Document ID: EPA-HQ-OW-2002-0049-0073
Agency: epa
Document Type: Supporting & Related Material
Title: 
Posted Date: 2003-03-19T05:00Z

CHAPTER
1
Background
WHAT
ARE
SEAGRASSES?

S
eagrasses
are
unique
marine
flowering
plants
of
which
there
are
approximately
60
species
worldwide
(
den
Hartog
1970,
Phillips
and
Menez
1988).
With
the
exception
of
some
species
that
occur
in
the
rocky
intertidal
zone,
they
grow
in
shallow,
subtidal
or
intertidal
unconsolidated
sediments.
Thus,
they
bind
millions
of
acres
of
shallow
water
sediments
in
the
coastal
waters
with
their
roots
and
rhizomes
while
simultaneously
baffling
waves
and
currents
with
their
leafy
canopy
(
Ginsberg
and
Lowenstam
1958,
Taylor
and
Lewis
1970,
den
Hartog
1971,
Fonseca
et
al.
1983,
Fonseca
1996a).
In
this
manner
the
canopy
inhibits
resuspension
of
fine
particles
and
traps
water­
column
borne
material
(
Ward
et
al.
1984,
Short
and
Short
1984),
clearing
the
water
column
This
cleansing
effect
extends
to
water
column
nutrients
as
well.
Nutrient
uptake
by
seagrass
blades
and
their
associated
epiphytes
and
macroalgae
as
well
as
roots
incorporate
dissolved
nutrients
into
plant
biomass,
which
can
improve
water
quality
(
Harlin
and
Thorne­
Miller
1981).
The
baffling
effect
of
the
canopy
on
sediment
stabilization
is
enhanced
by
the
presence
of
a
robust
root
and
rhizome
mat,
although
the
relative
contribution
of
the
mat
has
not
been
isolated
from
canopy
baffling
in
its
role
of
sediment
stabilization
(
Fonseca
1996a).
The
physical
stability,
reduced
mixing
and
shelter
provided
by
the
complex
seagrass
structure
provides
the
basis
for
a
highly
productive
ecosystem
(
Wood
et
al.
1969).
Overall
the
importance
of
seagrasses
and
their
role
in
many
coastal
ecosystems
has
been
extensively
docu­

1
mented
(
see
reviews
by
Thayer
et
al.
1975,
Phillips
1982,
Zieman
1982a,
Thayer
et
al.
1984,
Zieman
and
Zieman
1989)
and
the
nature
of
their
general
function
and
high
resource
value
are
no
longer
an
issue.

Seagrasses
occur
in
all
coastal
states
of
the
U.
S.
with
the
apparent
exception
of
Georgia
and
South
Carolina
where
freshwater
inflow,
high
turbidity
and
tidal
amplitude
combine
to
prevent
their
occurrence.
There
are
at
a
minimum
thirteen
species
of
seagrass
currently
recognized
to
occur
in
U.
S.
waters
(
Table
1.1).
The
presence
of
a
fourteenth
species,
Zostera
asiatica
on
the
West
Coast
remains
a
subject
of
debate
(
Phillips
and
Wyllie­
Echeverria
1990).
We
will
not
include
in
this
discussion
seagrass
species
occurring
in
U.
S.
possessions
in
the
Pacific
Ocean
because
little
is
known
about
their
status;
through
NMFS
Southwest
Regional
Office
reports,
we
know
that
Enhalus
acoroides
and
Halodule
uninervis
occur
on
Rota
Island
and
Saipan
Island
in
the
Pacific
Territories.
Also,
Phillips
and
Menez
(
1988)
list
Halophila
ovalis
and
Halophila
minor
(
fifteenth
and
sixteenth
species)
as
species
that
occur
in
Hawaii.
Halophila
hawaiiana
is
also
reportedly
present
on
Hawaii
(
K.
Bridges,
Univ.
Hawaii,
pers.
com.).
Drawings
of
the
major
U.
S.
species
are
given
in
Figure
1.1.
One
species,
Halophila
johnsonni
was
only
recently
described
as
a
separate
species
despite
its
occurrence
in
the
heavily­
studied
region
of
southeast
Florida.
Because
of
its
limited
distribution,
this
species
is
currently
under
consideration
for
listing
as
a
threatened
species
as
defined
by
the
Endangered
Species
Act.
Another
species,
Zostera
japonica
was
recently
introduced
to
the
Pacific
Northwest
.
It
is
spreading
and
tends
to
colonize
shallow
intertidal
flats,
converting
them
from
their
historical
ecological
status
as
mudflats
to
intertidal
eelgrass
habitat
(
Harrison
and
Bigley
1982,
Pawlak
1994).

Although
recognized
for
their
value
where
they
occur,
the
distribution
of
seagrass
is
not
as
well
known
as
it
should
be
for
proper
management
(
Wyllie­
Echeverria
et
al.
1994a).
Moreover,
knowledge
of
population­
level
temporal
dynamics
is
only
rudimentary
at
best.
We
know
that
at
least
90
percent
of
the
southeast
United
States
seagrass
acreage
(~
1.1
million
hectares)
exists
in
the
Gulf
of
Mexico
(
Orth
and
Van
Montfrans,
1990).
But
nationally,
the
distribution
and
abundance
of
two
genera
in
particular
have
been
overlooked.
The
full
extent
and
function
of
the
reported
~
400,000
hectares
of
seasonal
Halophila
beds
off
the
west
coast
of
Florida
(
Iverson
and
Bittaker
1986)
is
unknown.
Similarly
the
distribution
of
the
Hawaiian
Halophila
is
not
reported.
Also,
very
little
is
known
about
local
distribution
(
distribution
meaning
localized,
specific
locations
of
beds,
not
the
range
of
a
species)
of
a
unique
West
Coast
dominant,
the
rocky
intertidal
Phyllospadix
spp.,
although
work
has
been
done
regarding
its
population
ecology
(
Turner
1985,
Turner
and
Lucas
1985).
The
distribution
of
seagrass
on
the
West
Coast,
including
both
Alaska
and
Hawaii,
has
not
been
systematically
compiled
to
the
degree
seagrasses
have
on
the
east
and
Gulf
coasts
2
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
3
Table
1.1.
List
of
seagrass
by
family,
genus
and
species,
and
common
names
(
if
given)
that
are
found
in
the
United
States
and
adjacent
waters.
Species
marked
with
(?)
are
not
fully
documented
as
occurring
in
U.
S.
waters.

Family,
Genus,
and
Species
Common
Namea
Hydrocharitaceae
Enhalus
acoroides
Royle
Halophila
decipiens
Ostenfeld
paddle
grass
Halophila
engelmanni
Ascherson
star
grass
Halophila
hawaiiana
Doty
and
Stone
Hawaiian
seagrassa
Halophila
johnsonii
Eiseman
Johnson's
seagrass
Halophila
minor
(
Zollinger)
den
Hartog?
unknown
Halophila
ovalis
(
R.
Brown)
Hooker
f.?
unknown
Thalassia
testudinum
Konig
turtlegrass
Potamogetonaceae
Halodule
wrightii
Ascherson
shoalgrass
Halodule
uninervis
?

Phyllospadix
scouleri
Hook
Scouler's
seagrass
Phyllospadix
torreyi
S.
Watson
Torrey's
seagrass
Phyllospadix
serrulatus
Ruprecht
et
Ascherson
surfgrass
Ruppia
maritima
L.
widgeon
grass
Syringodium
filiforme
Kutz
manatee
grass
Zostera
japonica
Ascherson
et
Graebner
Japanese
eelgrass
Zostera
marina
L.
eelgrass
Zostera
asiatica?
Asian
eelgrass
a
Italics
on
common
names
indicate
suggested
common
names;
R.
Phillips,
Battelle
Laboratories,
Richland,
Wa.,
pers.
com.
4
Figure
1.1.
Drawings
of
most
seagrasses
found
in
U.
S.
waters
(
taken
from
Phillips
and
Menez
1988
and
Fonseca
1994).
All
scale
bars
are
set
at
2cm
and
thus
vary
with
seagrass
species.
A=
Zostera
marina;
B=
Zostera
japonica;
C=
Ruppia
maritima;
D=
Halodule
wrightii;
E=
Syringodium
filiforme;
F=
Thalassia
testudinum;
G=
Halophila
engelmanni
H=
Halophila
decipiens;
I=
Halophila
johnsonni;
J=
Phyllospadix
serrulatus;
K=
Phyllospadix
torreyi;
L=
Phyllospadix
scouleri.
5
Figure
1.1.
continued.
6
Figure
1.1.
continued.
although
the
general
range
of
species'
distributions
has
been
reported
(
Wyllie­
Echeverria
and
Phillips
1994).

Historically,
emphasis
been
placed
on
aspects
of
seagrass
primary
and,
to
a
lesser
degree,
secondary
production
attributes
(
see
descriptions
in
Zieman
1982a,
Phillips
1984,
Thayer
et
al.
1984).
Extensive
information
is
available
regarding
light
and
nutrient
requirements
of
seagrasses
(
Kenworthy
and
Haunert
1991,
Dennison
et
al.
1993,
respectively).
Seagrasses
are
flowering
plants
and
much
attention
has
been
paid
to
the
mechanics
of
pollination
and
seed
dispersal
(
see
review
by
Cox
1993
and
references
therein)
but
much
less
is
known
about
the
role
of
seeding
in
bed
maintenance
or
colonization
of
new
areas
(
Kenworthy
et
al.
1980,
Harrison
1993,
Orth
et
al.
1994).
With
the
exception
of
some
recent
studies
(
Duarte
et
al.
1994,
Durako
1994)
and
previous
transplanting
data
sets
(
Fonseca
et
al.
1987c),
demographic
studies
have
been
sorely
neglected
in
this
country
yet
this
is
a
topic
area
where
managers
ask
many
questions:
How
quickly
will
a
seagrass
bed
recover
from
a
given
impact?
Is
planting
necessary?
Given
intrinsic
recovery
rates
and
transplanting
success,
how
do
we
compute
replacement
ratios
or
estimate
interim
loss?
Should
we
be
concerned
about
genetic
diversity
of
the
population?
These
questions
are
only
now
being
addressed.
Chapter
1:
Background
°
7
Figure
1.1.
continued.
8
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
DEFINING
SEAGRASS
HABITAT
Seagrass
beds
exist
in
a
wide
variety
of
physical
settings
that
lead
to
different
coverage
patterns.
The
problem
is
coming
up
with
a
consistent
definition
of
what
constitutes
a
seagrass
bed.
Although
small
patches
may
themselves
have
significant
resource
value,
how
does
one
assess
the
collection
of
patches
and
determine
the
boundaries
of
a
seagrass
habitat?
Seagrasses
exhibit
a
variety
of
growth
strategies
and
coverage
patterns
which
occur
from
rocky
and
soft­
bottom
intertidal
habitats
to
depths
of
at
least
40
meters.
Some
species
can
rely
heavily
on
seeding
to
ensure
yearto
year
survival
(
e.
g.,
H.
decipiens
and
possibly
H.
engelmanni)
meaning
that
surveys
during
winter
months
would
need
to
include
sediment
seed
bank
assessments
to
accurately
define
the
presence
of
a
seagrass
bed.
Moreover,
some
species,
such
as
Z.
marina,
can
exist
either
as
perennials
or
annuals,
again
requiring
very
different
assessment
strategies,
varying
between
seed
bank
and
vegetative
material
depending
upon
time
of
year.
Clear
knowledge
of
seagrass
population
ecology
is
a
requirement
for
effective
management
and
planting;
that
is,
one­
time
snapshot
inventories
are
a
very,
very
poor
basis
upon
which
to
delineate
seagrass
habitat.

Seagrass
beds
move.
Depending
on
the
species
and
the
physical
setting,
the
rate
at
which
portions
of
the
seafloor
switch
from
vegetated
to
unvegetated
may
vary
on
the
scale
of
days
or
decades,
meaning
that
the
amount
of
open
seafloor
required
to
maintain
patchy
seagrass
beds
is
greater
than
the
coverage
by
the
seagrass
itself
at
any
one
point
in
time
(
Figure
1.2),
sometimes
by
a
factor
of
two
(
i.
e.,
over
time,
the
movement
of
seagrass
beds
means
that
they
will
soon
occupy
at
least
twice
the
presently
unvegetated
bottom
evident
at
any
one
survey
time).
Thus,
if
unvegetated
areas
among
existing
patches
of
seagrass
are
converted
to
channels,
the
long­
term
(
within
four
years,
unpubl.
data)
baseline
acreage
of
seagrass
in
the
vicinity
of
the
converted
habitat,
will
decline.
Therefore,
seagrass
habitat
must
be
recognized
as
including
not
only
continuous
cover
beds,
but
chronically
patchy
habitat;
a
policy
that
requires
considering
the
(
presently)
unvegetated
spaces
between
seagrass
patches
as
seagrass
habitat
as
well.
Management
of
seagrass
resources
therefore
depends
on
understanding
the
spatial
and
temporal
dynamics
of
seagrass
coverage.

One
of
the
biggest
problems
regarding
delineation
of
seagrass
habitat
relates
to
the
choice
of
sampling
scale
during
the
process
of
inventory,
especially
prior
to
a
planned
impact
to
a
seagrass
bed
(
see
section,"
Spatial
Scale
and
its
Role
in
Defining
Seagrass
Habitat,"
below).
Scale
is
roughly
defined
here
as
the
variation
of
pattern
as
a
function
of
the
range
and
resolution
of
examination.
The
scale
at
which
assessments
of
seagrass
coverage
take
place
varies
tremendously,
depending
on
some
Chapter
1:
Background
°
9
covariate
of
acreage,
interest
and
time
available
to
conduct
surveys.
In
contrast,
after
a
planting
is
installed,
monitoring
of
seagrass
plantings
is
less
prone
to
scale
problems
as
direct
count
methods
are
usually
employed
and
statistical
sub­
sampling
protocols
can
be
instituted
to
ensure
adequate
sampling
intensity.
However,
assessment
of
existing
natural
seagrass
and
post­
coalescent
seagrass
plantings
takes
place
at
many
spatial
scales
and
this
leads
to
very
different
values
of
seagrass
abundance.
If
aerial
photographs
are
used,
the
altitude
of
the
airplane,
the
camera
lens,
film,
solar
angle,
water
turbidity,
and
wind
waves
affect
the
ability
to
detect
seagrass
beds,
particularly
at
the
lower
end
of
their
depth
distribution.
Similarly,
if
one
chooses
to
survey
a
potential
impact
site
from
the
deck
of
a
small
boat
then
wavelets,
reflectance,
turbidity
and
an
individual's
search
image
all
influence
ability
to
assess
seagrass
abundance
Aerial
photography
such
as
that
recommended
by
the
NOAA
Coastal
Change
Analysis
Program
(
C­
CAP)
(
Dobson
et
al.
1995),
has
a
minimum
mapping
unit
of
0.03
ha.
At
that
resolution
roughly
37
percent
of
the
permits
issued
for
alter­

Figure
1.2.
Plot
of
the
cumulative
area
of
bottom
covered
in
50
x
50m
survey
areas
over
time.
Y­
axis
=
cumulative
cover
assessed
by
adding
new
square
meters
of
cover
to
that
not
previously
covered
in
any
survey.
X­
axis
=
sampling
dates.
Each
line
type
represents
a
different
50
x
50m
site.
10
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
ation
of
submerged
aquatic
vegetation
habitats
could
not
be
detected.
Fortunately,
those
that
could
be
detected
with
0.03
ha
resolution
accounted
for
~
99
percent
of
the
acreage
impacted
(
Rivera
et
al.
1992).
Inherently
patchy
seagrass
beds
would
be
even
more
difficult
to
detect
and
quantify
at
a
spatial
resolution
less
than
0.03
ha
using
C­
CAP
techniques.
These
scales
<
0.03
ha
are
spatial
scales
that
questions
of
planting
unit
(
PU)
spacing
and
groupings
of
PU
must
be
addressed
(
see
section
on
"
Spacing
of
Planting
Units"),
and
persistent
seagrass
patches
can
be
produced
at
these
smaller
scales.

Fonseca
(
1989a)
suggested
that
at
the
1:
24,000
scale
of
aerial
photography
when
the
ratio
of
average
seagrass
patch
diameter
to
the
distance
between
patches
exceeds
50:
1,
seagrass
habitat
continuity
no
longer
fosters
cognitive
recognition
by
a
viewer
as
constituting
seagrass
habitat.
He
suggested
that
above
that
ratio
the
area
should
no
longer
be
considered
continuous
seagrass
habitat.
Clearly
this
ratio
is
scale
dependent
If
a
ratio
of
50
shoot
widths
to
the
distance
between
shoots
were
used,
then
many
seagrass
beds
on
the
West
Coast
and
in
the
northeast
where
individual
plants
are
very
large
(>
2
m
length)
would
no
longer
be
considered
seagrass
habitat
even
though
the
unit
area
biomass
might
be
comparable
to
other
seagrass
beds
in
the
country.
Unfortunately,
we
are
not
aware
of
any
quantitative
description
of
how
bed
boundaries
are
interpreted
(
i.
e.,
when
a
bed
is
drawn
as
one
large
polygon
or
many
small
polygons).
However,
variation
in
seagrass
bed
form
can
easily
be
visually
detected
from
low­
level
aerial
reconnaissance
(
Figure
1.3),
and
appears
to
be
correlated
with
exposure
to
waves
and
currents.
Under
wave
and
current
conditions
beds
can
take
extreme
forms;
Molinier
and
Picard
(
1952)
and
Fonseca
(
1996a)
described
vertical
walls
of
Posidonia
and
Zostera,
respectively,
revealing
the
extent
to
which
seagrass
could
reduce
erosion
and
enhance
sediment
accumulation.
Seagrass
patterns
also
change,
revealing
areas
of
seagrass
coverage
loss
and
gain
at
meter
scales
within
short
time
periods
(
months)
(
Figure
1.4)
attesting
to
the
consistent
ability
of
seagrasses
to
stabilize
sediments.
For
at
least
20­
30
years
after
Molinier
and
Picard's
work,
little
in
the
way
of
a
quantitative
association
of
seagrasses'
effects
on
water
motion
and,
conversely
the
effect
of
water
motion
on
seagrass
bed
development
took
place.
During
this
time,
interest
in
the
physical
processes
occurring
in
seagrass
beds
was
confined
largely
to
qualitative
descriptions
of
their
geological
role
and,
to
a
lesser
degree,
the
implications
of
this
geological
stability
on
animal
utilization.

It
is
unlikely
that
there
will
be
a
universal
standard
for
defining
seagrass
habitat.
Different
seagrass
species
form
beds
that
occupy
too
great
a
diversity
of
habitats
and
exhibit
such
a
range
of
life
history
strategies
that
a
universal
definition
would
almost
certainly
be
restrictive
and
unworkable.
Further,
published
data
on
seagrass
biomass,
density,
and
structural
complexity
(
e.
g.,
surface
area)
have
tended
to
be
collected
from
11
Figure
1.4.
Change
in
seagrass
bed
cover
in
a
wave­
exposed,
patchy
seagrass
bed
near
Beaufort,
NC.
Dark
circle=
m2
areas
with
no
change
in
cover
(
6­
month
period),
+
=
areas
of
seagrass
gain,
°
=
areas
of
loss
and
no
symbols
=
areas
of
unchanged
sand.
Figure
1.3.
Aerial
photograph
of
mixed
Halodule
wrightii,
Ruppia
maritima,
Syringodium
filiforme,
and
Thalassia
testudinum
beds
on
the
western
margin
of
Tampa
Bay,
Florida.
In
the
foreground
at
the
bayward
edge
of
the
shoal
are
what
appear
to
be
wave­
sculpted
beds
while
further
landward,
in
shallow
water
are
more
continuous
cover
bed.
Reduction
in
wave
energy
from
both
the
shelving
shoal
and
the
grass
itself
is
thought
to
be
responsible
for
the
resultant
seagrass
bed
landscape
pattern.
Taken
from
Fonseca
(
in
press).
12
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
seagrass
beds
that
form
large
unbroken
meadows.
Limited
comparative
information
on
bed
spatial
heterogeneity
is
available
from
the
full
range
of
habitats
or
landscape
patterns
that
seagrasses
form.
Therefore,
if
we
used
published
data
to
set
boundary
definitions
of
seagrass
beds,
it
is
quite
likely
they
would
tend
to
define
only
certain
seagrass
species
(
i.
e.,
commonly
studied
species
such
as
Z.
marina,
etc.)
in
certain
settings
(
e.
g.,
relatively
wave­
protected
and
low
current
speeds
which
yield
extensive,
non­
patchy
habitat).
Further,
because
data
collection
has
been
historically
biased
toward
beds
in
lower
energy
environments,
the
more
fragmented,
patchy
nature
of
higher
energy
seagrass
beds
would
be
an
element
of
seagrass
bed
structure
that
would
not
be
captured
in
such
a
universal
definition.
On­
site,
direct
surveys
of
local
undisturbed
seagrass
beds
in
similar
physical
settings,
or
better,
pre­
impact
surveys
of
the
status
of
a
seagrass
bed
remain
over
time
the
best
guidelines
for
delineating
seagrass
habitats.

What
we
suggest
is
that
managers
must
have
some
historical
perspective.
Onetime
surveys
are
completely
inadequate
data
(
i.
e.,
see
Figure
1.2)
upon
which
to
base
management
decisions
that
could
have
effects
for
years.
Bed
form
migration
(
sensu
Patriquin
1975,
Marba
et
al.
1994,
Marba
and
Duarte
1995),
presence
of
seed
banks,
annual
populations,
recent
nonpoint
source
anthropogenic
impacts
(
e.
g.,
decreased
water
clarity),
and
even
deliberate
removal
of
seagrasses
all
combine
to
cast
doubt
on
the
veracity
of
one­
time
surveys
(
i.
e.,
see
Figure
1.2).
For
evaluations
of
extant
beds,
even
seemingly
straightforward
information
such
as
shoot
density
can
be
misleading.
Data
such
as
shoot
density
are
sometimes
inversely
related
to
shoot
size,
meaning
that
shoot
densities
of
even
less
than
one
shoot
m­
2
may
be
significant,
especially
if
that
shoot
is
very
large.
Conversely,
populations
of
Halophila
spp.,
of
which
there
may
be
in
excess
of
half
a
million
hectares
in
the
Gulf
of
Mexico
and
Indian
River
Lagoon
(
Iverson
and
Bittaker
1986,
Continental
Shelf
Assoc.
1991,
Kenworthy
1992),
return
almost
exclusively
from
seed
every
spring
(
Williams
pers.
com.).
As
with
other
species
that
rely
heavily
upon
seeds
for
seasonal
recovery,
surveys
taken
during
months
where
aboveground
biomass
is
all
but
absent
and
that
do
not
incorporate
seed
bank
surveys
would
erroneously
conclude
the
area
did
not
support
seagrass.

SPATIAL
SCALE
AND
ITS
ROLE
IN
DEFINING
SEAGRASS
HABITAT
If
physical
processes
have
the
potential
to
affect
habitat
heterogeneity
in
seagrass
communities
then
there
is
the
potential
for
affecting
associated
fauna
(
Fonseca
and
Fisher
1986).
Seagrass
beds
composed
of
isolated,
dune­
like
patches
of
~
2
m
in
diameter
can
coalesce
within
several
growing
seasons
upon
elimination
of
waves
and
tidal
currents
(
pers
obs).
Despite
the
clear
relationship
of
water
motion
to
seagrass
Chapter
1:
Background
°
13
bed
form,
we
have
only
begun
to
evaluate
their
spatial
(
or
temporal)
organization
(
Virnstein
1995),
otherwise
seagrass
beds
have
consistently
been
treated
as
a
"
black
box"
at
the
landscape
scale.
To
build
on
information
accumulated
on
ecosystems
and
apply
this
information
to
seagrass
systems,
research
emphasis
must
include
not
only
the
normative
1
m
scale
study,
but
scales
that
are
relevant
to
mechanisms
that
contribute
to
the
formation,
maintenance,
and
function
of
whole
systems,
such
as
sediment
transport
pathways
or
an
organism's
range.

If
the
pattern
of
distribution
observed
in
seagrass
beds
is
the
result
of
physical
processes
whose
effects
vary
with
the
spatial
scale
of
examination,
then
it
follows
that
the
influence
of
bed
pattern
on
such
things
as
faunal
abundance
will,
in
turn,
vary
with
spatial
scale
as
well
(
sensu
Bian
and
Walsh
1993,
Fonseca
1996).
Therefore,
knowing
the
range
of
these
scales
is
potentially
valuable
if,
after
gathering
empirical
evidence,
one
can
infer
structural
attributes
at
other
scales
of
interest,
especially
scales
that
may
be
less
expensive
to
derive
(
e.
g.,
aerial
photography).

Resource
managers
must
realize
that
a
relationship
between
ecological
phenomena
and
the
spatial
scale
of
a
survey
is
real
and
sometimes
intuitive.
At
the
least,
such
relationships
are
a
statistical
reality
that
can
strongly
affect
interpretation
of
field
survey
data
(
Rossi
et
al.
1992,
Cao
and
Lam
1997).
The
notion
that
interactions
at
one
scale
(
spatial
or
temporal)
affect
that
which
is
expressed
on
another
scale
provides
the
basis
for
hypothesizing
scale­
dependent
effects.
Therefore,
spatial
and
temporal
patterns
seen
in
seagrass
ecosystems
are
the
result
of
physical
processes
acting
both
on
individual
plants
and
the
local
population
level
(
individual
patch).
Responses
of
individual
plants
to
water
motion
and
associated
phenomena
(
e.
g.,
sediment
particle
size)
may
be
cumulative
and
affect
seagrass
landscape
patterns
perceived
at
coarser
scales
of
resolution.
To
summarize,
examples
of
the
importance
of
deriving
scale
dependence
in
seagrass
beds
include
identification
of:

1.
The
scale
at
which
samples
taken
in
the
landscape
are
independent
of
one
another
and
improve
sampling
stratification,

2.
Their
effect
on
animal
utilization
and
distribution,
and
3.
The
relevant
scales
over
which
sedimentary
processes
are
controlled
providing
a
better
prediction
of
alterations
in
current
patterns,
interception
(
or
lack
thereof)
of
wave
energy,
and
sedimentary
processes
as
the
result
of
altering
the
seagrass
landscape.

One
result
of
recent
research
on
seagrass
landscape
patterns
is
that
there
are
ranges
of
spatial
scales
over
which
estimates
of
coverage
vary
as
the
result
of
the
scale
of
sampling
resolution
chosen
by
the
investigator
(
Fonseca
1996b).
Moreover,
for
14
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
seagrass
beds
in
North
Carolina
and
Tampa
Bay
that
experience
relative
wave
exposure
values
(
see
"
Constraints
Imposed
by
Physical
Setting
on
Planting
Operations,"
below)
greater
than
3
x
106
(
on
a
scale
that
runs
from
0
to
~
6
x
106)
any
estimate
of
seagrass
coverage
will
differ
depending
on
the
size
of
the
sampling
unit
and/
or
the
distance
separating
those
sample
units
at
scales
<
10m
(
Fonseca
1996b).
This
means
that
interpretation
of
any
factors
related
to
seagrass
bed
coverage
sampled
within
this
range
of
1­
10
m
will
be
different
among
any
studies
that
sampled
at
different
spatial
scales
(
i.
e.,
samples
taken
1
m
apart
versus,
for
example,
5
m
apart).
Therefore,
comparisons
among
studies
or
surveys,
even
of
the
same
bed,
will
differ
to
some
degree
simply
because
different
size
quadrats
were
used
and
not
necessarily
as
the
result
of
actual
differences
in
the
factor
being
compared.
Of
course,
comparisons
between
studies
can
be
different
because
different
numbers
of
samples
(
which
approximates
statistical
power)
are
taken.
Finally,
this
has
implications
for
the
integrity
of
sampling
schemes
because
any
samples
taken
in
this
range
of
scale
dependence
will
not
be
statistically
independent,
casting
doubt
on
the
validity
of
among­
study
or
among­
survey
comparisons
which
were
conducted
at
different
spatial
scales.
This
can
create
problems
for
interpretation
of
planting
success.

Scale
dependence
in
sampling
has
not
only
spatial
but
temporal
considerations.
We
raise
this
caution
regarding
temporal
scale
dependence
because
in
our
section
titled
"
Comparative
Analysis
of
Seagrass
Planting
Efforts"
we
found
that
many
projects
changed
assessment
frequency
during
the
course
of
the
monitoring
period.
In
fact,
we
too
recommend
a
change
in
assessment
protocol
depending
on
whether
it
is
being
conducted
before
or
after
coalescence
of
planting
units.
Therefore,
statistical
comparisons
should
be
made
with
caution
between
data
collected
from
pre­
and
post­
coalescence
because
such
comparisons
of
one
site
over
time
likely
violate
rules
of
sample
independence.
Because
many
planting
projects
cannot
escape
problems
with
sample
independence
over
time,
the
use
of
simple
descriptive
measures
(
such
as
area
covered
and
persistence)
as
standard
measurement
protocols
becomes
very
important
to
minimize
problems
with
comparative
analysis
among
studies
or
among
dates
within
studies.

Another
problem
with
spatially
heterogenous
(
i.
e.,
patchy
as
opposed
to
continuous
seagrass
beds
is
the
perception
of
their
comparative
ecological
function.
Spatially
heterogenous
seagrass
environments
in
North
Carolina
have
been
classified
as
"
scattered"
(
Carraway
and
Priddy
1983)
versus
continuous
cover
beds
that
are
termed
"
dense."
This
unfortunate
classification
inferred
a
lower
resource
value
despite
the
fact
that
the
former
landscape
pattern
covers
many
thousands
of
acres
of
estuarine
seafloor
in
North
Carolina,
has
shoot
densities
and
primary
production
equivalent
to
continuous
cover
beds,
has
significantly
higher
below­
ground
biomass
Chapter
1:
Background
°
15
than
continuous
beds,
and
often
supports
equal
densities
of
some
economically
valuable
species
such
as
pink
shrimp
(
Murphey
and
Fonseca
1995).

VULNERABILITY
AND
SUSCEPTIBILITY
OF
SEAGRASS
ECOSYSTEMS
Why
are
seagrasses
so
often
impacted
by
human
activity?
One
of
the
reasons
is
their
location
in
the
coastal
zone.
Because
of
their
relatively
high
(
compared
to
phytoplankton
light
requirements
(
Kenworthy
and
Haunert
1991)
they
occur
in
shallow
nearshore
waters,
a
situation
that
makes
them
extremely
susceptible
to
damage
by
human
activity
such
as
nutrient
loading
(
Short
and
Burdick
1996),
light
reduction
(
Dennsion
et
al.
1993,
Kenworthy
and
Fonseca
1996),
and
propeller
scarring
(
Sargent
et
al.
1995).
As
our
utilization
of
the
coastal
zone
grows
so
will
the
damage
to
seagrass
ecosystems
unless
proactive
steps
are
taken
to
avoid
those
impacts
and
successfully
mitigate
when
impacts
occur.
Because
they
are
now
universally
recognized
to
be
valuable
habitats,
efforts
to
mitigate
their
losses
have
been
underway
for
many
years.

It
is
critical
that
one
recognizes
that
seagrass
mortality,
whether
mechanically
induced,
such
as
dredging,
or
physiologically
induced
from
reduction
in
light
(
e.
g.,
docks,
turbidity),
often
happens
rapidly;
time
scales
for
loss
can
range
to
as
little
as
weeks
or
months.
Recruitment,
however,
does
not
typically
keep
pace,
yet
if
the
site
were
capable
of
supporting
continued
cover,
seagrass
may
recolonize
within
a
few
growing
seasons
(
Kenworthy
et
al.,
1980,
Harrison
1987,
Fonseca
et
al.
1990,
Thayer
et
al.
1994).
Recovery
via
natural
recruitment
is
a
demographic
process
with
tremendous
spatial
and
temporal
variation
(
e.
g.,
0
to
>
10,000
seeds
m­
1
for
Z.
marina
and
is
very
difficult
to
predict.
It
is
clear,
however,
that
seed
set
and
successful
germination
are
often
requisite
for
rapidly
(
1­
2
growing
seasons)
balancing
anthropogenically
induced
seagrass
mortality.
In
contrast,
vegetative
encroachment
may
take
many
years
(
Johannson
and
Lewis
1992)
or
even
longer,
as
is
suggested
by
the
lack
of
seagrass
recovery
in
portions
of
the
northeast
U.
S.
from
the
"
wasting
disease"
loss
of
the
1930'
s
(
sensu
Short
et
al.
1993).
The
point
here
is
that
there
are
fundamentally
different
time
scales
involved
in
population­
scale
losses
and
their
recovery.
Only
recently
have
investigations
begun
to
assess
the
population­
scale
processes
of
seagrass
bed
formation
and
maintenance
(
Orth
et
al.
1994).
In
fact,
scientists
have
no
clear
idea
what
constitutes
a
population
for
these
plants
or
what
population
processes
are
at
work
(
i.
e.,
existence
of
metapopulations,
sensu
Orth
et
al.
1994).
At
a
minimum
documentation
of
distribution
together
with
elucidation
of
demographic
process
must
be
a
research
priority.
HISTORICAL
IMPACTS
AND
LOSSES
We
have
mentioned
environmental
constraints
to
seagrass
planting
(
see
review
by
Phillips
1982),
but
there
are
many
other
management
constraints
that
determine
the
effectiveness
of
seagrass
planting.
One
is
the
degree
of
philosophical
alignment
among
federal,
state
and
local
agencies
whose
jurisdictions
include
seagrass
habitat.
The
U.
S.
Army
Corps
of
Engineers,
whose
function
includes
issuance
of
dredge
and
fill
permits,
sometimes
cannot
follow
recommendations
from
other
agencies
to
conserve
seagrass
habitat
(
Mager
and
Thayer
1986).
Conflicts
between
preservation
of
seagrass
(
and
many
other
wetland
habitats)
and
implementation
of
public­
interest
development
projects
must
be
balanced
by
resource
agencies
but
often
results
in
the
loss
of
seagrass
habitat
(
sensu
Race
and
Fonseca
1996).
The
loss
of
seagrass
habitat
is
sometimes
addressed
by
proposing
in­
kind
mitigation.
In
addition,
maintenance
dredging
projects,
particularly
those
associated
with
national
security,
are
often
considered
exempt
from
mitigation
requirements
although
in
instances
of
very
long
dredging
cycles
(
years
to
decades),
mitigative
actions
are
sometimes
implemented
to
minimize
immediate
impacts.
It
has
been
our
experience
that
as
more
information
is
presented
to
managers
regarding
the
functions
of
seagrass
ecosystems
and
the
difficulties
involved
in
mitigating
for
their
loss,
fewer
permitted
impacts
are
occurring
in
seagrass
beds.

Although
the
loss
of
seagrasses
due
to
dredging
has
been
significant
(
Taylor
and
Saloman
1968,
Onuf
1994),
it
is
likely
that
the
majority
of
seagrass
habitat
loss
does
not
result
directly
from
dredge­
and­
fill
activities.
More
recently,
direct
impacts
from
mooring
scars
(
F.
Short,
Jackson
Est.
Lab.,
Durham.
NH,
pers.
com.),
propeller
scars
(
Sargent
et
al.
1995),
jet
skis
(
Kreuer
pers.
com.)
and
vessel
wakes
(
pers.
obs.)
are
emerging
as
a
major
source
of
seagrass
habitat
loss.
For
some
species
of
seagrass
such
as
Thalassia
which
is
slow
spreading
(
Fonseca
et
al.
1987c),
physical
damage
is
extremely
long­
lasting
(
Zieman
1976,
Durako
et
al.
1992).
Short
et
al.
(
1993)
and
the
Chesapeake
Bay
Program
(
1995)
recognized
improvement
of
wastewater
treatment
surface
run­
off,
restrictions
on
certain
fish
and
shellfish
harvesting
techniques,
and
regulation
of
boat
traffic
as
key
elements
in
protecting
seagrass
beds.
Although
scallop
harvesting
has
been
shown
to
damage
seagrass
beds
(
Fonseca
et
al.
1984)
as
has
raking
(
Peterson
et
al.
1984)
and
prop­
dredging
for
clams
(
Peterson
et
al.
1987),
other
fishery
techniques
such
as
trawling
for
bait­
shrimp
with
specially­
designed
gear
can
have
little
apparent
effect
on
seagrass
although
by­
catch
mortality
is
severe
(
Meyer
et
al.
in
review).
Work
by
the
Chesapeake
Bay
Program
(
1995)
also
lists
(
blue)
crab
dredging
(
scraping)
as
a
significant
impact
on
eelgrass
beds.
Fishing
gear
impacts
to
seagrass
beds
must
be
examined
on
a
gear­
by­
gear
basis.
16
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
Reduction
in
water
quality,
including
water
clarity,
is
another
significant
agent
of
seagrass
loss
(
Dennison
et
al.
1993,
Gallegos
1994,
Onuf
1994,
Gallegos
and
Kenworthy
in
press).
Burkholder
et
al.
1992
and
Dennison
et
al.
(
1993),
like
Batiuk
et
al.
(
1992),
provided
general
guidance
on
maintaining
water
chemistry
to
support
healthy
seagrass
beds.
In
doing
so,
Dennison
et
al.
(
1993)
essentially
determined
the
converse
of
health
standards;
they
defined
some
critical
water
chemistry
conditions
at
which
harm
would
come
to
seagrass
beds
(
Table
1.2).
These
data,
and
those
promulgated
by
the
Chesapeake
Executive
Council
(
1989)
and
the
Chesapeake
Bay
Program
(
1995),
are
perhaps
the
only
quantitative
water
chemistry
information
for
managers
to
evaluate
the
health
of
seagrass
environments
at
this
time.
They
are
likely
useful
for
most
temperate
seagrass
ecosystems
and
likely
describe
levels
that
would
be
too
high
for
typically
oligotrophic
tropical
and
sub­
tropical
waters,
particularly
those
dominated
by
carbonate
sediments
(
sensu
Fourqurean
et
al.
1995).
However,
the
correlation
between
human
development
of
the
shoreline
and
seagrass
decline
is
clear
(
Short
and
Burdick
1996).

Although
seagrass
beds
are
dynamic
systems,
with
some
beds
persisting
essentially
unchanged
for
decades,
others
change
with
the
season
(
den
Hartog
1971,
Zieman
and
Wood
1975,
Phillips
1980a,
Fonseca
et
al.
1983,
Duarte
and
Sand­
Jensen
1990).
Some
changes
in
seagrass
communities
can
be
attributed
to
the
life
histories
of
individual
seagrass
species
(
e.
g.,
Halophila
spp.).
However,
natural
perturbations
Chapter
1:
Background
°
17
Table
1.2.
Chesapeake
Bay
submersed
aquatic
vegetation
habitat
requirements.
For
each
parameter,
the
maximal
growing
season
median
value
that
correlated
with
plant
survival
is
given
for
each
salinity
regime.
Growing
season
defined
as
April­
October,
except
for
polyhaline
(
March­
November).
Salinity
regimes
are
defined
as
tidal
fresh
=
0­
0.5
o/
oo,
Oligohaline
=
0.5­
5
o/
oo,
Mesohaline
=
5­
18
o/
oo,
Polyhaline
=
more
than
18
o/
oo.
(
Taken
from
Dennison
et
al.
1993).

Light
Total
Dissolved
Dissolved
attenuation
suspended
inorganic
inorganic
Salinity
coefficient
solids
Chlorophyll
nitrogen
phosphorus
regime
(
K
d
m­
1)
(
mg/
l)
a(
ug/
l)
(
uM)
(
uM)

Tidal
freshwater
2.0
15
15
­
0.67
Oligohaline
2.0
15
15
­
0.67
Mesohaline
1.5
15
15
10
0.33
Polyhaline
1.5
15
15
10
0.67
18
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
greatly
influence
the
distribution
of
seagrass
species.
Disease
has
been
widely
implicated
in
the
loss
of
seagrass
beds
since
the
pan­
Atlantic
decline
in
the
1930'
s
(
Rasmussen
1973,
Short
et
al.
1987,
Muehlstein
1989).
Through
this
time,
seagrass
declines
attributed
to
disease
have
added
significantly
to
fluctuations
in
seagrass
distribution
Physical
disruption
from
storms
and
shifting
channels
redefine
seagrass
bed
distribution
and
composition.
Seasonal
disturbances,
such
as
low
tides
which
expose
and
desiccate
beds
(
Phillips
1980a,
Thayer
et
al.
1984),
and
catastrophic
events,
such
as
hurricanes
(
Eleuterius
and
Miller
1976,
Livingston
1987),
can
dramatically
restructure
seagrass
beds
both
in
terms
of
bed
size
and
seagrass
species
composition.
We
have
found
that
reductions
in
seagrass
bed
coverage
as
the
result
of
storms
is
a
positive
function
of
how
exposed
to
wind­
generated
waves
a
bed
is
prior
to
a
storm;
rapid
loss
of
coverage
can
occur
within
a
period
of
hours
(
unpubl.
data),
reiterating
the
fact
that
one­
time
surveys
of
seagrass
coverage
can
be
misleading
as
to
the
potential
distribution
of
seagrass
in
a
water
body.

Biological
disturbance
of
seagrass
beds
by
a
variety
of
organisms
can
also
be
extensive.
Overgrazing
by
herbivores
such
as
urchins
has
also
affected
spatial
distribution
and
standing
stock
of
seagrass
beds
(
Camp
et
al.
1973).
Ice
scour
(
Robertson
and
Mann
1984)
and
extreme
cold
(
Lalumiere
et
al.
1994)
have
been
shown
to
control
Z.
marina
distribution
in
the
sub­
Arctic.
Also,
excessive
epiphytic
load
(
Sand­
Jensen
1977),
burrowing
shrimp
(
Suchanek
1983),
vagile
macrofauna
(
Valentine
and
Heck
1990,
Valentine
et
al.
1994),
green
algae
(
den
Hartog
1994a),
and
lugworms
(
Philippart
1994)
have
all
been
shown
to
limit
seagrass
distribution
(
but
see
Reusch
et
al.
1994;
fertilizer
enhancement
of
eelgrass
by
blue
mussel
biodeposition).
Rays
too
have
been
implicated
in
many
seagrass
planting
failures
(
Merkel
1988a,
Mote
Marine
Laboratory
and
Mangrove
Systems
Inc.
1989,
Fonseca
et
al.
1994)
and
may
even
contribute
to
the
maintenance
of
natural
bed
patchiness
(
Townsend
and
Fonseca
in
1998).
These
are,
however,
natural
processes.
Similarly,
some
dieoffs
of
seagrass
such
as
the
"
wasting
disease"
of
the
eelgrass
(
Z.
marina)
in
the
North
Atlantic
during
the
1930'
s
(
Short
et
al.
1988)
and
the
current
demise
of
T.
testudinum
in
Florida
Bay
have
been
attributed
to
a
pathogenic
form
of
a
marine
slime
mold,
Labyrinthula
zosterae
(
Robblee
et
al.
1991),
among
other
factors.
In
nature,
however,
the
outbreak
of
this
fungi
has
not
been
easy
to
classify
as
a
cause
of
seagrass
decline
as
opposed
to
being
a
by­
product
of
some
other
environmentally­
or
anthropogenically
derived
decline
in
the
quality
of
the
seagrass
habitat
(
sensu
den
Hartog
1996).

When
human
impacts
are
added
to
the
natural
stresses
imposed
on
seagrass
beds,
disastrous
losses
of
seagrass
can
occur.
Such
losses
have
been
documented
in
Australia
(
Kirkman
1981,
Cambridge
and
McComb
1984)
and
southeast
Asia
(
Fortes
1988).
In
the
U.
S.,
large
scale
losses
have
been
documented
in
the
Chesapeake
Bay
(
Orth
Chapter
1:
Background
°
19
and
Moore
1981)
and
in
the
Gulf
of
Mexico
(
Livingston
1987).
Significant
impacts
to
seagrass
beds
in
Tampa
Bay
were
documented
by
Taylor
and
Saloman
(
1968),
eventually
reaching
over
50
percent
of
the
historical
seagrass
cover
in
Tampa
Bay
(
Haddad
1989).
Similarly,
35
percent
of
the
seagrass
acreage
in
Sarasota
Bay
has
been
lost
as
well
as
29
percent
of
that
in
Charlotte
Harbor,
Florida,
and
76
percent
of
that
in
Mississippi
Sound
(
Eleuterius
1987).
Pulich
and
White
(
1991)
reported
a
loss
of
90
percent
in
Galveston
Bay,
Texas.
Thom
and
Hallum
(
1991)
report
similar
ranges
of
losses
from
Puget
Sound.
Large
losses
of
seagrass
have
also
been
reported
from
San
Francisco
and
San
Diego
Bays
(
Kitting
and
Wyllie­
Echeverria
1992),
the
Laguna
Madre
(
brown
tide,
Onuf
1994),
and
large­
scale
damage
from
propeller
scarring
has
been
reported
in
Florida
(
Sargent
et
al.
1995).

Loss
of
seagrass
cover
leads
to
several
undesirable
and
difficult­
to­
reverse
conditions
First,
the
sediment
binding
and
water
motion
baffling
effects
of
the
plants
themselves
are
lost
(
Fonseca
et
al.
1983,
Fonseca
and
Fisher
1986)
allowing
sediments
to
be
more
readily
resuspended
and
moved
(
e.
g.,
Florida
Bay,
Thayer
et
al.
1994).
The
physical
ramifications
include
increased
shoreline
erosion
and
water
column
turbidity
Seagrass
planted
in
areas
with
these
conditions
may
not
survive
due
to
light
limitation
from
the
elevated
turbidity.
Loss
of
seagrass,
of
course,
eliminates
all
important
associated
habitat
functions
(
Kikuchi
1980,
Peterson
1982).

Much
of
the
documented
seagrass
loss
is
due
to
human­
induced
reductions
in
water
transparency
(
Kenworthy
and
Haunert
1991,
Bulthuis
1994;
these
losses
are
often
not
included
with
other
wetland
or
even
seagrass
loss
statistics).
Only
in
the
last
few
years
has
it
become
clear
that
seagrasses
typically
require
light
intensities
reaching
the
leaves
of
at
least
15­
25
percent
of
the
light
which
has
penetrated
to
just
beneath
the
water
surface
(
Dennison
and
Alberte
1986,
Gallegos
1994,
Gallegos
and
Kenworthy
1996).
Moreover,
the
length
of
time
over
which
a
seagrass
plant
spends
at
photosynthetically­
saturating
light
intensities
too
has
been
shown
to
be
correlated
with
growth
and
survival
(
Dennison
and
Alberte
1985,
1986,
Zimmerman
et
al.
1991).
However,
water
transparency
standards
have
historically
been
based
on
requirements
of
phytoplankton
which
may
need
only
~
1
percent
of
incident
light
(
Kenworthy
and
Haunert
1991),
meaning
that
there
is
often
no
legal
mandate
for
requiring
improvement
of
water
transparency
to
support
seagrasses.
This
absence
of
technical
and
legal
mandates
makes
the
task
of
demonstrating
the
need
for
restoration
of
water
quality
to
support
seagrasses
difficult.

There
are
many
factors
that
act
to
reduce
water
column
transparency
(
sensu
Dennison
1987,
Dennison
et
al.
1993,
Gallegos
1994,
Gallegos
and
Kenworthy
1996).
Excess
suspended
solids
and
nutrients
which
enter
the
water
column
as
the
20
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
result
of
poor
watershed
management
combine
to
reduce
transmitted
light
below
that
of
natural
fluctuations,
increasing
vulnerability
to
local
population
extinctions.
Suspended
solids
and
water
color
changes
reduce
water
transparency
directly.
Nutrient
additions,
such
as
from
septic
systems
(
Burkholder
et
al.
1992,
Short
and
Burdick
1996),
accelerate
growth
of
light­
absorbing
algae
in
the
water
column
as
well
as
benthic
macroalgae
(
den
Hartog
1994a,
b)
and
that
growing
epiphytically
on
seagrass
blades
(
Sand­
Jensen
1977),
all
of
which
combine
to
reduce
light
availability
to
seagrasses.
Moreover,
the
seagrass
canopy
has
intrinsic
light
attenuation
effects
through
mutual
shading
(
Dennison
1987,
Enriquez
et
al.
1992)
by
the
individual
plants.

When
losses
have
occurred
due
to
decreased
light
availability,
often
only
changes
in
watershed
management
(
such
as
controlling
storm
water
and
sewage
discharges)
can
reverse
the
trend
of
decline.
Such
a
reversal
in
decline
is
rare
but
has
occurred
(
Johansson
and
Lewis
1992).
Transplanting
into
areas
experiencing
seagrass
loss
due
to
decreased
water
transparency
without
independent
improvements
in
water
quality
will
only
result
in
the
death
of
the
transplants.
This
is
especially
problematic
in
areas
where
water
turbidity
may
be
due
to
sediment
resuspension
which
arises
as
a
result
of
seagrass
already
lost
and
is
therefore
not
necessarily
a
current
watershed
management
problem.

Reduction
in
water
transparency
is
not
the
only
anthropogenic
source
of
seagrass
loss
(
see
Phillips
1982
for
an
early,
detailed
review).
Thermal
effluents
from
electric
power
plants
have
caused
extensive
losses
such
as
those
documented
at
the
Turkey
Point
station
in
Biscayne
Bay,
Florida
(
Zieman
and
Wood
1975)
as
well
as
that
associated
with
the
Stock
Island
(
Key
West)
station
(
pers.
obs.).
In
the
past,
dredgeand
fill­
associated
losses
were
commonly
associated
with
private
sector
development
but
more
recently,
many
losses
can
be
ascribed
to
public
interest
projects,
such
as
the
replacement
of
the
Florida
Keys
Bridges
(
Mangrove
Systems
Inc.
1985a,
Thayer
et
al.
1985).
In
addition,
the
rapidly
increasing
number
of
small
boats
in
coastal
waters
has
resulted
in
the
aforementioned
widespread
damage
from
propeller
scarring
(
Sargent
et
al.
1995).
Because
of
the
chronic
nature
of
propeller
scarring,
hull
impacts,
and,
more
recently
jet
ski
scour,
such
damage
is
likely
very
difficult
to
repair
by
planting
(
e.
g.,
ferry
boat
landings
in
Puget
Sound,
R.
Thom,
Battelle
Pacific
Northwest
Lab.,
Sequim,
Wa.),
Sargent
et
al.
(
1995)
recommend
a
four­
point
plan
to
reduce
scarring
in
moderately
and
severely
scarred
meadows
(
defined
under
their
criterion)
which
includes
(
1)
education
of
the
public
as
to
the
nature
and
scope
of
scarring
impacts,
especially
in
the
Thalassia
testudinum
beds
which
are
very
slow
to
recover
from
impacts,
(
2)
installing
channel
markers
as
aids
to
navigation,
(
3)
enforcing
state
and
federal
statutes
that
address
propeller
scarring
and
caused
by
propulsion
systems
dredging,
and
(
4)
establishment
of
limited­
motoring
zones
in
areas
where,
due
to
the
extreme
shallowness
of
beds,
impacts
from
propulsion
systems
would
be
unavoidable.
Chapter
1:
Background
°
21
A
SHORT
HISTORY
OF
SEAGRASS
MITIGATION
AND
RESTORATION
Addy's
(
1947)
basic
logic
was
to
match
planting
and
harvest
site
environments,
and
this
remains
a
fundamental
tenet
in
almost
all
seagrass
planting
today.
Aside
from
early
interest
by
Phillips
(
1960),
almost
30
years
elapsed
before
serious
attention
to
planting
seagrass
developed.
It
was
not
until
Eleuterius
(
1975),
van
Breedveld
(
1975),
Thorhaug
(
1976),
and
Churchill
et
al.
(
1978)
that
documents
again
began
to
emerge
presenting
seagrass
planting
in
a
guideline
format.
But
even
though
suitable
planting
methods
have
long
existed,
the
track
record
for
successful
mitigation
of
impacts
to
seagrass
beds
remains
variable
(
see
review
by
Phillips
1982).
Some
spectacular
failures
of
seagrass
planting
(
Stein
1984)
have
created
a
lasting
impression
that
restoration
of
seagrass
beds
is
still
an
experimental
management
tool.
Yet
there
have
also
been
many
successful
plantings
(
e.
g.,
Thayer
et
al.
1985).
Seagrass
beds
have
often
been
successfully
planted
and
have
come
to
perform
much
as
naturally­
propagated
beds
(
Homziak
et
al.
1982,
McLaughlin
et
al.
1983,
Fonseca
et
al.
1996b).
Still
it
has
not
been
clear
what
factors
are
the
most
important
to
address
to
ensure
planting
success
We
had
previously
thought
that
seagrass
planting
was,
as
Ronald
Phillips
put
it,
"
a
two­
edged
sword"
(
R.
Phillips,
Battelle
Labs,
Richmond,
Wa.,
pers.
comm.),
providing
a
means
of
ameliorating
habitat
losses
but
perhaps
encouraging
habitat
destruction
through
the
mere
existence
of
a
possible
remedial
technique.
In
our
opinion
a
more
conservative
trend
has
emerged.
As
resource
managers
and
developers
have
become
educated
as
to
the
value
of
seagrass
systems
and
the
realities
of
their
costly
repair,
more
emphasis
appears
to
now
be
placed
on
impact
avoidance
and
minimization

Much
emphasis
was
placed
on
technique
development
in
the
late
1970s
and
early
1980s
(
see
reviews
by
Phillips
1980,
1982,
Lewis
1987,
Fonseca
et
al.
1988,
Thom
1990),
but
relatively
little
attention
was
given
to
developing
a
management
framework
within
which
these
techniques
could
be
effectively
implemented.
As
a
result,
most
seagrass
mitigation
projects
failed
to
achieve
the
goal
of
1:
1
habitat
replacement
(
i.
e.,
offset
a
net
loss
of
seagrass
habitat:
sensu
Fonseca
et
al.
1987c,
Fonseca
1989a,
but
see
Merkel
1988a,
b),
nor
have
they
consistently
addressed
whether
functional
equivalency
has
been
achieved
(
often
a
permit
requirement).

Phillips
(
1980b)
published
seagrass
planting
guidelines
that
relied
on
elevation
in
the
tidal
zone,
current
speed,
salinity,
soil
type
(
sandy,
combination,
or
cohesive)
and
seagrass
species.
Decision
keys
for
each
coast
of
the
U.
S.
were
compiled.
However,
with
additional
research
some
of
Phillips'
(
1980b)
threshold
criteria
should
be
changed.
He
accepted
current
speeds
up
to
1.82
m
s­
1
whereas
we
would
strongly
caution
against
planting
in
current
speeds
exceeding
0.5
m
s­
1
(
see
below).
Further,
22
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
Phillips
indicated
that
planting
in
sandy
sediments
was
a
cause
for
rejection
of
a
planting
site,
but
we
have
found
excellent
success
in
sandy
sediments
(
Lewis
1987,
Fonseca
et
al.
1987a,
b,
c).
Zimmerman
et
al.
(
1991)
argue
that
factors
increasing
root
and
rhizome
anoxia
such
as
cohesive
soils
recommended
by
Phillips
put
seagrass
(
at
least
when
using
bare­
root
planting
methods)
under
severe
physiological
stress,
a
factor
to
be
especially
avoided
during
planting
operations.
Similarly,
Merkel
(
1992)
recommended
planting
on
sandy
sediments
on
the
West
Coast
and
avoiding
consolidated
clays
and
mudstones
(
although
he
[
correctly]
noted
that
rhizome
extension
is
slower
in
coarse
sediments).
More
recently,
detailed
information
on
habitat
requirements
for
seagrass
(
and
other
submerged
aquatic
vegetation,
SAV)
has
emerged,
but
only
in
well­
studied
areas.
Notable
is
the
work
ongoing
in
the
Chesapeake
Bay.
Batiuk
et
al.
(
1992;
see
also
Dennison
et
al.
1993)
provide
a
detailed
synthesis
of
water
quality
requirements
for
SAV
(
Table
1.2).
Based
on
experimentation
and
strong
correlative
evidence
of
these
water
quality
parameters
and
SAV
distribution,
they
also
developed
a
series
of
target
water
quality
conditions
that
would
have
to
be
met
to
expand
SAV
distribution
by
allowing
it
to
colonize
greater
depths.
This
study
should
serve
as
a
model
approach
to
investigate
seagrass
restoration
efforts
in
other
areas.
The
applicability
of
these
data
to
other
areas
is
discussed
in
greater
detail
under
the
section
entitled
"
Light
Requirements
for
Transplanting."

Merkel
(
1992)
has
developed
a
field
manual
for
planting
eelgrass
on
the
West
Coast
that
includes
planning
protocols
and
detailed
guidance
on
planting
execution
that
is
otherwise
generally
lacking
in
the
literature.
Aspects
of
Merkel's
report
will
be
reviewed
throughout
this
document.
Fonseca
(
1989a,
1992)
published
what
were
essentially
Agency
checklists
for
planning
and
evaluating
seagrass
plantings;
the
design
of
those
checklists
were
the
basis
for
the
more
comprehensive,
yet
regionally­
specific
guidelines
published
later
(
Fonseca
1994).
The
planning,
planting,
and
monitoring
sections
of
this
document
were
adapted
from
Fonseca
1994:
"
A
Guide
to
Planting
Seagrasses
in
the
Gulf
of
Mexico."
Lockwood
(
1990)
published
criteria
for
placing
marinas
in
eelgrass
habitat
that
extolled
impact
minimization
as
the
only
guideline
for
mitigation.
Based
on
case
reviews
of
seagrass
mitigation
projects
(
Thayer
et
al.
1985),
Thayer
et
al.
(
1990)
published
a
preliminary
decision
matrix
that
incorporated
site
selection
criteria
as
well
as
environmental
conditions
required
for
the
growth
of
specific
seagrass
species.

In
general,
studies
of
seagrass
restoration
and
management
have
only
recently
become
a
focus
of
attention
(
e.
g.,
Chesapeake
Executive
Council
1989)
and
more
recently,
funding.
NOAA's
Coastal
Ocean
Program
has
focused
on
these
issues
for
both
seagrass
and
saltmarsh
through
its
Estuarine
Habitat
Program,
C­
CAP,
and
Decision
Analyses
Series.
In
conducting
our
study,
we
have
found
the
information
base
for
seagrass
management
difficult
to
locate.
For
example,
a
survey
of
published
Chapter
1:
Background
°
23
literature
since
1985
using
BIOSIS
 
revealed
that
there
were
655
published
works
on
seagrass.
This
search
of
the
open
literature
reveals
that
over
the
last
five
years
most
of
the
focus
in
seagrass
research
has
been
on
aspects
of
the
plant's
physiology.
This
is
typical
of
seagrass
research
over
the
last
quarter
century
where
interest
in
plant
physiology
and
seagrass
bed­
associated
fauna
have
dominated
the
open
literature.
Crossreferencing
"
seagrass"
with
"
restoration"
found
nine
references
while
"
mitigation"
provided
one
reference.
From
the
literature
we
accumulated
directly
from
journals
and
solicitation
of
colleagues,
we
found
that
approximately
half
was
found
outside
the
open
literature.
The
literature
on
the
subject
of
seagrass
bed
restoration
and
mitigation
is
found
in
the
grey
literature
and
is
often
not
subject
to
the
rigors
of
peer
review
(
but
see
Batiuk
et
al.
1992).
Another
large
body
of
information
lies
in
unpublished
project
reports,
the
quality
of
which
are
highly
variable.
We
feel
that
the
trend
to
generate
information
on
seagrass
restoration
and
mitigation
for
dissemination
in
forums
other
than
the
open
literature
has
been
one
of
the
major
reasons
that
seagrass
restoration
and
mitigation
is
perceived
as
an
experimental
tool,
when
it
could
be
an
established
management
practice.

What
are
the
problems
managers
face
in
restoring
seagrass
beds?
Chief
among
these
problems
is
the
tendency
to
plant
seagrass
in
areas
where
there
is
no
prior
history
of
their
existence
(
Fredette
et
al.
1985;
unless
of
course
the
site
was
created
for
the
purposes
of
planting
seagrass).
The
chronic
absence
of
seagrass
from
a
site,
especially
when
there
are
propagule
sources
nearby,
usually
indicates
that
the
site
cannot
consistently
support
seagrasses.
Ensuring
sufficient
light,
moderate
nutrient
loads
(
Batiuk
et
al.
1992,
Dennison
et
al.
1993,
Kenworthy
and
Fonseca
1996,
Short
and
Burdick
1996)
and
protecting
plantings
from
disturbance
are
major
considerations
for
developing
a
persistent
seagrass
bed.
Planting
stock
must
be
chosen
so
that
there
are
sufficient
young
shoots
and
growing
meristems
to
make
up
for
mortality,
a
ratio
that
changes
dramatically
depending
on
what
portion
of
a
seagrass
bed
is
examined,
the
species,
as
well
as
time
of
year.
Most
seagrasses
are
comparatively
short­
lived
and
have
high
natural
mortality
rates,
and
suitable
growing
conditions
are
needed
to
allow
new
shoot
generation
to
compensate
for
this
mortality.
Thus,
development
and
incorporation
of
seagrass
demographic
information
into
the
management
process
is
a
high
priority
area
for
research.
There
are
many
other
caveats
that
must
be
imposed
to
expect
successful
restoration
of
seagrass
beds.
These
will
be
discussed
later
both
in
general
terms
and
specifically
by
region
around
the
country.

Having
argued
that
seagrass
mitigation
is
no
longer
experimental
and
should
be
considered
an
established
management
tool,
why
then
place
such
a
priority
on
conservation
The
reason
is
that
while
techniques
and
protocols
exist
that
can
produce
persistent
seagrass
beds,
the
history
of
the
field
shows
that
guidance
and
protocols
are
often
inconsistently
applied.
This
has
resulted
in
spectacular
large­
scale
planting
fail­
24
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
ures
(
e.
g.,
the
aforementioned
Port
of
Miami
expansion
project:
a
multi­
million
dollar
~
200
acre
seagrass
mitigation
which
produced
only
a
few
acres;
Stein
1984).
The
fact
that
much
information
on
this
subject
is
conveyed
through
the
grey
literature,
which
does
not
always
circulate
widely,
has
resulted
in
repetititve
mistakes,
such
as
selection
of
inappropriate
planting
sites.

MANAGEMENT
CONTEXT
FOR
MITIGATION
AND
RESTORATION
OF
SEAGRASS
ECOSYSTEMS
Scientists
and
managers
are
always
faced
with
uncertainty
in
decisions
regarding
ecosystem
management.
As
pointed
out
by
Vitousek
(
1994)
for
global
environmental
change
issues,
scientists
know
with
certainty
that
changes
are
occurring
and
that
they
are
human­
caused.
What
scientists
cannot
do
is
always
predict
the
particular
consequences
of
a
given
human
activity
on
the
environment.
However,
some
trends
are
obvious
and
the
consequences
of
inaction
can
be
logically
derived.
It
is
irrefutable
that
extensive
loss
of
seagrass
resources
have
occurred
in
this
country
(
see
previous
section),
but
what
are
the
management
options
for
halting
and
reversing
this
decline?

We
have
compiled
this
synthesis
of
seagrass
restoration
in
an
attempt
to
identify
reasons
for
failures
and
successes
which
will
then
allow
managers
to
improve
the
odds
of
success
in
restoring
seagrass
ecosystems.
By
acting
to
mitigate,
restore
and
maintain
these
resources,
managers
can
offset
collateral
decline
of
many
ecological
functions
which
we
as
a
society
hold
important
(
erosion
control,
water
filtration,
fisheries
production,
and
associated
aesthetics).
However,
as
the
human
population
grows
it
is
highly
likely
that
losses
of
these
unique
plant
communities
will
continue
(
e.
g.,
Sargent
et
al.
1995).
There
are
no
ecological
substitutes
for
their
role
in
coastal
ecosystems.

The
critical
role
that
seagrasses
play
in
many
coastal
environments,
coupled
with
their
extensive
losses,
have
created
widespread
support
for
their
conservation
and
restoration.
The
"
no­
net­
loss"
policy
promulgated
by
the
Executive
Branch
provided
an
additional
impetus
to
consider
seagrass
conservation
and
restoration.
Meanwhile,
numerous
policy
changes
have
occurred
at
the
state
and
local
levels
over
the
last
ten
years
to
support
no­
net­
loss
of
habitat.
Therefore,
as
an
informationbased
system
of
judging
the
value
of
seagrass
ecosystems
has
emerged
over
the
last
decade,
the
question
is
no
longer
whether
seagrasses
should
be
protected,
but
how?
When
all
avenues
of
protection
have
failed
(
e.
g.,
sequencing;
the
US
Army
Corps
of
Engineers­
EPA
sequence
of
first
seeking
impact
avoidance
and
minimization,
and
then
compensatory
mitigation,
the
latter
being
composed
of
some
combination
of
enhancement,
restoration,
creation
and
under
rare
circumstances,
simply
preservation
then
active
planting
may
be
the
only
option
to
avoid
a
permanent
net
loss
of
seagrass.

In
order
to
proceed
with
discussions
of
management
issues,
some
terminological
clarification
is
needed.
We
will
utilize
the
terminologies
of
Fonseca
(
1994)
which
are
reprinted
in
amended
form
in
Appendix
A.
Particularly,
we
wish
to
draw
the
reader's
attention
to
the
differentiation
among
the
terms
"
restoration"
and
"
mitigation
They
are
not
interchangeable
terms.
Mitigation
refers
to
activities
related
to
permits
(
particularly
sec.
404
of
the
Clean
Water
Act)
and
embodies
a
sequence
of
avoidance,
minimization
and
ultimately,
if
needed,
compensatory
mitigation,
whereas
restoration
is
simply
returning
a
site
to
a
previous
condition.
Restoration
as
used
here
does
not
apply
to
permit­
associated
planting
projects.
We
will
also
differentiate
the
terms
"
transplanting"
and
"
planting."
Transplanting
is
a
subset
of
planting
in
that
here
it
refers
to
harvesting
of
existing
plants
whereas
planting
can
involve
cultured
plants,
seeding,
or
any
number
of
methods.
The
terms
restoration
and
mitigation
set
very
different
constraints
on
the
establishment
of
performance
criteria
and
the
evaluation
of
compliance
(
i.
e.,
success).
Lewis
(
1989)
defines
and
differentiates
restoration
and
mitigation
as
follows:

RESTORATION
 
"
Returned
from
a
disturbed
or
totally
altered
condition
to
a
previously
existing
natural,
or
altered
condition
by
some
action.
Restoration
refers
to
the
return
of
a
pre­
existing
condition."

MITIGATION
 
"...
the
actual
restoration,
creation,
or
enhancement
of
(
functionally
equivalent,
authors'
note)
wetlands
to
compensate
for
permitted
wetland
losses."

The
term
"
mitigation"
can
be
used
without
any
modifiers
but
is
often
applied
to
situations
more
aptly
termed
"
compensatory
mitigation."
Restoration
is
a
term
which
generally
applies
only
to
planting
activities
which
are
not
being
counted
against
the
destruction
of
existing
habitat.
Rather,
restoration
embraces
the
concept
that
anything
we
can
do
to
right
a
past
loss,
a
loss
for
which
there
may
be
no
litigative
recourse
to
seek
damage
recovery,
is
a
plus
to
set
against
the
Nation's
balance
sheet
for
no
net
loss,
but
not
against
that
of
a
project
with
a
pending
permit
to
eliminate
seagrass.
From
a
management
perspective,
restoration
for
the
sake
of
restoration
only
(
properly
planned
and
professionally
executed),
should
be
vigorously
pursued
because
it
will,
if
one
utilizes
the
above
definition,
bring
a
community
back
toward
previously
existing
conditions
(
i.
e.,
it
generally
cannot
make
the
situation
worse).
Chapter
1:
Background
°
25
PITFALLS
IN
THE
MITIGATION
AND
RESTORATION
PROCESS
Compensatory
mitigation
is
a
process
of
questionable
merit
(
Race
and
Fonseca
1996).
Unlike
restoration
projects
which
are
not
necessarily
initiated
under
the
404
permit
process,
the
circumstances
under
which
a
compensatory
mitigation
are
initiated
have
a
large
potential
to
make
matters
worse,
because
compensatory
mitigation
usually
involves
the
destruction
of
existing
habitat.
The
existing
habitat
is
or
has
been
traded
for
the
promise
of
replacement
habitat.
With
restoration,
we
are
dealing
with
a
past
loss
for
which
the
responsible
party
may
or
may
not
be
identifiable.
With
compensatory
mitigation,
the
agent
of
loss
and
the
responsible
party
are
known
and
sometimes
a
decision
(
likely
controversial)
might
be
made
to
trade
existing
habitat
for
replacement
habitat.
Of
course
when
a
injury
occurs
to
a
seagrass
bed
outside
of
the
permit
process
the
loss
of
seagrass
habitat
occurred
without
a
secure
means
of
mitigating
for
its
loss.
However,
whether
an
injury
is
deliberate
or
not,
if
existing
habitat
is
lost,
an
often
tangled
negotiation
process
follows
to
determine
the
means
by
which
compensation
for
that
loss
will
be
made.
In
many
instances,
the
negotiation
process
can
be
prolonged,
delaying
restoration
and
resulting
in
larger
impacts
than
might
occur
if
restoration
had
begun
sooner.

There
are,
however,
a
number
of
management
decisions
that
can
be
made
within
the
permit
process
to
ameliorate
a
loss
in
habitat
and
better
approaches
the
goal
of
no­
net­
habitat­
loss.
Mitigation
in
its
broader
definition
typically
also
includes
impact
avoidance
and
minimization
(
the
latter
term
unfortunately
implying
an
acceptable
net
loss
of
acreage).
In
practice,
avoidance
and
minimization
are
sometimes
difficult
to
achieve.
The
existence
of
techniques
to
transplant
seagrass
has
often
been
used
to
justify
the
destruction
of
existing,
productive
habitat
(
pers.
obs.).
But
as
pointed
out
earlier,
this
approach
has
consistently
produced
a
net
loss
of
habitat.
This
net
loss
of
habitat
occurs
for
a
number
of
reasons,
and
the
permit­
associated
activities
that
destroy
seagrass
beds
in
the
first
place
typically
are
long
lasting
(
i.
e.,
creation
of
channels
bridges,
bulkheads).
Those
activities
also
often
do
not
allow
enough
area
for
onsite
planting
to
offset
the
loss
of
habitat.
If
planting
is
considered
at
a
location
not
on
the
original
impact
site
(
off­
site
restoration
or
mitigation),
that
site
would
preferably
not
be
an
area
that
itself
has
lost
seagrass
to
some
other
impact.
This
is
a
subtle
point
that
is
often
overlooked
because
of
the
often
costly
(
in
time
and
effort)
site
history
data
that
must
be
obtained
to
make
a
quantitative
evaluation
of
no­
net­
loss.
The
problem
works
like
this:
if
one
permits
a
loss
of
seagrass
for
some
form
of
coastal
development
(
e.
g.,
­
1
acre)
and
plants
an
equivalent
area
(+
1
acre)
onto
a
site
which
had
previously
lost
seagrass
(
e.
g.,
­
1
acre)
but
was
not
associated
with
the
project
at
26
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
Chapter
1:
Background
°
27
hand,
then
the
net
change
in
habitat
is:
(­
1
+
­
1)
+
1
=
­
1
acre.
All
that
was
accomplished
was
the
repair
of
the
original
problem
at
the
planting
site,
but
it
does
not
address
the
loss
at
the
new,
most
recently
impacted
site.
While
there
would
be
no
net
loss
from
immediate,
present
day
acreage,
the
lack
of
consideration
of
past
losses
results
in
a
net
loss
on
a
recent
historical
time
scale.
The
critical
question
here
is
at
what
point
in
the
past
do
we
choose
to
represent
baseline
seagrass
acreage?
Moreover,
what
if
a
site
chosen
for
planting
does
not
currently
support
seagrass?
In
the
absence
of
site
history
information,
one
must
then
ask
why
it
does
not
presently
support
seagrass.
This
often
indicates
some
inherent
difficulty
in
colonization
or
persistence
of
seagrass.
The
events
influencing
the
colonization
process
are
sometimes
difficult
to
document
because
they
are
often
aperiodic,
acute
events
(
e.
g.,
extreme
low
tides,
storms,
migrating
rays
excavating
the
bottom).
Naturally
unvegetated
seafloor
should
not
be
substituted
for
vegetated
bottom
as
this
typically
creates
only
a
transient
seagrass
bed
and
alters,
not
necessarily
improves,
existing
habitat
functions.
The
take­
home
message
is
that
if
one
contemplates
off­
site
compensatory
mitigation,
there
are
usually
few,
if
any
sites
available
that:
(
a)
can
support
seagrass
growth,
and
if
they
do;
(
b)
do
not
involve
habitat
substitution;
or
(
c)
do
not
satisfy
the
no­
net­
loss
goal.
This
is
not
to
say
that
previously
damaged
sites
should
never
be
used
for
mitigation
or
restoration,
they
just
must
be
accurately
represented
in
any
no­
net­
loss
accounting.
As
pointed
out
by
Short
(
Jackson
Estuarine
Lab.,
Durham,
N.
H.,
pers.
com.)
in
reference
to
the
above
description
of
trade­
offs,
if
no
mitigation
is
done
on
a
previously
damaged
site,
one
ends
with
a
­
2
acre
net
loss
of
habitat
instead
of
­
1
acre
of
loss.

REGIONAL
BREAKDOWN
OF
PERMIT
ACTIVITIES
DEALING
WITH
SEAGRASS
MITIGATION
Under
a
Memorandum
of
Understanding
with
the
U.
S.
Army
Corps
of
Engineers,
the
National
Marine
Fisheries
Service,
Office
of
Habitat
Protection,
comments
on
development
permit
requests
under
Section
404
of
the
Clean
Water
Act
and
Section
10
of
the
Rivers
and
Harbors
Act.
While
seagrass
restoration
has
been
conducted
on
an
experimental
scale
along
all
coasts
and
within
all
coastal
regions
of
the
U.
S.,
actual
mitigation
of
impacts
resulting
from
Corps
of
Engineers­
permitted
activities
has
been
relatively
small,
and
has
been
greatest
in
the
NMFS
Southeast
and
Southwest
Regions.
A
summary
of
NMFS­
recommended
and
acted
upon
mitigation
actions
by
NMFS
Region,
based
on
reports
received
from
NMFS
Regional
Offices
as
of
early
1996,
is
provided
below
(
note
that
these
regions
do
not
match
the
ecoregions
described
later
in
the
text).
28
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
NORTHEAST
(
NE)
REGION
(
Maine
through
Virginia)

Seagrass
mitigation
in
the
Northeast
Region
of
NMFS
is
in
its
infancy
and,
while
permits
have
been
reviewed
which
deal
with
seagrass
habitat,
few
actions
are
ongoing.
In
1991
NMFS
began
recommending
seagrass
mitigation
for
projects
without
practicable
and
feasible
alternatives
that
would
damage
seagrass
habitat.
At
the
time
of
this
report
mitigation
actions
have
been
considered
in
New
Jersey,
Maine,
and
New
Hampshire,
but
site
selection
and
test
planting
for
a
3­
acre
mitigation
in
the
Piscataqua
River
(
N.
H.)
is
the
only
ongoing
permit­
related
mitigation
which
NMFS
has
been
involved
in
making
recommendations.
This
has
included
not
only
transplanting
but
also
consideration
of
alteration
of
bottom
topography
to
achieve
appropriate
planting
depths
for
eelgrass.
Proposals
are
currently
being
discussed
for
a
10­
30
acre
eelgrass
mitigation
in
the
upper
Penobscot
Bay
(
Maine).

In
addition
to
supporting
the
experimental
transplanting
work
that
is
ongoing
in
each
of
the
NE
states,
the
NMFS
Regional
Office
has
taken
a
proactive
approach
to
seagrass
habitat
protection.
This
has
included
involvement
in
the
development
of
seagrass
management
policies,
development
of
seagrass
survey
guidelines,
encouragement
for
interagency
mapping
of
seagrasses
including
involvement
of
the
NOAA
Coastal
Change
Analysis
Program
(
C­
CAP)
mapping
efforts,
and
the
convening
of
information
transfer
and
education
meetings
for
state
and
federal
agencies
on
seagrass
ecology
and
transplanting
technology.

SOUTHEAST
(
SE)
REGION
(
North
Carolina
through
Texas
including
the
U.
S.
Virgin
Islands
and
Puerto
Rico)

In
the
SE
Region
Habitat
Protection
Offices
reported
seagrass
mitigation
in
North
Carolina,
Florida,
and
Texas.
In
each
state
the
permits
were
obtained
or
under
consideration
primarily
for
channel
maintenance
or
development
related
to
onshore
construction.
In
Texas,
however,
permit
activities
relate
primarily
to
petroleum
pipeline
construction
and
mitigation
of
illegal
prop­
dredging
activities.
With
the
exception
of
Texas
there
has
been
little
or
no
monitoring
or
follow­
up
to
assess
the
degree
of
success
of
the
projects.
In
addition
to
permit­
related
activities
noted
below,
field
offices
of
the
NMFS
have
participated
in
similar
management
activities
noted
for
the
NE
as
a
means
of
educating
state
and
federal
agencies
and
potential
developers
of
the
ecology
and
sensitivity
of
seagrass
species
and
habitats.

Between
1985
and
1994
the
Habitat
Protection
Field
Office
in
Beaufort,
North
Carolina,
recommended
seagrass
mitigation
on
5
permits.
The
direct
seagrass
dam­
age
(
i.
e.,
removal
of
habitat)
ranged
from
0.23
to
2.0
acres,
and
the
ratio
of
seagrass
planted
to
that
lost
ranged
from
1:
1
to
3:
1.
However,
in
one
case
there
was
no
onsite
or
in­
kind
alternative,
and
oyster
reef
creation
was
accepted
as
an
alternative.
The
seagrasses
involved
were
Zostera
marina
and
Halodule
wrightii
primarily,
but
Ruppia
maritima
also
was
recommended
to
be
transplanted
in
one
instance.
During
the
9­
year
period
between
1985
and
1994,
a
total
of
3.25
acres
of
seagrass
habitat
were
permitted
to
be
destroyed
with
a
requested
mitigation
of
4.74
acres
of
seagrass
transplantation
Evaluation
of
the
mitigation
sites
has
been
carried
out
in
two
cases,
one
demonstrating
success
and
one
demonstrating
failure.

Florida
has
the
largest
extent
of
seagrasses
in
the
contiguous
U.
S.,
followed
by
North
Carolina
and
Texas
(
see
earlier
discussions).
Between
1978
and
1994
a
total
of
167
acres
of
seagrass
habitat
have
been
requested
for
mitigation
by
the
NMFS
Habitat
Protection
Field
Office
in
Panama
City,
Florida.
These
permit
requests
have
generally
been
the
result
of
new
channel
construction
and
port
development
and
have
ranged
in
mitigation
acreage
from
0.09
to
~
200
acres.
This
latter
was
the
result
of
a
permit
for
additional
development
of
the
Port
of
Miami.
Thalassia
testudinum,
Halodule
wrightii,
and
Halophila
engelmanii
have
been
involved
in
the
recommended
mitigation.
Based
on
reports
from
the
Panama
City
Field
Office,
the
degree
of
success
of
these
permit­
related
mitigation
has
been
generally
poor
and
in
many
cases,
unknown.

The
Galveston,
Texas
Field
Office
of
the
NMFS
Habitat
Protection
Division
reported
that
there
have
been
6
major
seagrass
mitigation
activities,
almost
all
in
the
Laguna
Madre,
between
1985­
1991.
A
total
of
107
acres
of
seagrass
habitat
have
either
been
recommended
for
creation
or
restoration.
These
have
included
filling
of
unused
pipeline
channels
and
associated
re­
contouring
of
the
bathymetry
to
downgrading
of
dredge
material
islands.
In
some
instances,
natural
recovery
of
the
site(
s)
has
been
recommended
while
in
others
transplanting
has
occurred.
The
species
involved
in
natural
recovery
have
been
Halodule
wrightii,
Halophila
engelmanni,
and
Ruppia
maritima,
whereas
Halodule
has
been
the
species
of
transplant
choice.
In
most
instances,
oil
companies
have
hired
private
concerns
to
monitor
the
mitigation
sites
or
staff
from
the
Galveston
Field
Office
have
had
the
opportunity
to
visit
the
mitigation
sites.
It
appears
that
site
selection
and
proper
bathymetric
contouring
has
occurred
because
the
Field
Office
reports
that
with
the
exception
of
22
acres,
there
has
been
a
mitigation
site
coverage
by
seagrasses
of
between
40­
99
percent
within
a
3
year
period
by
either
natural
or
transplanted
methods.
Some
planted
seagrass
sites
in
Florida
and
Texas
are
currently
being
evaluated
by
National
Marine
Fisheries
Service
staff
for
seagrass
and
faunal
recovery.
Chapter
1:
Background
°
29
NORTHWEST
(
NW)
(
Oregon,
Washington)
AND
ALASKA
While
research
on
experimental
restoration
approaches
have
been
or
are
being
carried
out
in
these
two
NMFS
Regions,
both
Regional
Offices
have
been
involved
to
only
a
very
limited
degree
in
seagrass
mitigation
and
restoration.
For
a
summary
of
eelgrass
transplanting
projects
in
the
Pacific
northwest
see
Thom
(
1990).

SOUTHWEST
(
SW)
(
California,
Hawaii,
and
Pacific
Territories)

Similar
to
other
field
and
Regional
Offices
in
the
SE
and
NE,
the
Southwest
Regional
Office
has
participated
in
seagrass
habitat
management
at
both
the
permit
as
well
as
research
and
educational
levels.
They
have
held
state­
federal
seminars
involving
the
scientific
community
in
discussions
on
the
ecological
value,
sensitivity,
and
restoration
of
seagrasses.
In
1991
an
eelgrass
mitigation
policy
was
drawn
up
and
adopted
by
NMFS,
the
U.
S.
Fish
and
Wildlife
Service
and
the
California
Department
of
Fish
and
Game
that
includes
recommended
transplanting
approaches,
monitoring
approaches,
and
measures
of
success
that
should
be
considered
(
see
local
evidence
for
seagrass
function
in
Hoffman
1986).

From
1976
through
1993
the
SW
Region
recommended
eelgrass
mitigation
on
25
permits
in
California
while
2
were
recommended
for
Enhalus
acoroides
and
Halodule
uninervis
on
Rota
Island
and
Saipan
Island
in
the
Pacific
Territories.
With
the
exception
of
6
permits,
most
mitigation
projects
have
not
exceeded
0.1
hectare;
the
remaining
6
ranged
from
0.8­
3.8
hectares.
Twenty
sites
have
been
visited
where
the
mitigation
activity
had
been
completed,
11
of
which
are
considered
a
success
by
Regional
Office
staff
while
4
have
shown
a
continued
decrease
in
seagrass
coverage
and
the
remainder
have
shown
no
change
in
coverage.
Overall,
the
success
rate
of
seagrass
planting
in
this
region
has
been
high
(
Hoffman
pers.
com.).

COMPARATIVE
ANALYSIS
OF
SEAGRASS
PLANTING
EFFORTS
At
this
time
we
are
not
aware
of
any
previous
analysis
of
seagrass
planting
effort
in
the
U.
S.
that
used
a
comparative
method.
Therefore,
we
documented
the
status
of
seagrass
planting
projects
from
around
the
country
by
soliciting
information
on
planting
activities
from
many
individuals
of
whom
we
were
aware
had
conducted
seagrass
plantings.
In
addition,
we
requested
that
all
National
Marine
Fisheries
Service
Regional
Offices
provide
us
with
listings
of
all
seagrass
mitigation
projects
for
which
they
had
reviewed
permits
under
their
statutory
authority.
We
also
conducted
site
visits,
especially
on
the
West
Coast
where
we
were
less
familiar
with
plant­
30
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
Chapter
1:
Background
°
31
ing
activities
in
order
to
collect
additional
planting
information.
Finally
we
then
compiled
all
references
on
this
subject
which
we
could
acquire;
this
included
review
of
all
literature
cited
by
reports
and
papers
we
collected.

This
is
not
a
complete
survey
and
is
complete
only
through
1995.
We
undoubtedly
have
missed
some
individuals
and/
or
planting
projects.
Some
persons
did
not
respond
to
our
queries.
Again
the
absence
of
this
work
from
the
peer­
reviewed
literature
made
it
difficult
to
find
the
information.
The
value
of
this
survey
then
is
heuristic,
but
addresses
questions
such
as
"
where
has
effort
generally
been
expended
What
data
have
been
collected?
What
techniques
have
been
used?
How
were
sites
selected
and
how
was
compliance
and/
or
performance
of
plantings
determined?
How
consistently
has
planting
technology
been
applied?

We
also
broke
down
the
survey
by
ecoregions
which
we
have
defined
for
the
purpose
of
isolating
practices
and
caveats
peculiar
to
different
parts
of
the
country.
Ecoregion
is
also
the
basis
for
the
creation
of
modules
where
recommendations
for
planning,
planting
and
monitoring
are
specifically
discussed
for
each
ecoregion.
In
addition,
our
original
intent
was
to
collect
information
on
coverage
rates
and
shoot
addition
of
individual
planting
units
(
PU)
from
around
the
country.
Any
differences
in
species'
coverage
and
shoot
addition
rates
would
aid
in
the
definition
of
ecological
regions
for
management.
However,
as
our
information
collection
progressed,
it
became
clear
that
there
were
insufficient
data
from
most
parts
of
the
country
to
conduct
these
coverage
and
shoot
rate
change
analyses.
Therefore,
we
have
divided
the
coastal
regions
of
the
country
based
on
our
knowledge
of
growing
season.
The
ecoregions
for
this
report
are
as
follows:

NORTHEAST
 
Maine
through
New
Jersey:
known
species
present
=
Zostera
marina
and
Ruppia
maritima.

MID­
ATLANTIC
 
Delaware
through
North
Carolina:
known
species
present
=
Halodule
wrightii,
Ruppia
maritima
and
Zostera
marina.

GULF
OF
MEXICO
AND
THE
FLORIDA
EAST
COAST
 
Mexico
to
Cape
Sable
and
north
of
Jupiter
Inlet
to
Cape
Canaveral:
known
species
present
=
Halodule
wrightii,
Halophila
decipiens,
Halophila
engelmanni,
Halophila
johnsonni,
Ruppia
maritima,
Syringodium
filiforme,
and
Thalassia
testudinum.

SOUTH
FLORIDA
AND
THE
CARIBBEAN
 
South
of
Jupiter
Inlet
to
Cape
Sable
and
P.
R.
and
USVI:
known
species
present
=
Halodule
wrightii,
Halophila
decipiens,
Halophila
engelmanni,
Halophila
johnsonni,
Ruppia
maritima,
Syringodium
filiforme,
and
Thalassia
testudinum.
32
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
CONTERMINOUSWEST
COAST
 
California
to
Washington:
known
species
present
=
Phyllospadix
scouleri,
Phyllospadix
serralatus,
Phyllospadix
torreyi,
Ruppia
maritima,
Zostera
japonica,
Zostera
marina.

ALASKA
 
Zostera
marina
and
Phyllospadix
spp.(
at
least
P.
serralatus).

HAWAII
AND
PACIFIC
TERRITORIES
 
known
species
present
=
Halophila
hawaiiana
(
K.
Bridges,
pers.
com.).

We
compiled
a
collection
of
138
documents
ranging
from
published,
peerreviewed
papers
to
project
reports.
Some
of
these
documents
were
reviews
or
guidelines
of
how
to
transplant
seagrass;
some
were
feasibility
studies;
some
were
laboratory
or
mesocosm
experiments
directed
at
enhancing
transplant
technology.
Each
document
was
categorized
several
ways.
We
first
determined
where
the
document
originated
Roughly
46
percent
of
the
documents
were
found
in
the
white
literature,
29
percent
were
unpublished
reports,
22
percent
in
grey
literature
and
~
3
percent
were
theses
(
Table
1.3).
All
together,
these
papers
reported
on
the
fate
of
over
686,000
planting
units
of
seagrass,
totaling
~
78
ha
of
field
acreage,
that
have
been
monitored.

Over
time
the
publication
rate
of
documents
concerning
seagrass
planting
have
increased.
We
found
less
than
1
percent
of
the
documents
published
prior
the
1960s.
In
the
1960s
we
found
2
percent
of
the
documents;
in
the
70s
21
percent;
in
the
80s
46
percent;
and
so
far
in
the
90s,
28
percent
of
the
documents.
At
this
rate
the
1990s
will
produce
the
greatest
amount
of
documents
on
the
subject
of
seagrass
planting.
Some
of
this
increase
in
publication
rate
may
be
that
more
recently
created
docu­

Table
1.3.
Percent
of
documents
on
seagrass
planting
compiled
by
literature
type.

Literature
Type
Percent
of
Documents
of
this
Type
White
Literature
46
Report
29
Gray
Literature
22
Theses
3
White
literature
=
peer
reviewed
journal
articles.
Report
=
not
peer
reviewed.
Gray
literature
=
not
in
a
library
circulated
journal,
may
or
may
not
be
peer
reviewed.
Theses
=
masters
thesis
or
doctoral
dissertation.
Chapter
1:
Background
°
33
ments
are
easier
to
locate,
but
it
seems
more
likely
that
interest
in
the
subject
has
grown.

The
purpose
of
the
documents
varied
widely.
The
largest
group
was
fieldresearch
oriented
which
comprised
~
57
percent
of
the
total.
Next
were
review
documents
(
29
percent),
followed
by
laboratory
experiments
(~
10
percent),
and
a
miscellaneous
group
(
6
percent),
which
included
feasibility
assessments,
economic
analyses
project
summaries,
and
recovery
assessments.
In
addition,
there
were
three
planting
associated
theses.
Most
laboratory
experiments
and
review
documents
were
published
in
the
peer­
reviewed
literature,
while
only
half
of
the
documents
presenting
new
planting
data
were
in
the
peer­
reviewed
literature.

Of
the
ecoregions
we
constructed,
most
documents
originated
from
either
the
West
Coast
(~
26
percent)
or
the
Gulf
of
Mexico
(
also
~
26
percent)
(
Table
1.4).
South
Florida
and
U.
S.
Caribbean
territories
produced
~
19
percent,
mid­
Atlantic
region
~
18
percent,
the
northeast
U.
S.
~
9
percent,
and
Alaska
~
2
percent.
Studies
from
other
countries
(
Australia,
France,
Great
Britain,
Italy)
were
also
reviewed
but
not
utilized
in
computation
of
summary
statistics.

The
greatest
number
of
planting
units
have
been
installed
in
the
South
Florida
ecoregion
(
Table
1.5),
followed
by
the
northeast,
West
Coast,
and
mid­
Atlantic
states
(
the
latter
three
regions
being
almost
equal
in
number
of
planting
units
install),
the
Gulf
of
Mexico,
and
lastly,
Alaska.

Table
1.4.
Percentage
of
complied
documents
on
seagrass
planting
presenting
field
transplanting
studies,
listed
by
ecoregion.
Values
are
percent
of
total.
Table
does
not
include
studies
from
outside
the
U.
S.,
guidelines,
reviews,
or
studies
involving
freshwater
plantings.
See
section
on
"
Regional
Breakdown
of
Permit
Activities
Dealing
with
Seagrass
Mitigation,"
above,
for
regional
boundaries.

Region
Percent
of
Documents
Found
in
Region
West
26
Gulf
25
South
Florida
19
Mid­
Atlantic
18
Northeast
9
Alaska
2
34
Table
1.5.
Reported
area
of
planted
seagrass
in
square
meters
and
number
of
planting
units
deployed
in
field
studies
by
region
and
species.

Region
Species
Area
M2
No.
PUs
ALASKA
Zostera
marina
?
40
GULF
Cymodosa
manitoruma
?
150
Halodule
beaudetteib
?
150
Halodule
wrightii
8,421
17,956
Ruppia
maritima
1
36
Syringodium
filiforme
591
2,336
Thalassia
testudinum
735
1,087
Zostera
marina
2,025
5,000
MID­
ATLANTIC
Halodule
wrightii
2,442
3,924
Ruppia
maritima
56
450
Zostera
marina
63,987
26,960
NORTHEAST
Zostera
marina
18,449
82,560
SOUTH
FLORIDA
Halodule
wrightii
227,639
161,503
Syringodium
filiforme
17,417
20,364
Thalassia
testudinum
332,770
332,239
WEST
Phyllospadix
torreyi
?
300
Zostera
marina
102,395
31,262
Alaska
=
entire
coast
of
Alaska
(
Ak.)
Gulf
Coast
=
Gulf
of
Mexico
to
Cape
Sable,
Fl.
and
the
Florida
East
Coast
North
of
Jupiter
Inlet
to
Cape
Canaveral
(
Tex.,
La.,
Miss.,
Ala.,
Fl.)
Mid­
Atlantic
=
Delmarva
Peninsula
to
North
Carolina
(
Del.
Va.,
Md.,
N.
C.)
Northeast
=
Maine
to
New
Jersey
(
Maine,
R.
I.,
N.
H.,
Mass.,
Conn.
N.
Y.,
N.
J.)
South
Florida
=
South
of
Jupiter
Inlet
to
Cape
Sable,
Puerto
Rico
and
the
U.
S.
Virgin
Islands
(
Fl.,
P.
R.,
U.
S.
V.
I.)
West
=
Washington
to
California
(
Wa.,
Ore.,
Calif.).

aProbably
Syringodium
filiforme.
bProbably
Halodule
wrightii.

PU
=
planting
units.
?
=
Insufficient
data
to
calculate
the
area.
The
use
of
different
planting
methods
by
ecoregion
and
seagrass
species
was
also
evaluated
(
Table
1.6).
We
constructed
fourteen
categories
of
planting
methods,
one
of
which
was
an
"
other"
category
that
contained
a
number
of
methods
not
widely
used
and
includes
studies
[
a
category
of
"
unknown"]
where
the
method
of
planting
was
not
described.
Of
the
fourteen
categories,
plugs
or
staples
were
the
most
common
~
40
percent
of
the
plantings
were
done
using
one
of
these
methods.
The
next
most
common
was
bare
root­
unanchored
sprigs
(
15
percent),
anchors
of
some
sort
(
8
percent),
followed
by
turfs
(
7
percent)
and
peatpots,
biodegradable
mesh,
seedlings
and
seeds
(
all
at
~
5­
6
percent
each).
Unusual
or
unknown
methods
accounted
for
[
were
employed
in]
~
2
percent
of
the
plantings.
Grids,
seed
tapes,
bagged
plants
and
attachment
to
boulders,
with
and
without
mesh
grids,
and
passive
seagrass
fragment
capture
were
used
in
the
remaining
~
4
percent
of
the
plantings.
The
Gulf
of
Mexico
ecoregion
had
the
greatest
number
of
planting
categories
(
11),
followed
by
the
West
Coast
(
10),
south
Florida
(
7),
and
the
mid­
Atlantic
states
(
5).

We
also
compiled
the
frequency
of
planting
methods
used
by
seagrass
species
(
Table
1.7).
Thalassia
testudinum,
Zostera
marina
and
Phyllospadix
spp.
have
been
transplanted
mostly
using
techniques
that
involve
removal
of
the
native
sediment
from
the
root­
rhizome
matrix
(
in
the
case
of
Phyllospadix,
there
may
have
been
no
sediment
to
remove
in
the
first
place).
The
remaining
three
species
listed
in
Table
1.7
have
been
transplanted
using
sediment­
free
and
sediment­
included
methods
in
about
equal
proportion.
Three
species,
H.
wrightii,
T.
testudinum,
and
Z.
marina
accounted
for
95
percent
of
the
planting
units
put
in
the
bottom
(
26,
21,
and
48
percent,
respectively
S.
filiforme
composed
the
remaining
3
percent
of
the
PU
while
two
other
species
composed
~
0.00013
percent
of
the
total
number
(
one
paper
reported
Halodule
beaudetti
and
Cymodocea
nodosa
as
occurring
in
the
Gulf
of
Mexico
but
we
suspect
these
were
either
H.
wrightii
and/
or
Ruppia
maritima).
Acreage
of
planting
by
species
closely
followed
percentages
for
PU
(
Table
1.9).
Some
seagrass
species
that
have
broad
distribution
have
received
comparatively
little
attention
to
that
given
Halodule,
Thalassia
and
Zostera.
For
example,
few
studies
have
been
done
regarding
Phyllospadix
spp.
planting
(
Phillips
et
al.
1992),
and
these
involve
attachment
to
large
rocks.
Aside
from
Phillips
et
al.
(
1992),
little
else
is
known
regarding
Phyllospadix
spp.
planting
techniques
even
though
this
species
ranges
along
the
entire
U.
S.
West
Coast
(
Phillips
1979,
Wyllie­
Echeverria
and
Phillips
1994).
Turner
(
1985)
provided
important
data
regarding
inherent
stability
and
recovery
of
natural
stands
that
have
at
least
heuristic
value
for
restoration
in
that
the
dynamic
aspect
of
the
community
can
be
recognized
and
incorporated
into
planning
(
see
Chapter
2,
Planning).
Similarly
Ruppia
maritima,
which
occurs
in
every
ecoregion,
and
Halophila,
of
which
there
may
be
(
based
on
an
incomplete
survey)
half
a
million
hectares
off
the
West
Coast
of
Florida
alone
(
Iverson
and
Bittaker
1986),
have
received
virtually
no
study
as
to
their
Chapter
1:
Background
°
35
36
Table
1.6.
Percentage
of
all
transplanting
methods
by
ecoregion.
Values
are
percent
of
total.
This
table
does
not
include
studies
from
outside
the
U.
S.,
guidelines,
reviews,
or
studies
involving
freshwater
plantings.

Method
Region
Alaska
Gulf
Mid­
Atlantic
Northeast
South
Florida
West
Plug
25
29
20
43
25
12
Peatpot
6
15
8
Turf
19
Mesh
6
15
3
Grid
3
Seedling
3
16
4
Seeds
25
5
13
4
Anchor
25
6
9
16
Sprig
25
11
29
19
20
Seed
Tape
14
Staple
6
45
14
16
24
Boulder
4
MBoulder
4
Other
3
4
Alaska
=
entire
coast
of
Alaska
(
Ak.);
Gulf
Coast
=
Gulf
of
Mexico
to
Cape
Sable,
Fl.
and
the
Florida
East
Coast
North
of
Jupiter
Inlet
to
Cape
Canaveral
(
Tex.,
La.,
Miss,
Ala.,
Fl.)
Mid­
Atlantic
=
Delmarva
Peninsula
to
North
Carolina
(
Del.,
Va.,
Md.,
N.
C.)
Northeast
=
Maine
to
New
Jersey
(
Maine,
R.
I.,
N.
H.,
Mass.,
Conn.,
N.
Y.,
N.
J.
South
Florida
=
South
of
Jupiter
Inlet
to
Cape
Sable,
Puerto
Rico
and
the
U.
S.
Virgin
Islands
(
Fl.,
P.
R.,
U.
S.
V.
I.)
West
=
Washington
to
California
(
Wa.,
Ore.,
Calif.).

Planting
methods
are
defined
as
follows
(
categories
are
mutually
exclusive):

Plug
=
tubes
as
coring
devices
are
used
to
extract
the
plants
with
the
sediment
and
rhizomes
intact.
Staple
=
U­
shaped
metal
staples
with
attached
bare
root
(
no
sediment)
planting
units.
Sprig
=
bare
root
planting
units
(
without
staples
or
anchors).
Anchor
=
any
structure
used
to
keep
the
planting
units
in
the
sediment.
Turf
=
large
square
sods
of
seagrass
that
are
ussually
extracted
with
a
shovel
and
planted
as
is.
Peatpot
=
a
plug
of
seagrass
that
is
transplanted
into
a
biodegradable
compressed
peat
container.
Biodegradable
Mesh
=
seagrass
sewn
to
a
biodegradable
mesh
fabric
and
attached
to
the
sediment
surface
as
a
planting
unit.
Seedling
=
a
newly
sprouted
seed
with
one
short
shoot.
Seed
=
seeds
with
no
sign
of
shoots
sprouting.
Plastic
Mesh
Grids
=
similar
to
biodegradable
mesh
except
these
are
plastic
(
non­
biodegradable).
Seed
Tape
=
method
of
planting
seeds
using
tape
that
has
seeds
sticking
to
it;
the
tape
is
then
rolled
out
along
the
sediment
surface.
Boulder
=
Phyllospadix
torreyi
is
attached
to
boulders.
MBoulder
=
P.
torreyi
is
attached
to
mesh
and
then
attached
to
boulders.
Other
=
rarely
used
methods
and
includes
studies
where
the
method
was
not
stated
in
the
document.
37
Table
1.7.
Percentage
of
transplanting
methods
by
seagrass
species.
Dashed
line
separates
methods
that
transport
associated
sediments
(
above
line)
from
those
that
do
not
(
below
line).

Method
Species
Hw
Pt
Rm
Sf
Tt
Zm
Plug
32
42
11
21
Peatpot
9
25
8
6
Turf
7
25
3
2
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
Mesh
9
25
4
4
Grid
3
Seedling
21
2
Seed
14
6
Anchor
6
17
7
11
Sprig
3
17
25
19
Seed
Tape
2
Staple
23
25
17
7
23
Boulder
?
MBoulder
?
Unknown
3
2
Hw
=
Halodule
wrightii
Pt
=
Phyllospadix
torreyi
Rm
=
Ruppia
maritima
Am
=
Zostera
marina
Tt
=
Thalassia
testudinum.
Sf
=
Syringodium
filiforme.

Planting
Methods
are
defined
as
follows
(
categories
are
mutually
exclusive):

Plug
=
tubes
as
coring
devices
are
used
to
extract
the
plants
with
the
sediment
and
rhizomes
intact.
Staple
=
U­
shaped
metal
staples
with
attached
bare
root
(
no
sediment)
planting
units.
Sprig
=
bare
root
planting
units
(
without
staples
or
anchors).
Anchor
=
any
structure
used
to
keep
the
planting
units
in
the
sediment.
Turf
=
large
square
sods
of
seagrass
that
are
usually
extracted
with
a
shovel
and
planted
as
is.
Peatpot
=
a
plug
of
seagrass
that
is
transplanted
into
a
biodegradable
compressed
peat
container.
Biodegradable
mesh
=
seagrass
sewn
to
a
biodegradable
mesh
fabric
and
attached
to
the
sediment
surface
as
a
planting
unit.
Seedling
=
a
newly
sprouted
seed
with
one
short
shoot.
Seed
=
seeds
with
no
sign
of
shoots
sprouting.
Unknown
=
the
method
was
not
stated
in
the
document.
Plastic
mesh
grid
=
similar
to
biodegradable
mesh
except
these
are
plastic.
Seed
Tape
=
method
of
planting
seeds
using
tape
that
has
seeds
sticking
to
it;
the
tape
is
then
rolled
out
along
the
sediment
surface.
Boulders
=
P.
torreyi
is
attached
to
boulders.
MBoulders
=
P.
torreyi
is
attached
to
mesh
and
then
attached
to
boulders.
?
=
insufficient
data
to
calculate
a
percentage.
38
Table
1.8.
List
of
experimental
parameters
and
the
percentage
(
in
descending
order)
that
were
incorporated
as
a
data
collection
or
as
independent
variables
in
field
transplant
studies.
Pre­
survey
=
the
site
selected
as
a
transplant
site
was
surveyed
prior
to
transplanting
for
its
suitability
to
sustain
a
transplant.

Experimental
Parameters
Percent
of
Documents
Using
this
Parameter
Pre­
survey
of
site
62
Planting
method
45
Post­
survey
of
site
27
Depth
26
Cost
analysis
22
Fertilization
type
21
Season
21
Faunal
study
18
Planting
unit
spacing
17
Tidal
zone
15
Energy
regime
14
Donor
survey
12
Sediment
particle
size
9
Enclosure
8
Shoot
numbers
8
In
vitro
propagation
8
Genetics
6
Light
intensity
5
Bioturbation
3
Burial
recovery
3
Apicals
1
Salinity
1
Planting
method
=
different
methods
of
transplanting
were
tested
for
their
effectiveness.
Post­
survey
of
site
=
the
effect
of
transplanting
on
the
site
location
was
evaluated.
Depth
=
effects
of
different
depths
on
transplanting
success
was
determined.
Cost­
analysis
=
the
total
cost
of
the
transplanting
was
determined.
Fertilization
type
=
effects
of
fertilizers
on
transplanting
was
evaluated.
Season
=
effects
of
time
of
year
on
transplanting
was
evaluated.
Faunal
study
=
fauna
was
sampled
in
transplanted
beds.
Planting
unit
spacing
=
the
effects
of
different
spacing
of
planting
units
was
evaluated.
Tidal
zone
=
effects
of
different
tidal
zones
on
transplanting
was
examined.
Energy
regime
=
effects
of
energy
regime
on
transplanting.
Donor
survey
=
there
was
a
study
conducted
on
the
recovery
of
the
transplant
donor
bed.
Sediment
particle
size
=
effects
of
different
sediment
size
on
transplanting.
Enclosure
=
effects
of
enclosure
devices
on
transplanting.
Shoot
numbers
=
effects
of
different
planting
unit
shoot
numbers
on
transplanting
success.
In
vitro
propagation
=
growing
seagrass
in
the
laboratory
to
be
transplanted.
Genetics
=
genetic
experiments
on
transplanted
seagrass
were
conducted.
Light
intensity
=
effects
of
various
light
levels
on
transplanting
success.
Bioturbation
=
bioturbation
effects
on
transplanted
seagrass.
Burial
recovery
=
effects
of
sediment
burial
on
transplanted
seagrass.
Apicals
=
effects
of
the
presence,
absence,
or
different
numbers
of
apicals
in
planting
units.
Salinity
=
effects
of
different
salinity
on
transplanting.
Chapter
1:
Background
°
39
ecological
role
in
the
coastal
zone.
Although
there
has
been
little
in
the
way
of
focused
attention
on
development
of
planting
techniques
for
these
latter
two
species,
we
expect
that
existing
methods
such
as
plugs
or
peatpots
may
have
promise
(
see
Chapter
2,
Planting).

There
are
also
some
incidental
plantings
of
which
we
are
aware.
We
know
that
Halophila
decipiens
was
transplanted
in
15
meters
of
water
on
St.
Croix,
U.
S.
V.
I.
in
1986
(
authors
unpubl.
data)
using
mini­
staples
constructed
of
130­
pound
test
wire
leader;
plantings
spread
and
apparently
persisted
to
the
end
of
the
normal
growing
season.
Harrison
(
1990)
has
also
transplanted
Z.
marina
in
British
Columbia
using
unattached
shoots,
cores,
and
by
attaching
shoots
to
re­
bar
(
sensu
Kelly
et
al.
1971).
Phyllospadix
was
planted
in
the
Monterey
Bay
Aquarium,
California.
Indoor
small
Table
1.9.
List
of
ten
most
common
parameters
recorded
in
monitoring
of
transplant
studies
Some
studies
considered
more
than
one
parameter.

Monitoring
Parameter
Percent
of
Studies
with
this
Parameter
Irregular
frequency
monitoring
74
Percent
survival
(
PU)
65
Shoot
counts
55
Shoot
density
53
Percent
cover
47
Leaf
length
29
Leaf
width
12
Rhizome
length
6
Directm
mapping
3
Biomass
3
Irregular
Frequency
Monitoring
=
irregular
time
intervals
were
chosen
for
follow­
up
monitoring
of
a
transplant
site.
Percent
Survival
=
percent
of
planting
units
(
PUs)
that
survived
were
monitored.
Shoot
Counts
=
direct
counts
of
planting
unit
shoots
was
conducted.
Shoot
Density
=
density
of
the
planting
units
was
monitored.
Percent
Cover
=
time
zero
area
was
known
and
considered
100
percent
cover
so
that
future
areal
coverage
could
be
compared
as
a
percent
of
that
original
coverage.
Leaf
Length
=
leaf
lengths
were
measured
directly.
Leaf
Width
=
leaf
widths
were
measured
directly.
Rhizome
Length
=
total
length
of
living
rhizome.
Direct
Mapping
=
actual
mapping
of
the
planting
units
for
the
area
covered.
Biomass
=
weight
of
a
given
area
of
seagrass.
40
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
tank
exhibits
had
poor
survival
but
plantings
lodged
under
rocks
and
experiencing
mild
simulated
wave
conditions
in
the
aviary
persisted
for
several
years
(
Monterey
Bay
staff,
pers.
com.).
Similarly,
Thalassia
has
been
grown
in
the
coral
reef
exhibit
at
the
Smithsonian
Institution
in
Washington,
D.
C.
for
several
years
but
with
great
logistic
cost.

Not
only
did
purposes
of
these
papers
vary
widely,
so
did
the
design
parameters
of
the
studies.
Table
1.8
describes
the
various
parameters
that
were
manipulated
for
documents
reporting
new
field
planting
data
(
66
documents
total:
reviews
and
lab
experiments
excluded).
Twenty­
four
different
parameters
were
examined.
Preliminary
surveys
of
some
environmental
conditions
at
the
planting
site
was
the
most
common
design
feature:
~
62
percent
of
the
papers
performed
some
pre­
planting
evaluation
of
the
site.
Slightly
less
than
half
of
the
papers
tested
planting
methods.
Only
~
25
percent
of
the
papers
continued
to
survey
some
environmental
conditions
after
plantings
were
installed,
making
it
very
difficult
to
establish
any
linkage
between
plant
performance
and
episodic
events.
Planting
depth
(
a
rough
surrogate
for
light
availability,
but
also
potentially
related
to
frequency
of
emersion)
was
at
least
noted,
if
not
a
factor
tested
for
influence
on
plantings
in
approximately
30
percent
of
the
papers.
Tidal
zone
(
as
opposed
to
some
sea
level­
normalized
depth
measure)
was
also
noted
in
17
percent
of
the
papers,
but
these
data
were
not
as
specific
as
depth
data.
Together,
however,
water
depth
and
tidal
zone
considerations
were
in
47
percent
of
the
papers.
Cost
analyses,
comparisons
among
planting
season,
and
fertilizer
effects
were
aspects
of
project
design
in
~
20
percent
of
the
papers.
Comparative
faunal
assessments,
effects
of
PU
spacing,
physical
energy
on
the
site,
and
recovery
of
plants
at
the
donor
site
were
parts
of
project
designs
in
12­
18
percent
of
all
studies.
An
additional
12
parameters
were
examined
in
the
papers
we
reviewed
but
were
never
included
in
more
than
10
percent
of
the
papers.

What
is
interesting
here
is
not
so
much
what
was
either
manipulated
or
noted
but
the
proportions
of
what
was
not;
that
is,
data
that
were
considered
relevant
varied
tremendously
among
studies.
Thirty­
eight
percent
of
the
papers
did
not
consider
or
at
least
did
not
report
what
information
was
used
to
choose
a
planting
site.
Of
those
reporting,
33
percent
simply
used
the
criteria
of
no
vegetation
present
which
when
used
alone
has
been
previously
described
as
an
unacceptable
criteria
(
Fredette
et
al.
1985,
Fonseca
et
al.
1987c,
Fonseca
1989a,
1992,
1994)
because
selecting
unvegetated
areas
with
no
known
history
of
seagrass
cover
disregards
the
fact
that
any
one
of
several
mechanisms
may
be
at
work
maintaining
that
level
of
patchiness
(
e.
g.,
waves,
currents,
bioturbation).
There
is
a
rich
body
of
literature
on
the
role
of
habitat
heterogeneity
on
ecosystem
function
that
would
have
to
be
ignored
to
recommend
converting
naturally
unvegetated
areas
to
vegetated.
Thus,
in
addition
to
being
Chapter
1:
Background
°
41
a
high­
risk
planting
area,
planting
in
such
an
environment
temporarily
substitutes
one
habitat
type
for
another.
Therefore,
based
on
our
survey,
24
percent
used
what
we
consider
to
be
an
appropriate
site
selection
criteria
(
the
site
having
been
previously
vegetated
but
was
now
barren,
although
there
are
caveats
to
this
criteria;
see
Planning
chapter).
Approximately
19
percent
of
the
plantings
were
on
dredged
material
while
20
percent
were
on
unvegetated
spaces
adjacent
to
existing
seagrass.

Few
other
factors
were
consistently
integrated
into
design
plans.
Only
17
percent
surveyed
fauna
after
planting.
Twelve
percent
considered
the
impact
of
harvesting
on
donor
seagrass
beds
(
i.
e.,
monitored
recovery
of
donor
site).
Less
than
10
percent
of
the
studies
manipulated
time
0
shoot
number
in
a
planting
unit
(
generally
as
an
attempt
to
determine
optimal
planting
unit
size).

Interestingly,
two
parameters
besides
percent
PU
survival
that
we
have
long
recommended
as
being
critical
baseline
monitoring
data
(
Fonseca
et
al.
1982,
Fonseca
1989a,
1992,1994),
number
of
shoots
PU­
1
and
percent
cover
of
the
bottom,
were
only
in
~
53
and
~
47
percent
of
the
papers,
respectively.
We
have
recommended
these
forms
of
data
collection
because,
when
combined
they
describe
many
aspects
of
planting
viability.
In
contrast,
shoot
density,
a
parameter
over
which
there
is
little
control,
was
used
as
a
performance
criteria
in
~
51
percent
of
the
studies.
Recent
findings
(
Fonseca
et
al.
1996b)
also
suggest
that
macroepibenthic
faunal
abundance
in
planted
seagrass
beds
asymptotes
at
comparatively
(
to
natural
beds)
low
shoot
densities
(
as
little
as
one
third
of
natural
beds),
indicating
that
it
might
not
be
relevant
to
require
shoot
density
in
a
planted
bed
to
equal
that
of
natural
beds
to
support
faunal
densities
equivalent
to
natural
(
but
see
performance
criteria
suggested
by
Short
(
1993
p.
51).
Although
some
lower­
than­
ambient
threshold
shoot
density
may
be
suitable
for
generating
faunal
equivalence,
lower
shoot
densities
may
not
provide
a
sufficient
buffer
to
population
fluctuations
of
the
seagrasses
themselves.
Thus,
the
issue
of
demographic
status
of
the
seagrasses
of
restored
vs.
natural
beds
is
only
beginning
to
be
evaluated.

Most
disturbing
was
that
less
than
7
percent
of
the
papers
actually
provided
quantitative
data
on
two
of
the
most
critical
limiting
parameters
known
for
seagrass
planting
success,
light
regime
and
bioturbation.
Although
depth
and
tide
zone
were
frequently
recognized
as
important
factors,
the
absence
of
direct
measurements
of
light
means
that
depth
and
tide
zone
data
are
not
easily
extrapolated
because
we
do
not
know
the
transparency
of
the
water
column.
We
can
look
up
information
on
tidal
amplitude
and
periodicity,
but
the
interaction
of
light
and
tides
on
seagrass
growth
is
only
now
being
modeled
(
Zimmerman
et
al.
1994,
Dennison
and
Kirkman
1996,
Koch
and
Beer
1996),
although
these
papers
suggest
the
interaction
of
tidal
42
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
amplitude
and
light
availability
to
accurately
predict
site
suitability
based
on
transmissivity
data.

Apart
from
those
parameters
that
were
monitored
and/
or
manipulated,
a
total
of
ten
parameters
were
actually
utilized
as
measures
of
planting
performance
and/
or
success
(
Table
1.9).
The
percentage
of
PUs
surviving
was
the
most
common
criteria
but
was
reported
in
only
~
66
percent
of
the
papers.
However,
~
74
percent
of
the
papers
varied
the
frequency
of
monitoring
after
planting
over
the
course
of
their
respective
investigations
(
e.
g.,
contrast
fixed
interval
monitoring
with
a
study
that
conducts
monthly
sampling
for
the
first
year
then
shifts
to
biannual
monitoring
for
following
years).
The
duration
of
monitoring
in
the
papers
we
reviewed
ranged
from
zero
to
eight
years
with
a
mean
and
median
of
~
1.5
years.
During
these
monitoring
periods
the
frequency
of
monitoring
was
also
highly
variable,
again
ranging
from
zero
to
an
equivalent
of
30
times
y­
1.
The
average
frequency
of
monitoring
was
4.6
times
y­
1.

Fonseca
(
1989,
1994)
has
recommended
that
early,
frequent
(
usually
quarterly)
monitoring
be
performed
for
the
first
year
after
planting
followed
by
less
frequent
(
e.
g.,
biannual)
monitoring.
Despite
problems
with
changing
temporal
scales
in
analysis
(
see
section
"
Scale
and
its
Role
in
Defining
Seagrass
Habitat"),
we
continue
this
recommendation
because
many,
but
not
all
(
particularly
plantings
with
high
initial
loss
of
PUs)
of
our
successful
experimental
plantings
followed
a
sigmoidal
population
growth
curve;
initially
high,
exponential
growth
with
low
mortality
followed
by
a
balancing
of
natality
and
mortality
of
shoots
which
leads
to
an
asymptote
of
plant
density.
Past
recommendations
for
this
monitoring
strategy
(
Fonseca
1989a,
Fonseca
1992,
Fonseca
1994)
actually
agree
well
with
at
least
the
mean
monitoring
time
values
of
the
papers
reviewed.
Similar
frequencies
of
monitoring
were
recommended
by
Merkel
(
1992),
of
time
0,
3,
6,
12,
24,
and
36
months
but
with
an
additional
recommended
survey
at
60
months.
Choice
of
3
years
for
monitoring
resulted
largely
from
compromise
in
that
permit
monitoring
is
rare
(
Race
and
Fonseca
1996)
and
shorter
monitoring
periods
increase
the
possibility
of
acquiring
monitoring
compliance.
So,
for
a
given
planting,
how
long
should
monitoring
proceed
in
order
to
judge
planting
performance?
Taken
together
with
the
average
monitoring
period
of
4.3
y,
and
the
fact
that
only
10
percent
of
the
papers
we
surveyed
achieved
an
ideal
100
percent
cover,
indicates
that
previous
suggestions
of
3­
year
monitoring
by
Fonseca
(
1989a,
1992,
1994)
may
be
a
serious
underestimate
of
the
time
required
to
document
project
success;
times
in
excess
of
5
years
may
be
more
appropriate.

What
is
probably
the
most
documented
parameter
in
natural
beds,
seagrass
biomass
was
only
measured
in
~
3
percent
of
the
papers,
perhaps
because
it
is
a
destruc­
tive
sampling
technique.
Also,
several
measures
of
the
plant's
morphology
were
used
frequently
to
determine
planting
performance
(
Table
1.9).
We
view
these
criteria
suspiciously;
seagrasses
are
often
phenotypically
plastic,
and
variation
in
plant
shape
and
size
is
only
loosely
linked
to
functional
attributes
of
seagrass
beds
at
this
time
(
Bell
et
al.
1991,
Fonseca
et
al.
1996a)
although
morphology
has
been
linked
to
significant
genetic
differences
(
Fain
et
al.
1992).
We
find
it
disturbing
that
simple
parameters
such
as
survival
and
coverage
were
not
more
universally
recorded.
From
the
low
replication
of
criteria
among
studies,
it
is
no
wonder
that
quantitative
performance
and
compliance
thresholds,
when
they
appear
in
mitigation
plans,
vary
so
tremendously
(
Thayer
et
al.
1985).
Moreover,
some
papers
used
irreproducible
units
such
as
"
scoopfuls"
and
"
bucketfuls"
to
describe
sampling
units.
Such
vague
planning
criteria
should
not
be
used
by
resource
managers.

The
results
of
monitoring
efforts
have
revealed
some
unexpected
trends
regarding
success.
To
analyze
this,
we
chose
two
categories,
the
final
reported
percent
PU
survival
and
the
percent
of
the
target
area
covered
that
was
reported
at
the
conclusion
of
a
paper.
Of
53
papers
that
reported
percent
PU
survival,
the
median
percent
PU
survival
was
35
percent;
mean
42
percent;
standard
deviation
=
29.9;
coefficient
of
variation
=
70
with
a
distribution
heavily
skewed
to
lower
percent
survival
(
Sk
=
0.35),
suggesting
adoption
[
use]
of
the
median
value
(
Figure
1.5).
Roughly
5
percent
of
the
plantings
reported
100
percent
PU
survival.
We
found
27
papers
that
reported
percent
of
the
target
area
covered.
The
median
percent
area
covered
was
40
percent
and
was
closer
to
the
mean
percent
area
covered
of
42
percent;
a
standard
deviation
=
31.2;
but
still
with
a
high
coefficient
of
variation
=
75
and
a
distribution
again
skewed
to
lower
coverage
amounts
(
Sk
=
0.41)
(
Figure
1.6).
We
should
point
out
that
some
of
the
variance
in
the
data
also
results
from
areas
such
as
southern
California
enjoying
generally
very
high
success
rates
(
approaching
100
percent).
The
reasons
for
that
success
rate
may
have
to
do
with
quiescent
settings
for
planting,
high
experience
level
and
perhaps,
comparatively
low
bioturbation
levels.
However,
on
a
national
scale,
only
approximately
10
percent
of
the
plantings
achieved
100
percent
cover
within
the
monitoring
period.
Thus,
these
data
indicate
that
replanting
is
a
consistent
requirement
of
seagrass
operations
unless
substantial
initial
overplanting
is
conducted
to
compensate
for
anticipated
losses.
Moreover,
low
initial
survival
rates
may
explain
why
seagrass
plantings
often
produce
less
acreage
than
originally
planned,
suggesting
that
initial
PU
survival
levels
should
be
held
to
high
standards
to
help
ensure
achieving
target
acreage.

An
extreme
interpretation
of
these
findings
would
be
that
based
on
the
median
survival
(
a
planting
should
have
an
overplanting
ratio
of
approximately
3.0).
In
other
words,
if
you
wished
to
ensure
that
100
planting
units
will
survive,
300
should
be
Chapter
1:
Background
°
43
44
Figure
1.6.
Frequency
distribution
of
percent
area
covered
by
plantings
from
the
documents
surveyed
nationally.
Y­
axis
=
percentage
of
the
area
covered
values
falling
in
the
percent
area
cover
categories
on
the
X­
axis
(
10%
increments).
Figure
1.5.
Frequency
distribution
of
percent
planting
unit
survival
from
the
documents
surveyed
nationally
Y­
axis
=
percentage
of
the
survival
values
falling
in
the
percent
survival
categories
on
the
Xaxis
(
10%
increments).
Chapter
1:
Background
°
45
planted.
Similarly,
based
on
this
national
average,
to
ensure
the
required
area
of
seagrass
bed
to
be
generated,
a
replacement
ratio
of
~
2.5
units
of
area
planted
to
1
unit
of
area
lost
is
needed
to
meet
a
no­
net­
loss
criteria
(
i.
e.,
1:
1
replacement
ratio).
We
conclude
that
this
is
an
extreme
interpretation
because
many
plantings
used
to
compile
these
statistics
were
conducted
on
sites
that
would
have
been
expected
to
produce
patchy
seagrass
beds
in
any
event.
Also,
many
sites
were
chosen
that
violated
recommended
site
selection
criteria
which
would
skew
the
distribution
toward
low
survival
and
coverage.
If
site
selection
criteria
are
employed
as
described
later,
it
is
possible
that
these
replacement
ratios
could
be
made
much
lower.

Managers
must
be
cognizant
of
the
different
sources
of
planting
failures
and
judge
planting
proposals
under
strict
criteria.
The
practice
of
seagrass
bed
mitigation
should
not
be
questioned
based
on
a
failure
in
judgment
on
the
part
of
someone
who
performed
a
planting.
Such
human
failures
must
be
separated
from
failures
of
the
approach
as
a
whole
in
order
to
responsibly
assess
seagrass
planting
as
a
mitigative
tool
(
Fonseca
et
al.
1994).
The
key
is
to
determine
what
made
some
plantings
so
successful
and
others
so
marginal.

Monitoring
as
recommended
in
the
past
(
e.
g.,
Fonseca
1989a,
1992,
1994)
does
not
lend
itself
to
determination
of
agents
of
planting
loss.
Only
sophisticated
monitoring
equipment
with
high
frequency
recording
capacity
could
hope
to
detect
environmentally
induced
losses.
Acute
and
capricious
events
such
as
bioturbation
and
vandalism
are
even
more
difficult
to
determine
with
complete
certainty
(
although
use
of
exclosure
cages
may
go
far
in
suggesting
the
influence
of
bioturbation,
Merkel
1988a,
Fonseca
et
al.
1994).
Therefore,
the
agents
of
loss
among
these
studies
cannot
accurately
be
presented
as
a
ranked
set.
However,
based
on
our
observations
in
the
field,
one
might
speculate
that
most
failures
occur
from
improper
site
selection
(
see
criteria
for
site
selection,
below)
and
execution.
From
our
experience
and
conversations
with
others
(
not
to
mention
some
published
findings:
Mote
Marine
Lab.
&
Mangrove
Systems
Inc.
1989;
Merkel
1988a,
b;
Fonseca
et
al.
1994),
we
conclude
that
once
a
site
has
been
appropriately
selected
under
the
criteria
described
below
(
e.
g.,
previous
history
of
seagrass
cover,
etc.)
the
primary
agents
of
loss
vary
between
bioturbation,
acute
storm
events,
algal
smothering,
and
vandalism.

These
compilations
indicate
that
most
of
the
planting
experience
is
centered
in
the
southern
and
western
parts
of
the
U.
S.
Also
only
a
few
species
are
regularly
utilized
in
mitigation
projects.
Given
the
widespread
impacts
to
seagrass
ecosystems,
concern
that
the
absence
of
these
other
species
from
the
literature
indicates
that
impacts
to
those
species
goes
unnoticed.
Either
that
or
these
plant
communities
may
not
be
receiving
sufficient
protection
under
current
management
practices.
These
45
46
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
survey
data
also
indicate
that
there
has
been
extensive
experimentation
with
planting
methodologies,
and
it
appears
that
only
a
few
are
consistently
employed
and,
again,
only
for
a
few
species.
The
concentration
of
planting
effort
in
Florida
and
the
West
Coast
may
be
due
to
comparatively
high
development
pressures
in
these
areas.
Although
high
habitat
loss
rates
also
occur
in
the
mid­
Atlantic
states
ecoregion,
the
proximity
of
research
laboratories
that
have
historically
focused
on
seagrass
to
those
estuaries
may
also
explain
the
concentration
of
work
in
that
ecoregion.

ARE
PLANTED
SEAGRASS
BEDS
FUNCTIONALLY
EQUIVALENT
TO
NATURALLY­
OCCURRING
BEDS?

What
is
"
functional
equivalency"?
In
a
general
sense,
this
means
that
a
restored
or
mitigated
system
attains
functions
the
same
as
those
of
an
unimpacted
system
in
a
similar
setting.
Seagrass
beds
have
many
functions
(
sensuWood
et
al.
1969),
some
of
which
may
be
more
difficult
to
restore
than
others.
As
is
the
case
with
much
of
biology
the
answer
to
the
question
of
functional
equivalency
is
both
"
yes"
and
"
no."
We
tend
to
take
the
stance
that
if
an
area
has
recovered
equal
or
greater
acreage
than
that
which
was
lost,
and
that
area
persists
with
the
same
seagrass
species,
a
planted
seagrass
bed
can
become
equivalent,
but
not
identical
to
a
natural,
unimpacted
bed.
Our
stance
is
not
universally
accepted.
Equivalent
means
"
equal
to"
but
is
sometimes
taken
to
mean
"
identical."
However,
since
no
two
samples
of
any
natural
ecosystem
are
ever
truly
identical,
some
subjectivity
comes
into
play,
both
in
terms
of
the
degree
of
equivalence
and
the
appropriate
functions
to
measure.
The
problem
then
is
what
drives
the
subjectivity?
A
developer
may
interpret
functional
equivalency
of
their
mitigation
project
in
far
more
general
terms
than
a
trained
biologist.
What
then
are
the
relevant
parameters
by
which
to
document
equivalency?

According
to
our
comparative
analysis
of
the
literature,
thirty­
three
different
parameters
were
used
to
describe
success.
This
indicates
the
broad
definition
of
functional
equivalent
 
practitioners
obviously
target
many
different
factors
and
differ
in
their
opinions
when
ranking
importance
of
these
factors.
Moreover,
there
is
conflicting
guidance
from
the
literature
regarding
the
rate
at
which
planted
beds
take
on
attributes
of
natural,
undisturbed
beds.
Brown­
Peterson
(
1993)
and
Montagna
(
1993)
conclude
that
attributes
of
planted
seagrass
beds
were
still
not
equivalent
to
natural
ones
after
31
and
14­
17
years,
respectively.
Similarly,
Smith
et
al.
(
1988a)
found
that
planted
beds
did
not
provide
equivalent
bay
scallop
habitat
over
a
growing
season.
Hoffman
(
1988)
concluded
that
one­
year
old
Z.
marina
plantings
in
San
Diego
did
not
support
some
fauna
at
levels
exactly
equal
to
that
of
natural
beds,
although
some
of
differences
were
small.
In
contrast,
Nessmith
(
1980),
Homziak
et
Chapter
1:
Background
°
47
al.
(
1982),
Fonseca
et
al.
(
1990,
1996
a,
b),
and
Wyllie­
Echeverria
et
al.
(
1994b)
found
that
faunal
abundance
and
composition
in
planted
beds
approached
that
of
natural
beds
within
2­
3
years.

Much
of
this
discrepancy
among
studies
may
be
the
result
of
intrinsic
differences
among
natural
reference
sites
and
planted
areas,
the
organisms
chosen
to
evaluate
recovery,
and
different
worker's
interpretation
of
what
constitutes
a
difference.
Because
of
the
tremendous
variability
among
natural
beds,
we
question
the
efficacy
of
precise
numerical
comparisons
in
an
interpretation
of
planting
success;
comparisons
that
include
estimates
of
variance
might
be
more
appropriate.
For
example,
distance
and/
or
isolation
of
planted
sites
from
natural
beds
will
cause
some
differences
(
Bell
et
al.
1988,
1992).
Brown­
Peterson
(
1993)
compared
fish
communities
among
planted
and
reference
sites
but
the
sites
were
located
on
opposite
sides
of
a
barrier
island
lagoon.
Montagna
(
1993)
compared
beds
established
both
by
natural
recolonization
and
planting
in
scraped­
down
dredged
material
islands
with
relatively
open
areas.
Thus,
there
is
some
question
as
to
whether
differences
among
planted
and
natural
treatments
were
the
result
of
planting
or
of
innate
differences
due
to
the
physical
setting.
However,
Montagna
(
1993)
points
out
that
most
studies
suggesting
faunal
equivalency
have
focused
on
more
vagile
macrofauna
such
as
fish
whereas
certain
infauna
(
e.
g.,
clams)
may
not
colonize
as
quickly.
Kenworthy
et
al.
(
1980)
and
Homziak
et
al.
(
1982)
found
rapid
colonization
of
a
planted
Z.
marina
site
by
scallops
and
meiofauna
as
did
Wyllie­
Echeverria
et
al.
(
1994b)
for
salmon
prey
(
largely
meiofauna).
McLaughlin
et
al.
(
1983)
concluded
that
recolonization
by
a
wide
variety
of
macrofauna
occurred
in
planted
Thalassia
beds
within
only
a
few
years.
Similarly,
Fonseca
et
al.
(
1990)
found
that
after
experiencing
widespread
failure
of
a
planted
area,
the
same
site
then
naturally
colonized
by
seed
and
supported
a
macrofaunal
community
not
statistically
different
from
adjacent
planted
sites
within
sixmonths
of
the
onset
of
seed
germination.

More
recently,
Fonseca
et
al.
(
1996
a,
b)
found
that
H.
wrightii
and
S.
filiforme
beds
planted
on
0.5
m
centers
in
Tampa
Bay
developed
fish,
shrimp
and
crab
density
and
composition
statistically
indistinguishable
from
nearby
natural
sites
within
three
years.
One
interesting
aspect
of
that
work
was
the
relation
of
animal
density
to
plant
density
(
Figure
1.7).
The
seagrass
density
at
which
animal
density
in
planted
beds
equaled
(
p
<
0.05)
that
of
natural
beds
was
only
approximately
one­
third
of
the
mean
natural
bed
shoot
density.
That
density
can
be
obtained
within
one
year.
They
found
that
although
linear
models
could
account
for
approximately
65
percent
of
the
variance
of
animal
density
as
a
function
of
plant
density
over
time,
a
non­
linear,
asymptotic
relationship
between
natural­
log
transformed
animal
density
and
seagrass
areal
shoot
density
was
apparent
(
Figure
1.7).
Although
transformation
of
a
straight
line
48
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
will
yield
some
asymptotic
tendencies,
they
are
not
enough
to
account
for
the
pattern
observed.
The
fact
that
animal
density
had
an
asymptotic
relationship
with
shoot
density
implies
that
monitoring
shoot
density
over
time
may
be
an
inexpensive
diagnostic
parameter
for
determining
a
threshold
planting
success
in
terms
of
fauna.
Short
(
1993)
found
similarly
rapid
colonization
of
Z.
marina
plantings
in
New
Hampshire
by
a
wide
variety
of
fauna.
Therefore,
monitoring
shoot
density
over
time
would
be
much
less
costly
than
direct
measures
of
faunal
communities.
In
order
to
justify
use
of
only
plant
data
in
assessing
some
aspects
of
planting
success,
however
the
temporal
relationship
among
shoot
density
and
faunal
community
structure
must
be
collected
from
planted
beds
across
a
broad
geographic
range.

Although
some
ecological
attributes
may
return
quickly
after
planting
seagrass,
there
is
still
a
measurable
period
of
time
until
the
system
has
attained
full
function.
The
loss
of
ecosystem
production
in
the
time
between
when
a
seagrass
bed
is
dam­
Figure
1.7.
Natural­
log
transformed
faunal
density
plotted
against
areal
shoot
density.
Open
circles
observed
points.
Closed
circles=
predicted
points
using
an
asymptotic
function.
Vertical
line
(
VL)
"
shrimp"
or
"
fish"=
areal
shoot
density
at
which
densities
of
these
animals
first
became
not
significantly
different
among
planted
and
natural
versions
of
that
seagrass
species;"
natural
bed
density"=
the
studywide
average
areal
shoot
density
of
natural
beds
of
that
seagrass
species.
Taken
from
Fonseca
et
al.
(
1996b).
fish
shrimp
natural
bed
density
Chapter
1:
Background
°
49
aged
and
functions
are
restored
has
long
been
an
issue
of
concern
to
managers.
This
loss
of
production
has
been
termed
"
interim
loss."
The
manner
in
which
this
loss
has
been
calculated
varies
widely,
as
will
be
discussed
later
under
sections
on
planting
We
raise
the
point
here
because
in
many
instances
a
reckoning
of
functional
equivalency
among
planted
and
reference
sites
is
made
in
an
attempt
to
recoup
interim
loss
of
various
living
resources.

Seagrass
species
substitution
is
another
issue
that
has
bearing
on
the
question
of
functional
equivalency
among
planted
and
reference
beds.
Although
much
of
the
temperate
U.
S.
is
dominated
by
one
seagrass
species,
Z.
marina,
subtropical
areas
and
more
recently
the
Pacific
Northwest
must
contend
with
the
functional
ramifications
of
substituting
one
seagrass
species
for
another
(
Pawlak
1994).
There
are
few
data
to
guide
this
decision.
Temporary
species
substitution
has
been
suggested
for
subtropical
species
(
Fonseca
et
al.
1987c)
where
faster­
spreading
species
such
as
H.
wrightii
and
S.
filiforme
are
planted
as
predecessors
to
recover
areas
previously
dominated
by
the
much
slower­
spreading
T.
testudinum.
In
that
circumstance
the
reasoning
was
that
interim
loss
of
ecosystem
functions
could
be
minimized
by
establishing
any
form
of
seagrass
coverage.
Based
on
their
work
in
Tampa
Bay,
Fonseca
et
al.
(
1996b)
have
suggested
that
differences
in
macroepibenthic
faunal
communities
among
these
three
subtropical
species
may
not
be
as
great
as
previously
inferred
(
Stoner
1983).
One
reason
for
the
differences
among
studies
could
be
that
Fonseca
et
al.
(
1996b)
focused
on
unit
area
of
seagrass
bed­
based
surveys
while
others
had
compared
animal
populations
among
seagrass
species
weighted
by
attributes
of
habitat
complexity,
such
as
leaf
surface
area.
While
acknowledging
that
such
faunal
assessments
should
not
be
construed
as
indicators
of
all
ecosystem
functions
(
e.
g.,
nutrient
cycling,
bed
stability
the
findings
of
Fonseca
et
al.
(
1996b)
support
the
notion
that
seagrass
species
substitution
to
ameliorate
interim
losses
of
some
ecological
attributes
may
be
a
legitimate
means
to
an
end
where
that
end
is
eventual
replacement
of
the
seagrass
species
that
was
damaged.

There
is
another
tack
to
take
in
assessing
whether
planted
seagrass
beds
provide
resource
functions
equivalent
to
those
they
are
intended
to
replace.
On
a
much
simpler
level,
we
are
not
aware
of
any
study
that
suggests
that
a
seagrass
bed
is
not
a
highly
productive
habitat.
Therefore,
we
may
infer
that
if
one
can
produce
a
desired
amount
of
seagrass
habitat
which
persists
over
time
that
many
ecosystem
functions
eventually
will
be
restored.
Whether
a
planted
bed
is
the
exact
replacement
for
another
seems
to
us
to
be
an
inappropriate
question.
It
will
depend
upon
what
one
considers
to
be
an
important
ecosystem
function
and
how
much
of
it
must
be
replaced
to
be
considered
"
equivalent."
Utilizing
generally
accepted
significance
levels
(
i.
e.,
p
<
0.05)
to
detect
differences
among
two
unique
portions
of
the
ecosystem
50
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
is
the
most
objective
and
scientifically
acceptable
means
of
testing
differences.
Of
course,
such
differences
are
certainly
affected
by
decisions
of
sampling
gear,
its
size,
the
bias
of
that
gear
towards
certain
fauna
and
size
classes,
and
temporal
and
spatial
density
of
sampling.
For
example,
the
differences
reported
by
Brown­
Peterson
(
1993)
and
Montagna
(
1993)
may
well
be
within
the
range
of
year­
to­
year
and
siteto
site
variation
if
longer
periods
of
time
were
sampled
and
across
wider
areas
(
the
choice
of
sampling
scale
being
a
powerful
determinate
of
our
perception)
(
e.
g.,
crab
densities:
Fonseca
et
al.
1996b).
Given
the
spatial
and
temporal
variance
in
animal
numbers,
we
feel
it
very
difficult
to
justify
additional
planting
based
on
relatively
small
differences
in
faunal
attributes
as
compared
to
unplanted
areas;
using
less
stringent
significance
levels
may
be
acceptable
as
well
(
i.
e.,
p
<
0.10).

In
general,
we
believe
that
planting
is
a
success
if
the
acreage
is
planted,
persists,
and
eventually
(
if
not
immediately)
leads
to
replacement
of
the
same
resource
functions
of
the
seagrass
species
that
were
damaged.
By
success
we
imply
functionally
equivalent
to
natural
beds.
Again,
we
stress
that
the
use
of
acreage
and
persistence
as
diagnostic
features
of
planting
success
needs
additional
geographic
replication.

As
mentioned
earlier,
this
view
of
assessment
of
planting
success
is,
of
course,
not
universally
held.
Short
(
1993)
established
other
criteria
for
mitigation
success
for
a
project
in
New
Hampshire.
There,
based
on
the
recovery
observed
in
his
planted
beds,
he
suggested
that
for
eelgrass
plantings
to
be
considered
initially
successful,
they
should
cover
30
percent
of
the
planted
area
in
one
year.
In
addition,
he
stipulated
that
40
percent
of
the
following
parameters
be
attained
by
plantings
within
one
year:
seagrass
primary
production,
shoot
density,
leaf
area,
percent
cover,
and
continuity
(
i.
e.,
meaning
that
40
percent
of
the
plantings
have
coalesced).
Moreover
he
stipulated
that
fish
and
infaunal
assemblages
constitute
25
percent
of
the
following
ecological
parameters:
presence
of
dominant
species
and
total
numerical
abundance
as
compared
to
nearby
natural
beds.
The
values
put
forth
by
Short
(
1993)
appear
to
coincide
with
findings
for
Tampa
Bay
(
Fonseca
et
al.
1996b)
in
that
faunal
recovery
will
be
closely
linked
with
planting
success
and
persistence.

Another
reason
for
using
plants
rather
than
fauna
as
a
metric
of
planting
success
is
that
one
can
envision
scenarios
where
faunal
recruitment
to
a
bed
could
be
inhibited
by
recent
natural
events
such
as
storms
or
local
pollution
sources
that
may
or
may
not
equally
affect
a
reference
site.
Conversely,
we
know
of
no
evidence
where
an
otherwise
unimpacted
natural
seagrass
bed
has
not
supported
high
faunal
density
and
diversity.
Thus,
if
the
faunal/
seagrass
relationship
is
not
necessarily
reciprocal,
use
of
fauna
alone
may
not
be
as
easily
tracked
as
the
plants
themselves,
which
may
be
suitable
to
infer
the
development
of
collateral
faunal
resource
functions.
At
the
very
least,
following
plants
alone
is
arguably
less
expensive
than
faunal
data
collection
and
may
represent
the
most
realistic
data
collection
effort
except
in
heavily
subsidized
restoration
efforts.
A
pragmatic
reason
for
accepting
simple
measures
of
acreage
and
persistence
is
that
few
plantings
are
well­
monitored
and
enforcement
of
non­
compliance
with
permit
conditions
is
sporadic
(
sensu
Race
and
Fonseca
1996).
The
parallel
between
seagrass
presence
and
fauna
we
feel
is
strong
enough
to
accept
some
metric
of
seagrass
alone
as
a
viable
indicator
of
functional
recovery.
Although
additional
research
must
be
conducted
to
strengthen
the
seagrass­
faunal
function
link
on
a
broader
geographic
basis,
our
research
suggests
that
monitoring
the
seagrass
itself
is
a
very
useful
option
for
assessing
restoration/
mitigation
success.

Combinations
of
parameters
have
also
been
suggested
as
an
appropriate
metric
for
gauging
success
of
seagrass
plantings
(
F.
Short,
Jackson
Estuarine
Lab.,
Durham,
NH.,
pers
com.).
Combining
factors
that
are
known
to
affect
faunal
abundance,
such
as
shoot
size,
density
and
water
depth
(
taken
together
as
a
measure
of
habitat
complexity
may
be
a
way
to
provide
a
better
means
of
indirectly
comparing
functions
among
planting,
impact,
and
reference
sites,
especially
when
they
occur
in
different
geographic
and/
or
physical
settings.
Short
(
Jackson
Estuarine
Lab.,
Durham,
N.
H.
pers.
com.)
has
proposed
canopy
volume:

(
shoot
density
*
canopy
height;
[
m/
m2=
#
shoots/
m2
*
m/
shoot])

because
it
is
a
value
that
should
change
more
slowly
than
shoot
density
alone
and
thus
be
more
tightly
coupled
to
a
greater
range
of
ecosystem
attributes
beyond
faunal
development
(
e.
g.,
current
speed
reduction,
change
in
sediment
composition,
nutrient
cycling).
Moreover,
the
canopy
volume
metric
can
be
obtained
with
nondestructive
methods
and,
unlike
shoot
density,
would
likely
not
exhibit
overcompensation
responses
with
some
seagrass
species
(
e.
g.,
Syringodium
filiforme:
Williams
1990,
Fonseca
et
al.
1994)
which
do
not
accurately
reflect
long­
term
recovery
from
injury
to
a
seagrass
bed.

Another
factor
in
assessing
functional
equivalency
is
habitat
size.
We
do
not
know
if
there
is
any
relationship
between
the
size
or
shape
of
a
seagrass
bed
and
its
functional
attributes;
this
is
true
for
both
planted
beds
and
natural
beds.
This
is
an
area
of
study
needing
much
additional
work.
From
our
experience,
however,
even
very
small
patches
1­
2
m2
of
seagrass
in
the
Beaufort,
N.
C.
area
has
significantly
greater
numbers
of
fish,
shrimp,
and
crabs
than
found
in
adjacent
sand
areas
(
unpubl.
data).
Moreover,
Murphey
and
Fonseca
(
1995)
found
that
on
a
unit
area
seagrass
basis,
even
beds
in
the
range
of
30­
40
percent
cover
had
penaeid
shrimp
densities
virtually
indis­
Chapter
1:
Background
°
51
52
°
Guidelines
for
the
Conservation
and
Restoration
of
Seagrasses
tinguishable
from
that
of
continuous
cover,
100
percent
beds.
These
data
taken
together
with
years
of
personal
observation
of
seagrass
beds
(
both
planted
and
natural
have
left
us
with
the
impression
that
even
these
very
small,
isolated
patches
provide
resource
functions
comprising
much
of
that
observed
in
more
extensive,
unbroken
coverage
beds
on
a
unit
area
basis.
Thus,
even
very
small
patches
of
seagrass
deserve
protection.