenworthy WJ, Thayer GW. 1998. Guidelines for the Conservation and Restoration of
Seagrasses in the United States and Adjacent Waters. NOAA Coastal Ocean Program Decision Analysis
Series, No. 12. NOAA Coastal Ocean Office, Silver Spring, MD 222 pp.
Harwell MC, Orth RJ. 1999. Eelgrass (Zostera marina L.) see protection for field experiments and implications
for large-scale restoration. Aquatic Botany 64: 54-61
Granger SL, Traver MS, Nixon SW. 2000. Propagation of Zostera marina L. from seed. Chp. 107, pp. 4-5. In
C.R.C. Sheppard (ed.) Seas at the Millennium: An Environmental Evaluation. Vol. III, Global Issues and
Processes. Pergamon, Amsterdam.
Short FT, Burdick DM. 2005. Eelgrass restoration site selection model. CD-ROM and manual. CICEET,
University of New Hampshire, Durham, NH.
Short FT, Burdick DM, Short CA, Davis RC, Morgan PA. 2000. Developing success criteria for restored
eelgrass, salt marsh and mud flat habitats. Ecological Engineering 15: 239-252.
Short FT, Davis RC, Kopp BS, Short CA, Burdick DM. 2002a. Site selection model for optimal transplanting
of eelgrass, Zostera marina L., in the Northeastern U.S. Marine Ecology Progress Series 227: 253-267.
Short FT, Kopp BS, Gaeckle J, Tamaki H. 2002b. Seagrass ecology and estuarine mitigation: a low-cost
method for eelgrass restoration. Proceedings of International Commemorative Sympsium 70th
Anniversary of the Japan Society of Fisheries Science. II. Fisheries Science. 68: 1759-1762.
Short FT, Short CA, Burdick-Whitney CL. 2002c. A Manual for Community-Based Eelgrass Restoration.
Report to the NOAA Restoration Center. Jackson Estuarine Laboratory, University of New Hampshire,
Durham, NH. 54 pp.
FTS, JLG, DMB (Jackson Estuarine Laboratory, University of New Hampshire, 85 Adams Point Road,
Durham, NH 03824: fred.short@unh.edu); RCD (Exponent, 4 Computer Drive West, Suite 201, Albany, NY
12205); BSK (United States Geological Survey, 196 Whitten Road, Augusta, Maine 04330)
67
❖
SEAGRASS SCARRING IN TAMPA BAY:
IMPACT ANALYSIS AND MANAGEMENT OPTIONS
J.F. Stowers, E. Fehrmann & A. Squires
ABSTRACT
There is little argument when discussing the value of seagrasses. These areas have extremely high
productivity and diversity and are frequented by endangered species such as the manatee.
Scarring of seagrass beds, mainly by boat propellers, has occurred throughout Tampa Bay and
as a result, many groups have taken steps to document the impacts and regulate access within
areas of seagrass coverage.
Pinellas County has been active in seagrass protection for over a decade with success in both
regulatory and experimental processes. In addition to the seagrass protection ordinance, a study
was commissioned to determine if nutrient injection was effective in enhancing regrowth of
seagrass into prop scars. Continued cooperation between a coalition of representatives from
government, educational institutions and environmental interest organizations as well as user
groups from both the recreational and commercial interests will be required for continued
success.
INTRODUCTION
Many studies have been performed documenting the value of seagrasses. These studies have
shown the extremely high productivity and diversity of both finfish and shellfish that utilize
these areas as both a nursery and refuge. Predator species are naturally drawn to seagrass beds
due to the prey species density, which in turn attract sportfishers seeking a challenge. These areas
are also frequented by endangered species such as the manatee. It has long been known that
scarring of seagrass beds, mainly by boat propellers, has occurred throughout Tampa Bay. As a
result, state and local governments, as well as educational institutions have taken steps to
document the impacts and regulate access within areas of seagrass coverage.
Documentation of the actual damage incurred can be a costly and labor-intensive effort but a
combination of aerial photography, photointerpretation, and extensive field verification can
result in very accurate estimates of seagrass damage. Seagrass scarring has become more pervasive
as more boats are registered and used in the Tampa Bay area. Technical reports by the FDEPFlorida Marine Research Institute indicate that moderate/severe scarring in Tampa Bay averages
nearly 30% of the total coverage by seagrass, some of the worst rates in the state (Sargent et
al.1995). Other studies have shown that when scarring becomes severe, the majority of the
habitat and water quality functions are lost and the whole bed may lose the ability to regenerate
and cease to exist (Sargent et al. 1995, Ehringer 1999). Finfish and shellfish production declines,
which in turn can severely affect the local commercial harvest economy as well as the
recreational fishery.
FT. DESOTO SEAGRASS PROTECTION EFFORTS
Pinellas County became concerned with seagrass scarring and cumulative impacts due to
boat propeller scarring in the mid- to late 1980s. Pinellas County’s initiatives began in 1990
and involved a coalition of regulatory and citizen representatives. These included both
commercial and recreational fishing interests. Many meetings were held to discuss the
issues to build a consensus about a solid action plan for the Ft. DeSoto area that would build
support as well as provide the needed resource protection. The group had reached a
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Stowers, Fehrmann & Squires
consensus by the end of 1991 and an ordinance was drafted and adopted in the beginning
of 1992 (Ordinance 92-11, since codified under later iterations in the Pinellas County
Code).
The ordinance provided that the Ft. DeSoto management area be divided into zones that:
• eliminated the use of internal combustion engines (exclusion zones)
• allowed use of engines, but imposed penalties for damage to seagrass (caution zones)
• required idle speed (allowed for engine use in exclusion zones to gain access to
features such as campsites)
• had no protection (control areas)
The ordinance provided that the zones be clearly marked (Figure 1) and that the County
monitor the zones for 5 years to determine the effectiveness of the management plan. The
original ordinance also included a “sunset clause” that required it to be renewed each year.
This proved to be a non-issue as opposition to the ordinance had disappeared. The sunset
clause was removed when the ordinance was renewed in 1993.
Figure 1. Typical sign located at area boat ramps and marinas.
The County believed that the best course of action was to take low altitude aerial
photographs of the Ft. DeSoto Management Area and then have them digitized and
interpreted by a seagrass specialist. Aerials were flown in 1992 prior to installation of signs
to provide a “baseline.” A second set of aerials was flown in 1992 right after sign installation.
Thereafter aerials were flown annually until 2000.
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Seagrass Scarring in Tampa Bay
Dr. Nicholas Ehringer of Hillsborough Community College (HCC) was retained to
digitize the aerials and interpret the results. The scar data were field truthed to provide
accuracy. The digitized images were downloaded into the County’s Geographic Information
System (GIS) (Figure 2).
Figure 2. Example map of digitized seagrass scars.
RESULTS
The scar rate had suffered a large increase prior to the installation of signs in the Caution
Zone compared with baseline data. Upon placement of signage in the Management Area
(Figure 3), the rate of increase of new scars was considerably reduced in the Caution and
Exclusion zones as compared with the control area (Table 1). (Note: it was a reduction of
the rate of increase in the early years, not a reduction in the scar rate.) (Ehringer, 2000)
The new scar rate remained fairly consistent over the next several years in spite of heavily
increased use (according to the perception of frequent users and the Sheriff’s marine unit),
apparently due to a proactive public relations campaign and expanded signage at area
marinas.
The scar rate in the Caution Zone peaked in 1996 at the same time that upwards of 35% of
the signs and 50% of the buoys were lost, damaged or relocated due to storms. Buoys
disappeared due to anchor failure and many of the signs broke off pilings due to the galvanic
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Stowers, Fehrmann & Squires
reaction between the steel bolts, aluminum signs and bird droppings. This appeared to
impact the Caution Zone more than the Exclusion Zone.
Figure 3. Example of warning sign.
Table 1. Number of scars, before and after installation of signs.
Boat restricted
Seagrass caution
Non-restricted
LATE SUMMER
1992
SIGNS INSTALLED
SPRING 1993
93
98
19
167
216
66
FALL 1993
SCARS
NEW SCARS
207
262
123
40
46
57
It is believed that the new sign attachment method and the replacement of buoys with
pilings beginning in 1997 had resulted in a downward shift of the scar rate due to more
complete informational coverage. Hiring of full time law enforcement officers with shallow
draft boats surely contributed to reduced scarring due to increased officer visibility by the
public. Beginning in 1996, both the Caution and Exclusion Zones experienced large
reductions in the scar rate. Unfortunately, the scar rate in the unprotected (control) area
continued to rise (Figure 4).
Keys to Ordinance Success
Pinellas County feels that the factors contributing to the success of the program include the
need to:
• Document the problem thoroughly and highlight the value of the resource.
Environmental quality has actual monetary value in addition to its intrinsic value.
This can be used to further convince opponents to support the proposed activities.
• Avoid assigning blame and “pointing fingers.” Psychological barriers become
instantly erected when accusations are leveled at opposing parties. These barriers
become increasingly difficult to overcome as discussions progress.
72
Seagrass Scarring in Tampa Bay
Figure 4. Linear feet of prop scars, 1993–2000.
• Get “buy in” from all users. Get public input early and try to incorporate concerns
from the users. Fully explain the goals of the program and how these goals will be
measured.
• Follow through on “promises” made to users. Failure to perform tasks or
agreements will make it nearly impossible to get “buy in” for future projects and
could possibly lead to reversal of the ordinance.
• Provide feedback to the users. The public as well as the original members of the
team must be kept informed of success or failure of the actions as well as possible
future decisions. Use the media to promote effectiveness when possible.
• Adjust the program based on results. Don’t be afraid to make changes if the data
shows it is the prudent thing to do.
Additional Research
As part of the Howard Park beach renourishment, Pinellas County proposed a
replanting/research project as mitigation. The mitigation plan involved the removal of 0.32
acre of seagrass from Fred Howard Park and the transplanting of the seagrass into the Fort
Desoto Management area. The transplanted seagrass was placed in prop scars in order to
repair boat propeller damage. The plan had several aspects as follows:
Area III of Fort Desoto had 48,365 linear feet of prop scars (0.93 acre). In this area nutrients
and plant growth regulators were injected into the prop scars to stimulate the growth of new
seagrass into existing prop scars without disturbing the grass beds that surrounded the prop
scars. Annual photographs of the site taken in the fall of each year were used to ascertain the
overall growth of seagrass into the prop scars. In selected sites within the area, small PVC
pipes were placed into the prop scars at one-meter intervals. The number of new shoots per
meter were compared to linear transects that had not been injected.
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Stowers, Fehrmann & Squires
Approximately 3,000 square feet of seagrass were dug up with sediment from Fred Howard
Park and replanted into prop scars at Area V of Fort Desoto. The method of removal
involved the digging up of sections of seagrass in squares of 10 inches by 10 inches that
included 8 inches of sediment. The seagrass plugs were transported to Fort Desoto in
styrofoam boxes and gently placed into prop scars keeping the sediment intact with the
rhizomes. For evaluation purposes, transects along the prop scars were set up as in section
#1 above.
About 10,000 square feet of seagrass at Fred Howard Park were removed by machine. The
seagrass was removed from the site with a small backhoe and placed in a strainer to separate
the seagrass from the sediment. The seagrass was transported to Fort Desoto in plastic
drums that kept the seagrass in fresh marine water. This seagrass was stimulated with plant
growth regulators prior to planting by hand in the prop scars. Areas II and VI were the sites
for planting the seagrass. The same evaluation system was used for this seagrass as with #1
above.
The remaining 939 square feet of seagrass (harvested from floating sprigs) was transplanted
into a seagrass nursery that had already been set up in Ruskin. The seagrass in the nursery
was stimulated with plant growth regulators to promote new shoots. This seagrass was kept
at the nursery and transplanted into sites at Fort Desoto in 1997 and in 1998 into sites where
previous plantings had failed.
The results indicated that injection of growth hormone and nutrient into scars where no
seagrass was planted was the most effective method of growing seagrass. Seagrass
transplanted with sediment was inefficient and had a very low survival rate in this particular
situation, and planted sprigs exhibited mixed results (Table 2)(Ehringer, 2000).
Table 2. Results of replanting/research mitigation project.
METHOD
Hand transplanted
Sediment transplanted
Seagrass planter
Field nursery
Scar injections
ORIGINAL (sq. ft.)
FINAL (sq. ft.)
500
3,190
3,925
1,000
100
971
3,980
500
26,104
TOTAL
31,655
Future Directions
It has been recommended in past studies that we expand protection to include areas not
currently under protection (The non-restricted control area and the area east of the island
of Shell Key). The County Commission approved this additional protection after the
presentation at the Seagrass Conference. (Seagrass protection for the Weedon Island
Preserve was added with an ordinance amendment in 1996.)
It was also recommended that we reduce the Exclusion Zones and redesignate the areas as
Caution Zones based upon the findings that the zones are statistically similar in protecting
74
Seagrass Scarring in Tampa Bay
seagrass.1 This redesignation was also approved by the County Commission after the
presentation at the Seagrass Conference. This action is consistent with our findings that the
ordinance success relies on “adjusting the program” and “following through on promises.”
Based upon studies, the Board of County Commissioners redesignated some of the zones
and added protective zones effective November 2000.
A sign maintenance program and enforcement presence is critical to the long-term success
of the protection program. Lack of signs was quoted as one of the main reasons for noncompliance and directly affected the ability of the compliance officers to issue fines for
violating the ordinance.
A proactive public information campaign is a key to success. The public in general is much
more likely to abide by and support the ordinance if they are well informed of the reasons
for the ordinance and can visualize the protection zones.
It is prudent to research and support seagrass planting and restoration efforts to prevent long
term problems. It is a goal of Pinellas County to get new seagrass beds established in areas
that should support growth based upon favorable growing conditions but where none
currently exist.
SUMMARY
To help reduce and avoid seagrass degradation, several local governments have undertaken
programs to manage the use of the areas to the benefit of both the citizens and the resource.
These programs have generated both controversy and praise. Regulators and political figures are
placed in the position of trying to form an alliance of users that are many times at odds with each
other. Education and compromise is used as well as persuasive arguments to gain consensus on
protecting the resource for the long-term benefit of all citizens.
There has been much success in the Tampa Bay area but additional initiatives are required if
seagrass beds are to thrive. Recent questions have centered on whether the “exclusion zones”
should have been redesignated as “caution zones” and whether the “caution zones” could be
expanded to now unprotected areas (Redesignation has been approved by the County
Commissioners, Figure 5.) In addition, the benefits and drawbacks of seagrass scar repair
(injections) and the initiation of new seagrass beds (transplanting) must be addressed. The future
approaches to seagrass protection and restoration must be formed by a strong coalition of
representatives from government, educational institutions and environmental interest
organizations as well as user groups from both recreational and commercial interests.
Project costs are always a consideration when planning or implementing seagrass protection
or restoration efforts. Pinellas County has made a significant financial investment of over
$2 million in seagrass protection efforts. A summary of approximate costs is as follows:
Project Costs (since 1992):
Aerials
Interpretation
$10,000/yr – 7 years
$15,000/yr – 7 years
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Stowers, Fehrmann & Squires
Staff time
Pilings/marking/signs
Enforcement:
Approx. $100,000
$220,000
1994: $157,000 plus $48,000 start-up costs
associated with equipment
Approx. $172,000 per year operational costs through FY2002
Figure 5. Shell Key Preserve resource, public use, and aquatic regulatory zones.
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Seagrass Scarring in Tampa Bay
While $2 million is a substantial sum of money, Pinellas County considers this a costeffective means of maintaining healthy seagrass. Restoration efforts are many times more
costly than prevention of degradation. A summary of acreage protected is listed below:
AREA OF SEA GRASS PROTECTED
ACRES
Ft DeSoto/Shell Key
Shallow Water/Caution
Combustion Motor Exclusion
1691
403
Weedon Island Preserve
Combustion Motor Exclusion
Slow Speed/ Minimum Wake
356
1058
Total Acres Protected (approx.)
3500+
This leads to a figure of approximately $571 per acre of seagrass protected, quite a bargain
when compared to seagrass planting and restoration efforts that involve machines, divers,
tools and manpower.
REFERENCES
Ehringer JN. 2000. Results of Analysis of Prop Scar damage at the Fort Desoto Aquatic Habitat Management
Area: Film Analysis From1993 to 2000. Hillsborough Community College/Brandon, Tampa, FL
Ehringer JN. 2000. Final Report on Seagrass Removal From Fred Howard Park Pinellas County, Florida.
Hillsborough Community College/Brandon, Tampa, FL
Ehringer JN. 1999. New and Innovative Techniques for Seagrass Restoration. Hillsborough Community
College/Brandon, Tampa, FL
Sargent FJ, Leary TJ, Crewz DW, Kruer CR. 1995. Scarring of Florida’s Seagrasses: Assessment and
Management Options, FMRI Technical Report TR-1. St. Petersburg, FL
JFS, EF, AS (Pinellas County Board of County Commissioners, 512 S. Ft. Harrison Avenue, Clearwater, FL
33756)
77
❖
ASSESSMENT OF A CONSTRUCTION-RELATED
EELGRASS RESTORATION IN NEW JERSEY
P.A X. Bologna & M.S. Sinnema
ABSTRACT
Dual power cables were installed across a shallow bay in New Jersey in 1999. During
construction activities a portion of one of the cables was misaligned and needed to be removed
and replaced within the construction corridor. During these activities a significant eelgrass
(Zostera marina) loss was recorded. In an attempt to restore the site to pre-construction levels, an
experimental seagrass restoration was conducted within the realigned portion of the construction
window. Eelgrass and widgeon grass (Ruppia maritima) were transplanted in fall 2000 at the site
using peat-pot and bundled-stapled planting unit techniques. Initial results in May 2001 showed
significant growth and survival of eelgrass (>75%) as well as flowering and seed production for
each technique, but minimal growth and survival for widgeon grass (<35%). Subsequently,
significant light reduction (>90% ambient) at the site, due to turbidity and brown-tide, produced
a hostile environment for the growth and survival of restored eelgrass. As such, by fall 2001 no
remaining eelgrass remained on the site. While the lack of long-term restoration success was
disappointing, valuable data on restoration timing and technique was gained for coastal New
Jersey.
INTRODUCTION
Seagrass communities are common in coastal tropical and temperate regions as well as
portions of the sub-Arctic. Seagrass structure is important in coastal regions because it
dampens wave energy and reduces water velocity (Fonseca et al. 1982). The reduction of
flow associated with grass beds increases particle deposition (Almasi et al. 1987) and the
extensive root-rhizome mat may bind particles, thereby stabilizing sediments (Thayer et al.
1984, Fonseca and Fisher 1986). Seagrass beds, therefore, act as sediment traps and may
retain finer sediments than unvegetated regions around them (Orth 1977), thereby reducing
turbidity. The overall structure of seagrass communities covers a spectrum of plant species
composition and spatial coverage. In general, seagrass habitats are often distributed as a
mosaic of vegetated cover interspersed with varying degrees of unvegetated sediments (see
Larkum and den Hartog 1989, Robbins and Bell 1994, Marba and Duarte 1995). These
habitat mosaics have variable shoot density, species composition, canopy height and plant
biomass (Bell and Westoby 1986). Therefore, seagrass habitat architecture can be defined
at many spatial and temporal scales (Robbins and Bell 1994), and defining the extent and
physical arrangement of the landscape may be essential for addressing ecological questions
(Holling 1992, Levin 1992, Bologna and Heck 2002).
Degradation of seagrass habitats is occurring worldwide and is a result of direct and indirect
human activities (Walker and McComb 1992, Short and Burdick 1996). Consequently,
increasing economic development has placed a significant strain on many coastal
ecosystems. Direct impacts on these systems include waterfront development, bridge
construction, dredging activities for channel maintenance, construction activities related to
power and communication cables, as well as recreational use of these waters (e.g., boating
impacts). Indirect effects are best exemplified by excessive nutrient and sediment loading
from terriginous point and non-point sources (Valiela et al. 1990, Short and Burdick 1996).
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Bologna & Sinnema
In 1996, the Magnuson-Stevens Fishery Conservation and Management Act established
guidelines for sustaining the long-term viability of living resources. Within the regulation
the Fisheries Management Councils identify essential fish habitat and provide guidelines
to minimize negative impacts. Seagrasses have been identified as essential fish habitat. The
National Marine Fisheries Service (Department of Commerce, National Oceanic and
Atmospheric Administration) has declared that eelgrass (Zostera marina) is essential habitat
for several commercially important species including summer flounder (Paralyichthyes
dentatis) and winter flounder (Pseudopleuronectes americanus). With the elevation of the habitat
status of eelgrass, no net loss of this habitat should occur through direct commercial
activities. Since construction activities negatively impact the density and spatial coverage of
eelgrass in coastal systems, permitting activities and long-term monitoring are needed to
ensure full recovery of submerged aquatic vegetation to pre-impact levels. Unfortunately,
during the monitoring phase many sites do not recover in the allocated time frame and as
a result, these areas need to be re-vegetated.
During the last 20 years great advances have occurred in seagrass restoration/mitigation
techniques (see Fonseca et al. 1998). However, no successful examples exist for New Jersey
(Reid et al. 1993). Some of these limitations to restoration in New Jersey may relate to
timing of field activities and techniques. Existing recommendations for New Jersey
restoration efforts suggested that activities should occur during the spring, in accordance
with populations located in more northerly habitats. However, in New Jersey’s estuarine
waters these recommendations may result in the placement of plants in a highly stressed
situation with warm water, poor water clarity, and the potential for overgrowth by algae
occurring during the summer (Bologna et al. 2001). Based on previous research (Bologna
et al. 2000), a decision was made to perform restoration activities in the fall instead of in the
spring. The rationale behind this decision related to cooler water temperatures and
increased water clarity (i.e., greater light availability). By initiating restoration efforts in the
fall, it was surmised that the plants would have an opportunity to take root before winter
storms, propagate through vegetative growth, and take full advantage of the early spring
growing season. This would then also include flowering and seed production, both of which
are essential to long-term re-establishment of Zostera marina. To assess the relative
effectiveness of restoration, an experimental restoration planting occurred utilizing five
treatments varying in technique and plant spacing. Four treatments consisted of Z. marina
plants alone, while the fifth incorporated a dual planting of Z. marina and Ruppia maritima
(widgeon grass).
SITE LOCATION AND PROJECT BACKGROUND
The restoration project was conducted in Manahawkin Bay, New Jersey, USA (Fig. 1). This
was the site of a submerged power cable placement construction project consisting of two
cables being laid within a narrow construction window in 1999. The installation of the
cables was performed by plow method; however approximately 38 meters of the northern
cable was misaligned within the project construction corridor. When the misaligned
northern cable began to encroach upon the southern cable construction corridor,
construction was stopped and the cable was then removed and aligned in the appropriate
location by diver hand-jetting methodology. The disturbance associated with the mislaid
80
Construction-Related Eelgrass Restoration
cable was limited to the six-meter wide construction easement. Based upon as-built
construction plans, it was calculated that the area impacted by the hand jetting consisted of
approximately 0.032 hectare (Fig. 2a).
Seaside Heights
Barnegat Inlet
Manahawkin Bay*
Shelter Island
Figure 1. Regional map of New Jersey indicating the Restoration Site, Manahawkin Bay. Three donor sites
were used in the restoration and include Shelter Island (Zostera marina donor), Barnegat Inlet (Z. marina donor)
and Seaside Heights (Ruppia maritima donor).
Because this region was vegetated by Zostera marina, submerged aquatic vegetation surveys
were conducted pre- and post-construction activities. Surveys indicated a significant loss of
eelgrass throughout the project area (Table 1). While the misaligned cable construction
activities would account for the loss of eelgrass within this particular area, it was not the
probable cause of the complete loss of eelgrass within the project area including control line
samples outside of the construction window. Possible causes of eelgrass loss include: 1) a
portion could be attributed to the combination of brown-tide (Aureococcus anophagefferens)
shading (Schuster et al. 2000) and high temperatures documented during the summer of
1999, which appeared to cause a significant increase in eelgrass wasting disease from another
site in the region during this time frame (Bologna et al. 2000); and 2) there have been
instances in Little Egg Harbor (1998) where massive amounts of drift algae have smothered
eelgrass and entirely eliminated eelgrass biomass (both above and below ground) from
healthy eelgrass beds (Bologna et al. 2001).
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Bologna & Sinnema
a
14 m
Disturbed Region (0.32 ha)
Restoration Planting Area (210m)
Experimental Planting Schematic
Peat Pot 1m
Zostera-Ruppia 1m
Stapled 1m
Peat Pot 2m
b
Stapled 2m
Figure 2. On-Site Restoration Plan. A: Experimental restoration plot area. Each grid within the
plot is 7m x 7m. the disturbed area from cable re-alignment within the construction corridor is
cross-hatched and impacted 25 of 60 planting grids. B: Restoration treatment plot identification.
Each of the five treatments used in restoration activities is identified in the legend. Specific plot
treatments are represented within the planting grid schematic.
Table 1. Comparison of eelgrass surveys from pre-construction through post-construction activities (June
1999 through October 2001). Values represent average shoot counts m-2 (± 1 SD) among all surveys for the
identified sampling lines. North Construction Corridor represents the region where the northernmost power
cable was inserted into the substrate. South Construction Corridor represents the southernmost power cable
insertion line. The Control Survey line represents the average of two 9.1m off-set survey lines north and south
of the construction lines.
NORTH CONSTRUCTION SOUTH CONSTRUCTION
CORRIDOR 1
CORRIDOR 2
CONTROL
OFF-SET LINE
Pre-Construction Survey
June 1999
50.0 ± 7.6
44.7 ± 5.9
29.3 ± 3.6
Post-Construction Survey
September 1999
0
0
0.11 ± 0.33
Post-Construction Survey
May 2000
2.8 ± 5.3
12.1 ± 8.7
4.0 ± 4.7
0
0.02 ± 0.15
0.02 ± 0.15
Post-Restoration Survey
October 2001
82
Construction-Related Eelgrass Restoration
MATERIALS AND METHODS
Site Restoration
Five experimental Zostera marina planting techniques were used to assess their relative
success in a disturbed site in New Jersey. They included: 1) peat-pot planting unit (PU)
technique, 1-m spacing; 2) peat-pot PU technique, 2-m spacing; 3) stapled, bundled PU
technique, 1-m spacing; 4) stapled, bundled PU technique, 2-m spacing; and 5) Zostera
marina-Ruppia maritima mixed planting, peat-pot PU technique, 1-m alternate spacing.
These treatments are standard techniques for transplanting submerged aquatic vegetation
and are reviewed by Fonseca et al. (1998). The restoration planting area was 2940 m2 (14m
× 210m) and coincided with the region disturbed by hand-jetting (Fig. 2a). This region was
subdivided into two blocks (east and west) that consisted of thirty 7m × 7m experimental
plots (Fig. 2b). Within each block, planting treatments were randomly assigned and
replicated six times.
Restoration activities commenced on October 11, 2000 and were completed on October 20,
2000. During this time 592 Zostera marina peat-pot PU, 424 Z. marina stapled PU, and 168
Ruppia maritima peat-pot PU were planted. Field conditions were amenable for collection
and direct planting on the same day. As a result, PU collection occurred during the morning
with subsequent on-site restoration activities being conducted. No planting units were
allowed to sit overnight to be used in subsequent days. If additional planting units remained
after in-water planting had finished, a diver would take the remaining units and plant them
within the region, but outside the identified planting area.
To accomplish restoration activities, plant donor sites were identified from previous surveys
of submerged aquatic vegetation (Bologna et al. 2000, Bologna unpubl. data). The first site
was near Barnegat Inlet in the northern portion of Barnegat Bay and the second was near
Shelter Island in Little Egg Harbor (see Fig. 1). These sites were chosen because of their
abundant and healthy Zostera marina populations. Both of these sites were similar in
proximity to an oceanic inlet and have relatively high tidal flushing rates. Although the
removal of Z. marina was a realized impact to the bed, the exceptional health of the donor
sites showed that recovery occurred quickly through vegetative re-growth into the open area
as well as winter seed recruitment. The collection of Ruppia maritima occurred from a plant
population near Seaside Heights (Fig. 1).
Spring 2001 Restoration Monitoring
During May 2001, surveys were conducted on half of the planted experimental restoration
plots (n=30). Fifteen sample restoration plots were sampled from both the east and west
blocks. Within each block, each planting technique was assessed by visual identification and
enumeration of planting units from three randomly selected plots of each of the five
treatments (n=15 samples/block, n=3 samples/technique). Planting success was then
determined by calculating the percent survival of planting units and compared among
treatments. Since plots were randomly assigned within each matrix and randomly sampled
during the monitoring program, data were pooled from both blocks to assess overall
treatment success. Data were analyzed using one-way ANOVA with α = 0.05. Percentage
data (i.e., percent survival) were arcsine transformed prior to analysis. Subsequently, data
83
Bologna & Sinnema
were pooled based on planting technique into either peat-pot or bundled-stapled categories.
Data were then analyzed in the above manner to determine whether differences existed in
PU survival between primary planting techniques (α = 0.05).
Fall 2001 Monitoring
The fall 2001 monitoring event was conducted October 8–11, 2001. Two phases of
monitoring occurred to determine 1) the overall spatial coverage of submerged aquatic
vegetation in the construction region coinciding with original pre- and post-construction
surveys, and 2) the relative survival of transplants.
Spatial Coverage
Survey transects were conducted to compare pre-construction (1999) shoot densities with
current (2001) shoot densities. Construction corridor transects were located 6.1m apart and
parallel to the cable centerlines, and the offset control transects were spaced 9.1m from and
parallel to, the plow transects. Sampling stations were placed at 9.1m intervals along each
transect. PVC conduit was set into the substrate at each sample station with a correlating
station number for diver/sampler identification. Exact locations were determined by using
a Topcon™ Total Station. Quadrats (1 m2) were placed on the bottom by diver, and
oriented by compass (N) at each sampling station. The divers proceeded to record coverage
data, based upon the number of 25-cm grids (16 total) that contained at least one live stem
of seagrass, to determine presence. Additionally, divers detailed and recorded the total count
of live seagrass stems within three randomly selected 25-cm grids. These grids were
predetermined utilizing a random numbers table.
Restoration Monitoring
A sample of restoration plots was staked for underwater survey. These restoration plots were
previously monitored during the spring 2001 restoration monitoring event. Fifteen plots
(three of each experimental treatments) were delineated for survey. During the course of
the transect surveys (Spatial Coverage, above), it was apparent that little vegetation was
present within the region. Several random transects were also investigated outside of the
construction region. These random transects were located to the north of the study area,
and similarly contained little to no vegetation. As a result of these in situ observations, only
six restoration plots were surveyed in detail to assess whether any planting units remained
and included two plots containing the Zostera marina-Ruppia maritima treatments and one
plot from the other four treatment combinations containing only Z. marina.
Environmental Data
Light Availability
Data were collected on August 15, 2001 to assess light availability on the site. Using paired
Licor® spherical light sensors, ambient light and in-water light values were measured.
Sensors were allowed to undergo stabilization and then recorded light values
(photosynthetically active radiation). During the time period of collection, a depth of 1.6
meters was recorded. Logged data were then downloaded to a personal computer and a
relative light reduction was calculated based on the paired light meter data. Specifically,
84
Construction-Related Eelgrass Restoration
percent light available was calculated for each sample pair based on the light intensity in air
compared to light intensity in the water at 15 cm from the bottom.
Water Temperature
From August 15th through September 6th 2001, a temperature data-logger was placed at the
site to record water temperature to assess potential detrimental effects of extreme summer
temperature on the survival of Zostera marina. It is recognized that extreme temperatures
(>30° C) can be lethal to Z. marina. However, even moderate temperature readings (>27°
C) may reduce growth. Data were recorded at one-minute intervals for the time period.
RESULTS
Spring 2001 Monitoring
Results from the restoration monitoring indicate considerable survival for each Zostera
marina planting technique, with survival greatest for peat-pot PUs (Table 2). While the
range of identified survival rates was broad (50–100%), average survival was at or above 75%.
For the bundled-stapled PUs, survival rates were greater for the 1m spacing plots compared
to the 2m spacing plots (see Table 2). While no significant differences were seen among
individual planting treatments (F4,25 = 1.7, P>0.18), pooled analysis comparing the
techniques showed that Z. marina survival was significantly greater from treatments using
peat-pots compared to bundled-stapled PUs (F1,28= 4.8, P<0.037). While Z. marina survival
was substantial, Ruppia maritima survival was relatively low (Table 2). Additionally, while
R. maritima PUs had survived, they showed no appreciable growth or expansion beyond the
peat pots.
Table 2. Restoration technique and planting unit spacing assessment. Values represent the average percent
survival of planting units for each of the five planting methodologies used in the study (mean ± 1 SD). The
range of survival represents the actual calculated survival rate ranges for planting units in individual plots
within each technique.
PLANTING
METHODOLOGY
# PLOTS
Peat Pot, 1m spacing
Peat Pot, 2m Spacing
Bundled, Stapled, 1m spacing
Bundled, Stapled, 2m spacing
Peat Pot, 1 m spacing, Zostera- Ruppia mix
Zostera marina
Ruppia maritima
6
6
6
6
6
AVERAGE
SURVIVAL
RANGE OF
SURVIVAL
74% ± 11
77% ± 18
70% ± 15
55% ± 23
59–90%
50–100%
49–90%
19–75%
79% ± 14
35% ± 11
64–100%
29–54%
While conducting the planting technique assessment surveys, it was apparent that many of
the PUs were undergoing reproduction. Reproductive Zostera marina shoots were common
in most PUs and were abundant (>5) in some of the units. While these data were not
formally gathered during the survey, observation suggested that significant reproduction,
growth, and vegetative expansion had occurred for all Z. marina PUs.
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Bologna & Sinnema
Fall 2001 Monitoring
Spatial Coverage
During the on-line sampling event in October 2001, little to no live seagrass was
encountered (Table 1). In fact, only 3 of the 168 quadrat samples collected contained any
Zostera marina, and in each of these cases it was a single, poorly anchored shoot. During the
course of this monitoring event, several random transects (n=3) were also investigated
outside of the study area. These transects were located to the north of the study area and
similarly, contained no vegetation.
Restoration Monitoring
Based on the information gathered from the May 2001 monitoring, fifteen restoration plots
were originally selected to assess long-term PU survival. However, due to the results of the
on-line spatial coverage sampling (above), we visually inspected nine of the restoration plots
and then randomly selected six restoration plots to be assessed in detail. One plot of each
of the four Zostera marina treatments was monitored and two Zostera marina-Ruppia maritima
plots were sampled (n=6). Monitoring results showed that no Z. marina planting units
survived in any of the treatments. In one of the Zostera-Ruppia plots a single R. maritima
growing region was noted, but it is uncertain whether this was a specific restored planting
patch or whether this was an incidental colonization of R. maritima from outside the
construction area. Our uncertainty revolves around the location of the patch, as it was
located in what should have been a Z. marina planting quadrat. It is possible that the spacing
of PUs was disrupted in this restoration plot, whereby a R. maritima planting unit was placed
incorrectly in the matrix and was observed during the monitoring survey. Regardless, no
long-term survival of Z. marina transplants occurred at the site.
Environmental Data
Light Availability
On August 15, 2001, 214 paired light measurements were collected from ambient air and
in water at 15 cm above the bottom at a depth of 1.6 meters. Based on these data,
photosynthetically active radiation was reduced to 7.66% ambient light (range 3.7–16.2%
ambient). This represents a significant reduction in light availability (>90% reduction) at
this site and would pose a significant light stress on the plants.
Water Temperature
Collected water temperature data showed that between August 15 and September 6, 2001,
water temperature averaged 25.9°C. However, temperature ranged between 22.6 and 28.6°
C. These high summer temperatures may pose a potential threat to both the growth and
survival of Zostera marina, but not Ruppia maritima.
Brown-tide
Sampling conducted by the New Jersey Department of Environmental Protection showed
a significant brown-tide (Aureococcus anophagefferens) bloom at the site during 2001. Data
presented by Downs-Gastrich at the Barnegat Bay Estuary Program Monitoring Workshop
(Downs-Gastrich 2001) showed significant bloom conditions in 2001 from May through
July at site 1703C (Manahawkin Bay), which is adjacent to the restoration site. In May 2001,
86
Construction-Related Eelgrass Restoration
cell counts of brown-tide organisms were already at Category 3 bloom levels (>200,000
cells/ml) and were greater than 1,000,000 cells/ml. Brown-tide continued to persist at this
site through July 5th, with maximum values recorded on June 25th, 2001 when cell counts
exceeded 1,800,000 cells/ml (Downs-Gastrich 2001). With this dramatic bloom in browntide organisms, substantial light stress was placed on submerged aquatic vegetation.
DISCUSSION
During this project substantial data were collected regarding the restoration of seagrasses
in New Jersey. While the prediction was made that Zostera marina plantings should occur
in the fall, no previous data existed on this. Through the restoration efforts we were able to
see significant survival and growth of transplants on the site (Tables 2, 3), and, perhaps
more importantly, we saw significant reproduction in those same transplants. Ultimately,
successful restoration of any plant community must demonstrate sexual reproduction, and
our observations of reproductive shoots bearing seeds in May 2001 demonstrated this.
Consequently, these data suggest that for New Jersey, the fall probably represents the
optimal conditions for Z. marina transplanting efforts. Despite the resulting loss during the
summer due to extrinsic light attenuation at the site, understanding the principles of
restoration timing in New Jersey is a key element in future mitigation and restoration
efforts.
Table 3. Initial planting unit survival comparisons between experimental plots planted outside and within the
disturbed cable area. Values represent average survival of planting units for each restoration technique and
spacing. Outside Survival represents the average survival of planting units in plots sampled outside of the
mislaid (disturbed) cable region, while Within Survival represents the average survival of planting units in the
specific region where the cable was moved from the original plow placement to the correct on-line position.
The number of individual plots used for each assessment is provided. Only the Peat Pot 2m spacing technique
was not equally represented in the Outside-Within comparisons. *Compiled site averages of initial survival
comparisons between Outside and Within (P < 0.003).
PLANTING
METHODOLOGY
PLOTS
ANALYZED
Peat Pot, 1m spacing
Peat Pot, 2m spacing
Bundled, Stapled, 1m spacing
Bundled, Stapled, 2m spacing
Peat Pot, 1m spacing, Zostera- Ruppia mix
Zostera marina
Ruppia maritima
Site Zostera marina average*
OUTSIDE
SURVIVAL
WITHIN
SURVIVAL
3–3
4–2
3–3
3–3
81%
81%
81%
73%
68%
67.5%
59.7%
37.7%
3–3
3–3
81%
32%
76%
39%
16–14
79.7%
61.5%
While the fall was the preferred timing for Zostera marina, it was not favorable for Ruppia
maritima. Ruppia maritima survival was low (35%) and it did not grow appreciably.
Additionally, the lack of growth and long-term survival of R. maritima suggests that it too
received sufficient light stress to significantly reduce survival. While this information is
important in narrowing the window for R. maritima transplanting, overall water quality
within an area will determine long-term success.
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Bologna & Sinnema
The plant spacing and methodology data provide important information for future
restoration efforts. The results from this work suggest that the peat-pot method was
preferable, as was closer plant spacing (Table 2). However, fine sediments characterized this
site and the peat-pot methodology may not be the best approach for all restoration activities;
especially those with greater dynamic water movement (e.g., waves, tides). The success of
the bundled-stapled planting units suggests that these may be more important as a
technique in higher energy sites (sensu Fonseca et al. 1998).
While survival differences existed among methodologies, the greater determinant factor for
PU survival was “within, compared to outside,” the hand-jet disturbed region. To assess the
relative impact of the placement, removal, and re-alignment of the cable; data were
restricted to reflect plots either within this hand-jetted area or outside (Table 3). Based on
this distinction, PU survival was significantly greater from plots outside the disturbed
region compared to survival within this region (F1,28=10.2, P<0.003). In all cases, survival
was greater in regions where the line was laid correctly and averaged 79.7%. Conversely,
average survival in the impacted region was 61.5% for Z. marina and was further reduced
based on the technique. Peat-pot methodology resulted in a 68% survival while survival for
the bundled-stapled units was only 49% in the disturbed region. Based on these data, it
would appear that the best technique for restoration under these conditions would be the
peat-pot technique (Table 3).
It has been shown that a reduction in ambient light produces significant reductions in
survival and growth of seagrasses (Goodman et al. 1995, Short et al. 1995). The identified
90% reduction in light measured in August could easily account for the lack of seagrass on
this site. This change in water clarity was not entirely due to brown-tide, as the developed
bloom had dissipated by this time (Downes-Gastrich 2001). Consequently, these reductions
in light during August must represent a significant, non-bloom turbidity issue. Overall,
during the spring and summer it appears that significant light attenuation is occurring and
may prohibit Zostera marina from actively growing within this region, regardless of any
disturbance events. Lathrop et al. (2001) described a seagrass model incorporating light
attenuation in relation to seagrass survival depth for New Jersey. Based on this model, they
suggest a 1.2m depth (MLW) for which SAV would survive and grow. Based on our
hydrographic survey, much of the identified region’s depth is within, or exceeds, this
theoretical limit (EDG 2001). However, there are other extenuating light stresses on this
site.
Normally, substantially elevated structures produce limited shading effects in aquatic
communities. However, the relative size and orientation of a cross-bay bridge may provide
additional light stress for Zostera marina. The east-west orientation of the bridge creates a
shading shadow on the north side, based on the southerly sun position. As a result, shading
from the bridge occurs on a daily basis. While alone this probably represents a minimal
shading impact, the cumulative effects of turbidity, brown-tide, depth, and bridge shading
provide a significant large-scale light limitation, which may limit any significant plant
growth.
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Construction-Related Eelgrass Restoration
If the site elevation was reduced to less than two feet at low tide, a possibility exists that
Zostera marina would again survive. However, there is no guarantee that restoration would
be successful. Changes in elevation may create greater human disturbance through boat
engine-sediment interactions (i.e., prop scars and prop wash), which would mechanically
destroy the integrity of the grass bed and increase sediment resuspension, thereby increasing
turbidity and reducing light availability during the summer. While this technique has been
used in other regions of the United States (e.g., Texas), it is primarily used as a method to
dispose of dredge spoil material and not as a habitat enhancement technique.
CONCLUSION
In order to fully understand any eelgrass community and the factors that influence the
health and survival of plants within construction corridors, external control plots would
provide a critical link. Specifically, if control plots outside of this study area had been
monitored in 1999, prior to construction, then the potential impacts of extrinsic
environmental factors could have been assessed at the post-construction monitoring event.
In this case, during the summer of 1999 when construction activities were occurring, the
Barnegat Bay system endured a prolonged brown-tide bloom (Schuster et al. 2000). This
bloom had a centralized intensity in the upper reaches of Little Egg Harbor and
Manahawkin Bay. We know from previous studies that this can significantly impact
seagrasses (Dennison et al. 1989) and other living resources (Bricelj and Lonsdale 1997). As
such, adequate pre-construction control plots might have made it possible to discern the
impacts of construction activities versus uncontrollable environmental factors at the postconstruction monitoring event. In the future, it would be germane to require adequate
external control plots for construction activities that occur in seagrass and other critical
habitats to determine the proper course of action. However, current environmental
regulations do not specify distant external controls to be monitored during construction
activities, nor do they require control plots for mitigation activities. While adequate external
controls are preferable from a scientific point of view, they are a realized economic cost to
businesses. Consequently, these economic interests often drive the scope of projects and we,
as scientists, must strive to conduct the best research we can, given these restrictions.
Perhaps the most important finding of our research involves the preferred fall seasonal
timing of restoration activities in New Jersey for Zostera marina. As coastal development
continues to increase and impacts submerged aquatic vegetation, we must understand the
essential timing of restoration activities to ensure future success.
LITERATURE CITED
Almasi M, Hoskin C, Reed J, Milo J. 1987. Effects of natural and artificial Thalassia on rates of sedimentation.
J. Sedim. Petrol. 57:901–906.
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decapods. J. Exp. Mar. Biol. Ecol. 104:275–295.
Bologna P, Heck K. 2002. Impact of habitat edges on density and secondary production of seagrass-associated
fauna. Estuaries. 25:1033–1044.
Bologna P, Wilbur A, Able K. 2001. Reproduction, population structure, and recruitment limitation in a bay
scallop (Argopecten irradians Lamarck) population from New Jersey, USA. J. Shellfish Res. .20:89–96.
Bologna P, Lathrop R, Bowers P, & Able K. 2000. Assessment of submerged aquatic vegetation in Little Egg
Harbor, New Jersey. Technical Report 2000-11. Institute of Marine and Coastal Sciences, Rutgers, the
State University of New Jersey. New Brunswick, New Jersey. 30 p.
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Bricelj M, Lonsdale D. 1997. Aureococcus anophagefferens: causes and ecological consequences of brown tides in
U.S. mid-Atlantic coastal waters. Limnol. Oceanogr. 42:1023–1038.
Dennison W, Marshall G, Wigand C. 1989. Effect of “brown tide” shading on eelgrass (Zostera marina L.)
distributions. p. 675-692. In E. Cosper, V. Bricelj, and E. Carpenter (eds.), Novel Phytoplankton Blooms.
Springer-Verlag, New York.
Downs-Gastrich M. 2001. New Jersey Brown Tide Assessment Project. Barnegat Bay National Estuary
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Fonseca MS, Fisher JS. 1986. A comparison of canopy friction and sediment movement between four species
of seagrass with reference to their ecology and restoration. Mar. Ecol. Prog. Ser. 29:15–22.
Fonseca MS, Kenworthy JW, Thayer GW. 1998. Guidelines for the Conservation and Restoration of
Seagrasses in the United States and Adjacent Waters. NOAA Coastal Ocean Program Decision Analysis
Series No. 12. NOAA Coastal Ocean Office, Silver Spring, MD 222 pp.
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flow. Est. and Coastal Sci. 15:351–364.
Goodman J, Moore K, Dennison W. 1995. Photosynthetic responses of eelgrass (Zostera marina) to light and
sediment sulfide in a shallow barrier island lagoon. Aquat. Bot. 50:37–48.
Holling,CS. 1992. Cross-scale morphology, geometry, and dynamics of ecosystems. Ecol. Monogr.
62:447–502.
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A. J. McComb, and S. A. Shepherd (eds.), Biology of seagrasses: a treatise on the biology of seagrasses with
special reference to the Australia region. Elsevier Science Publishers, Amsterdam. 842 pp.
Lathrop R, Styles R, Seitzinger S, Bognar J. 2001. Use of GIS mapping and modeling approaches to examine
the spatial distribution of seagrasses in Barnegat Bay, New Jersey. Estuaries 24:904–916.
Levin SA. 1992. The problem of pattern and scale in ecology. Ecology 73:1943–1967.
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migration. J. of Ecol. 83:381–389
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(ed.), Ecology of Marine Benthos. University of South Carolina Press, Columbia. 467 pp.
Reid R, MacKenzie C, Vitaliano J. 1993. A failed attempt to re-establish eelgrass in Raritan Bay (New
York/New Jersey). NOAA/NMFS/NEFSC: Woods Hole, MA. [Northeast Fisheries Science Center] Ref.
Doc. 93-27.
Robbins BD, Bell SS. 1994. Seagrass landscapes: a terrestrial approach to the marine subtidal environment.
Trends Ecol. Evol. 9:301–304.
Schuster R, Feerst E, Olsen P. 2000. Annual summary of phytoplankton blooms and related conditions in the
New Jersey coastal waters summer of 1999. Water Monitoring Report, Bureau of Marine Water
Monitoring, New Jersey Department of Environmental Protection 30 p.
Short F, Burdick D. 1996. Quantifying eelgrass habitat loss in relation to housing development and nitrogen
loading in Waquoit Bay, Massachusetts. Estuaries 19:730–739.
Short F, Burdick D, Kaldy J. 1995. Mesocosm experiments quantify the effects of eutrophication on eelgrass,
Zostera marina. Limnol. Oceanogr. 40:740–749.
Thayer GW, Kenworthy JW, Fonseca M. 1984. The ecology of eelgrass meadows of the Atlantic coast: a
community profile. U.S. Fish Wildl. Serv. Biol. Ser. Prog. FWS/OBS-84/02, 147 p.
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PB (Montclair State University, Department of Biology and Molecular Biology, Science Hall, Montclair, NJ
07043; MS (Birdsall Engineering, Inc., 611 Industrial Way West, Eatontown, New Jersey 07724-2213.
ms@birdsall.com)
Author contact: bolognap@mail.montclair.edu
90
EXPERIMENTAL HALODULE WRIGHTII AND SYRINGODIUM
FILIFORME TRANSPLANTING IN HILLSBOROUGH BAY, FLORIDA
W. Avery & R. Johansson
ABSTRACT
In 1987, the City of Tampa, Bay Study Group (BSG) transplanted ca. 13m2 of Halodule
wrightii into several areas of Hillsborough Bay. About 2.3m2 of H. wrightii was planted as bare
root units. The rest of the material was planted as intact sod units at seven locations around
the perimeter of the bay. Over 90% of the planted material was lost within the first year.
However, the areal coverage for the remaining bare root units and sod units increased to ca.
300m2 and 1100m2, respectively. In 1988, the BSG transplanted ca. 2.5m2 of H. wrightii and
7.5m2 of Syringodium filiforme as bare root units into southeastern Hillsborough Bay. This
planting attempt failed within one year. A second attempt using ca. 2m2 of H. wrightii and 2m2
of S. filiforme planted as sod units also failed. Loss of transplants may be attributed to sediment
transport, high drift macroalgae biomass, wave energy, rainfall events, and bioturbation.
Personnel expenditures for these projects were ca. 2300 man-hours.
INTRODUCTION
Eutrophic conditions in Hillsborough Bay were substantially reduced during the late
1970s and early 1980s through aggressive management of nutrient sources, primarily
originating from point sources (Johansson 1991, Johansson and Lewis 1992). Soon
following the nutrient reductions, important indicators of estuarine water quality (e.g.
phytoplankton biomass and water column light penetration) improved in Hillsborough
Bay and other areas of Tampa Bay as well (Boler 2001, Johansson 2002). By 1984,
phytoplankton biomass, as measured by chlorophyll-a, was about half the levels found
only a few years earlier. Water clarity, measured from Secchi disk depth, increased at a
similar magnitude. Also, drift macroalgae accumulations on the shallow sand flats
decreased dramatically in Hillsborough Bay during the mid and late 1980s (Avery 1997;
Johansson 2005).
Apparently in response to decreased competition from the phytoplankton for available
light and possibly less competition from the macroalgae for suitable substrate, sparse
new growth of Halodule wrightii (shoalgrass) was noted on the shallow sand-flats in
southeastern Hillsborough Bay in the mid-1980s (R. Lewis pers. comm.). The new
seagrass growth was seen after many years of seagrass coverage apparently being absent
in this section of Tampa Bay (Avery 1991). The new growth implied that this area and
other intertidal areas of Hillsborough Bay may have achieved adequate conditions to
support continuous seagrass growth.
To test the theory of sufficient conditions for seagrass growth and to provide vegetative
source material to areas lacking naturally recolonizing seagrass, the City of Tampa, Bay
Study Group (BSG) initiated, in 1987, a series of test plantings at multiple locations in
Hillsborough Bay. The initial efforts were coordinated with the Tampa Bay
Experimental Seagrass Planting Project organized by the Florida Marine Research
Institute (FMRI) and the National Marine Fisheries Service (NMFS). H. wrightii source
material for these initial plantings were harvested from an area of Old Tampa Bay that
was destined to be impacted by a Florida Department of Transportation road widening
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Avery & Johansson
project at the Courtney Campbell Causeway. Subsequent test plantings, including both
H. wrightii and Syringodium filiforme (manatee grass), were undertaken by the BSG and
utilized source materials from Port Manatee in lower Tampa Bay.
Finally, the convener of the Seagrass Restoration Workshop requested that each
presenter provide, if possible, an account of costs associated with the transplanting
projects. In this report, we have estimated the cost in terms of total man-hours used
from the initiation of the project to the point in time when maximum transplanted
seagrass coverage was attained.
METHODS
Site Selection
Generally, transplant site selection met the criteria described by Fonseca et. al. (1998).
However, the criteria of using elevations of nearby seagrass beds as a reference for
planting depth was not possible in all areas due to lack of naturally occurring seagrass.
Therefore, in areas lacking reference beds, planting sites were selected encompassing a
range of elevations. Finally, observations made by Phillips (1962) describing the
historical species distribution in Hillsborough Bay contributed to our decision for
selection of species specific transplant sites.
Courtney Campbell Donor Site
In June and July 1987, ca. 13m2 of H. wrightii was harvested by shovel from the Courtney
Campbell Causeway donor site (27.971º N, 82.578º W, Figure 1) and moved to
Hillsborough Bay by BSG personnel. Transplanting was conducted using bare root and
sod units.
Bare Root Units: Approximately 2.3m2 of donor material were used to prepare 860
bare root units for transplanting. Each unit was assembled similar to the method
described by Fonseca et al. (1987, 1994) and contained, on average, 15 short shoots and
three apical meristems. Shoots and meristems were secured to a six-inch U-shaped steel
staples with wire ties. Units were placed in coolers filled with water from the donor site
until transplanting.
A 10×20m site adjacent to MacDill AFB in western Hillsborough Bay (Area 8, Figure 2)
was selected for planting. The site elevation was determined with a rod and level
referencing a nearby H. wrightii bed. Units were planted on 0.5m centers by hand using
the staple to anchor each unit to the sediment.
Sod Units: Approximately 330 H. wrightii sod units with intact sediment and covered by
wet burlap were transported by boat from the donor site to seven planting areas in
Hillsborough Bay (Figure 2). Each unit measured ca. 300cm2 and contained an average
of 170 short shoots and 23 apical meristems. Units were planted at elevations of
50–80cm below mean tide level (MTL) in planting Area 1 and 25–30cm below MTL in
planting areas 2–7. The site elevation was determined with a rod and level.
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Experimental Transplanting in Hillsborough Bay
Figure 1. Location (arrows) of the Courtney Campbell Causeway and Port Manatee donor sites.
Within each planting area, planting sites were marked at 50m intervals with a PVC pole.
Each planting site was comprised of two sod units that were planted ca. 1m of either side
of the PVC pole on a line roughly perpendicular to the shoreline. The number of sod
units placed in each planting area is summarized in Table 1.
Port Manatee Donor Site
During 1988 and 1989, BSG personnel moved H. wrightii and S. filiforme from Port
Manatee (27.631º N, 82.573º W, Figure 1) to two 10×20m planting blocks in
southeastern Hillsborough Bay (Area 1, Figure 2). In May 1988, ca. 7.5m2 of S. filiforme
93
Avery & Johansson
and 2.5m2 H. wrightii were moved as bare root units. In June 1989, ca. 2m2 of S. filiforme
and 2m2 H. wrightii were moved as sod units.
Figure 2. Location of planting areas 1–8 in Hillsborough Bay.
Bare Root Units: Average S. filiforme bare root units were made of 17 short shoots and
1 apical meristem. About 860 units (Table 1) were planted on 0.5m centers in one of the
planting blocks. In the second planting block, about 420 S. filiforme and 440 H. wrightii
bare root units (Table 1) were planted on 0.5m centers in alternating monospecific rows.
The average H. wrightii unit consisted of 31 short shoots and 6 apical meristems.
Sod Units: In 1989, the same planting blocks used in 1988 were planted as
monospecific plots. One planting block was planted with 66 S. filiforme sod units with an
average unit consisting of 87 short shoots and 5 apical meristems. The second planting
94
Experimental Transplanting in Hillsborough Bay
block was planted with 66 H. wrightii sod units with an average unit consisting of 163
short shoots and 32 apical meristems (Table 1).
Table 1. Number of Halodule wrightii (Hw) and Syringodium filiforme (Sf) units placed in each planting area
between 1987 and 1989.
AREA
1987 BARE ROOT
1987 SOD
1988 BARE ROOT
1989 SOD
1
2
3
4
5
6
7
8
—
—
—
—
—
—
—
860 Hw
168 Hw
66 Hw
12 Hw
6 Hw
20 Hw
12 Hw
42 Hw
—
441 Hw, 1281 Sf
—
—
—
—
—
—
—
66 Hw, 66 Sf
—
—
—
—
—
—
—
Transplant Monitoring
Transplants were monitored in the spring, summer, and fall to determine areal coverage,
short shoot density, and canopy height. Ancillary data included observations on epiphyte
loading, epiphyte description, and seagrass appearance. Monitoring was terminated at a
specific planting site when the transplant coverage coalesced with seagrass adjacent to the
site.
The NMFS also monitored the bare root planting blocks for development of plant
components (Fonseca et al. 1996a) and faunal recruitment (Fonseca et al. 1996b). The
floral and faunal characteristics were compared to nearby H. wrightii or Caulerpa prolifera
beds.
RESULTS AND DISCUSSION
Bare Root Units
1987 Planting
H. wrightii bare root areal coverage decreased from the initial 2.3m2 planted in 1987 to
about 1.8m2 in the spring of 1988 (Figure 3) as nearly 65% of the units were lost. Of the
remaining units, ca. 10% increased in areal coverage. A substantial increase in areal
coverage was not seen until the spring of 1989. Maximum coverage of 291m2 was
attained in the summer of 1991. Following the maximum, areal coverage decreased in
each subsequent survey. By 1994, no H. wrightii was found within the planting block.
1988 Planting
In the monospecific S. filiforme planting block, areal coverage tripled from the initial 5m2
planted in May 1988 to just over 16m2 in July (Figure 4). However, coverage declined
over 40% by October and the site was virtually barren by January 1989. There was no
coverage present in May 1989.
A similar expansion in areal coverage was seen with the S. filiforme within the mixed
planting block as the initial coverage doubled in the first two months of planting and
95
Avery & Johansson
then declined almost 40% by October 1988 (Figure 4). By May 1989, no S. filiforme was
present.
300
250
Areal Coverage, m
2
200
150
100
50
0
1987
1988
1989
1990
1991
1992
1993
1994
Year
Figure 3. Seasonal areal coverage of Halodule wrightii bare root units in Area 8, 1987–1994.
In contrast to the growth pattern seen with the S. filiforme plantings, H. wrightii coverage
within the mixed planting block declined each monitoring period succeeding the initial
planting (Figure 4). As with the S. filiforme, all H. wrightii transplants were absent in May
1989.
Sod Block Units
1987 Planting
Between the planting during the summer of 1987 and the fall of 1988, areal coverage for
the H. wrightii sod units decreased from 10.7m2 to 2.6m2 as nearly 80% of the sods did
not persist (Figure 5). However, as the remaining sod units became established, the rate
of loss decreased and areal coverage began to increase. Coverage increased from about
38m2 in the spring of 1989 to nearly 1100m2 in the spring of 1993 (Figure 5). Sod units
persisted in Areas 2, 5 and 7 in 1993. Monitoring was discontinued after the summer of
1994 as the sod unit coverage became indistinguishable from naturally developing H.
wrightii coverage in most areas.
1989 Planting
The May 1989 planting of H. wrightii and S. filiforme was not monitored on a seasonal
schedule. However, nine months following the planting, all material was absent.
96
Experimental Transplanting in Hillsborough Bay
18
15
Areal Coverage m2
12
Hw mixed
Sf mixed
9
Sf mono
6
3
0
May-88 Jun-88
Jul-88
Aug-88 Sep-88 Oct-88 Nov-88 Dec-88 Jan-89 Feb-89 Mar-89 Apr-89 May-89
Month
Figure 4. Seasonal areal coverage within the Syringodium filiforme (Sf) monospecific bare root planting block
and the Syringodium filiforme / Halodule wrightii (Hw) mixed bare root unit planting block.
100%
1200
90%
80%
70%
800
60%
50%
600
Percent Survival
Areal Coverage, m2
1000
Coverage, m2
Percent Survival
40%
400
30%
20%
200
10%
19
94
19
93
19
92
19
91
19
90
19
89
0%
19
88
19
87
0
Year
Figure 5. Seasonal areal coverage and percent survival of the Halodule wrightii sod planting units, 1987–1994.
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Avery & Johansson
Potential Factors Affecting Transplants
The primary objectives of the transplant effort were to identify areas of Hillsborough
Bay suitable for seagrass recolonization and to facilitate vegetative growth through the
establishment of source material. Further, the planting project was initiated with the
premise that the abatement of eutrophic indicators such as high chlorophyll and
macroalgae biomass had resulted in improved conditions sufficient for seagrass
recolonization. As the transplant coverage expanded nearly two orders of magnitude
between 1987 and 1992, it became apparent that several factors discussed below had the
potential to limit the ultimate success of the transplant project.
Between 1992 and 1993, a 40–50cm high sand ridge passed through the bare root
plantings in Area 8, which resulted in transplant burial followed by a 60% reduction in
coverage. Comparable sediment transport leading to seagrass burial has been reported by
Duarte and Sand-Jensen (1990). They observed plant mortality in the seagrass Cymodocea
nodosa following the passage of highly mobile sand-waves.
Abundant macroalgae biomass has been reported to negatively impact seagrass coverage
(Den Hartog 1994). Persistent macroalgae biomass frequently exceeding 40gdwtm-2 was
seen in northeastern Hillsborough Bay (Area 2) through 1995 (Avery 1997). Macroalgae
biomass of this magnitude may shade seagrass and thereby reduce available light for
seagrass production (Dennison and Alberte 1982). Further, drifting macroalgae mats
have been observed to scour the sediment and seagrass meadows in Hillsborough Bay
(pers. obs.). In Area 2, shading and abrasion by macroalgae may have contributed to the
loss of nearly all of the sod units. Finally, hypoxic conditions have also been noted
within areas of decomposing macroalgae in Hillsborough Bay. Decomposing macroalgae
has been shown to create a reduced sediment redox potential (Zimmerman and
Montgomery 1984) resulting in increased sulfide concentrations that may affect seagrass
viability (Pulich 1983, Carlson et al. 1994, Goodman et al. 1995).
Wind and ship generated wave energy may also have contributed to transplant loss at
several Hillsborough Bay transplant sites. Koch (2001) presented a synopsis of direct and
indirect impacts from waves on seagrass beds. Direct impacts included alterations in
landscape that caused removal of plant material. Indirect impacts resulting in seagrass
loss included sediment resuspension, changes in sediment grain size, and the water
column mixing. A hydrodynamic model developed by Fonseca et al. (2002) suggests that
wind wave energy may impact seagrass distribution in Tampa Bay, however, ship
generated waves were not included in their investigations. Wave crests approaching one
meter in height have been observed impacting shallow flats in Hillsborough Bay
following the passage of ship traffic.
In 1988, a 25-year rainfall event rapidly reduced the salinity from 25PSU to near zero in
several areas of Hillsborough Bay. Low salinity as a result of this event may have had a
detrimental effect on plantings, especially those located close to areas with discharges
from stormwater pipes and tidal creeks. H. wrightii may tolerate salinity ranging from
5PSU (McMahan 1968) to over 70PSU (McMillan and Moseley 1967), however,
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Experimental Transplanting in Hillsborough Bay
mortality occurs as the salinity drops to less than 4PSU (McMahan 1968). Other
potential impacts from the storm event include sediment erosion and deposition of
debris that covered the sod units. Further, increased concentrations of water column
dissolved organic matter that may have decreased the amount of available light for
seagrass followed the rain event (Boler 1990).
Finally, transplanted material may have also been lost due to bioturbation. Fonseca et al.
(1996a) noted that bioturbation exclosure cages greatly improved survival of plantings in
western Middle Tampa Bay. We observed signs of apparent bioturbation, which may
have impacted transplants, in many Hillsborough Bay planting areas.
Personnel Expenditures
The costs associated with the transplanting projects totaled about 2300 man-hours
including the planning, transplant and monitoring phases. Over 60% of the time was
used for collection and transplanting of seagrass material (Figure 6). Time used for
monitoring ranged from 130 to 202 man-hours per year. Comparing project time to the
maximum areal coverage attained in 1992, nearly two man-hours were utilized for each
square meter of seagrass grown.
Figure 6. Man-hours vs. Halodule wrightii areal coverage, 1987–1994.
CONCLUSIONS
When the Tampa Bay Experimental Seagrass Planting Project was initiated in 1987,
water quality conditions appeared to be adequate to support persistent seagrass growth
on most of the intertidal and shallow subtidal areas in Hillsborough Bay. However, most
areas of Hillsborough Bay lacked seagrass coverage at that time. The transplant effort
successfully provided initial H. wrightii source material to these areas. For example,
transplant area 2 in northeastern Hillsborough Bay was devoid of seagrass coverage in
1987 but now has ca. 1.1ha of H. wrightii (City of Tampa 2003).
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Avery & Johansson
It is less clear if the project was successful in identifying areas suitable for seagrass
growth. Although the light climate appeared sufficient to allow seagrass growth in the
selected transplant areas, other factors appeared to impede transplant survival. High
macroalgae biomass and sediment transport was observed at several transplanting sites
and may have hindered the survival and expansion of the planting units. Also, wave
impacts and bioturbation may have negatively affected the plantings.
The final calculated areal coverage of the transplanted H. wrightii material at the
termination of monitoring in 1994, may not reflect the actual contribution of the
transplanting effort to the total seagrass coverage in Hillsborough Bay at that time and
later. Although many transplants did not persist at a specific planting site, some of the
original material may have been redistributed and promoted vegetative growth in
surrounding areas.
ACKNOWLEDGMENTS
We gratefully acknowledge the following: Mike Burwell, Bridget Kelly, Theresa Meyer, Gene Pinson and
Andy Squires for their contributions during the planting and monitoring phases of the project; and Frank
Courtney for facilitating the permitting process. Finally, we thank Kerry Hennenfent, Patricia McNeese
and Hugh Kirkman for their review of the manuscript.
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Zimmerman CF, Montgomery JR. 1984. Effects of a decomposing drift algal mat on sediment pore water
nutrient concentrations in a Florida seagrass bed. Marine Ecology 19:299–302.
Authors’ e-mail: walt.avery@ci.tampa.fl.us; roger.johansson@ci.tampa.fl.us
101
❖
COSTS AND SUCCESS OF LARGE-SCALE EELGRASS
(ZOSTERA MARINA L.) PLANTINGS IN NEW ENGLAND
(NEW HAMPSHIRE AND MAINE)
R.C. Davis, J.T. Reel, F.T. Short & D. Montoya
ABSTRACT
The U.S. Army Corps of Engineers, New England District (USACE) was required to plant 5.5
acres of eelgrass (Zostera marina L.) as mitigation for impacts associated with maintenance
dredging to deepen an existing mooring field. The 5.5 acres were planted at four sites, including
2.5 acres within the dredging footprint. Eelgrass was collected from naturally occurring beds
nearby and transplanted using the Horizontal Rhizome method (HRM) and the TERFS™
(Transplanting Eelgrass Remotely with Frame Systems) method. Monitoring locations were
established at each site to monitor percent survival one month, six months, and one year postplanting.
A total of 65,359 planting units were installed over 5.575 acres at four sites with an average
planting density of 11,724 planting units per acre. At the end of one year, overall percent survival
was 51.4% resulting in the creation of approximately 4.0 acres of eelgrass habitat at two sites. Two
other sites were not successful and had little or no eelgrass surviving at the time of the final
monitoring. Overall cost for the project, not including the initial site selection, was
approximately $495,935 or $219,829 per hectare ($88,957 per acre) in 2002 dollars. Major factors
contributing to the cost included use of divers to complete the majority of the work in adverse
conditions (cold water, strong current) and use of wage rates required by USACE contracting
guidelines (per the Davis-Bacon Act). The per hectare and per planting unit costs for this project
were less than for a previous successful large-scale eelgrass planting project in New England with
comparable success, but more expensive than similar ongoing projects in the Chesapeake Bay,
primarily due to the use of divers.
INTRODUCTION
The U.S. Army Corps of Engineers, New England District (USACE) was authorized to
conduct maintenance dredging in Little Harbor, NH to increase the depth in a portion of
an existing mooring field. Because eelgrass (Zostera marina) existed in the 5.5-acre portion
of the mooring field to be dredged, state and federal regulatory agencies required the
USACE to plant 5.5 acres of eelgrass as mitigation for the impacts associated with the
dredging. The mitigation was required because eelgrass meadows create important habitat
and form a basis of primary production that supports ecologically and economically
important species (Thayer et al., 1984; Orth et al., 1984; Heck et al., 1995). Eelgrass, and
seagrasses in general, are an essential component of healthy estuarine and coastal ecosystems
(Fonseca et al., 1998). Eelgrass plants baffle wave energy (Fonseca and Cahalan, 1992),
creating a depositional environment and provide sediment stabilization (Ward et al., 1984).
The plants also filter and retain nutrients from the water column (Short and Short, 1984).
Transplanting seagrass has been used as a means for mitigating impacts to naturally
occurring seagrass beds due to coastal development for a number of decades (Fonseca et al.
1998).
The USACE selected four sites for transplanting eelgrass to mitigate for the impacts to the
existing beds in Little Harbor. These sites included 2.5 acres within the dredging footprint
in Little Harbor, and 3.0 acres offsite (1.5 acres at Kittery Point, ME, 1.0 acres at the Schiller
Terrace and 0.5 acres at Pierces Island, NH; Figure 1). This paper describes the methods
103
Davis, Reel, Short & Montoya
used to collect and transplant 5.5 acres of eelgrass and the results of the one-year postplanting monitoring effort. The costs for the overall project are presented and compared
with the costs of two other similar projects, one completed and one underway. The process
and costs for identifying the potential transplanting sites are not described in this paper.
Figure 1. Location of transplanting sites in New Hampshire and Maine. Dredging project occurred in Little
Harbor.
METHODS
Transplanting
Eelgrass Collection
Eelgrass was collected from existing beds in Little Harbor, NH and Fishing Island, ME.
Material collected from Little Harbor was transplanted in Little Harbor. Material collected
from the Fishing Island bed was transplanted at Pierces Island, Kittery Point, Little Harbor
and Schiller Terrace. The Fishing Island bed is predominately intertidal and has been
successfully used as donor bed for subtidal transplanting (Davis and Short, 1997). The
majority of the transplanted eelgrass was collected from the Fishing Island bed.
Two methods were employed to collect the eelgrass, based on water depth. When water
depth was greater than 3.0 feet, scuba divers harvested plants by swimming over the beds
and hand picking individual shoots or small groups of shoots out of the sediment. The
shoots were rinsed free of sediments and arranged with the rhizomes together and loosely
104
Large-Scale Eelgrass Plantings
bundled in groups of 50 to 100 and secured with a rubber band. In water depth less than 3.0
feet, eelgrass shoots were harvested by wading in the plant beds and hand picking individual
shoots. These shoots were arranged and bundled in the same manner as the diver-harvested
shoots. The bundles of harvested eelgrass shoots were stored in mesh dive bags during
collection. The bundles were moved to a lobster car and stored floating in the river until
they were ready for transplant. All eelgrass was transplanted within 48 hours of collection.
Eelgrass Planting
Eelgrass shoots were planted using the Horizontal Rhizome Method (Davis and Short,
1997, hereafter referred to as HRM) and the TERFS™ (Transplanting Eelgrass Remotely
with Frame Systems) method (Short et al, this volume). Both methods are also described
in National Oceanic and Atmospheric Administration’s Guidelines for the Conservation
and Restoration of Seagrasses in the United States and Adjacent Waters (Fonseca et al.
1998). Each HRM plot consisted of 49 planting units on one-foot centers in a 36-squarefoot area. A 6 × 6 foot PVC pipe frame with rope intersections every foot was placed on the
sediment surface and used as a planting grid for the HRM plots. Planting units were
installed at every intersection on the grid. A planting unit consisted of two eelgrass shoots
with the rhizomes overlapping (Figure 2A). Divers installed each planting unit by placing
it on the sediment surface and anchoring it in place, flush with the sediment surface, using
a bamboo skewer bent in half. Once all the planting units were installed, the planting grid
was gently lifted off the plants and placed in the next planting location. Grids were planted
in a checkerboard pattern.
A
B
Figure 2. Eelgrass transplanting methods. A) Horizontal Rhizome Method (HRM) in which two adult
eelgrass shoots, with overlapping rhizomes are anchored into the sediment using a bamboo skewer. See Davis
and Short, 1997 for further details. B) TERFS™ method in which eelgrass planting units with overlapping
rhizomes are tied to a weighted frame. See Short et al. in these proceedings for further details. Note: frames
were tied with both 50 and 25 planting units (25 planting units shown in the figure).
TERFS™ are approximately 2.7-foot square weighted wire frames with 50 or 25 planting
units attached (Figure 2B). The size of the wire frame is constant and when 25 planting
105
Davis, Reel, Short & Montoya
units are used, the spacing interval is approximately 5 inches on center. When 50 planting
units are used, the spacing interval is approximately 2.5 inches on center. The majority of
TERFS™ deployed for this project contained 50 planting units. Each planting unit consisted
of two eelgrass shoots with overlapping rhizomes tied to the frame with biodegradable paper
ties. Planting units were tied to the frames on shore, and the frames were transported to
transplant site by boat and placed in the water at the approximate planting locations. Divers
arranged the frames, generally in groups of four, around a center stake and ensured that the
frames were level and the rhizomes were on the substrate. The frames were left in place for
a minimum of 21 days to allow adequate time for the plants to root. At the end of this time,
the frames were gently lifted off the plants and removed from the site by divers. The
placement and removal of the frames can also be accomplished from the boat, without the
use of divers (see Short et al., this volume).
Protective fencing was installed at the Kittery Point and Schiller Terrace sites to baffle wave
energy and potentially reduce green crab bioturbation (Davis et al. 1998). Fences were
constructed by installing 1" diameter wooden stakes on 6.0-foot centers and attaching, with
cable ties, black tensor netting (4 feet tall) flush with the sediment surface. At Kittery Point,
the fencing was placed off of the deep edge of the site, parallel to the main river channel for
the entire length of the site. At Schiller Terrace, three small sections of fencing were placed
across the planting area, perpendicular to the main river channel. Protective fencing was not
used at Little Harbor or Pierces Island due to the potential for the fencing to interfere with
recreational and commercial boating. All fencing material was removed in late November
2001 to prevent the fences from trapping ice that could damage the transplanted eelgrass
(Davis and Short, 1997).
Monitoring
Seventy-four locations (36 HRM and 38 TERFS™) of the 526 total locations planted were
sampled for the one-month, six-month and one-year monitoring events to allow for
tracking the change in percent survival over time. The number of sample locations was
allocated among the four sites based on the area planted. During the first monitoring event
(one-month after transplanting), a ½" diameter screw anchor with a toggle buoy attached
was inserted in the middle of a group of two HRM plots or four TERFS™ to permanently
mark the sample location. During the one- and six-month monitoring events, percent
survival was determined by counting the number of shoots in two plots planted diagonal
to each other (Figure 3). At the one year monitoring event, percent survival could no longer
be determined because of shoot coalescence. Instead, the number of shoots in two 25-cmsquare subquadrats of a 1.0 square meter sampling grid were counted to estimate the shoot
density and the percent cover of the entire quadrant was estimated.
RESULTS
Transplanting
A total of 65,359 planting units were installed over 5.575 acres at four sites from July 20,
2001 to September 26, 2001 at an average planting density of 11,724 planting units per acre
(Table 1).
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Large-Scale Eelgrass Plantings
Figure 3. Monitoring design for installed eelgrass. During the one- and six-month monitoring events, percent
survival was determined by counting the number of shoots on two plots planted diagonal to each other.
Table 1. Number of planting units installed and area planted.
SITE
Little Harbor
Kittery Point
Pierces Island
Schiller Terrace
NO. PLANTING UNITS BY METHOD
HRM
TERFS™
18,130
10,143
2,009
7,252
9,625
9,900
3,800
4,500
107
AREA (ac.)
2.532
1.5
0.47
1.073
Davis, Reel, Short & Montoya
Little Harbor
A total of 27,755 planting units were transplanted within a 2.532-acre area in Little Harbor
at 568 locations from July 20, 2001 through August 17, 2001 (Table 1). A total of 370 HRM
plots and 198 TERFS™ were planted. One hundred eighty-seven (187) TERFS™ contained
50 planting units each, and 11 TERFS™ contained 25 planting units each. The resultant
planting density in Little Harbor was 10,962 shoots per acre.
Kittery Point
A total of 20,043 planting units were transplanted in a 1.5 acre area at Kittery Point at 415
locations from August 20, 2001 through September 26, 2001. The majority of the eelgrass
was transplanted from August 20, 2001 through August 28, 2001 using HRM. TERFS™
were placed at Kittery Point from July 20, 2001 through September 19, 2001, with majority
of the TERFS™ installed between August 17, 2001 and September 19, 2001. A total of 207
HRM plots and 208 TERFS™ were planted (Table 1). One hundred eighty-eight (188)
TERFS™ contained 50 planting units each and 20 TERFS™ contained 25 planting units
each. The latter TERFS™ were placed in the intertidal zone at approximately the same
elevation from which the donor material was collected. The resultant planting density at
Kittery Point was 13,362 planting units per acre.
Pierces Island
A total of 5,809 planting units were transplanted in a 0.47-acre area at Pierces Island at 117
locations. Eelgrass was planted using HRM at the Pierce Island site on three dates, August
23, August 30 and September 5, 2001. TERFS™ were placed at the Pierce Island site from
August 30, 2001 through September 19, 2001 with the majority of the TERFS™ installed
between August 30 and September 6, 2001. A total of 41 HRM plots and 76 TERFS™ were
planted (Table 1). All TERFS™ contained 50 planting units each. The resultant planting
density at the Pierce Island site was 12,359 planting units per acre.
Schiller Terrace
A total of 11,752 planting units were transplanted in a 1.073-acre area on the Schiller
Terrace at 238 locations from September 3, 2001 through September 18, 2001. A total of
148 HRM plots and 90 TERFS™ were planted (Table 1). All TERFS™ contained 50
planting units each. The resultant planting density at the Schiller Terrace was 10,952
planting units per acre.
Monitoring
The same 74 planted areas (36 HRM and 38 TERFS™) were sampled at the one-month,
six-month and one-year monitoring events to allow for tracking the change in percent
survival over time. Percent survival for the TERFS™ method ranged from 27% to 73%
(mean of 51.6%) one month after transplanting (Table 2). Percent survival for the HRM
method ranged from 22% to 76% (mean of 47.7%; Table 2). The total number of shoots
present at each site declined at the time of the one-month monitoring, then generally
increased during the next two monitoring events at the two sites that were ultimately
considered successful (Figure 4). This pattern was more clearly evident in the changes in
shoot density over time (Figure 5).
108
Number of Shoots
(in Thousands)
Large-Scale Eelgrass Plantings
80
70
Little Harbor
60
50
Pierces Island
Kittery Point
Schiller Terrace
40
30
20
10
0
Installed
One
Six
Twelve
Months After Transplanting
Number of Shoots
(in Thousands)
16
Little Harbor
14
Kittery Point
12
10
Pierces Island
Schiller Terrace
8
6
4
2
0
Installed
One
Six
Twelve
Months After Transplanting
Figure 4. Total number of shoots over time for the two methods: Horizontal Rhizome Method (HRM)
above, TERFS™ method below. Note: the y-axis scale varies.
After the one-year monitoring, overall percent survival was calculated by multiplying the
estimated percent survival for each method at a particular site by the number of planting
units installed by that method. When estimated percent survival exceeded 100% (due to
production of new shoots by the transplanted eelgrass), then percent survival was capped
at 100% and the number of shoots initially transplanted was used in the calculation of
overall percent survival. One year after transplanting, overall percent survival was 51.4%,
similar to the percent survival determined at the one-month monitoring (Table 2).
109
Shoots per Square Meter
Davis, Reel, Short & Montoya
150
Little Harbor
125
Kittery Point
100
Pierces Island
Schiller Terrace
75
50
25
0
Installed
One
Six
Twelve
Shoots per Square Meter
Months After Transplanting
300
Little Harbor
250
Kittery Point
200
Pierces Island
150
Schiller Terrace
100
50
0
Installed
One
Six
Twelve
Months After Transplanting
Figure 5. Changes in shoot density (number of shoots per m2) for eelgrass transplanted using the two
methods: Horizontal Rhizome Method (HRM) above, TERFS™ method below. Note: the y-axis scale varies.
Error bars are ± s.e.
Table 2. Percent survival one month after installation.
SITE
PLANTING METHOD
HRM
TERFS™
Little Harbor
Kittery Point
Pierces Island
Schiller Terrace
76 %
44 %
22 %
23 %
45 %
73 %
27 %
51 %
Overall average
54.2%
53.5%
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Large-Scale Eelgrass Plantings
DISCUSSION
Success
The 51.4 % survival at the one-year monitoring slightly exceeded the 50% survival criteria
required by the project’s permits. The 51.4% overall survival was due to the results from
Little Harbor and Kittery Point. Using the area planted at these two successful sites, the
project restored approximately 4.0 acres of eelgrass habitat.
The data indicated that the HRM method had higher percent survival than the TERFS™
method in Little Harbor. Qualitatively, the HRM method seems to have allowed for greater
expansion of the transplanted shoots at this site. Based on qualitative observations of the
areas surrounding the monitoring locations, the majority of transplants in Little Harbor
survived and were expanding.
At the Kittery Point site, the data indicated that the HRM method had a higher percent
survival after one year than the TERFS™ method, but this was due to one highly successful
HRM plot on the site. Overall, most of the locations planted using TERFS™ had surviving
shoots and eelgrass patches created by the TERFS™ were clearly visible from the surface.
At the Schiller Terrace site, no plants survived within the monitoring locations. Some
surviving shoots were observed at the site, but the majority of the transplanted material did
not survive. The situation was similar at the Pierces Island site, where no surviving
transplants were found within the monitoring locations, but sparse patches of eelgrass were
observed at the site. Whether these shoots survived from the initial transplanting or
recruited to the site from nearby eelgrass beds could not be determined.
Overall, both methods were successful in establishing eelgrass in Little Harbor and in small
patches at Kittery Point. These results suggest that site conditions, rather than transplanting
method, determine the ultimate outcome of the project. The HRM method was more easily
employed than the TERFS™ due to the logistics of the latter (i.e., transporting frames to the
deployment sites and retrieving them later). The HRM method also appears to allow for
greater expansion of, and increase in, the number of shoots (Figure 4). Therefore, the HRM
method is preferable to the TERFS™ method for transplanting eelgrass when large areas
need to be restored. However, the TERFS™ method was successful at establishing higher
shoot densities at Kittery Point and similar shoot densities at Little Harbor compared to the
HRM method (Figure 5). Shoot density is an important habitat attribute that is often
included in monitoring programs and has been related to important seagrass functions
(Fonseca et al., 2000). The TERFS™ method may be preferable on smaller projects that
include shoot density or any concomitant habitat functions as project goals.
Costs
The overall cost for the project was approximately $495,935.00 or $219,829 per hectare
($88,957 per acre) in 2002 dollars (Table 3). These costs do not include the initial site
selection efforts, any subsequent monitoring events that may be required by the project’s
permits, or USACE costs for project planning, administration and management. The major
factor that contributed to the overall cost was the use of divers to complete the majority of
the transplanting work. The water depth at the sites, which varied between –0.75 to 8.0
111
Davis, Reel, Short & Montoya
meters below mean low water, necessitated the use of divers, rather than relying solely on
wading, snorkeling or the use of TERFS™. The USACE contracting guidelines required
that the divers be compensated in accordance with the Davis-Bacon Act and paid prevailing
wages. Working in adverse conditions, with water temperature averaging 14°C and currents
as fast as 0.5 meters/second, decreased the length of dive times and increased the length of
surface intervals. Overall, the per hectare and per planting unit costs for this project were
less than for a previous successful large-scale eelgrass planting project in New England with
comparable success (Table 3). The New Hampshire Port Authority (NHPA) Mitigation
project was completed in the mid-1990s (Davis and Short 1997). Divers were used for the
NHPA project but were mostly undergraduate and graduate students and were not paid
“prevailing wage rates.” However, the costs for the NHPA project included the first three
years of an extensive monitoring program to quantify the use of eelgrass habitat by
macroinvertebrate and fish communities at both transplanted and reference beds. The
USACE project was more expensive than similar ongoing projects in the Chesapeake Bay.
The Chesapeake Bay projects (MD-SAV and VA-SAV in Table 3) involve planting eelgrass
and other species of SAV at sites in Maryland and Virginia waters in the lower Potomac
River. The sites are located in relatively shallow water and are being completed without the
use of divers. In addition, the planting density for the Chesapeake Bay projects is less than
that used in the New England projects. The costs reported here for the Chesapeake Bay
projects do not include costs associated with identifying and conducting test transplanting,
which were significant. The costs for the monitoring program to be completed for the
Chesapeake Bay projects are also not included in Table 3.
Table 3. Cost comparison for submerged aquatic vegetation transplanting projects.
PROJECT
NO. PLANTING
UNITS (PU)
Little Harbor, NH
NHPA Mitigation
MD – SAV
VA – SAV
65,359
104,690
150,000
15,000
HECTARES
(ACRES)
2.256 (5.575)
2.934 (7.25)
8.0937 (20.0)
0.80937 (2.0)
COSTS1
PER PU
PER HECTARE
$7.59
$8.92
$2.05
$3.40
$ 219,829.33 2
$ 318,335.89 3
$ 37,992.51 4
$ 63,011.97 4
1
Costs were converted into 2002 dollars using Consumer Price Index conversion factors.
Does not include site selection costs (minimal); monitoring very limited
3
Includes costs for significant 3 years of monitoring (see Davis and Short 1997)
4
No divers being used; does not include site selection costs (significant); does not include monitoring
costs (significant)
2
These results indicate the need to specifically state which project components are being
considered when reporting costs for SAV transplanting projects. In certain instances,
particularly when the use of divers is required, the actual installation costs can be
considerable and comprise a substantial portion of the overall project cost. In other cases,
site selection and monitoring can be the most costly project components, particularly if
multi-year test transplanting to select final planting sites, or multi-year monitoring is
required.
112
Large-Scale Eelgrass Plantings
REFERENCES
Davis RC, Short FT. 1997. Restoring eelgrass, Zostera marina L., habitat using a new transplanting technique:
the horizontal rhizome method. Aq. Bot. 59:1–15.
Davis RC, Short FT, Burdick DM. 1998. Quantifying the effects of bioturbation by green crabs (Carcinus
maenas) on eelgrass (Zostera marina) transplants using mesocosm experiments. Restoration Ecology
6(3):297–302.
Fonseca MS, Cahalan JA. 1992. A preliminary evaluation of wave attenuation by four species of seagrass. Est.
Coast. Shelf Sci. 29:501–507.
Fonseca MS, Kenworthy WJ, Thayer GW. 1998. Guidelines for the Conservation and Restoration of Seagrasses in the
United States and Adjacent Waters. NOAA Coastal Ocean Program Decision Analysis Series No. 12. NOAA
Coastal Ocean Office, Silver Spring, MD. 222 pp.
Fonseca MS, Julius BE, Kenworthy WJ. 2000. Integrating biology and economics in seagrass restoration: How
much is enough and why? Eco. Eng. 15:227–237.
Heck KL Jr., Able KW, Roman CT, Fahay MP. 1995. Composition, abundance, biomass and production of
macrofauna in a New England estuary: comparisons among eelgrass meadows and other nursery habitats.
Estuaries 18(2): 379–389.
Orth R J, Heck KL Jr, vans Montfrans J. 1984. Faunal communities in seagrass beds: a review of the influence
of plant structure and prey characteristics on predator-prey relationships. Estuaries 7(4A): 339–350.
Short F T, Short CA. 1984. The seagrass filter: purification of estuarine and coastal waters. Pp 395–413 in: V.S.
Kennedy (Editor), The Estuary as a Filter. Academic Press.
Short FT, Davis RC, Koop BS, Gaeckle SJ, Burdick DM. 2003. Using TERFS™ and Site Selection for
Improved Eelgrass (Zostera marina L.) Restoration Success.
Thayer GW, Kenworthy WJ, Fonseca MS. 1984. The ecology of eelgrass meadows of the Atlantic coast: a
community profile. U.S. Fish and Wildlife Service. FWS/OBS-84/24, 85 pp.
USACE 2001. Eelgrass Planting, Rye and New Castle, New Hampshire, Pierces Island, New Hampshire, and
Kittery Point, Maine. Construction Solicitation and Specifications DACW33-01-R-0016.
Ward LG, Kemp WM, Boynton WR. 1984. The influence of waves and seagrass communities on suspended
particulates in an estuarine embayment. Mar. Geo. 59: 85–103.
RCD (Quantitative Environmental Analysis, LLC, 80 Glen St., Glens Falls, NY 12801); JTR (Rummel,
Klepper & Kahl, LLP, 81 Mosher Street, Baltimore, MD 21217); FTS (Jackson Estuarine Laboratory,
University of New Hampshire, 85 Adams Point Road, Durham, NH 03824); DM (USACE New England
District, 696 Virginia Road, Concord, MA 01742-2751)
113
❖
BISCAYNE BAY SEAGRASSES AND RECENT RESTORATION EFFORTS
G.R. Milano & D.R. Deis
ABSTRACT
Regional modifications of freshwater inflow, and past dredging and filling practices associated
with the rapid urbanization of the Greater Miami area, have resulted in serious environmental
degradation to Biscayne Bay. Two human-made inlets through Miami Beach have altered
circulation and salinity regimes and associated bay communities. Low coastal wetlands have been
virtually eliminated in north Biscayne Bay, and have been covered with dredged bay bottom fill.
Dredge fill has been also placed on submerged bay bottom communities to make developable
land and causeways to offshore barrier islands. In addition, seawalls were commonly constructed
to contain the newly created land. Dredging of the north and central Biscayne Bay for spoil
emplacement and the creation of navigation channels resulted in numerous dredged areas of
varying size and depth, ranging from 2.1 meters (7 feet) to 9.2 (30 feet) in depth. Deep dredge
troughs and borrow areas in north Biscayne Bay have been shown in past studies to have poor
water quality, to be of limited habitat value, and to be a source of chronic turbidity compared to
natural bay bottom communities. Restoration of the dredged areas to shallower depths can result
in the reestablishment of seagrass or other benthic vegetation, enhance habitat value, and improve
water quality. The central bay area is a transition zone from the heavily urbanized northern basins
to the nearly un-dredged south bay area. Much of southern Biscayne Bay has good water quality
and has retained its relatively pristine habitat.
This paper summarizes recent (1980–present) seagrass mapping, monitoring, and restoration
efforts in Biscayne Bay. The Miami-Dade Department of Environmental Resources
Management (DERM) has successfully tested the feasibility of using maintenance dredging spoil
material to restore previously dredged areas in Biscayne Bay to natural contours.
INTRODUCTION
Biscayne Bay and its associated coastal ecosystems are some of Florida’s most valuable
natural resources. The bay provides habitat for a productive and diverse community of
tropical marine plants and animals. It offers a variety of commercial and recreational
opportunities to visitors and the over 2 million residents of Metropolitan Miami-Dade
County. The bay is a shallow subtropical estuary located on the southeast coast of Florida
(see Figure 1). Extending approximately 56 kilometers (35 miles) from north to south and
varying in width from less than 1.6 kilometers (one mile) to approximately 12.8 kilometers
(eight miles), it covers an area of 572 kilometers (220 square miles). The bay is bordered on
the west by the Greater Miami area and on the east by a series of barrier islands and
submerged vegetated banks, which separate the bay from the Atlantic Ocean. The bay is
shallow [less than 4 meters (13 feet)]) except in the dredged bottom areas, which range from 2.5
meters (8.2 feet) to 16.7 meters (50 feet) in depth. Prevailing winds are from the east-southeast,
and the bay is sheltered from oceanic swells by the offshore reef tract, barrier islands, and
vegetated mud banks. After tide, the second most important factor affecting circulation in
Biscayne Bay is wind. Freshwater naturally enters the bay through upland runoff, groundwater
seepage, and rainfall. In the mid-1900s, to control upland flooding, a network of canals was
created to discharge large pulses of freshwater into the bay during periods of heavy rainfall.
Seasonal water temperature ranges from 13ºC to 31ºC. Salinity (one meter depth [3.3 ft]) is
measured monthly at a total of 100 stations. Of the 75 stations located in saline areas, average
salinity was less than 33 ppt at 32 stations and greater than 36 ppt at 7 stations (Alleman et al,
1995). Wanless (1976) categorized the major bay sediment types as quartz, carbonate, calcareous
115
Milano & Deis
sand, calcareous mud, calcitic mud or peat, and quartzose calcareous. He also reported that over
50% of Biscayne Bay has less than 15 centimeters (6 inches) of sediment cover over the
limestone bedrock and natural sediment accumulation is confined primarily to the deeper midbay axis (Wanless, 1969). Water quality in Biscayne Bay meets state water quality standards.
Figure 1. Biscayne Bay, with limits of Biscayne National Park and Biscayne Bay Aquatic
Preserve.
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Biscayne Bay Seagrass Restoration
Biscayne Bay Resource Impacts
Rapid urbanization and associated coastal development over the last 100 years has severely
altered natural habitats in Biscayne Bay (Harlem, 1979). The northern third of the Bay (north
bay), which has been most severely impacted by development, is subdivided by six filled
causeways and a major seaport facility. Low coastal wetlands have been virtually eliminated in
north Biscayne Bay. Over 50% of the existing north bay bottom area is barren (Harlem, 1979;
Milano, 1983), caused by the creation of deep dredge holes and associated spoil emplacement,
and chronic elevated turbidity levels. High turbidity in the north Biscayne Bay area has been
correlated with resuspension of unconsolidated bay bottom and spoil island shorelines, eroding
margins of dredge banks and unvegetated bottom sediments (Wanless et al. 1984).
The central bay area is a transition zone from the heavily urbanized northern basins to the
almost undredged south bay area. Free exchange with ocean waters occur in this region through
a 14.5-kilometer (nine-mile) system of shallow vegetated mud banks.
The south bay area western shoreline rises more gradually than the northern regions, with
elevations of only 0.3 to 0.6 meters (one to two feet) above mean sea level for 1.6 kilometers (1
mile) or more inland from the shore. As a result of the resource protection regulations provided
through Biscayne National Park and local regulatory agencies, the south bay contains pristine
habitats due to the absence of heavy development.
Activities that disturb the bottom communities of the bay disrupt the balance between biological
and physical forces that maintain the bay’s water clarity and sediment stability. These activities
include dredging and filling operations, wave energy deflecting off seawalls, prop scarring and
scouring by recreational and commercial vessels, and bottom damage or disturbance by fishing
activities (Biscayne Bay Partnership Initiative, 2001).
A large body of scientific literature exists documenting the importance of coastal habitats to local
fisheries, food web relationships, habitat value, and as shoreline stabilizers (Idyll et al., 1968;
Odum et al., 1982; Lewis, 1990a). Seagrasses are important primary producers, sequestering
carbon, producing oxygen, and converting the sun’s energy into food and structure useful to
fish, invertebrates, and wildlife (Wood et al, 1969)
Biscayne Bay Restoration and Enhancement Program
Program Development
The natural qualities of Biscayne Bay, and the need to protect them have been recognized at
national, state, and local levels. In 1980, Congress created Biscayne National Park, originally
established in 1968 as a national monument, to preserve and protect tropical marine, terrestrial,
and amphibious life in relatively pristine portions of central and south Biscayne Bay and adjacent
environments (Figure 1). In 1974, in order to maintain the bay in an essentially natural
condition, the State of Florida passed the Biscayne Aquatic Preserve Act, and the Miami-Dade
County Commission declared Biscayne Bay an “Aquatic Preserve and Conservation Area” and
empowered the County Manager to develop a management plan for the bay. These efforts led
to the development of the “Biscayne Bay Management Plan” (Miami-Dade County Department
of Environmental resources Management and Miami-Dade County Planning Department,
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1980) and one of its principal implementation tools, the Biscayne Bay Restoration and
Enhancement Program. The Restoration and Enhancement Program, which was initiated in
1978, is funded through a variety of sources, and is locally administered through Miami-Dade
County Department of Environmental Resources Management (DERM). The primary goal of
the program is to restore, maintain, and improve the ecological, recreational, and aesthetic values
of the bay. Early efforts to develop specific strategies for Biscayne Bay resource management
were hampered by the lack of comprehensive, scientific data. Therefore, habitat restoration,
monitoring, and study of the bay became one of the key elements of the Biscayne Bay
Restoration and Enhancement Program.
Scientific Study and Habitat Restoration
The following investigations were conducted in the early 1980s as part of the Biscayne Bay
Restoration and Enhancement Program, and constituted the foundation upon which other
program components were built.
• Bottom Community Mapping
• Water Circulation in North Biscayne Bay
• Water and Sediment Quality
• Sources of Turbidity
• Benthic Sampling Program
• Fisheries Assessment
The data accumulated established a baseline against which to assess future changes in Biscayne
Bay. In addition, the program was guided initially by a Scientific/Technical Committee, which
compiled and ranked a list of bay-wide projects, which included the restoration of habitats in
Biscayne Bay, the filling of deep dredge holes in north Biscayne Bay, and the planting of
seagrasses.
Since 1988, DERM has restored and enhanced approximately 122 hectares (300 acres) of coastal
wetlands on public lands. Major wetlands restoration efforts have been conducted by DERM
at the following sites: Cape Florida State Park, Oleta River State Park, Highland Oaks Park,
North Virginia Key Preserve, Bear Cut Preserve, Florida International University (Biscayne Bay
campus), and Chicken Key Bird Rookery (Milano, 1999a, 1999b, 2001). DERM has also created
wetlands on dredge filled islands in north-central Biscayne Bay (Milano, 2000).
Additionally, DERM has created or restored over 24 hectares (60 acres) of maritime hammock
at public parcels throughout Biscayne Bay. Island restoration and enhancement activities are
underway to stabilize eroding shorelines, restore historical dune communities and wetlands,
eradicate exotic vegetation, and create wetlands, dune, coastal strand and tropical hardwood
hammock communities. DERM has successfully completed habitat restoration on 18 islands in
Biscayne Bay. Four habitat types (dune, coastal strand, tropical hardwood hammock, and
wetlands), consisting of approximately 90 species, have been established on natural and dredge
spoil islands. Dune/strand species have been planted at 15 islands and tropical hardwood
hammock communities have been established on seven dredge spoil islands in Biscayne Bay
(Milano, 2000).
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Biscayne Bay Seagrass Restoration
Seagrass Mapping and Distribution
Bottom communities in Biscayne Bay were mapped in 1983 (Milano, 1983) and updated in
1993, to document the current distribution of seagrasses, hard bottom, and other bottom types.
Bottom types were delineated through the use of high-resolution aerial photographs and inwater inspections. In addition, the inventory of the bottom communities of Biscayne National
Park were updated in the late 1990s (Lewis et al., 1999).
The seagrass community is the dominant bottom community type, covering approximately 64%
of the total bay. Seagrasses occur in shallow sand or mud covered areas where light is able to
penetrate to the bottom. Turtle grass (Thalassia testudinium)(=Thalassia) is the predominant
seagrass in Biscayne Bay, and is most abundant in central and south bay. Shoal grass (Halodule
wrightii)(=Halodule) and manatee grass (Syringodium filiforme)(=Syringodium) cover significant
portions of north Biscayne Bay, and shallow areas along western portions of central and south
Biscayne Bay.
Hard bottom communities comprise about 17% of the total bay bottom and are located
primarily in south bay. The most conspicuous organisms found in hard bottom communities
are soft corals and sponges. A total of approximately 15% of Biscayne Bay does not support
conspicuous plant or animal life. As stated earlier, north Biscayne Bay has been most seriously
impacted by development. Dredging has directly altered approximately 41% of the north bay,
and a total of 58% of north bay is devoid of submerged aquatic vegetation.
Seagrass Monitoring, 1985–1995
A long-term epibenthic bottom community monitoring program was initiated in 1985 to
establish a quantitative database for detecting trends and seasonal variability in Biscayne Bay
bottom communities. Initially, fifteen locations were picked to be representative of broad areas
of seagrass and hardbottom communities in Biscayne Bay. All of the sites are near one of
DERM’s water quality monitoring stations, which are monitored on a monthly basis. The
bottom community monitoring stations span the entire north–south extent of Biscayne Bay,
from Haulover Inlet in the north to the vicinity of the mouth of the C-111 canal in Manatee
Bay, in the south. The Manatee Bay stations were established in 1988, as a result of a very large
release of freshwater from the C-111 canal in the extreme south end of Miami-Dade County.
Permanent sampling transects were established at each station. The ends are marked with earth
anchors and attached sub-surface buoys, to aid in relocating the transect locations. Transects are
46 meters (150 feet) in length, with three permanent one square meter (3.3 foot) sampling
locations distributed along the transect to quantitatively sample seagrasses, soft corals, hard
corals, and algae. The species composition and relative abundance are recorded along each
transect. Fixed grids are located to quantify the density and diversity on each transect. A 1-meter
square (10.9 square feet) PVC (polyvinyl chloride) grid subdivided into 25 equal subunits is used
to define the area. At each grid, five of the subunits are randomly selected for counting. Seagrass
short shoots and blades are counted for Thalassia and Syringodium within these subunits. For
Halodule, only short shoots are counted. In addition, an estimate of standing crop is performed
for each station by collecting all aboveground biomass from three 0.04-meter square quadrats
adjacent to the permanent grids.
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Seagrass Monitoring, 1995–present
The current SAV monitoring design is comprised of fixed seagrass monitoring stations and
stratified random monitoring.
Fixed Stations: As stated above, a series of fixed transects were established in September 1985,
throughout the bay. Initially, sampling was conducted quarterly at 15 sites. Three additional sites
were added 1989, two in Manatee Bay and one in Barnes Sound. Currently, sampling is
conducted annually during the month of June at 10 of the original 15 sites. Monitoring of
stations located near Black Ledge and Turkey Point was discontinued in 1996. The three stations
added in 1989 were incorporated into DERM’s SAV monitoring program in northeast Florida
Bay, and sampling is currently conducted at these sites on a semiannual basis in May and
November.
Parameters collected include, seagrass shoot and blade density, standing crop biomass by species,
and seagrass composition along a 45-m transect. Shoot and blade density are determined at each
station by sampling a 0.2-m2 section at each of three fixed one-meter square grids.
Standing crop biomass is harvested from three 0.04-m2 areas at each station. Biomass samples
are segregated by species, rinsed in a mild HCl solution, then dried in an oven at 60ºC and
weighed.
Stratified Random Sampling: The monitoring network consists of 101 stratified random
sites sampled annually using the modified Braun-Blanquet cover-abundance scale (BBCA).
Overall cover for each species of seagrass, and total cover for all species is estimated using the
BBCA scale. Frequency, abundance, and density are calculated for each site. This method of
sampling is currently being used in Florida Bay and the Florida Keys National Marine
Sanctuary.
Biscayne Bay Large Scale Seagrass Restoration Efforts
Port of Miami Seagrass Mitigation Project
In October 1980, The U.S. Army Corps of Engineers (USACE) issued a dredge and fill permit
for expansion of the Miami Seaport Facility. As a special permit condition the Seaport was
required to plant 102 hectares (251 acres) of bay bottom with seagrasses to mitigate for damages
to 33 hectares (81acres) of grass beds. The detailed specifications of the planting and monitoring
program were prepared in October 1981. The program was divided into two phases: Phase I
included the planting and monitoring of one 10-hectare (25 acre) large-scale planting and
thirteen 0.4-hectare (1-acre) test plantings (Test Plots) intended to provide spatial, species,
planting methods, and other guidance for the planting of the remaining 86 hectares (213 acres)
in Phase II (Figure 2) (Dial and Deis, 1986).
The following survival rates were obtained from Connell Associates, 1984, the Biscayne Bay
Aquatic Preserve Management Plan report, Miami-Dade County Planning Department, 1986,
the 1995 Biscayne Bay SWIM Plan (Alleman et al, 1995), and Dial and Deis (1986).
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Biscayne Bay Seagrass Restoration
Figure 2. Seagrass planting efforts in Biscayne Bay, Florida, with locations of Port of
Miami Seagrass Mitigation Project 1981 through 1984 Phase I and II restoration efforts
and the 1988 Dredged Area Pilot Restoration Project. Stars – Port of Miami Seagrass
Mitigation Phase I planting efforts; triangles – Port of Miami Seagrass Mitigation Phase
II planting efforts; hexagon – Dredged Area Pilot Restoration Project.
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Milano & Deis
Phase I 13 Test Plots and a 10-hectare (25-acre) Planting Effort (Figure 2)
In 43% of the test plots, the degree of survival was rated as a total loss (Connell and Associates,
1984). Of those that survived, Thalassia shoots had the highest rate of survival (63%), followed
by Halodule shoots (46%) and Syringodium (9%). Halodule plugs, which were planted in six test
plots, had a 24% survival rate.
Planting success varied depending on the geographic location within Biscayne Bay. The most
successful sites were in clear water and not exposed to wave action. The goal of the seagrass
restoration program was to achieve an overall survival rate of 70%, but only 22% of the tested
plots achieved 70% survival (Dial and Deis, 1986).
The rate of survival in the 10-hectare (25-acre) planting off Mercy Hospital in central Biscayne
Bay was extremely low. After one year, the mean survival rate for Phase I was approximately
12%.
Phase II Large Scale Planting Efforts (Figure 2)
North Biscayne Bay 8-hectare (20-acre) Planting Site: A second phase of planting was
conducted in north Biscayne Bay, which demonstrated the highest rates of survival in Phase I.
This site is located between the NW 36 Street (Julia Tuttle) Causeway and the Venetian
Causeway (NW 15 Street) in north Biscayne Bay. 15 acres of Halodule shoots and 5 acres of
Thalassia shoots were planted during the summer of 1984. After a one year period, the mean
survival rate was approximately 12%.
Central Biscayne Bay 30-hectare 73-acre Planting Site: In the summer of 1985, 8 hectares
(25 acres) of seagrasses (primarily Thalassia and Syringodium) that were scheduled to be destroyed
by the Key Biscayne Beach Renourishment project, were transplanted to a 30-hectare (73-acre)
central Biscayne Bay site on 1.2-meter (4-foot) centers. Monitoring during the summer of 1986
revealed that the mean survival rate was 10% (Gaby and Langley, 1985).
Alternative Seaport Mitigation Plan
As stated earlier, the Miami-Dade County Seaport Department was required to complete a
mitigation program as a condition of a USACE regulatory permit. As of January 1988, the
Seaport had spent approximately $3,000,000, and the cost to fulfill the obligation of the balance
of the required seagrass planting, was estimated to be $1,200,000.
As a result of (Phase I and Phase II) very low survival rates and limited availability of suitable
planting sites, an alternative Seaport mitigation plan was proposed and approved by the USACE.
The USACE alternative mitigation plan consisted of the following habitat improvements and
activities, and were implemented by DERM:
• Continued monitoring of the previous phases of the seagrass planting efforts
• Wetlands restoration (5.3 hectares [13 acres]) at Oleta River State Park
• Biscayne Bay artificial reef construction
Additionally, Miami-Dade County DERM Class-1 Coastal Construction Permit required the
following mitigation components for impacts associated with the Seaport expansion activities:
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Biscayne Bay Seagrass Restoration
• Shoreline stabilization (riprap and mangrove planters)
• Inshore artificial reef construction
• Spoil Island Enhancement
Restoration of Seagrasses in Dredged Areas in North Biscayne Bay
In 1988, a study to evaluate alternative techniques for filling existing dredged areas in north
Biscayne Bay was initiated, and resulted in a three-phased pilot project in a 1.05-hectare (2.6acre) site located approximately 500 meters west of the western shore of Miami Beach at
Biscayne Point (NE 110 Street) (Figure 2).
Project Design and Development
The pilot project was developed to demonstrate the feasibility of using clean dredge spoil
material to restore previously dredged areas to natural depth contours, and to develop alternative
cost-effective, environmentally sound methods for spoil disposal. The following factors were
considered in the design of the pilot project to determine the spatial distribution of fill and the
volumes recommended for placement:
• Existing bottom conditions
• Long-term stability
• Environmental impact
• Cost-effectiveness
The selected dredge area is bordered on the north and south by shallow Syringodium seagrass
beds, and on the east and west by deeper bay bottom. Several spoil containment alternatives
were evaluated and eliminated due to impracticability, cost or environmental impact. Some of
these included sheet-pile dikes, earthen embankments, and concrete filled bags. The final spoil
containment system consisted of the construction of submerged rock dikes on the east and west
sides of the project site, where the water depths are greater. The water depth at the crest of the
dikes is –0.9 meters (–3 feet NGVD) (–0.6 meters –1.9 MLW]) and the slope of the
containment dike is 1 vertical: 2 horizontal.
The original plan was to fill the contained area with two types of fill. Approximately 1.2 meters
(4 feet) of clean dredge spoil material (9,175 cubic meters [12,000 cubic yards]) would be placed
into the 1-hectare (2.6-acre) depression followed by a 30.5-centimeter (12-inch) cap layer of
clean aragonite sand (2,300 cubic meters [3,000 cubic yards]). The cap sand layer was designed
into the project to provide a test of containment using coarse grain sands for future dredging
activities associated with the maintenance of the Atlantic Intra-coastal Waterway (AIW).
The objectives of this pilot project were twofold:
1. To restore north Biscayne Bay seagrass communities through the development and
implementation of techniques for the filling of deep areas in north Biscayne Bay with
clean dredge spoil material; and,
2. To identify cost-effective dredge spoil disposal alternatives, in order to eliminate the
need for AIW maintenance dredge spoil disposal on recently restored spoil islands, or
on submerged aquatic vegetation, within the USACE easement areas.
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Federal, state, and local environmental permits were obtained for the pilot project, and funding
assistance was provided through the Florida Inland Navigation District and the Miami-Dade
County Biscayne Bay Environmental Enhancement Trust Fund.
Implementation
The restoration project was constructed through three separate phases to provide optimum
flexibility in the design and implementation. The project was initiated in October 1991, and
completed in November 1994. Phase I consisted of the installation of lime-rock boulder
containment dikes on the eastern and western boundaries of the dredged area. Phase II involved
the filling of dredged area with clean dredge spoil material, from the Miami-Dade County
Seaport expansion project, to natural depths (–1.2 meters [–4 feet]). The original Phase II
construction contractor agreement was terminated due to noncompliance. As a result of limited
funds, Phase II was re-bid without the capping layer. The original Phase II contract fill material
was barged directly from the Seaport expansion dredging project, and the Phase II contract rebid fill material was barged from an upland staging site. Surface to bottom turbidity curtains
were positioned around the entire 1.05-hectare (2.6-acre) restoration site. Heavy equipment
deposited the fill material from the barge to the dredged area. Phase III consisted of the
transplanting of three species of seagrasses from nearby donor beds to the restoration site. The
following are detailed specifications included in the project:
• Only clean fill material was used for the restoration.
• To reduce turbidity and meet state water quality requirements, bedding materials and
lime-rocks were pre-washed prior to deployment.
• During the placement of all material, a surface to bottom (weighted) turbidity curtain
was positioned completely around the fill area, to contain fine materials within the work
site. The turbidity curtain was securely staked in position outside the edge of all seagrass
shoal margins adjacent to the fill area, and remained in place until physical stabilization
of the fill material.
• Weather permitting, fill material was placed and leveled using a standard excavator
during daylight hours only.
The project site was monitored for the following parameters: turbidity during construction;
seagrass density adjacent to the fill area pre-and post-construction; any changes in elevation of
the top of filled area by means of depth surveys; and success of experimental seagrass
transplantation. Turbidity levels were measured continuously during the construction period.
A detailed “as-built” was required to ensure fill quantities and final design elevation compliance.
Three species of seagrass, Halodule, Syringodium, and Thalassia, were planted at the site, using bare
root material harvested from approved nearby donor sites.
The turbidity curtains were effective in containing the turbidity on-site. Wind-driven currents
were found to reduce the effectiveness of the surface to bottom turbidity curtains. As a result,
fill deployment activities were not allowed to occur during wind events of 15 knots or greater.
The turbidity curtains were also found to be very effective in the containment of the deposited
fill. As a result, the limerock boulder containment dikes may not be a necessary project
component for future restoration efforts. In addition, the fill boundaries and the positioning of
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the turbidity curtains could be located closer to adjacent desirable seagrasses, this would
eliminate the resulting trough between the existing seagrasses and the restored seagrass area.
After 24 months from completion of construction, no significant changes were observed in the
elevation of the 1.05 hectare (2.6-acre) filled area, and no project-related impacts to nearby
habitats were detected. Table 1 includes project cost details for all three phases of the restoration
effort. Total cost for all three phases was $576,000 or $548,571/ hectare.
Table 1. Restoration of seagrasses in a north Biscayne Bay dredged area, project costs:
DESCRIPTION
UNITS
Containment Dike System:
Mobilization
lump sum
Silt Barrier
lump sum
Bedding Material
tons
Limerock Riprap
tons
Filter Fabric
square yards
Navigational Aids
each
QUANTITY
UNIT PRICE
TOTAL
1
1
750
2,500
3,500
4
$98,000
$30,000
$30
$35
$4
$2,000
$98,000
$30,000
$22,500
$87,500
$14,000
$8,000
Actual Cost
Fill Placement:
Mobilization
Filter Fabric
Bedding Stone
Limerock Riprap
Spoil Material
Capping Material
Silt Barrier
$260,000
lump sum
square yards
tons
tons
cubic yards
cubic yards
lump sum
1
600
100
375
9,000
6,000
1
Estimate Total (includes capping material)
Actual Cost (no capping material)
$73,975
$4
$25
$35
$18
$27
$30,000
$73,975
$2,400
$2,500
$13,125
$162,000
$162,000
$30,000
$446,000
$284,000
Seagrass Transplantation Cost (Actual)
$32,000
Total Cost For All Project Components
$576,000
Seagrass Transplanting and Monitoring
The area was planted in May–June 1994 using planting units of Thalassia, Syringodium, and
Halodule anchored with geo-textile staples on approximate 0.9-meter (3-foot) centers. The
planting units consisted of bare-root seagrass rhizomes with a minimum of three apical
meristems and minimum three culms behind the meristem. A total of 12,957 planting units
were installed over the area including 5,397 planting units of Halodule, 3,780 planting units of
Syringodium, and 3,780 planting units of Thalassia. The individual species were planted in species
plots with Halodule on the eastern side of the fill area, Syringodium in the center, and Thalassia on
the west.
The donor site was a large grassbed located south of the planting site in the basin between Julia
Tuttle Causeway and 79th Street Causeway. This grassbed contains areas where Halodule,
Syringodium, Thalassia grow in mixed beds and monoculture in sediment consisting primarily of
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Milano & Deis
calcium carbonate formed by Halimeda. Generally, it was easy to separate the plants from the
sediment. It was noted in the post-construction synopsis (Deis, 1994), that each species from
Halodule to Syringodium to Thalassia became progressively more difficult to acquire from the
donor site and to plant. Each plant produced more waste to develop a planting unit. It was
recommended that Halodule be the first choice for future plantings because of ease of collection,
less damage to the donor site, and quality of the planting unit providing an abundance of apical
meristems. Syringodium is difficult to plant because of the buoyancy of the leaves and rhizomes.
Thalassia should be planted only as sods. As discussed, the dredged material from the port used
to fill the site was not capped with aragonite as proposed. The result was sediment that was
coarse grained and rocky.
Miami-Dade DERM biologists surveyed the project site on June 20, 1995, approximately one
year after the planting (DERM, 1998). Survivability of planted units was measured only in the
Thalassia plot because the Syringodium and Halodule areas had coalesced such that individual
planted units were not visible. The survey found 64.8% of the planted Thalassia units survived
the first year.
Percent cover was measured using a 1-meter (3.3-foot) grid divided into 100 subunits. Table 2
provides the results of the survey. Halodule and Syringodium were found in all of the plots. The
Halodule and Syringodium plots each contained approximately 60% coverage.
Table 2. Percent cover of seagrass species by planted areas (plot) one year after planting at the north Biscayne Bay
site (from Miami-Dade DERM, 1998).
SEAGRASS SPECIES
THALASSIA PLOTS
SYRINGODIUM PLOTS
HALODULE PLOTS
Thalassia
Syringodium
Halodule
7.82
6.67
26.72
0
49.64
16.11
0
2.48
58.78
TOTAL
39.61
60.59
59.57
Red drift algae
Green algae
38.72
4.47
27.14
6.55
8.63
6.59
A qualitative survey of the site in June of 2002 found 30% patchy cover of all three species of
seagrasses on the site. We have no long-term quantitative data for the restoration site. As a result
of the rock groins and buffer areas for turbidity controls, the site remains an independent feature
in the bottom communities of this section of the bay. Other factors, e.g., the coarseness of the
material used to fill the site, may be contributing to patchiness in the seagrasses currently found
on the site. Long-term monitoring at a seagrass location within this basin of Biscayne Bay has
demonstrated a change in seagrass species from Syringodium to Halodule. This is sometimes
associated with changes in water quality within a location.
Seagrass Restoration and Management Opportunities in Biscayne Bay
During the early 1900s, more than 40% of north Biscayne Bay bottom communities, including
seagrass habitats, were dredged to provide fill for upland development. South Florida has
experienced tremendous population growth since then, and will continue to do so. As a result,
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Biscayne Bay Seagrass Restoration
Biscayne Bay environments face many challenges and threats to their present and future health.
Shallow seagrass beds are being degraded by recreational and commercial watercraft traffic.
Scarred and scoured seagrasses have been documented throughout the state of Florida, mostly
in shallow coastal waters less than 2 meters (6 feet) deep (Sargent et al. 1995). Experimental
techniques are being developed by governmental agencies to restore these vessel-related impacts
nationwide.
Seagrass planting has been generally more successful when restoration is conducted at sites
where seagrass communities existed, but were disturbed by physical impacts that can be
corrected or eliminated (Fonseca et al, 1998). The restoration of the structure of the seagrass
habitat is of primary importance. Seagrasses have been observed to naturally recruit into newly
restored coastal areas (Cape Florida State Park, North Virginia Key and the Chicken Key Bird
Rookery) in Biscayne Bay (Milano, 2001). Natural recolonization can occur vegetatively
(rhizome extension) or through seedling recruitment (Fonseca et al, 1998) if the proper
elevations for growth of the seagrass is achieved in the restoration. A number of seagrass
restoration opportunities, requiring the filling to natural contours, are being considered in north
Biscayne Bay. In addition, Biscayne National Park resource managers are presently initiating an
effort to catalogue all vessel-related impacts in south Biscayne Bay for potential future
restoration (R. Clark, Biscayne National Park, pers. comm.). Future south Biscayne Bay seagrass
restoration opportunities are limited to these shallow propeller scars and boat groundings.
Biscayne Bay seagrass restoration opportunities may be limited to the filling of previously
dredged areas in north Biscayne Bay to natural contours, and the restoration of marine vessel
propeller scars and boat groundings in shallow coastal waters. As illustrated in this review, the
project component costs (fill material, transportation, placement of fill, planting, project
monitoring, etc.) for seagrass restoration are dependent on site-specific environmental
conditions (water depth, currents, wave energy, etc.).
Sargent et al. (1995) have recommended that education is an essential part of any effort to make
all boaters understand the sensitive nature of shallow seagrass communities. Miami-Dade
County, South Florida Water Management District, and the Florida Fish and Wildlife
Conservation Commission have developed a boater’s guide with maps illustrating the location
of seagrass in Biscayne Bay. Additionally, on-going statewide boater education certification
programs should include information on seagrass protection.
In addition to the ongoing efforts to restore shallow coastal watercraft impacts and educate the
boating public, shallow water motorboat exclusion zones can be used as a management tool to
help protect and conserve seagrass habitats, provide manatee protection, and enhance boating
safety. Improved navigational signage is an additional tool that can be used to further these goals.
Inclusion of seagrass protection signage in appropriate conservation waters, such as critical
wildlife areas or national parks, would:
• Provide resource protection to shallow marine environments.
• Demonstrate effective methods of delineating sensitive marine communities.
• Provide an opportunity to develop effective enforcement and education strategies.
• Save long-term restoration and mitigation dollars.
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Milano & Deis
Exclusion zones, with the necessary enforcement, may be an effective management and natural
resource conservation tool to assist in delineating, protecting, and restoring shallow seagrass areas
from vessel related impacts.
This paper illustrates the high cost of seagrass restoration, especially in previously dredged areas.
Preservation of seagrasses is the most cost-effective approach. Preservation of south Florida
seagrasses can be accomplished through improved delivery and scheduling of freshwater run-off
to coastal areas, providing boater education programs, implementing resource protection zones,
and providing dedicated marine resource enforcement.
REFERENCES
Alleman RW, Bellmund SA, Black DW, Formati SE, Gove CA, Gulick LK. 1995. Biscayne Bay surface water
improvement and management plan. South Florida Water Management District, West Palm Beach, Florida.
Biscayne Bay Management Plan, Miami, Fl., Miami-Dade County Department of Environmental Resources
Management and the Miami-Dade County Planning Department, 1980.
Biscayne Bay Partnership Initiative Technical Document, 2001.
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Dial RS, Deis DR. 1986. Mitigation options for fish and wildlife resources affected by port and other waterdependent developments in Tampa Bay, Florida. U.S. Fish and Wildlife Service Biological Report 86(6). 150
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Fonseca MS, Kenworthy WJ, Thayer GW. 1998. Guidelines for the conservation and restoration of seagrasses in
the United States and adjacent waters. National National Marine Fisheries Service, Southeast Fisheries Science
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to 1976. Sea Grant Technical Bulletin. University of Miami, No. 40. 151 pp.
Idyll CP, Tabb DC, Yokel B. 1968. The value of estuaries to shrimp. Pp. 83 – 90 in Newsom JD, ed., Marsh and
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Lewis RR. 1990. Creation and restoration of coastal plain wetlands in Florida, Pp. 73–101 in Kusler JA, Kentula
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Lewis RR, Kruer CR, Hodgson AB. 1999. Seagrass Distribution in Biscayne National Park. Tech Rep. 99–388,
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13 pp+ maps.
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Milano GR. 2000. Island restoration and enhancement in Biscayne Bay, Florida. Pp.1–17 in P. Cannizarro PJ, ed.
Proceedings of the 26th Annual Conference on Ecosystem Restoration and Creation. Hillsborough
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Odum WE, McIvor CC, Smith TJ. 1982. The ecology of the mangroves of south Florida: A community profile.
U.S. Fish and Wildlife Service Office of Biological Services. FWS/OBS-81/24. Washington, DC.
Sargent FJ, Leary TJ, Crewz DW, Kruer CR. 1995. Scarring of Florida’s seagrasses: assessment and management
options. Florida Marine Research Institute Tech. Rep TR-1, St. Petersburg, Florida. 37 pp + appendices.
Wanless HR. 1969. Sediments of Biscayne Bay—distribution and depositional history. M.S. Thesis, University of
Miami, Coral Gables, Florida. 260p.
Wanless HR. 1976. Geologic setting and recent sediments of the Biscayne Bay region. P. 1-32, In: Thorhaug A,
Volker A., eds., Biscayne Bay : Past, Present, and Future. University of Miami, Sea Grant Special Report No.
5. 315 p.
Wanless HR, Cottrell D, Parkinson R, Burton E. 1984. Sources and circulation of turbidity, Biscayne Bay, FL. Final
report to Sea Grant and Dade County, 499 p.
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Simp. Intern. Lagunas Costeras. UNAM-UNESCO, pp. 495-502.
GRM (Miami-Dade Department of Environmental Resources Management [DERM], 33 SW. 2nd Avenue, Suite
1000, Miami, FL 33130-1540); DRD (PBS&J, 7785 Baymeadows Way, Suite 202, Jacksonville, FL 32256)
129
❖
TOPOGRAPHIC RESTORATION OF BOAT GROUNDING DAMAGE AT
THE LIGNUMVITAE SUBMERGED LAND MANAGEMENT AREA
P.L. McNeese, C.R. Kruer, W.J. Kenworthy,
A.C. Schwarzschild, P. Wells & J. Hobbs
ABSTRACT
This project involved topographic restoration of a 50-meter long eroding twin propeller scar on
a shallow seagrass flat within the Lignumvitae Key Submerged Land Management Area in the
Florida Keys. The primary ecological goal was to arrest continued erosion of the damage site. The
secondary goal was to initiate site recovery to the pre-damage condition of a Thalassia testudinumdominated seagrass community in soft carbonate mud. The project included filling of the scar
portion of the damage area with native limerock gravel material using a work barge conveyor belt
system, and installing bird roosting stakes for the purpose of natural fertilization. The work was
managed by the Florida Keys Environmental Restoration Trust Fund and utilized a marine
contractor, paid professional staff and in-kind resources. Actual dollar costs of the project and
estimated cost-equivalents for in-kind services are presented along with costs associated with
three years of monitoring and reporting. The monitoring data indicates that the gravel has
remained in place and the scar has stabilized. Calcareous and fleshy macroalgae have colonized
the gravel but seagrass recruitment has not occurred. The gravel will be capped with fine
sediment to accomplish the secondary goal of seagrass restoration.
INTRODUCTION
The Lignumvitae Key Submerged Land Management Area (LKMA) is located in
Islamorada, Florida Keys. This 10,000-acre management area consists of a series of shallow
seagrass banks bisected by meandering channels (Figure 1). The heavy boating activity and
predominance of a transient (visiting) boating population has made this area a focus of
boating impacts management through channel marking, limited-motor zoning and
education. Sargent et al. (1995) documented the magnitude and scope of the problem of
seagrass habitat loss from boating impacts throughout Florida and especially in the Florida
Keys.
When a vessel impacts the bottom the resulting injury may include any or all of the
following (Figure 2): propeller (prop) scars, grounding holes, and berms (Kenworthy et al.
2002). At LKMA vessel impacts are generally classified by location on the banks as either
interior injuries or edge injuries. Bank interiors are dominated by prop scar type injuries,
typically generated by small vessels. Bank edges experience more grounding injuries, often
from larger vessels, and are subject to continued erosion induced by tidal currents and the
wakes of passing vessels. These erosive forces have the effect of widening and deepening
existing injury footprints causing additional resource loss (Sargent et al. 1995; Whitfield et
al. 2002).
We were interested in applying practical restoration methods to address the initial damage
and subsequent erosion of bank edges at LKMA. Restoration of the pre-injury habitat is the
desired goal for all injury sites but for bank edge injuries continued peripheral damage
caused by site erosion must be arrested first.
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McNeese, Kruer, Kenworthy, Schwarzschild, Wells & Hobbs
Figure 1. Graphic depiction of Lignumvitae Key Submerged Land Management Area in Islamorada, Florida,
showing meandering channels and shallow seagrass flats (latter indicated by diagonal fill pattern).
Site Description
The project injury, located on Peterson Key Bank (Figure 2), consisted of a deep twin
propeller scar and blowhole within a habitat of mixed seagrass dominated by Thalassia
testudinum and Halodule wrighttii growing in soft carbonate mud. The original grounding
most likely occurred in 1993 (P.Wells, author’s observation). Within five years of the injury,
the twin propeller scar had eroded into a wider and deeper single scar. The blowhole had
not noticeably increased in size and had seagrass growing in the bottom of it. This project
involved restoration of the scar portion of the injury only.
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Topographic Restoration of Boat Grounding Damage
scar trench
NOAA
scar
berm
blowhole
Figure 2. Aerial photograph of the pre-restoration condition of the damage site (photograph taken December
1996) at Lignumvitae Key Submerged Land Management Area. The scar trench, blowhole and berm features
of the site can be distinguished. Also visible in this photograph are several other scars on the surrounding flat
including the NOAA Beaufort Laboratory study scar site.
The physical and biological characteristics of the injury site were used to provide a basis for
restoration design. Initial site data were collected on October 4, 1998. The injury consisted
of an eroded prop scar “trench” 52.7 m in length (Figure 2). The scar averaged 4.97 m in
width (n=6; range of 2.46 m to 6.67 m). The average depth over the entire scar (n=21) was
12 centimeters (cm) below the grade of the adjacent grassbed, while the centerline of the
scar averaged 30 cm in depth relative to adjacent natural grade. The bottom of the scar was
not symmetrical. On its south side, it was deeper, unvegetated, and had a steeper “wall.”
The north side was more gently sloped and had recruited with calcareous algae. Currents
were apparently being funneled more aggressively along the south wall of the scar.
Visible in the aerial photography of the site is a restored prop scar adjacent to the project
injury (Figure 2). This scar is the subject of an experiment being conducted by the National
Oceanic and Atmospheric Administration (NOAA) Beaufort Laboratory (Beaufort, N.C.)
initiated in 1995 to test the concept of restoration by “modified compressed succession”
(Fourqurean et al. 1995; Kenworthy et al. 2000).
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McNeese, Kruer, Kenworthy, Schwarzschild, Wells & Hobbs
Site Goals
The primary ecological goal was to arrest erosion of the scar through restoration of the
topographic profile and allow natural recovery of the substrate and the biotic community.
Although filling of dredged areas in the Keys has been performed successfully, filling of boat
grounding injuries had never been undertaken (C. Kruer, author’s observation). Based on
local knowledge and characteristics of this site, we hypothesized that filling of the injury
with a gravel material heavy enough to remain in place would restore the bank profile and
arrest erosion, thereby allowing the natural accumulation of fine sediments.
The primary management goal for this site was to explore the feasibility, cost-effectiveness
and repeat-ability of the stabilization technique used. The LKMA area contains dozens of
injury sites similar to the subject site and hundreds more are present in other shallow-water
areas of the Florida Keys. Restoration of all of these sites is not practical but restoration of
the most vulnerable sites may be feasible as part of a larger seagrass restoration program if:
• the construction method allows for quick mobilization and installation soon after
injury;
• the number of different stabilization/restoration treatments applied to a single site
could be kept to a minimum;
• the construction method is relatively simple and not extremely specialized;
• the construction method is cost-effective for use on a multiple-site basis; and
• the construction method could be easily bid, contracted, and supervised at the
LKMA staff level.
METHODS
In order to achieve the site restoration objectives, site planning and design were closely
coordinated with the method of construction. We involved a local marine contractor,
Adventure Environmental, Inc. of Key Largo, Florida to assist our team in site planning.
Due to the sensitivity of the resources surrounding the site conventional construction
equipment and methods could not be used. Several alternatives were considered to generate
the final design: placement of 1-inch to 1.25-inch (25 mm – 32 mm) diameter gravel size
to bring the scar feature topography to adjacent natural grade. We used a small barge
conveyer system specially designed and built by the contractor for this project. Based on the
calculated volume of the scar we estimated that 31.4 m3 or 41 cubic yards (cy) of material
was needed to fill it. We set a goal of filling the scar to within 0.25 m of adjacent grade.
A second design component was the addition of bird roosting stakes for the purpose of
encouraging fertilization of the scar substrate to enable compressed succession of the biotic
community to seagrasses. Fertilization with bird stakes has encouraged the rapid expansion
of H. wrightii in shallow prop scars elsewhere at LKMA (Kenworthy et al. 2000) including
directly adjacent to this site (Figure 2).
Construction
The work was scheduled for the period of highest tides during the last week of July 1999
to maximize the available water depth over the site. Turbidity curtains anchored with pvc
poles were installed on two sides of the trench at approximately 1 m to 1.5 m from the edge
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Topographic Restoration of Boat Grounding Damage
of the bare scar. Turbidity curtains were left in place throughout construction and were
inspected each morning before beginning work.
The fill material was washed native limerock gravel obtained from a local supplier and was
specified as 25 mm to 35 mm diameter. The supply company’s equivalent specification for
this size (1-inch to 1.25-inch) was called “Ballast Rock #4” but the actual rock delivered was
larger averaging 33 mm and ranging from 23 mm to 48 mm in diameter (n=57, sd=5.28
mm).
The gravel was deposited using a 33-foot barge fitted with two fiberglass sheets, attached
to a metal roller at the bow of the vessel, and stretched from bow to stern over a slick plastic
panel surface (Figure 3). The gravel material was loaded onto the barge at the LKMA dock
staging area with a small front-end loader. Approximately 0.86 m3 (1.12 cy) of material
could be accommodated in one barge load. The fully loaded barge had a draft of
approximately 12 inches. The material was transported approximately 1 mile from the
staging area to the site. The barge entered the scar from the east (channel) end, stopped at
the west end and then slowly backed out of the scar as the gravel was dumped. To deploy
the gravel, a worker positioned in the water used a hand crank to wind the fiberglass
sheeting around the shaft causing the gravel to drop off the vessel bow into the water
(Figure 3).
As the gravel was deposited in the scar, a second worker in the water distributed the material
evenly with hand tools. Quick settling of the rock during dumping made it apparent that
the original projected fill amount was underestimated and additional gravel was obtained
to finish the project. Upon completion of filling a total of 48 bird roosting stakes were set
on approximately 3 m centers in the scar (Figure 4). The finished bird stake array formed
three rows along the long axis of the scar and 16 rows across the scar. The stakes consisted
of 1.5-inch pvc set to extend approximately 20 cm above mean high water. No roosting
treatments were applied to the tops of the stakes (see Kenworthy et al. 2000) in an effort to
test a simple stake design. It was expected that cormorants and terns would roost on the
ends of the large-diameter pvc.
Monitoring
Monitoring was conducted at Time Zero (placement of fill), 1 year, 2 years and 3.5 years
after filling. The extent of monitoring was dictated by the availability of funding and
personnel (Table 1). Site conditions were documented upon completion of construction at
the Time Zero monitoring event on August 21, 1999. The depth of the fill surface was
measured using a meter tape stretched laterally across the width of the filled scar and
anchored into place on each side at the adjacent natural grade. A diver used the tape as a
guide to measure fill topography relative to natural grade. The vertical distance between
filled grade and the stretched tape was measured at 0.5 m intervals to the nearest 5 cm. This
“profile” measurement was performed every 5 m down the length of the filled scar for a
total of 10 transects. At the same time, the width of the fill surface area was recorded using
the stretched tape for each of the 10 transects. This fill surface profiling event was
performed to document “as-built” conditions and was not repeated in subsequent
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McNeese, Kruer, Kenworthy, Schwarzschild, Wells & Hobbs
monitoring. In order to document changes in the surface depth of gravel over time, a
permanent on-site measuring device was installed by tightly attaching orange plastic cable
ties 25 cm above the fill on each of the 48 bird stakes.
Figure 3. Barge conveyor system used to deploy gravel for scar restoration at Lignumvitae
Key Submerged Land Management Area.
Between completion of fill placement (August 1, 1999) and the time zero monitoring event
(August 21, 1999) strips of unvegetated sediment had appeared along each side of the scar
adjacent to the gravel fill. The thin layer of gravel deposited and spread to the edges of the
scar footprint proved insufficient to significantly retard current flow allowing erosion to
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Topographic Restoration of Boat Grounding Damage
continue along the sides of the scar (A. Schwarzschild, author’s observation). The widths
of the unvegetated strips, from the edge of the continuous seagrass bed to the edge of the
gravel, were recorded to the nearest cm at the 16 bird stake locations along each side of the
scar. Continuing erosion of these areas documented during Year 1 monitoring led us to
conclude that planting seagrass in the unvegetated strips might be warranted. The strips
were planted with H. wrightii in April 2001, 20 months from time zero (see “Planting”
section for description of methods).
Figure 4. Aerial photograph of the time zero condition of the restoration site (photograph taken
September 1999) at Lignumvitae Key Submerged Land Management Area. The scar feature is filled and
set with bird stakes.
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McNeese, Kruer, Kenworthy, Schwarzschild, Wells & Hobbs
Table 1. Monitoring of filled propeller scar site at Lignumvitae Key Submerged Land Management Area. For
each monitoring event, plus (+) signs indicate individual parameters monitored and minus (-) signs indicate
individual parameters not monitored. “Not applicable” (n/a) indicates that the parameter did not yet exist and
therefore could not be monitored during the applicable event.
PARAMETER MONITORED
Depth of gravel fill
Width of unvegetated side strips
Planting unit survival
Planting unit density
Bird roosting on stakes
Condition of gravel (qualitative)
Recruitment of gravel (quantitative)
Recruitment of gravel (qualitative)
MONITORING EVENT
Year 1
Year 2
Time Zero
+
+
n/a
n/a
+
+
n/a
n/a
+
+
n/a
n/a
+
+
+
+
+
+
+
+
+
+
+
Year 3.5
+
+
+
+
+
+
A general description of the condition and vegetative recruitment of the gravel in the scar
was recorded at time zero and at each subsequent monitoring event. A qualitative
characterization of the adjacent natural seagrass beds was also recorded. At the Year 1
monitoring event, presence/absence of vegetation and fouling organisms, identified to genus
when possible, was recorded in four 20 cm x 20 cm quadrat samples at 0.5 m distance
around each of the center row of bird stakes for a total of 64 samples. At the Year 2
monitoring event, qualitative coverage observations and genera present on the gravel were
recorded. For the Year 3.5 event, sampling of recruiting vegetation over the gravel was
quantified using the Braun-Blanquet (B-B) coverage abundance scale method (Fourqurean
et al. 2001). B-B values were recorded using a 0.25 m2 quadrat set at 15 stations along each
of two transects (n=30) located 0.5 m from each center stake. Five sites along each of two
adjacent control transects were also monitored for B-B values (n=10).
Bird use of the roosting stakes was recorded during the time zero monitoring event, during
each annual monitoring event, and at any intervening opportunity during additional site
visits. The total number of roosting individuals of each different species or group (e.g.,
terns, seagulls, etc.) were recorded as the site was approached.
Planting
Transplantation of H. wrightii into the site was performed on April 23, 2001. Planting
material was harvested from the adjacent experimental staked scar installed by NOAA
Beaufort Laboratory (see Kenworthy et al. 2000). Halodule wrightii bare root planting units
(PU) were made by bundling 15-25 shoots to a 15 cm long metal sod staple with papercoated biodegradable “garden” twist-ties (Fonseca et al. 1998, see specifically Figure 3.2,
notes (a), (b), (c), (d), (g), and (h) in that document).
A total of 242 PUs were installed in the unvegetated perimeter strips along each side of the
scar on 0.5 meter centers. Ninety-three (93) PUs were installed along the southern
unvegetated strip and 85 PUs along the northern edge. Sixty-four (64) PUs were also
installed in the gravel around the center row of stakes along the entire length of the scar.
Each of the 16 bird stakes in the center row had four PUs installed around it at 0.5 m from
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Topographic Restoration of Boat Grounding Damage
the stakes. Planting units in the gravel area were installed by two persons, one digging a
small hole in the gravel with a metal tree planting bar and the other immediately installing
the staple PU into the hole and gently letting the gravel fall back around it.
Planting units were checked at one week from installation and were all found to be
surviving, some having experienced minor effects of herbivory. PU survival was monitored
at Year 2 including presence/absence and shoot count taken for each surviving plant. Some
of the planting units had coalesced so monitoring of each PU was performed by placing a
20 cm x 20 cm quadrat over the approximate center of the PU. Shoot counts were made and
where shoot count density was very high (over 50 shoots) a 10 cm x 10 cm quadrat
subsample was counted. At Year 3.5, qualitative observations of PU survival were recorded
but no shoot counts were performed.
RESULTS
Construction
A total of 69.6 m3 (91 cy) of gravel was placed in the scar over a 2-week period. The gravel
was observed to settle and compact very quickly during construction probably due to the
underlying soft carbonate mud sediments. This phase required 114 hours over 11 days to
complete from mobilization to final demobilization and clean-up. A total of 81 barge loads
were transported to the site at the rate of 6 to 11 barge loads per day. The barge transported
an average of 1.02 m3 (1.34 cy) per load. Turbidity plumes were generated primarily by the
gravel dumping activity. Turbidity plumes tended to stay between the two side-configured
curtains especially on slack and low tides.
Time Zero
Time zero monitoring was conducted on August 21, 1999. The width of the finished gravelfilled trench ranged from 2.95 m to 7.55 m with an average of 5.13 m (n=16). The length
of the filled trench was 50 m. Gravel depths at Time Zero ranged from 0 cm to 35 cm below
the adjacent natural grade and averaged 9 cm below grade overall (n=180). The width of
the unvegetated strip north of the scar ranged from 0 m to 1.6 m and averaged 0.59 m
(n=16). On the south side the strip ranged from 0 m to 0.55 m and averaged 0.26 m.
Year 1
The difference between time zero and Year 1 orange cable tie (depth reference)
measurements for each individual station was compared (Table 2). The individual
differences averaged an overall increase (representing subsidence of the gravel) of 3.4 cm
over the entire scar (n=48). In the first year the width of the unvegetated strips had
increased dramatically. On the north side of the scar the strip ranged from 0.2 m to 6 m and
averaged 2.2 m (n=16). On the south side of the scar the strip ranged from 0.5 m wide to
11 m wide with an average width of 2.2 m (n=16). Within the first year algae of the genera
Batophora and Caulerpa had recruited throughout the gravel surface. Sponges and an
unidentified alga were also seen. Only five of the 64 stations sampled on the gravel did not
contain any recruitment of macrophytes or macroinvertebrates.
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McNeese, Kruer, Kenworthy, Schwarzschild, Wells & Hobbs
Table 2. Changes in surface topography of gravel material in the filled scar at Lignumvitae Key Submerged
Land Management Area. Measurements from fixed cable tie reference markers to gravel surface were taken
along the scar and averaged for the north, central and south row of bird stakes and then an overall average was
calculated. The figures in this table represent the differences in average measurements over the time period
shown. A positive number represents subsidence of the gravel surface (increased depth of measurement over
time) while a negative number represents a rise or accretion at the surface.
TIME PERIOD
AVERAGE CABLE TIE MEASUREMENT DIFFERENCE (cm)
Time Zero – Year 1
North
Central
South
3.8
1.6
4.8
Overall Average: 3.4
Year 1 – Year 2
North
Central
South
4.6
1.1
2.3
Overall Average: 2.7
Time Zero – Year 2
North
Central
South
8.4
2.6
7.2
Overall Average: 6.1
Year 2 – Year 3.5
North
Central
South
!1.6
!1.7
!0.6
Overall Average: !1.3
Time Zero – Year 3.5
North
Central
South
6.8
0.9
6.6
Overall Average: 4.8
Year 2
Continued gravel subsidence was evident from depth reference measurements averaging
2.7 cm from Year 1 to Year 2 and averaging 6.1 cm from time zero to Year 2. By Year 2, the
width of the unvegetated zones had apparently stabilized. On the north side of the scar the
width ranged from 0.15 m to 7.1 m and averaged 2.4 m (n=16). On the south side of the
scar the width ranged from 1.2 m wide to 24.4 m wide with an average width of 2.4 m
(n=16). Year 2 monitoring revealed recruitment of H. wrightii extending approximately 40
cm into the gravel scar from the adjacent NOAA Beaufort Laboratory study scar at the
northeast corner (Figure 5). Other vegetation recorded on the gravel surface throughout the
scar included macroalgae from the genus Penicillus, Halimeda, Caulerpa, and Laurencia. Of the
178 H. wrightii PUs planted in the unvegetated perimeter, 83 (50%) were still surviving after
4 months. Short-shoot counts ranged from 1 to 60 with an average count of 14 shoots per
PU. Most of the surviving PUs, a total of 50, were on the north side of the scar, even
though a lower number (85) had been originally installed there. All of the PUs in the gravel
substrate had failed.
Year 3.5
The difference between Year 2 and Year 3.5 depth reference measurements (Table 2)
averaged !1.3 cm over the entire scar (n=48) indicating stabilization and slight infilling in
the third year. The difference between time zero and Year 3.5 measurements averaged 4.8
cm (overall subsidence since time zero). The widths of the unvegetated strips were not
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Topographic Restoration of Boat Grounding Damage
measured in Year 3.5 because of an obvious change in the bank-wide vegetation pattern
evident from analysis of repeat aerial photography from 1996, 1998, 1999, 2000, and 2003.
A clear diminishment of vegetation density, especially seagrass, was evident across this
portion of Peterson Key bank and, relative to our restoration site, was migrating from the
east (channel) end to the west (blowhole) end. In fact, earlier aerial photography (1996
through 1999) shows a consistent dense grassbed around all but the very east end of the
injury site. The 2003 aerial photograph (Figure 6) shows a significant thinning out of
vegetation all around the site and around the adjacent NOAA Beaufort Laboratory study
scar. Groundtruthing confirmed both the thinning of vegetation and the loss of seagrass in
this area of Peterson Key Bank.
Figure 5. Ground photograph of filled scar site conditions at 1.5 years from time zero at Lignumvitae Key
Submerged Land Management Area. The scar feature is on the left hand side of the photograph and the
NOAA Beaufort Laboratory study scar is on the right hand side. Halodule wrightii from the NOAA scar can be
seen expanding into the filled scar in the foreground of the photograph.
Vegetation recruitment in the gravel was again dominated by algae including Batophora,
Laurencia and calcareous species. Table 3 shows the results of coverage-abundance scale
monitoring at Year 3.5 for algal cover in the scar and algal/seagrass cover on the adjacent
natural substrate. The scar had a heavy accumulation of drift algae during the Year 3.5
monitoring event (February 2003) and this pattern was confirmed as a very dark color
signature on 35-mm oblique aerial photography obtained one month earlier (Figure 6). The
H. wrightii observed at Year 2 in the northeast corner of the scar had receded back out and
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McNeese, Kruer, Kenworthy, Schwarzschild, Wells & Hobbs
only one transplanted PU survived into Year 3.5. These observations were consistent with
the concurrent thinning out of the natural surrounding seagrass beds adjacent to the scar.
Figure 6. Aerial photograph of Year 3.5 condition of the restoration site (photograph taken January 2003) at
Lignumvitae Key Submerged Land Management Area. The scar feature is covered with dense algae including
Laurencia drift algae. Thinning of seagrasses on the adjacent bank is also evident.
Bird Use of Stakes
Terns acclimated to the site immediately. On August 21, 1999 at the time zero monitoring
site visit, 6 terns were noted roosting on the newly set stakes. At the same time, doublecrested cormorants (Phalacrocorax auritus) were seen roosting on the pvc/wood-block stakes
in the adjacent NOAA Beaufort Laboratory study scar. Over the next 3.5 years of
monitoring, terns dominated at the site while cormorants used the stakes occasionally and
in lesser numbers. Over the 25 observations recorded through the Year 3.5 monitoring
event, an average of 51% of the stakes was occupied. The occupied stakes consisted of an
average of 92% terns and only 8% cormorants.
Costs
The work was managed by the Florida Keys Environmental Restoration Trust Fund
(FKERTF) and utilized a marine contractor, paid professional staff and in-kind services
provided by NOAA Beaufort Laboratory, University of Virginia, Florida Marine Research
Institute and LKMA park staff and equipment. Dollar costs expended by FKERTF were
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Topographic Restoration of Boat Grounding Damage
closely tracked while in-kind services have been estimated. Actual costs are reported
through the end of the Year 3.5 monitoring event (including data analysis) but do not
include compilation and analysis of combined monitoring. Costs are summarized in Table
4.
Table 3. Braun-Blanquet coverage abundance values for algae growing on the gravel substrate of the filled
scar at the Lignumvitae Key Submerged Land Management Area (LKMA) at the Year 3.5 monitoring event
(m = mean, sd – standard deviation). Thirty samples were obtained over the gravel fill and 5 control sample
sites were analyzed for algae and for seagrass. There was no seagrass in the gravel scar at Year 3.5.
GRAVEL TRANSECTS
NORTH
SOUTH
4
3
5
5
5
3
4
5
3
3
5
5
5
5
3
4
4
5
4
3
1
1
1
2
2
5
5
3
2
3
CONTROL MACROALGAE
NORTH
SOUTH
0.5
2
3
3
3
1
0.5
1
0.5
1
CONTROL SEAGRASS
NORTH
SOUTH
0
2
3
4
4
Total B-B macroalgae on gravel
m = 3.6
sd = 1.3
Total B-B macroalgae in controls
m = 1.5
sd = 1.1
Total B-B seagrass in controls
m = 2.5
sd = 1.2
2
3
2
2
3
The FKERTF is a not-for-profit trust fund held by Audubon of Florida dedicated solely to
habitat restoration and management in the Florida Keys, with a maximum operating
overhead at the time of this project of 12%. It was also able to obtain consulting and
technician services at about 50% to 65% lower than normal minimum fees for private sector
work. In-kind support of $12,690 comprised a significant amount (29.5%) of the base cost
of this project covering some vital functions such as oversight and monitoring.
The contract work for filling the scar was also relatively low, totaling about $150/barge load
of gravel handled, or about $112/ton. The contractor provided a lower rate mainly because
of his interest in testing the gravel offloading system he had developed specifically for this
type of work. He estimated after completion of the job that a fee of roughly $150/ton to
$200/ton would probably make future jobs cost effective for his operation. In summary, a
50% increase in the base cost of the project in 1999 would result in a total of $64,516.58.
This is considered a modest estimate of the true expected cost of such a project.
DISCUSSION
The target profile parameter of filling the scar to within 25 cm of the adjacent natural grade
was largely met. Our estimate of the amount of fill needed had been based on average field
measurements. Improved technology such as that being used by staff of the Florida Keys
National Marine Sanctuary (Kirsch et al. 2005) may allow for more accurate volume
143
McNeese, Kruer, Kenworthy, Schwarzschild, Wells & Hobbs
calculations in the future. The gravel compacted and stayed in place on site. The results of
repeat depth measurements show that the gravel did subside after initial filling but had
stabilized by the third year. The goal of arresting erosion of the scar has been accomplished
but the gravel is still visibly exposed and no significant natural sediment accumulation has
occurred over the scar’s surface. Based on experience with this site and with smaller
experimental sites conducted in the middle Florida Keys (J. Kenworthy, author’s
observation) we believe a smaller gravel size (“pea gravel” of about 0.25" diameter) would
work for these sites.
The expansion of the unvegetated strips along the sides of the gravel are surmised to be
from a combination of erosion and general decline of the adjacent natural seagrass bed. The
thinning of vegetation on this portion of Peterson Key Bank should continue to be
monitored and may significantly affect the recovery of seagrass at our project site over time.
This will allow evaluation of site success within the context of community-wide vegetation
changes (Fonseca et al. 1998).
With respect to bird stakes, terns contribute some nutrient value, but cormorants are
preferred since they are larger birds with a greater mass of excrement. The cormorants’
preference for the wood block roosting design recommended by Kenworthy et al. (2000)
was clearly demonstrated at our site since cormorants continued to use wood block roosts
in the adjacent NOAA study scar but were reluctant to use our “flat top” pvc roost design.
Therefore we do not recommend the flat pvc design used on this project as an alternative
to the wood block design.
Our management goal for the site was to explore the feasibility, cost-effectiveness and
repeat-ability of the stabilization technique used. For past Florida Keys restoration sites we
have found it useful to evaluate project costs in terms of fill material handled. The total
project cost of the filled scar at LKMA equates to about $512 per cubic yard of fill handled
or about $182 per square meter. Without seagrass transplantation and site monitoring the
filling portion of the job, including design, construction and oversight cost about $324 per
cubic yard or $115 per square meter. We found the filling method used to be practical and
appropriate for the special site conditions. Two key factors impact the cost of handling fill:
the size of the site and the hauling distance. The cost of construction generally increases
directly with decreasing site size, primarily because the contractor’s costs of mobilization
and commitment of equipment is higher for a small site. The cost may also increase directly
with increasing distance of fill hauling. We would recommend that costs be reduced by
combining multiple sites under one filling effort. This would decrease the mobilization and
oversight cost per individual site. Our project was basically a demonstration project
managed by a non-profit entity but still incurred a fairly high cost-equivalent. In order to
make future jobs cost-effective at LKMA, multiple site contracting is recommended.
Project Evaluation and Adaptive Management
We are now engaged in an adaptive management process of refinement of the secondary
project goal, that of recovery to seagrass. The challenge is to identify restoration techniques
that are practical from a cost-benefit standpoint by comparing the direct and indirect (e.g.,
144
Topographic Restoration of Boat Grounding Damage
managerial) costs in light of the projected and desired site recovery expectations. With the
current success of the site and with increasing use and practical application of “compressed
succession” techniques in the Florida Keys, especially at LKMA (see Kenworthy et al. 2000)
new techniques for speeding recovery of this site are available and relatively easy to
implement. A practical method for “capping” of the scar with biodegradable tubes of fine
sediment is currently being tested at sites in the Florida Keys, including LKMA (Kenworthy
et al. 2004) and is being applied to this site. The bird stakes have also been adapted to
incorporate the wood block roosting design described by Kenworthy et al. (2000). We have
shown thus far that stabilization and rapid vegetative colonization of these types of injuries
can be reasonably accomplished within the management environment regardless of
temporal seagrass habitat recovery expectations.
Table 4. Cash and in-kind costs associated with the scar filling project at the Lignumvitae Key Submerged
Land Management Area (LKMA). Actual dollar costs spanning the project period of 1998 through 2003 are
shown. Selected subcategories of the total cost in a major category are given in parentheses next to the
subcategory.
TASK (year expended)
CASH EXPENDITURE IN-KIND SERVICES
(cash equivalent)
Design (1998–1999)
Site investigations
Project coordination
Project design
Permitting
Project contracting
Construction (1999)
Gravel fill and
turbidity control
Project oversight, bird
stake installation and
time zero work
Seagrass transplantation
Monitoring (2000–2003)
Year 1
Year 2
Year 3.5
Project Totals
With FKERTF overhead
TOTAL
$4,256.63
$2,500
$6,756.63
$19,339.40
($12,144.49)
$3,300
$22,639.40
($4,805.00)
($2,650.00)
($2,389.91)
$6,725.02
($5,042.96)
($1,282.06)
($400.00)
$30,321.05
$33,959.58
($650.00)
$6,890.00
($1,300.00)
($500.00)
($5,090.00)
$12,690.00
$13,615.02
$43,011.05
$46,649.58
As a final note, we are concerned with the overall effect of apparent degradation of seagrass
in Peterson Key Bank on this project. As the restoration site continues to be modified and
monitored we must consider the probability that recovery to seagrass habitat may be up
against a natural process of bank-wide degradation. This continues to point to how little we
know of the behavior of these habitats over time and reminds us to keep the value of these
small restoration sites in the proper perspective. Restoration of these sites can play a
meaningful role, but prevention of damage must continue to be the first priority of any
seagrass protection program.
145
McNeese, Kruer, Kenworthy, Schwarzschild, Wells & Hobbs
LITERATURE CITED
Fonseca MS, Kenworthy WJ, Thayer GW. 1998. Guidelines for the Conservation and Restoration of Seagrasses
in the United States and Adjacent Waters. NOAA Coastal Ocean Program Decision Analysis Series No.
12. NOAA Coastal Ocean Office, Silver Spring, MD. 222 pp.
Fourqurean JW, Willsie A, Rose CD, Rutten LM. 2001. Spatial and temporal pattern in seagrass community
composition and productivity in South Florida. Mar. Biol. 138: 341-354.
Fourqurean, JW, Powell GVN, Kenworthy WJ, Zieman JC. 1995. The effects of long-term manipulation of
nutrient supply on competition between the seagrasses Thalassia testudinum and Halodule wrightii in Florida
Bay. Oikos 72:349-358.
Kenworthy WJ, Hammerstrom KK, Fonseca MS. 2004. Scientific evaluation of a sediment fill technique for
the restoration of motor vessel injuries in seagrass beds of the Florida Keys National Marine Sanctuary.
In: Development and Research Evaluation of Restoration Tools for Tropical Seagrasses in the United
States, Project report submitted to NOAA Damage Assessment Center and NOAA Office of National
Marine Sanctuaries and the Florida Keys National Marine Sanctuary, prepared by W. Judson Kenworthy.
45 pp.
Kenworthy WJ, Fonseca MS, Whitfield PE, Hammerstrom KK. 2002. Analysis of seagrass recovery in
experimental excavations and propeller scar disturbances in the Florida Keys National Marine Sanctuary.
J. Coastal Research 37:75-85.
Kenworthy WJ, Fonseca MS, Whitfield PE, Hammerstrom KK, Schwarzschild AC. 2000. A comparison of two
methods for enhancing the recovery of seagrasses into propeller scars: mechanical injection of a nutrient
and growth hormone solution vs. defecation by roosting seabirds. Final Report to the Florida Keys
Environmental Restoration Trust Fund. September 2000. 40 pp.
Kirsch KD, Barry KA, Fonseca MS, Whitfield PE, Meehan SR, Kenworthy WJ, Julius BE. 2005. The Mini-312
Program—An expedited damage assessment and restoration process for seagrasses in the Florida Keys
National Marine Sanctuary. J. Coastal Research S1(40):109–119.
Sargent FJ, Leary TJ, Crewz DW, Kruer CR. 1995. Scarring of Florida’s seagrasses: assessment and
management options. FMRI Tech. Rep. TR-1. Florida Marine Research Institute, St. Petersburg, Florida.
37 p. plus appendices.
Whitfield PE, Kenworthy WJ, Hammerstrom KK, Fonseca MS. 2002. The role of a hurricane in the expansion
of disturbances initiated by motor vessels on seagrass banks. J. Coastal Research, SI(37):86-99.
PLM: PO Box 450, Crystal River, FL 34423
146
CULTIVATION STUDIES OF THE HALOPHILA SEAGRASSES
H. JOHNSONII AND H. DECIPIENS
B. Baca, G. Stone & A. Sanchez-Gomez
ABSTRACT
Johnson’s seagrass, Halophila johnsonii, and paddle grass, H. decipiens, are the subject of ongoing
research because of their protected status, their rarity, and endangerment from coastal
development. The goal of this work was to determine the feasibility and best methods for
cultivation and transplantation, in order to protect and restore these species, and to mitigate for
their loss. H. johnsonii was cultivated in sixty aquaria beginning in the spring of 2001 and was first
planted in the field in fall of 2001. Cultivation in aquaria and the field continued through 2003
with successes in both cases, but with several failures in the field due to unstable (erosional)
substrates and siltation (field tests objectively compared all substrates and methods). H. decipiens
was collected from a local marina dredging project in October 2002 and was cultured in 60
aquaria plus tanks for reproduction and later replanting. Aquarium culture in winter 2002–2003
resulted in reductions in plant numbers, but this was anticipated based on previous winter field
and aquarium studies, and plant numbers increased by summer in excess of initial stocking
densities. A comprehensive, overwintering field study was also performed on both species in
2001–2002 and this showed that both species had reduced numbers in the winter months
(January and February). Braun-Blanquet ratings fell by ~50% and occurrence in quadrats fell by
~40%. These numbers returned to normal by spring of 2002.
INTRODUCTION
Johnson’s seagrass was placed on the State of Florida endangered species list (FGFWFC
1996) as a result of its rarity in coastal waters, and following the research of Eiseman (1980),
Eiseman and McMillan (1980) and Dawes et al. (1989). It is currently designated as federally
Threatened for the State (USFWS 1998). The localized species was separated from H.
decipiens which was found in deeper or more offshore waters. Culture work and field
observations indicated that it is a single sex species as no male flowers have been found. Its
transient nature results in its being absent during the colder months of the year. H. decipiens
Ostenfeld (paddle grass) is closely related to other members of the Hydrocharitaceae H.
johnsonii and H. englemannii (star grass) (Littler et al. 1989) by taxonomic features of leaves,
stems and flowers, and also by its small, delicate shape. It is a small, bright green seagrass
with pairs of minutely-toothed leaves arising from rhizome nodes. It grows to 5 cm (2 in)
tall, in water depths to 30 m (99 ft), and colonies can spread by rhizomes to form patches
to 1 m (3.3 ft) across (Littler et al. 1989). It has been noted that paddle grass is a pioneer and
rapid colonizer, making it ideal for planting projects (Fonseca et al. 1998, Josselyn et al.,
1986). The Halophila seagrasses have been the subject of very few culture studies, mostly the
work of McMillan (1976, 1978, 1980) and Fonseca et al. (1998).
METHODS
Aquarium Culture
H. johnsonii for aquarium culture work were first collected in spring 2001, from the Nova
Oceanographic Center marina entry channel (main donor site). They were installed in
groups of five within sixty 0.37-m2(2 ft × 2 ft), 114-L (30-gallon) glass aquaria (Fig. 1). The
system is flow-through, each aquarium having its own spray jet, with filtration by sand
filter. Sand filtration was not always used because of clogging problems. The bottom is
covered with approximately 3 cm (1.2 in) of clean beach sand. Photometer readings were
taken in the water above growing seagrasses and found to be approximately one half that of
ambient light; therefore, 50% rated shade cloth was used to cover the aquaria. H. johnsonii
was cultured in aquaria from spring 2001 through fall 2002.
147
Baca, Stone & Sanchez-Gomez
Figure 1. Seagrass aquaria used for cultivation of Johnson’s seagrass and paddle grass,
shown without heavy shade-cloth cover (inset shows cover).
H. decipiens plants for aquarium culture were collected in comparatively deep water (2.6 m,
8.5 ft) at Bahia Mar Marina in Ft. Lauderdale (Intracoastal Waterway). Plants were
transported in seawater to the aquaria where they were sorted into planting units (individual
sprigs with a pair of leaves). They were planted at 60 per aquarium (approx. 160/m2), and
with two layers of 50% rated shade cloth. Backup supplies (commonly from broken pieces
and leafless rhizomes left from collections) were cultured in 114-L (300-gal) plastic tanks.
Photometer readings were taken in the water above growing seagrasses and shade cloth was
used to approximate the light reduction in the aquaria (needed to keep down epiphyte
fouling). As light levels continued to drop over the winter only one layer of shade cloth was
used. In spring 2003 two layers of shade cloth were again used.
Near daily aquarium maintenance included vacuuming of algae, maintenance of water
inflow jets, sprinkling beach sand over plants, and scraping aquarium sides. Algae problems
were reduced with shade cloth and maintenance, and invertebrate and fish grazers were
controlled by periodic, heavy seawater flushing and stirring sediment. In cases where
seagrasses were completely decimated by grazers (direct observation), tanks were flushed
with fresh water followed by a 10-min soaking in 10% bleach and another freshwater flush.
These were then replanted with a backup supply located in the plastic tanks. Use of a sand
filter was also implemented in 2003 to prevent invasion by grazers.
Monthly water quality parameters collected were temperature, salinity, and light levels.
Salinity was measured with an AES® automatic temperature compensating refractometer,
calibrated to 0.0 parts per thousand (ppt) salinity with distilled water. Light was measured
with a submersible HoBo® light meter and data logger, and later with a GE® portable light
meter. Monthly counts and comparisons were made of all aquarium cultured material.
148
Cultivation Studies of Halophila spp.
Field Planting
The 2001/2002 field planting research used Johnson’s seagrass. The first main field planting
took place in July 2001 and these sites were monitored for several months and then reexamined one year later (June–July 2002).
Observations of cover and survival since initial plantings were made at 5 donor/recipient
sites having the following characteristics:
Site 1 — Low Energy / Moderate depth (1 m, 3 ft at dead low),
Site 2 — High Energy / Moderate depth (1 m, 3 ft at dead low),
Site 3 — High Energy / Deep water (2 m, 6 ft at dead low),
Site 4 — Low Energy / Shallow depth (0.3 m, 1 ft at low tide), and
Site 5 — Shallow depth (0.3 m, 1 ft at dead low)
For Johnson’s seagrass, planting units (pu) consist of one pair of leaves (one plant) with
associated rhizome and roots. These were collected by SCUBA or snorkel and transported
in water to the donor sites and the aquaria. Less than 10% of each donor bed was collected.
Planting units were placed in 3 replicates of ten each at each recipient site. Because of the
spreading and increased densities of some seagrass planting sites which occurred over the
one-year period from July 2001 to July 2002, planting sites were sampled for the present
study with 25-cm x 25-cm quadrats (625 cm2).
A test planting of paddle grass was conducted in November 2002 using aquarium-cultured
material but less than 5% of the plantings survived through the winter of 2002–2003,
presumably because of low temperatures, low light levels, and siltation from construction.
For this test planting, approximately 2,700 plants were removed from tanks at the Nova
Oceanographic Center (Fig. 1) which were previously collected in the north dredge area on
Oct. 17, 2002. Plants were removed from tanks using a hard-tine lawn rake. The plants were
transported in separate ice chests containing fresh seawater to the Bahia Mar site. Each
planting unit consisted of 5 plants (10 leaves) on one rhizome attached with an unfolded
large paper clip (Fig. 2). They were planted in marked 1-m2 quadrats. A total of 1,820 plants
(364 planting units) were installed in marked and unmarked quadrats. The remaining
unused plants (~25%) were transported back to NSU and replanted in a tank because the
planting units were too short or broken up.
Most plantings of paddle grass occurred in 2003. The first was in May 2003 at the north and
south sections of Bahia Mar Marina. Plants were planted in marked quadrats, along
transects, in three main channels of this marina. A total of 2,160 plants (432 planting units)
were installed in the north basin using the 5-plant pu with large, unfolded paper clips. The
unique aspect of this planting was the profusion of seeds (approximately 3/pu) in the
planting units. A total of 2,434 plants (406 planting units) were installed in the south basin
using a 6-plant pu (the idea that more plants per pu would increase success). The final
plantings took place in canals at four locations at Broward County parks. A total of 1,320
paddle grasses were planted, in 2003 June, in 5-plant units. Most planting areas were
selected on the basis of having Johnson’s seagrass growing upslope.
149
Baca, Stone & Sanchez-Gomez
Figure 2. Typical paddle grass planting unit: 5 plants (10 leaves) secured by unfolded paper
clip.
Field Studies
The primary field studies on H. decipiens and H. johnsonii were performed along the
Intracoastal Waterway in West Lake Park (Broward County Parks and Recreation Division)
in the vicinity of Port Everglades, Hollywood, Florida. The area is a shelf extending to the
edge of the channel. Seven reaches were spatially selected along the shoreline and three
transects were run at each reach, beginning from the shoreline and extending into deeper
water (approx. 10 ft, 3 m). On each transect, three 1-m2 quadrats were randomly placed and
permanently marked. Within each quadrat 16 sub-quadrats, each measuring 25 cm x 25 cm,
were located, and eight were sampled in detail. A total of 504 sub-quadrats were sampled
monthly, beginning in December 2001. Braun Blanquet (1985) and other cover measures
were collected.
RESULTS
Aquarium Studies
Results of the H. johnsonii aquarium studies are given in Table 1. As shown, plants decreased
in number in the winter but began increasing as the temperature increased. Overall survival
was about 100%, but survival and increase in low density culture (mean 31/tank or 84/m2)
was 155%. An examination of aquaria with poor survival showed “contamination” by
invertebrates (crustaceans and annelids) and these were believed to be the main source of
mortality (confirmed with later paddle grass studies).
Beginning in mid-July 2002, and seen in previous years, a dense growth of blue green algae
covered small seagrass species and any hard substrates in area waters. This also occurred in
the aquaria. This cover has not affected seagrass growth thus far in the aquaria and it is not
150
Cultivation Studies of Halophila spp.
known whether effects occur in the field. General observations are that this cover of algae
is not present in winter, spring, or early summer months.
Results of paddle grass aquarium cultivation are given in Table 2. As shown, plants
experienced a reduction in the winter 2002–2003, but recovery began taking place rapidly
by the spring. Counts do not include 2 large tanks which totaled over 700 total plants in
May 2003.
Table 1. H. johnsonii aquarium culture study results; planting date 2/15/02.
DENSITY
DATE
MEAN % SURVIVAL
Low (<40/tank or <107.6/m2)
2/15/02
3/21/02
7/24/02
100
75.6
154.6
Medium (40–49/tank or 107.6–131.9/m2)
2/15/02
3/21/02
7/24/02
100
58.8
98.1
High (>50/tank or 134.6/m2)
2/15/02
3/21/02
7/24/02
100
43.3
70.1
Mean survival = 99.1%
Table 2. Paddle grass aquarium results (percent survival) for n=60 aquaria, by sample dates, beginning with
the planting date.
DATE
10/1/02
12/17/02
1/20/03
3/20/03
4/28/03
5/21/03
3,600
2,132
1,704
999
2,039
3,615
#/AQUARIUM
60
36
28
17
34
60
% SURVIVAL
100
59
47
28
57
100
TOTAL #
Transplant Studies
Counts of H. johnsonii planting units, by site, at the beginning of the study (fall 2001) and
1 year later, are given Table 3. Over the year, total seagrass numbers increased 369%, from
450 plants installed to 1,659 remaining and recruited. However, Site 2 never grew seagrasses
since the first planting, and seagrasses died off at Sites 1 and 3. Field observations showed
that Site 1 experienced heavy sedimentation over the winter of 2001–2002 and sediment at
Sites 2 and 3 had completely washed away, leaving rock and gravel substrates. Sites 4 and
5 did very well, having spread to adjacent areas. Although the third replicate of a planting
location at Site 4 had no surviving plants, a fourth replicate taken at another planted location
of Site 4 planting site had a count of 496/m2.
151
Baca, Stone & Sanchez-Gomez
The main donor site near Site 1 was also censused and counts/m2 are given in Table 4. As
shown, the donor site was in good condition, averaging over 1,000 plants/m2 (1,000 pairs
of leaves).
Table 3. Results of H. johnsonii plantings over the one-year study (measures in #/m2).
SITE
1
2
3
4
5
DAYS AFTER
PLANTING
1
REPLICATE
2
3
TOTAL
0
361
0
361
0
361
0
361
0
361
30
0
30
0
30
0
30
112
30
288
30
0
30
0
30
0
30
203
30
656
30
0
30
0
30
0
30
0
30
400
90
0
90
0
90
0
90
315
90
1,344
TOTAL
400
859
400
1,659
Table 4. Census of H. johnsonii donor site population (#/m2) one year after plant removal.
REPLICATE 1
REPLICATE 2
REPLICATE 3
REPLICATE 4
672
1,792
1,440
928
Mean=1,208, SD=504
Paddle grass results are preliminary, but good growth was seen in a few locations at Bahia
Mar and at one park location (DeGroff Park). Best results occurred in high flushing areas
of the marina and in the highest current areas of the canals. Planting downslope of
Johnson’s seagrass was not the best approach because this species was found to invade and
dominate the paddle grass sites in the summer.
Field Studies
Seven months of data were analyzed (December 2001–July 2002), including some from the
coldest periods for south Florida. A summary of Braun-Blanquet (B-B) data is given in
Table 5. Transects sometimes contained other species of seagrass besides H. johnsonii: H.
decipiens (predominant) and Halodule wrightii (rare or uncommon shoal grass in the
transects). Therefore B-B collection data are shown with H. decipiens and rare H. wrightii
lumped together versus H. johnsonii alone. As shown in the B-B key, mean cover for H.
decipiens showed numerous shoots but with <5% cover (possibly a common occurrence in
small seagrass species). In contrast, H. johnsonii plots generally had a few shoots with very
little cover (B-B range 0.4–0.7).
Further comparisons were made between H. johnsonii and H. decipiens, using number of subquadrats containing each species (8 sub-quadrats per quadrat × 3 quadrats per transect =
152
Cultivation Studies of Halophila spp.
24 maximum per transect). As shown in Table 6, approximately 15% of the sub-quadrats
contained H. johnsonii in summer and 31% contained H. decipiens in summer. In contrast
to casual field observations, no trends were evident in seagrass occurrence relative to depths
(T1 being the shallowest and T3 being the deepest).
Table 5. Seagrass field study results, plant abundance B-B measures. Key to B-B measures: 0.1—solitary shoot
with small cover; 0.5—few shoots with small cover; 1.0—numerous shoots but less than 5% cover; 2.0—any
number of shoots but with 5–25% cover.
12/2001
1/2002
DATE
2/2002
4–5/2002
6–7/2002
H. decipiens (with rare
Halodule wrightii)
Mean B-B
SD
1.2
0.7
1.1
0.6
1.2
0.8
1.6
0.7
1.4
0.8
H. johnsonii only
Mean B-B
SD
0.4
0.6
0.4
0.5
0.5
0.8
0.7
0.8
0.6
0.7
SEAGRASS
Table 6. Seasonal occurrence (#subquadrats with Halophila seagrasses).
SEAGRASS
12/2001
1/2002
DATE
2/2002
4–5/2002
6–7/2002
H. johnsonii
Mean
SD
2.8
4.4
1.8
2.7
1.8
3.4
3.4
4.1
3.1
3.7
H. decipiens
Mean
SD
9.1
5.7
7.2
5.6
5.2
4.1
7.6
5.4
7.5
5.6
DISCUSSION AND CONCLUSIONS
The results of these studies were positive for the concept of holding pre-dredge Halophila
seagrasses in aquaria and tanks for transplantation later. However, besides further testing,
a number of modifications and guidelines for aquarium culture should be implemented, as
follows:
• Avoid collection and planting during the colder months (December–February)
• Use constant flow-through systems
• Use filtered water for flushing
• Sterilize sand before using it as a substrate
• Use a heated head tank to feed aquaria during coldest months
• Implement constant maintenance as described herein
Transplantation looks promising, with good results obtained for H. johnsonii. Results with
H. decipiens indicate potential for success, and its rapid growth in aquaria shows promise for
transplantation of this species.
153
Baca, Stone & Sanchez-Gomez
ACKNOWLEDGMENTS
The authors thank Florida Department of Transportation for providing the funds for the Johnson’s seagrass
work, and Boca Resorts, Inc. for funding the Bahia Mar Marina work.
LITERATURE CITED
Braun-Blanquet J. 1985. Plant Sociology: The study of plant communities. C.D. Fuller and H.S. Conrad (eds.,
transl.). Hafner, London.
Dawes CJ, Lobban CS, Tomasko DA. 1989. A comparison of the physiological ecology of the seagrasses
Halophila decipiens Ostenfeld and H. johnsonii Eiseman from Florida. Aquatic Botany 33, No. 1-2, pp.
149–154.
Eiseman NJ. 1980. An illustrated guide to the sea grasses of the Indian River region of Florida. Harbor Branch
Foundation, Inc., Tech Rep. No. 31.
Eiseman NJ, McMillan C. 1980. A new species of seagrass, Halophila johnsonii, from the Atlantic coast of
Florida. Aquatic Botany 9, No. 1, pp. 15–19.
FGFWFC. 1996. Florida’s endangered species, threatened species and species of special concern. Florida Game
and Fresh Water Fish Commission, Tallahassee, FL.
Fonseca MS, Kenworthy WJ, Thayer GW. 1998. Guidelines for the Conservation and Restoration of
Seagrasses of the United States and Adjacent Waters. NOAA Coastal Ocean Program Decision Analysis
Series No. 12. NOAA Costal Ocean Program, Silver Springs, MD, 222pp.
Josselyn M, Fonseca M, Nielsen T, Larson R. 1986. Biomass production and decomposition of a deep water
seagrass, Halophila decipiens Ostenf. Aquatic Botany 25, pp. 47–61.
Littler DC, M.M. Littler MM, Bucher KE, Norris JN. 1989. Marine Plants of the Caribbean. Smithsonian
Institution Press, Wash., D.C. 263pp.
McMillan C. 1976. Experimental studies on flowering and reproduction in seagrasses. Aquatic Botany 2, pp.
87–92.
McMillan C. 1978. Morphogeographic variation under controlled conditions in five seagrasses, Thalassis
testudinum, Halodule wrightii, Syringodium filiforme, Halophila engelmannii, and Zostera marina from Kenya.
Aquatic Botany 4, pp. 169–189.
McMillan C. 1980. Flowering under controlled conditions by Cymodocea serrulata, Halophila stipulacea,
Syringodium isoetifolium, Zostera capensis, and Thalassia hemprichii from Kenya. Aquatic Botany 8, pp. 323–336.
USFWS. 1998. U.S. Fish & Wildlife Service Endangered Species Home Page:
http://ecos.gov/species_profile/
BB (CSA South, Inc., Dania Beach, FL); GS (Bermello-Ajamil & Partners, Inc., Miami, FL); AS-G (Nova
Southeastern University Oceanographic Center, Dania Beach, FL)
Author contact e-mail: baca@csasouth.com
154
SUITABILITY OF ALTERNATIVE SAV MEASUREMENTS
AS AN INDICATOR OF WATER QUALITY EFFECTS
LOWER ST. JOHNS RIVER, FLORIDA
A.M. Steinmetz, D. Dobberfuhl & N. Trahan
ABSTRACT
The St. Johns River Water Management District has been monitoring submerged aquatic
vegetation (SAV) in the lower St. Johns River (LSJR) since 1995. Response of the SAV
community to water quality variability has been assessed using SAV percent cover and occurrence
measurements. However, percent cover and occurrence data did not represent field observed
changes within SAV beds. Geostatistical analysis was explored as an alternative measure of SAV
bed dynamics and response to water quality variability. SAV percent cover, occurrence, and
geostatistical estimates were compared to water quality parameters at two sites in the LSJR.
Results of the comparisons indicated that percent cover and occurrence reflected only large-scale
changes in the SAV community and were insufficient in representing small or short-term
changes. Geostatistical estimates characterized not only small alterations in the SAV community
at each of the sites, but also illustrated grassbed structure dynamics and sensitivity to other water
quality parameters not detected using percent cover and occurrence data. Geostatistical analysis
may be a more sensitive tool for measuring not only the response of SAV to water quality, but
also grassbed structure dynamics and potential impairment.
INTRODUCTION
The SAV monitoring program for the LSJR first began in 1995 by the St. Johns River Water
Management District. Since that time, SAV and associated water quality have been
monitored at a large number of sites across a range of spatial and temporal scales to provide
baseline data on the extent of SAV communities and their response to water quality changes.
The LSJR is a blackwater, tidal estuary dominated primarily by the freshwater macrophyte,
Vallisneria americana. Estimates from 1998 aerial photography indicate there were
approximately 760 ha of SAV in the LSJR. Because of the darkly-stained water, monitoring
protocols and SAV measurements within the LSJR are limited with the majority of the data
collected as percent cover and percent occurrence.
However, these metrics did not appear to adequately capture field-observed SAV changes
or reflect the potential effects of water quality variability. Often the collected data indicated
relatively stable conditions while anecdotal field observations were contradictory, suggesting
that discernible changes had occurred. In essence, the collected data were not capturing
subtle changes in SAV beds that were being interpreted from anecdotal observations as
either an impairment or improvement. This may reflect a shortfall of the sampling
methodology as others have found that broad spatial sampling can be misleading (Fonseca
et al. 2002). After examining the data, it became apparent that the cover and occurrence
measurements were not reflecting important changes in patch size and distribution of SAV
within the bed. To overcome this limitation, geostatistical analysis was conducted to
quantify bed structure in terms of patchiness and distribution of plants. Geostatistical
analysis has the advantage of quantifying the ecologically important patch dynamics that
have previously gone undetected. An additional objective was to relate geospatial parameters
to water quality changes. Finally, SAV measurements of percent cover, occurrence, and
geospatial parameters were compared for their respective sensitivity to water quality
parameters.
155
Steinmetz, Dobberfuhl & Trahan
METHODS
SAV data were collected at two sites in a freshwater region (Rice Creek North) of the river
and in an oligohaline region (Buckman Bridge; Fig. 1). Cover and occurrence data were
collected along five transects that extended to the outer edge of the SAV bed. Transects were
set twelve meters apart and placed perpendicular to the shore. Cover data were collected
using the terrestrial line intercept method. This method considers SAV to be present if root
and/or foliar cover intersect the transect tape. Percent cover was then estimated by dividing
the total meters of SAV by the total length of the SAV bed. Presence of more than one
species along a transect can result in greater than 100% coverage. Occurrence data were
collected according to an estimated percent occurrence cover class using a quarter-meter
quadrat at one-meter intervals along transects. Cover classes were 0%, 1–33%, 33– 66%, and
66–100%.
Figure 1. Location of the submerged aquatic vegetation (SAV) study sites in the lower St. Johns River.
Water quality data (i.e., nutrients, suspended solids, and field parameters) were collected
biweekly at each site using surface grab samples. Predicted light attenuation (Kd) was
calculated from the LSJR Optical Water Quality Model developed by Chuck Gallegos of the
Smithsonian Environmental Research Center (Gallegos 2002).
ESRI’s ARCGIS software (ArcMap & Geostatistical Analyst Extension) were used to
generate interpolated SAV surfaces of bed density including bed patchiness and structure.
156
Alternative SAV Measurements as Water Quality Indicators
Data points of percent occurrence from September 2001 through December 2002 were
interpolated using ordinary kriging by entering the midpoint of percent occurrence for each
cover class, measured at each meter along each of the five transects. A spherical model was
judged to be the best fit to the semivariogram. Anisotropy was enabled because it slightly
improved the fit and increased the root-mean-square-standardized errors (Fig. 2).
Geospatial parameters generated from the interpolations were used in stepwise regressions
to identify relationships between bed structure and water quality parameters. Geostatistics
used in the analyses included sill, partial sill, nugget, and minor and major range. Partial sill
and nugget are spatial and non-spatial components of bed patchiness (sill). Minor range
represents bed patch size perpendicular to shore and major range represents bed patch size
parallel to shore. An ANOVA was also performed to examine differences in bed structure
between sites.
Figure 2. Semivariogram without (left) and with (right) anisotropy. Root-mean-square standardized without
anisotropy: 0.9607. Root-mean-square standardized with anisotropy: 0.9771.
RESULTS
Changes in SAV percent cover at the Rice Creek North site appeared to be influenced by
Kd and color (Figs. 3 and 4). Relationships between changes in SAV cover and other water
quality parameters were not detected. A notable decline in SAV cover of approximately 29%
was apparent after October 2001 when both Kd and color increased. Correspondingly, an
increase in SAV cover did not occur until May 2002 through August 2002, when Kd and
color values declined. An immediate decline in SAV cover occurred again as Kd and color
values increased from September 2002 to November 2002. Declines in SAV cover (Fig. 3)
appeared to correspond well with significant SAV declines in the interpolated SAV surfaces
from the geostatistical estimates (Fig. 5) specifically during October 2001, November 2001
and March 2002. On the contrary, SAV cover did not appear to reflect significant increases
in SAV in April 2002 as shown by the interpolated SAV surfaces. However, SAV percent
cover data and the interpolated SAV surfaces were in agreement illustrating a large increase
in SAV in August 2002.
SAV percent occurrence at the Buckman Bridge site appeared to respond to fluctuations in
Kd and salinity (Figs. 6 and 7) during 2000 and 2001 when drought reduced river flow. As
157
Steinmetz, Dobberfuhl & Trahan
a result, there was a significant decline in SAV occurrence since the site was first monitored
in May 1999 (Fig. 5a). SAV began recovering from July 2001 to September 2001 with only
slight increases in subsequent months. However, SAV interpolated surfaces illustrated a
much larger increase in SAV from May 2002 through December 2002 that was not apparent
from the percent occurrence data (Fig. 8).
Rice Creek North - Percent SAV Cover
Mean SAV Percent Cover
160%
Error bars indicate SE.
140%
120%
100%
80%
60%
40%
20%
N
ov
D 00
ec
Ja 00
nFe 01
bM 01
ar
-0
Ap 1
r-0
M 1
ay
Ju 01
n0
Ju 1
l-0
Au 1
gSe 01
p0
O 1
ct
-0
N 1
ov
D 01
ec
Ja 01
nFe 02
bM 02
ar
-0
Ap 2
r-0
M 2
ay
Ju 02
n0
Ju 2
l-0
Au 2
gSe 02
p0
O 2
ct
-0
N 2
ov
D 02
ec
-0
2
0%
Date Monitored
Figure 3. Rice Creek North percent submerged aquatic vegetation (SAV) cover. Greater than 100% coverage
occurs when more than one species of SAV is present along a transect.
Rice Creek North - Color and Predicted Kd
Color
Predicted Kd
600
5
400
4
300
3
200
2
1
0
0
N
ov
-0
D 0
ec
-0
Ja 0
n0
Fe 1
b0
M 1
ar
-0
Ap 1
r-0
M 1
ay
-0
Ju 1
n0
Ju 1
l-0
Au 1
g0
Se 1
p0
O 1
ct
-0
N 1
ov
-0
D 1
ec
-0
Ja 1
n0
Fe 2
b0
M 2
ar
-0
Ap 2
rM 02
ay
-0
Ju 2
n0
Ju 2
l-0
Au 2
g0
Se 2
p0
O 2
ct
-0
N 2
ov
-0
D 2
ec
-0
2
100
Date
Figure 4. Rice Creek North color and predicted Kd.
158
Predicted Kd (m-1)
6
500
Color (cpu)
7
Alternative SAV Measurements as Water Quality Indicators
Figure 5. Rice Creek North submerged aquatic vegetation (SAV) interpolated surface from SAV percent
occurrence data points.
Buckman Bridge - Percent SAV Occurrence
Error bars indicate SE.
90%
80%
70%
60%
50%
40%
30%
20%
10%
0%
M
ay
-9
9
Ju
l-9
Se 9
p9
N 9
ov
-9
Ja 9
n0
M 0
ar
-0
M 0
ay
-0
0
Ju
l-0
Se 0
p0
N 0
ov
-0
Ja 0
n0
M 1
ar
-0
M 1
ay
-0
1
Ju
l-0
Se 1
p0
N 1
ov
-0
Ja 1
n0
M 2
ar
-0
M 2
ay
-0
2
Ju
l-0
Se 2
p0
N 2
ov
-0
2
Mean Percent SAV Occurrence
100%
Date Monitored
Figure 6. Buckman Bridge percent submerged aquatic vegetation (SAV) occurrence.
Other measured water quality parameters, although not related to percent cover or
occurrence, were detected in stepwise regressions using geospatial parameters (Table 1).
Bed patchiness (sill) was related to conductivity and total phosphorus at Rice Creek North,
dissolved inorganic nitrogen and dissolved inorganic phosphorus at Buckman Bridge, and
Kd at both sites. Patch size (range) was related to dissolved inorganic phosphorus and water
temperature at Buckman and dissolved inorganic nitrogen and Kd at Rice Creek North.
159
Steinmetz, Dobberfuhl & Trahan
Buckman Bridge - Salinity and Predicted Kd
6
5
20
4
15
3
10
n-
-0
ec
D
Ju
2
2
M
ay
-0
-0
ov
N
Ap
r-0
1
0
-0
ct
O
M
ar
-0
99
gAu
03
0
1
0
0
1
99
5
bFe
2
-1
25
Salinity (ppt)
7
Salinity (ppt)
Predicted Kd
Predicted Kd (m )
30
Date
Figure 7. Buckman Bridge salinity and predicted Kd.
Figure 8. Buckman Bridge submerged aquatic vegetation (SAV) interpolated surface from SAV percent
occurrence data points.
Comparing the two study sites using an ANOVA revealed different spatial characteristics
(Table 2). Nugget was significantly different between sites at the p <0.05 level while sill
and major range were significantly different at the p <0.10 level. The nugget, a measure of
localized variability unexplained by the semivariance, suggests that small-scale variation
differs between sites (Table 2). While the geospatial analyses did not detect the influence of
salinity (Table 1), percent occurrence data suggests that it may be a controlling factor in the
oligohaline reach (i.e., Buckman Bridge site) during low-flow conditions.
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Alternative SAV Measurements as Water Quality Indicators
Table 1. Results of regression equations from stepwise regression analyses between semivariogram parameters
and water quality variables (dissolved ammonium [NH4-D], total Kjeldahl nitrogen [TKN-T], total Kjeldahl
nitrogen dissolbed [TKN-D], conductivity [COND], total phosphorus [TP-T], light attenuation [Kd],
dissolved nitrate and nitrite [Nox-D], dissolved orthosphosphate [PO4-D], water temperature [WTEMP].
Partial sill – spatial patchiness; nugget – nonspatial patchiness; sill – spatial and nonspatial patchiness; marjor
range – patch size parallel to shore; minor range – patch size perpendicular to shore.
RICE CREEK
Partial sill
Sill
Nugget
Major range
Minor range
= –76 – 3477 NH4-D + 433 TKN-T
= –775 + 1397 COND + 2560 TP-T + 174 Kd
= 574 – 363 TKN-D + 3581 NH4-D
= 105 + 970 NH4-D – 27.4 Kd
= 68.7 + 809 NH4-D – 19.2 Kd
r2 = 75.2%
r2 = 76.6%
r2 = 77.5%
r2 = 87.2%
r2 = 88.3%
p< 0.0001
p< 0.0001
p< 0.0001
p< 0.0001
p< 0.0001
r2 = 80.6%
r2 = 75.6%
r2 = 61.6%
r2 = 57.0%
r2 = 64.8%
p< 0.0001
p = 0.001
p = 0.008
p = 0.001
p = 0.001
BUCKMAN BRIDGE
Partial sill
Sill
Nugget
Major range
Minor range
= 594 + 0.669 COLOR – 927 Nox-D – 16.3 Kd
= 672 – 867 Nox-D +4946 PO4-D – 18.7 Kd
= 1887 – 193 TKN-T + 2051 PO4-D – 200 pH
= 4.94 + 0.558 WTEMP
= –3.47 +71.7 PO4-D + 0.536 WTEMP
Table 2. Results of an analysis of variance of semivariogram parameters between the two study sites.
PARAMETER
Partial sill
Sill
Nugget
Major range
Minor range
RICE CREEK
MEAN
BUCKMAN BRIDGE
MEAN
F
P VALUE
421.0
713.7
292.7
30.6
20.2
390.5
605.5
215.0
17.5
12.7
0.33
3.17
4.63
3.28
1.60
0.57
0.08
0.04
0.08
0.22
DISCUSSION
It was somewhat surprising that cover and occurrence data did not reflect the magnitude of
changes observed in the field because they are inherently spatial. This could be indicative
of a problem with the scale of measurement versus process. However, patch size and
patchiness could not be calculated from transect data without the use of geostatistical
procedures. This result may be, in part, related to different distributions of patch sizes, patch
orientation, and proportional patch size changes and how our current transects capture, or
fail to capture, these features. For example, the interpolated surfaces at the Rice Creek
North site demonstrated a considerable expansion of the grassbed from March to April
2002, whereas coverage appeared to only increase by approximately 4%. From our field
observations, a lesser number of larger patches generally appeared to be more robust, with
respect to invertebrate colonization and plant condition, than smaller, fragmented patches.
Indeed, larger patches likely provide greater ecological benefits in terms of resource and
habitat. (West and King 1996, Jeppesen et al. 1998).
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Steinmetz, Dobberfuhl & Trahan
Clearly, areal losses would be occurring that were not reflected by the data. Unfortunately,
much of the ecologically relevant small-scale changes in a bed undergoing early stages of
degradation will go undetected when using our current cover and occurrence
measurements. Generally, an indicator of bed condition is not particularly useful if the
resource has to essentially disappear before the data indicate a problem.
SAV cover and occurrence data appeared to mirror large-scale changes and response to water
quality conditions well, but did not appear to offer sufficient precision to detect small or
short-term changes in the SAV community. For instance, color at the Rice Creek North site
had returned to lower levels from April 2002 to August 2002, but the corresponding cover
data showed only slight increases in coverage during this time. Conversely, the
corresponding interpolated SAV surfaces illustrated a significant degree of bed expansion
and density increases during the same period. Similarly, interpolated SAV surfaces at the
Buckman Bridge site also showed a measurable change and expansion within the bed in the
months after December 2001 that were not reflected in the occurrence data during periods
of decreased salinity. Thus, geospatial parameters appear to better reflect small temporal and
spatial scale changes at the patch level.
The geostatistical estimates appeared to be a better indicator of both small and large-scale
changes in SAV at the bed-level, in addition to providing information on bed structure
dynamics. The use of geospatial parameters as a measurement of bed condition appears
promising. Changes in water quality variables related to nutrient and light availability were
significantly correlated to bed structure at these two sites (Table 1) suggesting that the beds
are showing a greater response to water quality changes than has been previously seen. It
should be noted that this result is somewhat preliminary as the regressions were based on
only a 16 month time series and do not encompass the majority of the available SAV data.
Additionally, relationships with water quality emerged that were not detected when using
cover or occurrence data, despite a longer period of record. Therefore, geostatistical analysis
also appears to be a more sensitive technique to monitor SAV condition in the river. This
spatial approach can be easily adapted to most existing data sets for both freshwater and
coastal systems.
Geospatial parameters also reveal differences in patch structure between the two study sites.
Comparison of Rice Creek and Buckman Bridge sites shows significant (p <0.10)
differences in patchiness (sill) and patch size (range; Table 2). Rice Creek has a larger patch
size but is also considerably patchier than is Buckman Bridge. Both of these sites are
generally dominated by V. americana but the Buckman site experiences higher and more
frequent exposure to elevated salinities while Rice Creek has more darkly stained water. It
remains to be seen how strongly these water quality differences, as well as differences in
local environmental characteristics, influence patch dynamics. However, the implication is
that relatively strong differences exist between sites with respect to grassbed morphology
and response to water quality changes. These site-specific responses will have to be further
evaluated before water quality and SAV restoration targets are established.
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Alternative SAV Measurements as Water Quality Indicators
LITERATURE CITED
Fonseca, M, Whitefield PE, Kelly NM, Bell SS. 2002. Modeling seagrass landscape pattern and associated
ecological attributes. Ecol. Appl. 12: 218–237.
Gallegos CL. 2002. Development of an Optical Water Quality Model for the Lower St. Johns River —Final
Report for the St. Johns River Water Management District. Smithsonian Environmental Research Center,
Edgewater, Maryland.
Jeppesen E, Lauridsen TL, Kairesalo T, Perrow MR. 1998. Impact of submerged macrophytes on fishzooplankton interactions in lakes. In Jeppesen E, Sondergaard M, Christoffersen (eds.), The Structuring
Role of Submerged Macrophytes in Lakes. Ecological Studies 131: 91–114.
West RJ, King RJ. 1996. Marine, brackish, and fish communities in the vegetated and bare shallows of an
Australian coastal river. Estuaries 19: 31–41.
AMS (BCI Engineers & Scientists, Inc., on-site consultant to St. Johns River Water Management District, PO
Box 1429, Palatka, FL 32178-1429: asteinmetz@sjrwmd.com); DD (St. Johns River Water Management
District, PO Box 1429, Palatka, FL 32178-1429); NT (Jones, Edmunds, & Associates, on-site consultant to St.
Johns River Water Management District)
163
❖
IMPLEMENTATION, REGULATORY AND RESEARCH PRIORITIES
FOR SEAGRASS RESTORATION:
RESULTS FROM THE SEAGRASS RESTORATION WORKSHOP,
MARCH 11–12, 2003
H. Greening
A total of 260 scientists, resource managers and regulators attended a Seagrass and
Submerged Habitat Restoration Workshop, jointly convened by the four Florida National
Estuary Programs and held at Mote Marine Laboratory in Sarasota, Florida on March 11–13,
2003. Presentations from more than 70 scientists, engineers and resource managers
experienced in the restoration of estuarine and coastal seagrass, hard bottom and tidal creek
habitats, impacts from dredging and prop scars, and artificial reefs were included in the
three-day workshop. Participants represented a wide variety of interests and entities, as
shown here (percentage of the total number of registered participants):
Academic/laboratories
Local governments
State agencies
Federal agencies
Private sector
NGOs
12%
12%
29%
15%
23%
23%
The papers in this volume were presented at The Seagrass Restoration Workshop, held on
March 11th and 12th. Several overall messages emerged from the presentations and
discussions during the seagrass session, including the following:
1. Adequate site location is critical for successful seagrass planting. If seagrasses are
absent from an area, it is crucial to determine why they are not currently there
before planting.
2. Seagrass planting in Florida can be a viable restoration option for small-scale
restoration or mitigation (such as in prop scars or following damage from
groundings), given care and appropriate site location.
3. Most effective large-scale seagrass restoration in Florida (such as needed to reach
long-term seagrass management and restoration goals for estuaries) has been
accomplished through methods other than planting, such as water quality
improvements or wave energy reductions.
In addition to presentations and posters, participants were asked to identify and prioritize
critical gaps in our current understanding of seagrass restoration, and in the implementation
of seagrass restoration techniques. Top-ranked issues identified by the participants fell into
three categories (implementation issues, regulatory issues, and research needs). A full list
of issues and the number of “votes” each received is shown in Table 1. The top-ranked
issues identified in each of these categories included the following:
165
Greening
SEAGRASS:
Implementation Issues
• Standardized monitoring methods
• Coordination of water quality and SAV restoration efforts
• Effective methods for seagrass planting in disturbed high energy zones
• Long-term monitoring of restored areas and control sites
• Natural recovery vs. planting- what are the differences, benefits, issues of each?
• Cost-effective ways to generate transplants without using donor beds
Regulatory Issues
• Long-term monitoring to determine biotic community restoration, not just seagrass
survival
• Consistent standards for success criteria
• Mitigation ratios: how much is enough?
Research/Data Needs
• Determine if natural recruitment is more efficient than planting to restore seagrass
systems
• On-line access to seagrass restoration/transplant projects in gray, older literature,
permitting and monitoring reports
• Seasonal tracking of natural beds
• Identify critical water quality levels for seagrass recolonization and sustainable
populations
Clearly, monitoring the effectiveness of seagrass restoration and planting is considered a top
priority, as well as determining the relative efficiency of planting versus other restoration
methods such as water quality improvement and natural recruitment. The Tampa Bay
Estuary Program will be working with the other conveners and participants in the workshop
to address priority issues over the next several years.
Table 1. Identified issues and rankings from participants in the Seagrass Restoration Workshop. Workshop
participants were asked to identify issues during the workshop. Participants then “voted” on priority issues
(each participant was allotted ten votes) from the collated list. Total number of votes received for each issue
is indicated in brackets.
IMPLEMENTATION ISSUES
[49] Standardized monitoring methods
[46] Coordination of water quality and SAV restoration efforts
[45] Effective methods for seagrass planting in disturbed high energy zones
[44] Long-term monitoring of restored areas and control sites
[39] Natural recovery vs. planting- what are the differences, benefits, issues of each?
[38] Cost-effective ways to generate transplants without using donor beds
[27] Cost-effective ways to control stormwater as prevention
[14] Effective control of bioturbation
[13] Plant material sources
[12] Determine conditions under which it works to harvest seeds and broadcast directly
[12] Enhanced local site selection criteria
166
Priorities: Results from the Workshop
(Table 1 continued)
[11] Offshore bar restoration as a seagrass restoration technique
[10] Improved low cost restoration methods
[10] Cost per acre to replace seagrass—total costs, not just construction
[10] Determination of best methodology for grass planting
[8]
Site selection criteria
[7]
Successful and economic mapping methods
[6]
Funding for implementation
[4]
Site selection: classification of zones by potential impact or climate change
[2]
Evaluation of prop scar planting methodologies
[1]
R&D costs are high for new planting methods
[1]
Cost effective methods for seagrass planting- comparison
REGULATORY ISSUES
[50] Long-term monitoring to determine full biotic community restoration, not just seagrass
survival
[44] Consistent standards for success criteria
[34] Mitigation ratios: how much is enough?
[26] Enforce protections of newly restored areas
[26] Quantify secondary impacts to seagrass as a result of dredging impacts
[27] Interagency coordination in permitting
[20] How to determine when functional losses have been replaced
[18] Defined and standardized regulatory success criteria
[16] Rule-making, enforcement increased in aquatic preserves
[10] How does the scientific method (controls, etc) fit into the regulatory structure of permitting
[10] How to assess temporal lag in success criteria
[9]
Evaluate passive recovery as an acceptable strategy for regulatory requirements
[4]
Effects of Minimum Flows and Levels on SAV extent
TAMPA BAY
[11] Impacts of Piney Point discharge on seagrasses in Bishop Harbor and Lower Tampa Bay
[7]
Develop hypotheses relating loss of seagrass in Tampa Bay to water quality changes during
El Nino; test with existing data throughout Tampa Bay
[6]
Use site-suitability index approach to select transplant sites in Tampa Bay
[2]
Develop limiting criteria for successful restoration applicable to Tampa Bay
RESEARCH/DATA NEEDS
[42] Decision matrix or model to determine if natural recruitment is more efficient than planting
to restore seagrass systems
[40] On-line access to seagrass restoration/transplant projects in gray, older literature, permitting
and monitoring reports
[36] Seasonal tracking of natural beds
[35] Identify critical water quality levels for seagrass recolonization and sustainable populations
[33] Effect of epiphytes on seagrass restoration
[32] Role and interference of macroalgae in seagrass systems
[29] Physical changes: impacts on seagrass expansion- berms, bars breakwaters
[23] Identify sediment characteristics critical for seagrass recruitment
[19] More research on Syringodium and Halophila spp.
[18] Biology of sexual/asexual reproduction and colonization
[17] Habitat value of edge (importance of heterogeneity)
[17] Does “compressed succession” actually occur in restoration sites?
[16] Economic value of seagrass habitat
[12] Funds available for restoration- how to find funds for baseline distribution
[12] Effect of “anthropogenic” -derived wave energy
167
Greening
(Table 1 concluded)
[13] Seed ecology issues
[13] Genetic information for recovering beds- clones or not?
[11] Improved forecasting of seagrass population ecology
[11] Resilience of restored sites to natural disturbance
[10] Economic valuation system that properly considers and values seagrasses
[8]
“Micro” (patch-size) dynamics, sedimentation, and nutrient availability
[7]
Seagrass species/restoration—success evaluation
[7]
Identify differences between inshore and offshore water quality
[5]
Better, cheaper propagation methods
[5]
Evaluation of lag time between planting and growth of transplants
[5]
Comparable techniques comparisons, at same sites and times
[5]
Understanding the natural variability and natural recruitment processes
[4]
Limiting factors to seagrass restoration success
[4]
Lack of flowering and seed production in Halodule in Florida
[3]
Species shifts as water quality improves and coverage increases
[2]
Offsetting effects of SAV recruitment in light of wetland loss
[2]
Historical distribution data
[2]
Third-party review of all attemps at seagrass restoration
[2]
Water quality levels affecting different species
[2]
Potential for lab growth of seeds of species not typically sexually reproductive (i.e., Thalassia)
[1]
Recovery time when injured by fuel/oil pollution
[1]
Impact of over-abundance of grasses
[1]
Document reproductive effects in distribution
[1]
Optical model development
HG (Tampa Bay Estuary Program, 100 8th Ave SE, St. Petersburg, FL 33701)
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WRAP-UP OF SEAGRASS RESTORATION:
SUCCESS, FAILURE AND LESSONS ABOUT THE COSTS OF BOTH
M.S. Fonseca
INTRODUCTION
My goal in introductory and wrap-up talks was to address the topic of “Lessons Learned.”
To meet this goal, I have broken the discussion down into six topical areas that characterize
what I believe are the areas of seagrass restoration that need further discussion and clarity:
•
•
•
•
•
•
Planting methods
Site selection
Seagrass function
Monitoring
Cost of restoration
Future of SAV restoration
These will be addressed below, but first I wanted to make a few observations regarding the
presentations in the workshop.
Many of the examples given in the Workshop were derived from eelgrass (Zostera marina L.)
communities. If a way could be found to introduce eelgrass to Tampa Bay, many of the
problems that vex restoration of subtropical seagrasses would be overcome. That is because
with eelgrass, each shoot is an apical, meaning that the transplantation of any individual
shoot carries the explicit potential for vegetative colonization of the bottom, whereas with
all the Floridian species, that is not the case. Eelgrass is also both a colonizer and climax
species. Although all the Floridian species can colonize and maintain monospecific beds,
there is often a species succession, particularly in response to disturbance. This characteristic
has prompted the adoption of “compressed succession” (to my knowledge attributed to M.
Moffler) in implementation of restoration projects where colonizing species, such as
Halodule wrightii or Syringodium filiforme are introduced first with later additional Thalassia
testudinum in an attempt to shorten the time to the establishment of competitive dominant
(T. testudinum). Eelgrass also has a high rate of vegetative colonization that compares
favorably with the subtropical colonizers (H. wrightii and S. filiforme) and can be restored
using both whole plants and seed. Moreover, this introduction would undoubtedly provide
an economic boom as northern scientists flocked to warm water in the winter to continue
their studies. Unfortunately, with global warming it is unlikely that the distribution of
eelgrass will spread south to cure the ills of Floridian seagrass restoration.
By way of a reality check, I also wanted to comment on the unrealistic level of expectation
associated with seagrass restoration. These expectations are often most out of line with
reality for those who have only recently entered into resource management positions as
those persons often lack the experience of dealing with these projects and have not received
appropriate training.
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Expectations for restoration success are far higher than are warranted by the track record.
We conducted a survey in 1998 of all the extant projects that we could find at the time and
found that successful establishment of seagrass cover occurred in ~50% of the projects.
Thus, we are faced with a classic “half-full, half empty” choice. Another, perhaps more
poignant example of reality may be gained by comparing seagrass restoration probability
with marketplace speculation on crop futures. During the year, investors speculate on the
outcome of large agricultural commodities. Despite collective millennia of terrestrial
agricultural practice, on fields where we can exert controls over a wide variety of ecological
factors, it is still a profitable exercise to speculate on success. Given the comparative lack of
controls that we can exert on wild, open systems—which characterize seagrass restoration
efforts—we cannot come anywhere close to guaranteeing success, and the 50% begins, at
least to my eyes, to look like the glass is half full, not half empty (i.e., we are doing pretty
well, but do not expect or believe in miracles).
Finally, and most important to my mind, was the low level of scholarship that often
accompanies the conceptualization of restoration projects. While that was NOT evident so
much at a workshop like this one, in practice it has been my observation that this is too
often the case. Little homework is done before restoration projects are concocted, and often
by personnel that know little about the ecosystem. Findings are not published in easily
retrievable venues. This combination of poor homework and avoidance of the vetting role
of peer review has produced a tremendous amount of low quality and redundant work.
PLANTING METHODS
A frequent absence of the scientific method (e.g., Popperian hypothesis testing) has often
stalled or misled the field. The absence of controls in association with planting projects
remains a tremendous liability— one so obvious that often we do not see it or implement
it (e.g., I neglected to voice a strong need for it in our recent Guidelines for Mitigation and
Restoration of Seagrass in the United States and Adjacent Waters [Fonseca et al. 1998]). I asked the
workshop attendees to examine the various presentations for their adherence to the basics
of the scientific method. In conversations with workshop participants there appeared to be
consensus that while some were lacking, many displayed appropriate attention to this area
of concern—a critical step forward for assessment and implementation of planting methods.
As for planting methods themselves, there were a tremendous variety of methods discussed
and frankly, most of them work. What works for an individual practitioner, together with
the skill level of the workforce, is key and we saw papers in the workshop where labor
organization was clearly a key to success. The current methods range from passive methods
(Clark’s drift-capture approach) to shovels with holes to heavily mechanized (Anderson’s
planting machines), with traditional hand-planting methods in abundance.
Again, I advise that any method should carefully utilize the growth strategy of the plant
(e.g., make sure rhizome apicals with sufficient associated short shoots and rhizome are
utilized in planting units as previously demonstrated). More work is needed on seeding
methods nationwide, particularly with respect to questions regarding what are the limits of
techniques for a given species and how can these be used to achieve large scale restoration
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(e.g. Orth’s eelgrass seeding in Chesapeake Bay). I strongly caution the workshop
participants to apply the scientific method to new techniques—to require rigorous testing
and monitoring—and remember that a method is “guilty until proven innocent”—in other
words, should be assumed to NOT work until proven by statistically valid testing to be
otherwise. Large projects have been undertaken with unproven technology and failed as a
result of insufficient testing of the methods before project-level application.
Nonetheless, I maintain that most of the issues associated with methodological
consideration are actually market issues. Planting methods that do not establish a track
record of success will be used only for a short time before their history catches up with
them. Therefore, introduction and testing of new methods is a valuable market exercise.
Particular concern should exist regarding the use of untested new methods in permits and
restoration projects when public funds are used; such actions should be correctly termed
“research” and subject to the timely and full scrutiny of peer review panels before setting
of project goals, responsibilities, and logistics.
Finally, the use of test plots was shown throughout the workshop to be a vital step, not only
in testing planting methods, but also in site selection (e.g., Short’s PTSI and TSI). Frequent
use of project scaling was implicitly revealed in many presentations to be an effective vetting
device that should save money.
SITE SELECTION
Site selection errors constitute the single greatest error that I see in the restoration process,
as evidenced by the high frequency of error in this aspect of planning in the projects given
me for review. Many of these errors could be avoided if practitioners (and reviewers of
plans) made better use of the population ecology and plant’s growth strategy in planning a
project. This simply means that some species spread faster than others and can be
augmented by simple facilitation techniques; thus, simple calculations of planting density
and recovery projections can and should be made.
Short’s Preliminary Transplant Suitability Index (PTSI) and TSI are amazingly among the
first generic tools that have been systematically developed for site evaluation since Phillips’
work in the early 1980s (which had several significant flaws). It would appear that the
development of similar indices for the subtropical species is an area ripe for immediate
research.
One small footnote is that planners should recognize that the application of planting
methods and the selection of sites will vary with the nature of injury, particularly whether
there is an identified responsible party (RP) or whether the restoration effort if in response
to non-RP situations, as is the case with non-point source pollutants. In the latter case, the
restoration monitoring and success criteria may be relaxed to some degree as getting
something is better than nothing, so long as it passes an acceptable cost to benefit
comparison in favor of the public. For the former, exacting monitoring and success criteria
must be employed—which means site selection is crucial—in order to assess lost interim
resource services and compute realistic replacement ratios.
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I also listed some simple metrics for site selection that I submit should be addressed before
moving on to the more comprehensive PTSI and TSI approaches. First, ask the questions:
• If seagrass is not there now, why not?
• What makes you think you can do better than nature?
• Do you have the data to back up your optimism?
• What are sources of disturbance? Water quality? Waves? Bioturbation?
• How will these be controlled?
• How well have they been evaluated?
In addition, I suggested that selection of a site should pass the following indicators (taken
from the aforementioned Fonseca et al. 1998). In addition to seeing restoration successful
at similar sites, a site should:
• Have depth is similar to nearby natural beds
• Be anthropogenic ally disturbed
• Not be subject to chronic storm disturbance
• Not be undergoing rapid and extensive natural recolonization
• Not be among patches of existing seagrass
• Have sufficient acreage to achieve goals
• Have similar quality habitat restored as was lost
SEAGRASS FUNCTION
Curiously, the workshop voted heavily to place research on the function of restored seagrass
beds near the top of the priority list. I strongly disagreed with this consensus and assert that
“If you build it, they will come.” Some of the most compelling data that supports this
notion was published out of Tampa Bay (including the fact that animals displayed
colonization peaks early in the development of planted beds, at densities only a of mature
plant coverage). Given that there has never (to my knowledge) been a paper published that
failed to demonstrate that any seagrass that persists for a least a year, whether naturally or
artificially colonized, was not full of animals, stabilized sediments and had high levels of
primary production, then why spend a lot of money proving that yet again? In addition, in
U.S. Federal Court cases the value of seagrass ecosystems does not have to be proven (it is
a given), so why struggle to document this over and over?
MONITORING
Like site selection, monitoring was shown in the workshop to fall into two general
categories: for situations with an RP and situations without an RP. Monitoring, whether it
be of the environmental conditions contributing to the performance of a planting or
monitoring of the planting itself, simple metrics may apply for the non-RP condition but
projects with clients or RPs were consistently shown in the workshop to have specific
standards (along with consequences for non-compliance). Non-scientists often despise
these exacting standards, but if the situation involves court proceedings, admissibility (to
court) standards demand a high bar.
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Wrap-Up
There seemed to be convergence on simple success criteria, such as seafloor coverage and
its persistence over time. Other measures are useful in a research context but often do not
provide information that can be used to institute interim corrective restoration measures,
or represent factors that cannot really be controlled and such metrics were (fortunately)
conspicuous in their absence during the workshop.
COST OF RESTORATION
Again, there was a strong convergence of cost data on restoration projects. Gone were the
estimates of turnkey seagrass restoration for a few thousand dollars an acre. My colleagues
and I have computed that seagrass restoration costs for a 1.5-acre subtropical (0.607 ha)
project could be broken down into five general components:
COMPONENT
PERCENT OF PROJECT COSTS
Map & groundtruth
Planting
Monitoring
Contractor
Government oversight
5.5
18.5
58.7
8.3
9.1
The total cost ranged between $360,000 and $590,000 (1996), or $240,000–$393,000 per
acre.
Another aspect of computing cost is the determination of success. What is the cost of
restoration vs. success and how do you measure your return? I argued that an assessment
of lost interim resource services and discounting of those services over time, as done by
NOAA in Natural Resource Damage Assessment cases, is the most defensible strategy at
this time. For a detailed review of this, see Fonseca et al. 2000.
Success, however, may not always be apparent. Again, evaluation of success can only be fully
determined by working within the context of the variation of the natural system. For
example, Walt Avery presented information on patch size over time (a normal distribution),
which to my eyes, was essentially the life history of a patch in that environment. While these
data demonstrate the problem with selecting spaces among existing, patchy seagrass as a
suitable restoration site (the system will be pulsed with added production for a few years,
but no long-term addition to the resource base was realized), it also demonstrated that the
introduction of seagrass to a propagule-limited site was critical to short circuit the
colonization process. Although the parent patch died out, it generated daughter patches that
persist—providing a consistent albeit spatially transient resource base for the area—which
constitutes a successful restoration of cover to that area as the cover is naturally patchy. This
also provided some response to Susan Bell’s question about quantifying contributions
beyond the border of the original planting; control plots—or in this case, systematic
observation of areas outside the plot boundaries—were required to provide a full assessment
of the contribution of the planting effort.
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SUMMARY OF LESSONS LEARNED
• Conduct work that is conducive to publication of the findings;
• Do your homework (read, follow up on references in that reading and perform
literature searches);
• Treat the project as if it were a for-profit business, with real criteria for costs and
products;
• Use an agricultural approach: Don’t plant where the plants cannot grow (begs good site
evaluation criteria like Fred Short is using and has been applied at the landscape scale
in the Chesapeake Bay);
• Measure your results with simple metrics and robust sampling designs;
• Recognize natural variability and set realistic expectations;
• Photo document the work;
• DO YOUR HOMEWORK: “those who forget the lessons of history are doomed to repeat it” (George
Santayana).
FUTURE OF SAV RESTORATION
Based on what I saw at the workshop, I had the following closing observations regarding the
future needs of seagrass or SAV restoration in general and associate them with examples to
be found with presentations in the workshop:
• We need more non-gray haired people, i.e., we need recruitment to the field (pers.
obs.);
• Greater attention is needed to over come the inertia preventing a synthetic approach;
the parts are all here, but each project must put them together: use extant information
on site selection and success criteria, proven methods, and statistically valid
monitoring which can link cause and effect (e.g. Poirrier: Lake Ponchatrain);
• Use: “a scientific method involving the formulation of theories or hypotheses from
which singular statements (predictions) are deduced that can be tested; deductive
method” (e.g. Carlson, et al.; definition taken from Lincoln et al. A Dictionary of
Ecology, Evolution and Systematics);
• Conduct long term monitoring of natural beds and planting efforts to create
forecasting baseline: provide standard methods to managers (e.g. Avery &
Johansson; Ott, Morris, Wade, et al; Wilbur’s conundrum);
• Consider multiple species (phyla) restoration (Chesapeake Bay and L. Ponchatrain
• Continue technique development (Clark; Montin and Dennis):
• Test new methods (e.g. Anderson planting machines vs. other methods; Hall et al);
• Improve documentation of physiological stress (e.g. Cuba);
• Focus on good organization (e.g., Montin & Dennis);
• Deal with disturbance: high energy and biological disturbance (many studies here);
• Improve ecological forecasting (beyond TSI) such that there are quantitative links
that allow restoration performance to be gauged by an evaluation from local
environmental monitoring (e.g. Tomasko; Ott);
• Keep reminding the public that seagrasses are a “canary” for the estuary (e.g. Wade);
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Wrap-Up
• Train our managers properly—up front, not on the job, so I don’t come back in
another five years and tell the same stories again. Resist Yogi Berra’s: “it’s déjà vu all
over again.”
LITERATURE CITED
Fonseca MS, Kenworthy WJ, Thayer GW. 1998. Guidelines for Mitigation and Restoration of Seagrass in the United
States and Adjacent Waters NOAA COP/Decision Analysis Series. 222p; http://shrimp.ccfhrb.
noaa.gov/library/digital.html)
Fonseca MS, Julius BE, Kenworthy WJ. 2000. Integrating biology and economics in seagrass restoration: How
much is enough and why? Ecological Engineering 15:227–237
MSF (NOAA, NOS, Center for Coastal Fisheries and Habitat Research, 101 Pivers Island Road, Beaufort,
NC 28516-9722)
Opinions expressed herein are those of the author and do not reflect the official position of NOAA.
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