LUGO, ARIEL E., SANDRA BROWN, AND MARK M. BRINSON

Limnol. Oceanogr., 33(4, part 2), 1988, 894-909
0 1988, by the American
Society of Limnology
and Oceanography,
Inc.
Forested wetlands in freshwater and salt-water environments
Ariel E. Lugo
Institute of Tropical Forestry, USDA Forest Service, Southern Forest Experiment
Rio Piedras, Puerto Rico 00928-2500
Station,
Sandra Brown
Department
of Forestry,
University
of Illinois,
Urbana 6 180 I
Mark M. Brinson
Department
of Biology,
East Carolina
University,
Greenville,
North
Carolina
27858
Abstract
A review of data from over 50 freshwater and about 50 salt-water sites revealed that freshwater
and salt-water forested wetlands exhibit parallel responses to hydrologic factors. Greater ecosystem
complexity and productivity
are associated with higher hydrologic energy and more fertile conditions (riverine > fringe 1 basin > dwarf = scrub). However, structural complexity is greater in
freshwater forested wetlands than in salt-water forested wetlands. Net primary productivity,
litter
fall, and export of organic matter are higher in salt-water forested wetlands. These differences raise
questions about the efficiency with which nutrients are used in forested wetlands. Available data
suggest that nutrient-use efficiency by litter fall and litter turnover are higher in tidal salt-water
wetlands than in freshwater wetlands,
Forested wetlands occupy more than
330 x lo6 ha from tropical to boreal latitudes and from seashores to montane cloud
forests (Matthews and Fung 1987). Evidence reviewed by Lugo et al. (1988a) shows
that forested wetlands contribute significantly to the productivity
of coastal and
freshwater fisheries, regulate the quality of
runoff, influence the water budget of then
watersheds, and have measurable effects on
global phenomena. In spite of the ecological
significance of forested wetlands, only a few
studies describe their structure and function. The massive literature on mangroves
(Chapman 1976; Rollet 198 1) is mostly floristic, and the review of Lugo and Snedaker
( 1974) on forest dynamics remains current.
The literature on freshwater forested wetlands is more sparse than that on mangroves
(Lug0 1984; Lugo et al. 1988a); only recently have ecologists addressed the functional aspects of these ecosystems (e.g.
Schlesinger 1978; Mitsch et al. 1977; Brown
I98 1; Ewe1 and Odum 1984; Frangi and
Lugo 1985). The internal dynamics of wetland forests are usually ignored in limnological studies (e.g. Furtado and Mori 1982)
and in studies of tropical lagoons and estuaries (Ayala Castanares and Phleger 1969).
Clearly, forested wetlands deserve more
ecollogical attention, particularly in terms of
their nutrient, carbon, and water dynamics.
1:n this paper, we test the hypothesis that
the structure and processes of forested wetlands are regulated by a few environmental
factors (termed “core factors”), whose effects can be modified by a larger number of
secondary environmental
factors. We accomplish this by comparing the structure
and function of freshwater forested wetlan#ds with those of salt-water forested wetlan’ds. The distinction between freshwater
and salt-water is useful but not necessary to
advance the understanding of these forests.
Unifying principles of wetland ecology exist, and these can be used to accelerate research and overcome the scientific neglect
that has limited the understanding of forested wetlands.
Methods
Only original literature sources were used
in this review (Table 1). Because the structure and function of wetland forests are affected by the age of the stand, we summarized data for stands that were either mature
or in late stages of succession. However,
many of the mangrove forests of the Caribbean (a large subset of our data base) are
impacted by hurricanes, which maintain
894
895
Forestid wetlands
Table 1. Forested wetland study sites used in this review.
(see text) and by latitude within a group.
Forest type or dominant
species
Alnus
Quercus robur
Q. robur
Shallow-water beaver pond
Mixed floodplain forest (Fraxinus nima)
Populus sargentii
Mixed hardwood (4 stands)
Bottomland forest (Acer saccharinum)
Bottomland forest (A. saccharinum)
Bottomland forest (A. saccharinum)
Mixed hardwood (3 stands)
Mixed hardwood (5 stands)
Nyssa- Taxodium
Mixed hardwood
Nyssa aquatica
N. aquatica
N. aquatica
Nyssa- Taxodium
Taxodium distichum
T. distichum
Mixed hardwood and TaxodiumNyssa
Taxodium-Nyssa
(3 sites)
Mixed hardwood and Ny.ssa- Taxodium
Nyssa sylvatica
Nyssa- Taxodium
Liquidambar-Celtis,
Carya-Fraxinus,
N. aquatica-N. sylvatica, and Nyssa- Taxodium (4 stands)
T. distichum
T. distichum
T. distichum
Pterocarpus oflcinalis (3 stands)
Palm floodplain
Mixed hardwood
Eugenia swamp
Igapo forest
Inundation
forests
Avicennia marina
Rhizophora mangle
R. mangIe
R. mangle
R. mangle
Laguncularia
racemosa
R. mangle
R. mangle
R. mangle
Rhizophora sp.
Pelliciera rhizophorae
Rhizophora sp.
Rhizophora apiculata
Sites are grouped according to hydrologic
Location
Site
No.
Riverine-freshwater
Washington (48”N)
Czechoslovakia (48”N)
Czechoslovakia (48”N)
Alberta (5 1”N)
Michigan (42”N)
regime
Rcfcrence
Fonda 1974
Vyskot 1976a
Vvskot 1976b
Hbdkinson 1975
Merritt and Lawson
1979
6
7
8
Lindauer 19 8 3
Buell and Wistendahl 195 5
Brown and Peterson 1983
Illinois (4O”N)
Illinois (40”N)
Virginia (39”N)
New Jersey (39”N)
Illinois (37”N)
North Carolina (36”N)
North Carolina (36”N)
North Carolina (36”N)
North Carolina (36”N)
Alabama (32”N)
Florida (30”N)
Florida (3O”N)
Louisiana (30”N)
9
10
11
12
13
14
15
16
17
18
19
20
21
Johnson and Bell 1976
Peterson and Rolfe 1982
Hupp 1982
Ehrenfeld and Gulick 198 1
Mitsch et al. 1977
Mulholland
1979
Brinson et al. 1980
Brinson et al. 198 1
Brinson 1977
Hall and Penfound 1943
Brown 1981
Nessel 1978
Conner and Day 1976
Louisiana
Louisiana
(3O”N)
(30”N)
22
23
Conner et al. 198 I
White 1983
Louisiana (30”N)
Louisiana (30”N)
Florida (3O”N)
24
25
26
Hall and Penfound 1939a
Hall and Penfound 19393
Elder and Cairns 1982
Florida (26”N)
Florida (26”N)
Florida (26”N)
Puerto Rico (17”-19”N)
Puerto Rico (18”N)
Panama (9”N)
Malaysia (3”N)
Manaus, Amazonas, Brazil
(3”N)
Central Amazonia, Brazil
27
28
29
30
31
32
33
34
Burns 1978
Duever et al. 1984
Duever et al. 1975
Alvarez-Lopez
1988
Frangi and Lugo 198 5
Golley et al. 1975
Furtado et al. 1980
Adis et al. 1979
35
Irmler and Furch 1980
36
37
38
39
40
41
42
43
44
45
46
47
48
Briggs 1977
Lugo and Snedaker 1974
Pool et al. 1977
Snedaker and Brown 198 1
Sell 1977
Pool et al. 1977
Pool et al. 1975
Pool et al. 1977
Martinez et al. 1979
Hernandez et al. 1980
Pool et al. 1977
Golley et al. 1975
Ong et al. 1980a,b
Colorado (38”-41”N)
New Jersey (41”N)
Illinois (40”N)
Riverine-salt
water
Australia (34%)
Florida (26”N)
Florida (26”N)
Florida (26”N)
Florida (26”N)
Mexico (23”N)
Puerto Rico (18”N)
Puerto Rico (18”N)
Puerto Rico (18”N)
Colombia (11”N)
Costa Rica (lOoN)
Panama (8”N)
Malaysia (6”N)
896
Lug0 et al.
Table 1. Continued.
-~
Forest type or dominant
species
R. apiculata
Rhizophora sp.
Site
No.
Location
Malaysia (6”N)
Colombia (3”N)
Fringe-salt
Mangroves*
Mixed
Florida
Mixed (2 sites)
Mixedmonospecific
(5 sites)
Monospecific
Mixed
Mixed
Mixed
Monospecific (2 sites)
Mixed/monospecific
(2 sites)
Monospecific (4 sites)
Mixed/monospccific
(7 sites)
Monospecific
Mixed/monospecific
(4 sites)
Mixed/monospecific
(3 sites)
Mixed
Mixed
Mixed (7 sites; ages, 5 yr-mature)
Monospecific
Monospecific (2 sites)
Ong et al. 1979 --Hernandez and Mullen
51
52
53
54
55
56
57
58
59
60
61
Twilley 1982; Lugo and Snedaker 1974
Pool et al. 1977
Lugo and Snedaker 1975
Heald 1971
Pool et al. 1977
J. H. Day, Jr. pers. comm.
Pool et al. 1977
Martinez et al. 1979
Martinez et al. 1979
Levine 198 1
Martinez et al. 1979
62
63
Golley et al. 1962
Martinez et al. 1979
64
Pool et al. 1977
65
Pool et al. 1977
66
67
68
69
Christensen 1978
Ong et al. 1980a,b, 1982
Bunt 1982
Schaeffer-Novelli
and Cintron
1980
Reader and Stewart 1972
Reader and Stewart 1972
Bay 1967
Bay 1967
Verry 1975
Parker and Schneider 1975
Schwintzer 1977
Reincrs and Reiners 1970; Reiners 1972
Reiners and Reiners 1970; Reiners 1972
Dabel and Day 1977; Day
1982; Gomez and Day 1982
Dabel and Day 1977; Day
1982; Gomez and Day 1982
Dabel and Day 1977; Day
1982; Gomez and Day 1982
Dabel and Day 1977; Day
1982; Gomez and Day 1982
Schlesinger 1978
Best 1984
Deghi et al. 1980; Brown 198 1;
Dierberg and Ewe1 1984
Brown et al. 1984
(22”55’N)
Florida (25’50’N)
Southern Florida (2S”-26”N)
Southern Florida (25”-26”N)
Mexico (22”30’N)
Mexico (1 S”40’N)
Puerto Rico ( 18” 15’1~)
Puerto Rico ( 18” 15’N)
Puerto Rico ( 18” 14’1~)
Puerto Rico (1 S”08’N)
South coast, Puerto Rico
(17’58’N)
Puerto Rico (17’58’N)
South coast, Puerto Rico
(17”55’-57’N)
South coast, Puerto Rico
(17’57’N)
Santa Rosa, Costa Rica
(lO”45’N)
Phuket, Thailand (8(N)
Peninsular Malaysia (6”N)
Australia (8”-15’S)
Florinapolis,
Brazil (27”5O’S)
Basin-freshwater
Manitoba (5O”N)
Manitoba (50”N)
Minnesota (47’30’N)
Minnesota (47’30’N)
Minnesota (47’30’N)
Michigan (46’30’N)
Michigan (45’30’N)
Minnesota (45”N)
70
71
72
73
74
75
76
77
Marginal
Minnesota
78
(45”N)
Virginia
(37”N)
79
Virginia
(3 7”N)
80
Virginia
(37”N)
81
Mixed hardwoods
Virginia
(3 7”N)
82
T. distichum var. nutans
T. distichum var. nulans
T. distichum var. nutans (5 sites)
Georgia (3 I “N)
Georgia (3 1ON)
Florida (29.5”N)
83
84
85
Scrub T. distichum
sites)
Florida
86
T. distichum
swamp
Acer- Nyssa
Chamaecyparis
thyoides swamp
var. nutans (2
(26”N)
1979
water
I’icea mariana bog
Muskeg
P. mariana bog
Groundwater bog
I? mariana bog
Alnus rugosa swamp
Bog forest
Thuja occidentalis swamp
fen
Reference
49
50
Forested wetlands
897
Table 1. Continued.
Forest type or dominant
Basin-salt
Mangroves?
Mixed (3 sites)
Monospccific
(2-4 sites)
Mixed
Mixed (5 sites)
Monospccific/mixed
Mixed (2 sites)
* Monospecific
and Australia;
in S.E. Asia.
t Monospecific
Site
No.
Location
species
water
South Florida
(26”N)
South Florida
(26”N)
Mexico (23”N)
Puerto Rico (18”N)
Puerto Rico (18”N)
Brazil (25’S)
(I 4 sites)
Reference
Cintron et al. 1985; Twilley et
al. 1986
88 Lugo and Snedaker 1974; Twilley et al. 1986
89 Pool et al. 1977
90 Pool et al. 1975, 1977
9 1 Cintron et al. 1985
92 Cintron et al. 1985
87
fringe mangroves are dominated
by R. mangle in Florida and tropical America and by R. apiculata or Rhizophoru spp. in S.E. Asia
mixed fringe mangroves also contain Avicennia gerrninans and/or L. raccmosa in tropical America and Florida and Bruguiera spp.
basin mangroves
are dominated
by
A. germinans; mixed stands also contain
them in a younger state. The sample area
in most studies ranged from about 0.05 ha
to about 0.2 ha. Variation in area sampled
affects the estimates of the number of species
since species richness and area are related.
However, in forested wetlands, where low
species numbers growing in discrete zones
are the rule, the effect of sampled area on
species richness is expected to be less important than in upland forests. Measurements of forest structure (including trees
with diameters at breast height ~2.5 cm),
biomass (harvest and allometry),
aboveground biomass production, litter fall, litter
standing stocks, and litter decomposition
followed the standard methods outlined in
Newbould (1967). Gas exchange methods
for estimates of stand respiration, gross primary productivity,
net primary productivity, and transpiration
are described clsewhere (Brown 198 1; Lugo and Snedaker
1975).
We used hydrologic factors to stratify our
comparison of salt-water and freshwater
forested wetlands. Only subjective appraisals of the nutritional
environments of forested wetlands were possible because of generally inadequate information. The problem
of classifying forested wetlands was cnormous, even when the wetland environment
was simplified to four core factors (one nutritional and three hydrological variables:
Table 2) because the possible combinations
of factors and intensities were numerous and
largely unquantified. We thus adopted the
R. mangle and L. racemosa.
classification suggested for mangroves by
Lugo and Snedaker (1974) and modified by
Cintron et al. (1985). This system incorporates the four core factors but is based on
geomorphology and recognizes three wetland groupings. These are the riverine, basin, and fringe types (Table 2).
The core factors that define the fundamental niche of a wetland forest can be classified as hydrological and nutritional (Lug0
et al. 1988b). Hydrologic factors include the
kinetic energy of flow (e.g. waves, tides, water
runoff), the predominant direction of water
flows (whether water flows through the wetland unidirectionally
as in river floodplains,
bidirectionally
as in fringe forests, or fluctuates vertically as in basin forests), and the
hydroperiod which includes both duration
and frequency and can range from occasionally saturated soil to long-term flooding. Nutritional
factors refer to the nutrient
quality of the site, including sediments and
waters, which ranges from extremely nutrient-poor to nutrient-rich.
In general, trees will not form forested
wetlands if the hydrologic energy is excessive as in high-energy coastal zones, or if
the hydropcriod is excessively long and the
water too deep. Trees require flood-free
conditions to reproduce and water shallow
enough to prevent waterlogging of seedlings
or adventitious
gas exchange structures
(Hook and Crawford
1978; Kozlowski
1984aJ). Some mangrove species can regenerate in areas with long hydroperiods and
Lug0 et al.
898
Table 2. Core environmental
factors for different types of forested wetlands.
ranking from high (1) to low (4) and are based on the perception of the authors.
Core environmental
~Hydroperiod
--
-Wetland
tYPc
Kinetic
energy of
water flow
Freshwater
Riverine
3
Fringe
4-5
Basin
Predominant
direction
of water flow
Duration
Parallel to forest
3
1
5
Perpendicular to
forest
Vertical fluctuation
2
Salt water
Riverine
2
Parallel to forest
2
Fringe
1
1
Basin
4
Perpendicular to
forest
Vertical fluctuation
2
waters as deep as 1 m, but in order to survive they must have continuous tidal water
motion and seeds that germinate on the trees
prior to dispersal. Any stabilization of the
hydroperiod or reduction in tidal flow results in massive tree mortality (Lug0 et al.
I98 1; Lugo and Brown 1984; Jimknez et al.
1985). Unfortunately,
hydrological factors
that delimit the forested wetland are inadcquately quantified, and the tolerances of
forested wetland species have seldom been
verified experimentally.
Nutritional
conditions do not appear to
limit the establishment or distribution
of
forested wetlands. Wetland forests can grow
on rich alluvial soils (e.g. many floodplains
in the southeastern United States: Brinson
1988) or on nutrient-poor sites [e.g. ombrogenous peat forests (Moore and Bellamy
1974; Brown 1988) and white siliceous sands
in south Florida (Brown 198 l), Puerto Rico
(Figueroa et al. 1984), and the Venezuelan
Amazon
(Klinge
and Herrera
198 3)].
Floodwaters of forested wetlands can be
equally diverse in their nutritional
quality
(Moore and Bellamy 1974; Furtado and
Mori 1982; Ewe1 and Odum 1984; Brinson
1988; Brown 1988). However, nutritional
factors interact with hydrological
factors
and, in some cases, override the effect of
hydrology. Nutrients regulate the accumu-
Numbers
represent relative
faclors
--
Frequency
Seasonal rains or
floolds
Seasonal with lake
level
Seasonal rains
At least monthly
spring tides
Daily
At least monthly
spring tides
Nutritional
Nutrient-rich
factors
(1)
Nutrient-rich
or poor depending
on lake waters (3)
Nutrient-rich
or poor depending
on water source (3)
Nutrient-rich
Usually
(2)
nutrient-poor
(4)
Nutrient-rich
or poor depending
on sediments (3)
--
lation of biomass in a forest (Vitousek and
Reiners 1975) and the structure of leaves
(i.e. sclerophylly: Loveless 1962; Small 1973;
Schlesinger and Chabot 1977). Under severe nutrient limitations,
stunted vegetation (dwarfness or scrubiness) occurs.
We believe that analysis of the data presented here will not benefit from elaborate
statistical analysis; there are too many gaps
and uncontrolled variables to make such an
approach profitable. However, by presenting stand data by geomorphological
grouping and separating salt-water from freshwater stands, a significant step is made to
reduce variability
and discover trends that
cart be discussed. The trends that we describe should be viewed as hypotheses. We
hope that this presentation will stimulate
more comprehensive research over a broader range of wetland conditions so that these
trends or hypotheses can be better tested.
For the purpose of our review, we will
consider forested wetlands to be freshwater
if mangroves are absent, since it is difficult
to separate salt-water and freshwater wetlands by soil salinity. For example, AlvarezLopez (1988) suggested that 10’%0was an
upper limit for the salinity tolerance of Pterocarpus oficinalis (considered a freshwater
species) growing in Caribbean coastal zones.
However, mangroves may grow at salinities
Forestep! wetlands
< 1OYm(Carter et al. 1973), but they are not
considered freshwater mangroves even when
located many kilometers inland (Brinson et
al. 1974; Lugo 198 1). Moreover, in temperate latitudes, mangrove forests would be
excluded and replaced by salt marshes at
salinities < 1O%O.
Biotic responses to freshwater and
salt- water wetland environments
All results are summarized in Table 3 (averages, ranges, and sample size), and the
trends described below refer to the data in
this table.
Forest structure-The average number of
tree species decreases in the order riverine,
basin, and fringe in both freshwater and saltwater forested wetlands. However, the range
in the number of species in a given wetland
type is a function of other site factors, such
as hydroperiod and hydrologic energy. The
number of tree species decreases with increasing intensity of hydroperiod and hydrologic energy (cf. Tables 2 and 3). Salinity
also decreases species richness; mangroves
have a lower number of tree species than
freshwater wetland forests. Alvarez-Lopez
( 1988) found that coastal riverine Pterocarpus forests with low salinity (3?&) had fewer
tree species than basin or riverine forests in
freshwater.
No clear pattern emerges with respect to
basal area of wetland forests. Basal area has
a higher average in freshwater basin forests
than in salt-water forests, and there was no
difference between basin and riverine freshwater forests. Riverine forests appear to have
comparable basal areas irrespective of the
presence or absence of salinity. However,
within salt-water wetlands, basal area decreases at higher soil salinities (Cintrbn et
al. 1978). Comparable data for fringe forests
are not available.
Tree density is higher in basin forests than
in riverine forests, with higher densities in
both types of mangrove forests. Age may
explain some of these differences because
younger stands have higher tree densities,
and most mangrove data are from the Caribbean where hurricanes maintain young
forests. In spite of the age differences, the
high tree density in fringe mangroves suggests that increasing hydrologic energy af-
899
fects this parameter. In fact, the width of
the tangle of fringing mangrove trees on
tropical coastlines increases with increasing
wave energy (Cintron et al. 1978). The high
tree densities in basin forests (both freshwater and salt-water) cannot be explained
by hydrologic energy, as these systems grow
in environments with the lowest hydrologic
energy (Table 2). In these cases, aeration is
a problem, and it is possible that a greater
trunk surface area (high basal area produced
by many small-diameter trees) facilitates gas
exchange through the spongy barks and buttresses of many of the trees that grow in
basin conditions (Hook and Scholtens 1978;
Brown et al. 1979).
Data on biomass of forested wetlands are
very limited, and comparisons are only possible in riverine wetlands. Freshwater riverine wetlands have more aboveground biomass than riverine mangroves, which in turn
have a larger amount of belowground biomass than other wetlands types (Table 3).
However, the excavation of roots and particularly the separation of live and dead roots
is difficult. This methodological
problem
may inflate the estimate of live root biomass
in these ecosystems.
Leaves of freshwater and salt-water wetland forests also show differences. Although
there is little difference in leaf biomass between these systems, the leaf area index
(LAI) is much smaller in salt-water systems.
The LA1 of 11 salt-water stands averaged
2.4 (range 0.2-4.4: Golley et al. 1962; Burns
pers. comm.; Carter et al. 1973; Pool et al.
1975), while that of 14 freshwater stands
averaged 4.9 (range 2.0-8.5: Holdridge
et al. 1963; Williams et al. 1972; Brunig
1970; Brown 1981; Frangi and Lugo 1985).
The result is a lower specific leaf weight
(leaf biomass/LA1
= g cmm2 leaf surface) for freshwater forested wetlands than
for mangroves. These results may be interpreted as a greater degree of leaf sclerophylly
in salt-water forested wetlands than in
freshwater ones. Leaf sclerophylly in wetland forests has been attributed to low nutrients (Small 1973), high salinity (Saenger
1982), and water relations (Brunig 197 l), or
a combination of factors (Lug0 et al. 1988~).
Of these causal factors, salinity is the only
one unique to mangroves. Leaf sclerophylly,
900
Lug0 et al.
Table 3. Summary of averages and ranges of values for structural indices, standing stocks, and rates for
forested wetlands. Data are from the sites in Table 2. (Number of sites in parentheses.)
Riverine
Parameter
(units)
Structural indices
Avg. No. species
Range
Sites
Density (No. ha-‘)
Range
Sites
Basal arca (m* ha-‘)
Range
Sites
Standing stocks
Aboveground biomass
(t ha-l)
Leaves
Range
Sites
Total aboveground
biomass
Range
Sites
Belowground
(t ha-‘)
Range
Sites
biomass
Litter (t ha-‘)
Range
Sites
Rates
Gross primary productivity (g m-2 d-‘)
Site
24-h plant respiration
(g m-* d-’ 1
Site
Fresh
Fringe
Salt
Fresh
Basin
Salt
Fresh
Salt
2.6(16)
l-4
44,46
-_
-_
--
1.7(32)
l-3
51, 52, 556 I, 63-65
6(17)
1-14
70-74, 7686
2(5)
l-3
87, 89-92
2,131(16)
400-4,670
44, 46
--_
--
4,005(33)
440-l 6,760
51,52, 5561, 63-65
2,834(14)
1,440-7,820
70, 71, 7686
3,580(5)
2,468-5,130
87, 89-92
33.2( 16)
I 1.5-96.4
44,46
-_
---
22.2(33)
6.0-43.0
51, 52, 5561, 63-65
39.9( 15)
9.5-70.8
72-74, 7686
18.5(5)
15.2-23.2
87, 89-92
5.6(8)
2.8-12.1
2, 19, 20,
27, 28,
31, 32
5.6(3)
3.6-9.5
37,47
----
-
5.0(12)
2.3-10.8
70, 71, 75,
77-83,
85, 86
-
242(17)
79-608
3, 9, 12, 14,
19-21,
27, 28,
31, 32
170(4)
98-279
36, 37, 47
-_
-_
--
111.5(7)
49-159
53, 62, 66
163.6(13)
4-345
70, 71, 75,
77-83,
85, 86
-
43(9)
172(2)
--
50(L)
17.6(7)
-
12-84
3, 14, 16,
20, 27,
28, 31, 32
154-190
36,47
---
62
7.8-3 1.O
70, 71, 7982, 86
-
-
50.4(4)
22.7-102.1
37,47
----
3 1.8(6)
0.3-98.4
53,66
4.9(7)
4.0-5.5
77-82, 86
3.1(5)
0.6-3.6
87, 88
52.1(l)
24.0( 1)
--
13.5(l)
25.3(l)
18.0(l)
19
46.4( 1)
37
11.4(l)
---
51
3.8( 1)
85
2 1.9(l)
88
12.4(l)
19
37
--
51
85
88
8.3(30)
l-23
1, 6-8, 11,
12, 14,
15, 18,
19, 21,
23, 24,
25, 30
1,076(29)
7 l-2,730
1, 2, 6-8,
12, 14,
15, 18,
19, 21,
23-25, 30
37.8(32)
12.0-92.3
1, 2, 6-8,
11, 12,
14, 15,
18, 19,
21, 2325, 30
Forested wetlands
Table 3.
901
Continued.
Riverine
Pnramcter
(units)
Net primary productivity (g m-2 d-*)
Site
Litter fall (1 ha-’ yr-I)
Range
Sites
Wood biomass production (t ha-l yr-I)
Range
Sites
Total aboveground
biomass production (t ha-’ yr-‘)
Range
Sites
Litter decomposition
(kg v-‘1
Range
Sites
Nutrients in litter fall
(kg ha-’ yr-I)
Nitrogen
Range
Sites
Phosphorus
Range
Sites
Potassium
Range
Sites
Calcium
Range
Sites
Magnesium
Range
Sites
Within-stand
Nitrogen
Range
Sites
nutrient-use
Fringe
Fresh
Salt
Fresh
Basin
Salt
Fresh
Salt
5.7(l)
12.6(l)
9.7( 1)
3.4( 1)
5.6( 1)
19
5.7(16)
3.2-l 7.0
3, 8, 13, 14,
19,2022, 27,
28,31
6.94(16)
37
13.0(8)
11.2-16.0
39, 40, 42,
45,48-50
51
7.1(11)
4.3-9.5
60, 67, 68
85
5.3(10)
2.5-7.6
77-83, 85,
86
88
6.6(8)
4.8-9.7
87, 88, 90
-
-
3.25(8)
-
1.77-l 7.88
3, 8, 13, 14,
19-22,
27, 28, 31
12.65(16)
-
-
0.48-5.41
75, 77, 78,
83, 85, 86
-
10.2(l)
-
5.96(10)
-
49
-
-
0.72-10.29
70, 71, 75,
77, 78, 83,
85, 86
-
-
-
2.5(9)
0.75(7)
3.4(7)
0.21-4.95
4, 5, 10, 17,
20, 27,
29, 31, 35
-
-
0.85-8.39
53, 54, 56,
62, 69
0.25-1.87
79-82, 85,
86
1.5-6.0
88
78.6(4)
60.1-86.7
10, 15, 34,
35
-
-
46.5(11)
15-86
60, 67, 68
3 1.0(7)
20.8-45.0
77-83
62.0(3)
53-70
87, 88
3.66(5)
1.5-8.1
10, 15, 31,
34, 35
22.3(4)
17.0-30.6
10, 15, 34,
35
70.8(4)
29.9-l 29.2
10, 15, 34,
35
-
-
4.7(8)
1.2-9.0
67, 68
2.8(8)
1.2-6.5
77-83, 85
-
-
-
-
5.4(5)
3.3-9.5
79-83
-
-
-
46.3(7)
23-9 1
77-83
-
18.6(4)
7.4-37.2
10, 15, 34,
35
-
-
7.7(7)
4.7-l 2.0
77-83
-
150(7)
90-202
48,49, 67
-
90(7)
80-l 20
77-83
123(3)
60-l 65
87, 88
6.68-21.36
3, 8, 13, 14,
19-22,
27, 28, 31
0.93(32)
efficiency
-
-
96.6(7)
65-l 25
67
-
ratio*
85(8)
48-l 16
10, 15, 3335
-
179(11)
90-294
60, 67, 68
902
Lug0 et al.
Table 3.
Continued.
Riverine
-Parameter
Fringe
-
(units)
Fresh
Salt
Fresh
Basin
Salt
--
Fresh
Salt
Phosphorus
Range
Sites
3,1 OO(‘7)
600-4,900
10, 15, 31,
33, 34, 47
1,300(5)
910-1,630
48, 49, 67
----
1,865(8)
9 14-3,567
67, 68
1,100(8)
633-2,527
77-83, 85
-
Calcium
Range
Sites
80(5)
39-222
10, 15, 3234
5.6( 1)
99(5)
65-103
48, 49, 67
...-
80(7)
65-103
67
70(7)
47-85
77-83
-
-
.-
-
3.2(2)
2.2(3)
2.6-3.8
85, 86
l-6-2.8
88
Evapotranspiration
(mm d-l)
Range
Sites
* LAler
fall : nutrient
content
in litter
-
-
.-
-
19
-
.-
-
fall (from
Vitousck
1984).
however, is not unique to mangroves because freshwater wetlands growing on nutrient-poor
soils also exhibit leaf sclero-.
phylly (Sobrado and Medina 1980; Larsen
1982; Lugo et al. 1988c).
Primary productivity and evapotranspiration-Gross primary productivity
measured with similar gas exchange methods is
higher in freshwater forested wetlands than
in mangroves when corresponding types of
sites are compared. However, 24-h plant
respiration is also higher in freshwater wetland forests; thus, net primary productivity
is higher in salt-water forested wetlands. The
higher respiration in the freshwater wetland
forests may be due to the greater biomass
maintained in these older forests.
Information on annual aboveground biomass production is sparse; thus, strict comparison is unreliable. However, riverine forests are always more productive
(about
twofold higher) than the other forested wetland types for which information
is available. Obviously,
some forested wetlands
may be highly productive (mainly riverine
types), but others are not.
Evapotranspiration
decreases among
freshwater forest types in the same order as
the gross primary productivity
and may be
higher than in salt-water forests. Basin forested wetlands have been shown to use water
efficiently
during production
of organic
matter and to reduce water loss relative to
open water (Brown 1981; Brunig 197 1).
Litter dynamics- The average rate of litter fall is higher in salt-water than in fresh-
water wetlands. The rates for salt-water forests follow the order expected from a ranking
based on hydrologic energy and nutritional
factors (riverine > fringe > basin). The faster leaf turnover in mangroves has been sugges.ted to be a mechanism to eliminate salts
thalt accumulate in leaves, but there is no
exlperimental
evidence to support this
(Clough et al. 1982). Pool et al. (1975) suggested that leaf turnover in mangroves was
rekated to water turnover. However, dwarf
(small-size trees with normal leaf size) and
scrub (small-size trees with small leaves)
forests and certain fringe forests on rocky
shores have slow leaf turnovers regardless
of salinity or hydrologic conditions (Lug0
and Sncdaker 1974; Brown 198 1; Lugo
19886). These types of forests appear to be
responding to nutritional limitations or other stressors.
There is no clear trend regarding litter
standing crop. More litter accumulates in
freshwater basin forests than in salt-water
counterparts, but salt-water riverine and
ftinge forests accumulate large amounts of
litter. Litter standing crop in mangroves is
variable because tidal forces may either export or import litter, depending on conditions. Although the predominant direction
of transport is away from the wetland (see
below), litter pileups are common in certain
fringe mangroves (Lug0 1988b).
Rates of litter decomposition (measured
by the loss in mass from standard decomposition bags) are much higher in salt-water
th;an in freshwater forests. Litter fragmen-
Forested wetlands
tation and transport by tidal forces may explain some of these high decomposition
constants. However, these processes contribute to litter turnover, which is clearly
higher in salt-water than in freshwater environments.
Mangrove forests also have higher total
organic carbon (TOC) concentrations
in
their waters [range of 9-22 mg liter-’ (Twilley 1985) vs. 1 l-l 5 mg liter-l in freshwater
forests (Day et al. 1977; Mulholland
and
Kuenzler 1979); sites 14, 2 1, and 5 1 in Table l] and higher rates of carbon export than
freshwater forested wetlands. For example,
four studies in mangroves averaged 198 g
C m-2 yr-l of TOC export (SE = 105: Golley
et al. 1962; Heald 197 1; Boto and Bunt 198 1;
Twilley 1985) vs. a mean of 11.6 g C m-2
yr-l for 17 studies in freshwater forested
wetlands (SE = 2.3: Lugo et al. 1988~).
Twilley (1985) showed that tides transported 66% of the carbon exported from a
basin mangrove wetland. The balance was
transported by freshwater runoff. He also
showed that organic carbon export in mangroves was a function of cumulative tidal
amplitude. This is a clear demonstration of
the contribution of tidal energy to the energy
signature of basin forests and raises questions about how frequently flushed wetlands
conserve nutrients.
Nutrient dynamics- The ratio of mass of
organic matter to mass of nutrients in litter
fall provides some insight into how much
carbon is returned to the forest floor per unit
of nutrient mass. If this ratio is known for
live tissue, it is possible to infer reabsorption of nutrients prior to leaf fall. Vitousek
(1984) termed the ratio of mass of litter fall
to mass of nutrients in litter fall the “withinstand nutrient-use efficiency” (WNE). He
suggested that such a ratio could be used to
compare possible nutrient cycling strategies
of ecosystems because when a nutrient became limiting
to the system, the corresponding ratio would increase in absolute
value. Using this ratio for comparison, it
appears that mangroves return more organic matter to the forest floor per unit of N
and Ca nutrient return than do freshwater
wetland forests. The pattern of P is not clear,
but the WNE ratio for the fringe mangrove
forest is notably high.
903
High ratios of WNE suggest high retranslocation rates in plants, which implies conservative use of nutrients. Such conservative use of nutrients must be of high adaptive
value in forested wetlands exposed to frequent flushing. In fact, Twilley et al. (1986)
documented increases in the WNE in black
mangrove stands subjected to more frequent tides.
Constraints on comparisons
On a global scale, forested wetlands are
extremely diverse in terms of species composition, physiognomy of vegetation, and
location, and there is no agreement on how
to classify this global diversity. A review of
the classification schemes of forested wetlands revealed that dozens of schemes, some
with more than 50 forest types, have been
used (Lug0 1988a). For this reason, comparisons are extremely difficult. Factors that
can affect the results of comparisons include
latitude, climate, topography, substrate, age,
and even site-specific environmental anomalies. For example, forested wetland vegetation is very sensitive to environmental
gradients which induce plant zonations and
compound the problem of comparing across
or between zones (Snedaker 1982; Lugo
1988b). The problem is exacerbated by the
poor data base, which limits the breadth of
possible comparisons and the use of statistics.
We have made several simplifying
assumptions
when drawing comparisons.
First, we recognize that forested wetlands
are stressed ecosystems. This facilitates
comparisons because flooding, anoxia, or
high hydrologic energy stress wetlands and
override some of the geographical differences that are expected for upland ecosystems. For example, monoculture forests are
more likely to occur in wetlands of tropical,
temperate, or boreal zones than they are in
well-drained
lands. Second, we recognize
that stressors affect wetland forests in proportion to their intensity and vary in their
effect depending on whether they act alone
or in synergy (Lug0 1978). Furthermore,
stressors have different effects on wetland
forests depending upon which sector of the
ecosystem they attack (Lug0 et al. 198 1).
These considerations allowed us to separate
904
Lug0 et al.
core factors from other factors that, although important, are not the determinants
of the “fundamental niche” of forested wetlands (Lug0 et al. 1988b). Third, we sought
wetland groupings that could be identified
easily in the field and could be used to generalize about system structure and dynamics. To accomplish this, we assumed that
the kinetic energy and direction of water
flow, the duration of flooding, and nutrient
supply are the four core factors governing
wetland
responses. These factors are
embedded in three types of wetlands: riverine, basin, and fringe forests, defined by
geomorphological
characteristics
(sensu
Lugo and Snedaker 1974).
The hypothesis being tested here is that
the structure and dynamics of forested wetlands are a function of the physical environment. Factors in the physical environment act in subsidy-stress fashion, and for
this reason it is important to define the complete spectrum of environmental forces that
converge on a given wetland ecosystem. This
spectrum of environmental
forces has been
termed the “biotope” or “energy signature”
in recognition of its subtle variation from
one location to another and its role in driving ecosystem structure
and dynamics
(Odum 1983). An ideal comparison between freshwater and salt-water forested
wetlands would thus involve systems with
the same energy signature except for salinity.
Other factors constraining the comparisons are secondary environmental
factors
that modify the forested wetland environment (and also complete the energy signature). The presence of salt causes changes
in biotic response without fundamentally
changing the overall response of wetland
forests. The same appears to be true of frost,
fire, human stressors, and various climatic
factors. The main effects of the frostline on
forested wetlands appear to be biogeographic and a modifier of other stressors. Mangrove wetlands, for example, are replaced
by salt marshes above the frostline (Lug0
and Paterson-Zucca 1977). Holdridge et al.
( I 97 1) reported more tree species in tropical
freshwater forested wetlands than in analogous forest types above the frostline in
Florida (Ewe1 and Odum 1984). Lugo and
Brown (1984) found that forest tolerance to
chronic flooding decreased with decreasing
nurnber of frost-free days. Further comparisons of wetlands above and below the frostline require consideration
of forests with
similar core factors.
One of the fundamental differences between salt-water and freshwater wetlands is
the greater abundance of sulfate in marine
waters. The sulfate concentration
of fullstrength seawater is 0.92 g S liter-l, while
the world average for rivers is over two orders of magnitude lower, 3.7 mg S liter-l
(Livingstone 1963). The significance of this
difference is twofold. First, sulfate is an electron acceptor for microbial decomposition
of organic matter when oxygen is limiting.
Because of the greater abundance of sulfate
in salt-water wetlands, decomposition is less
likely to be limited by anaerobic conditions
than in freshwater wetlands. This is a partial
explanation for higher decomposition rates
in salt-water wetlands. The second effect is
that hydrogen sulfide, an end product of sulfate reduction, is toxic to many organisms,
thus requiring adaptations for avoiding or
surviving high levels of this gas. In addition
to osmotic stress from salinity, sulfide toxicity is an additional stressor that may help
to segregate plant species in the two wetland
types.
Climatic factors can be used to broadly
categorize forested wetlands. For example,
in life zones (sensu Holdridge 1967) with
high ratios of potential evapotranspiration
to precipitation (PET : P > I), forested wetlands are usually riparian or dependent on
gro,undwater (Brown et al. 1979) and, if marinse, they are likely to be fringe forests (Cintron et al. 1978). Wetlands in these regions
may also be coupled to water sources from
othler life zones with PET : P < 1. Examples
can be found in arid riparian areas with
rivers that drain humid montane regions.
As PET : P becomes << 1, the diversity,
structural complexity, and primary productivity of wetland forest types increases. For
example, basin wetland forests grow in moist
and wet coastal areas, inland areas, and on
mountain slopes where soil saturation is
common. Although significant differences
can be observed in the structure, species
composition, and rates of processes in forested wetlands from different life zones (cf.
Holdridge et al. 197 1 for freshwater and
Forested wetlands
MacNae 1968, Chapman 1976, and Clough
1982 for mangroves), the comparisons must
also be stratified by core environmental factors because any of those factors can easily
override a climatic effect. For example, a
subtropical, moist mangrove forest growing
on acid peats has structure (one species and
trees 1 m high) and rates of leaf fall and
metabolism that are similar to a cypress
scrub forest located in a warm, temperate,
moist life zone in Florida (Brown 198 1). In
this example, nutritional
factors override
climatic considerations.
Summary of trends and
research needs
This review of literature on forested wetland structure and function reveals that it
is possible to group these forests using easyto-identify
geomorphological
criteria that
segregate the hydrologic portion of their energy signature. Accordingly, data show that
structural complexity and rate of ecosystem
processes usually follow the order riverine > fringe L basin (Table 3). Basin and
fringe forests are usually closely ranked and
change positions in a number of parameters
such as tree species richness. A critical research need in this field is a more systematic
quantification
of the environmental
factors
that compose the energy signature of forested wetlands. This is particularly true of
hydrologic and nutritional factors, which are
seldom measured in descriptive
studies.
Very little progress in the analysis of structure and dynamics of forested wetlands will
be possible without more comprehensive
data on environmental
factors. Hydroperiod, hydrologic energy, and nutrient supply
covary in ways that make it difficult to
quantify a total effect.
We have shown that saline and freshwater
forested wetlands have parallel responses to
hydrologic forces. However, salinity has significant effects on forest structure and function. Brown and Lugo (1982) found that the
complexity of vegetation, measured with the
Holdridge index, decreased 1.25 units per
1%0increase in salinity between 35 and 65Ym.
Salinity simplifies ecosystem structure and
at high levels poses physiological problems
to trees that lack adaptations to tolerate salt
(Wainwright
1984). Salinity stress, in synergy with tidal factors, affects carbon and
905
nutrient dynamics of wetland forests by exposing these systems to potentially
high
losses through leaching and transport. Organic matter allocation in mangroves is different from that of their freshwater counterparts; mangroves respire less per unit area
of ecosystem and have higher net primary
production, particularly in the form of litter
fall.
The response of mangroves to the problem of organic and inorganic matter conservation includes leaf sclerophylly and high
WNE (i.e. production of litter with low nutrient concentrations and high C : N ratios).
Microbial degradation must occur at different rates in freshwater and salt-water forests
with litter of different nutritional
quality.
Although litter disappears faster from decompostion bags in mangroves, it is not
known if microbial decomposition follows
the same trend. We expect that microbial
decay is slow in forested wetlands with
sclerophyllous leaves, regardless of salinity
regime, because such leaves are likely to
contain less labile material.
Another implication of the differences in
litter quality in forested wetlands is the role
that detritus may have in downstream ecosystems. Low-quality
detritus will require
longer periods of microbial action before it
is upgraded for use by higher trophic levels
in the food web, Twilley et al. (1986) reported that the high C : N detritus of red
mangroves was upgraded from a ratio of 98
to 33-43 in 240 d, while that of the black
mangrove changed from a ratio of 47 to 1723 in the same time period. Similar changes
in C : N ratio may occur in freshwater wetland forests (Brinson 1977). Because both
freshwater and salt-water wetlands are
coupled to fisheries, much research is needed to understand the importance of time
lags between litter fall and use by high-level
consumers, as well as the relative role of
nutrition and hydrology in the inducement
of these time lags.
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