WETLANDS, Vol. 28, No. 2, June 2008, pp. 473–486
’ 2008, The Society of Wetland Scientists
THE EFFECTS OF RIVER FLOODWATERS ON FLOODPLAIN WETLAND WATER
QUALITY AND DIATOM ASSEMBLAGES
Christine L. Weilhoefer1, Yangdong Pan, and Sara Eppard
Environmental Sciences and Resources, Portland State University
Portland, Oregon, USA 97207
1
Present address:
Western Ecology Division, US EPA
2111 SE Marine Science Drive
Newport, Oregon, USA 97365
E-mail: [email protected]
Abstract: We investigated the effects of river floodpulses on the water chemistry and diatom
assemblages in a floodplain wetland. During the two year study period (November 2003–September
2005), the river and wetland exhibited three periods of surface hydrologic connectivity. The impacts of
flooding depended on flood magnitude and duration. Both the long/high magnitude and short/high
magnitude floods thoroughly mixed river and wetland waters, with conductivity, total nitrogen, and total
phosphorus in the wetland decreasing to levels similar to the river. In contrast, the short/low magnitude
flood did not mix water chemistry. Wetland conductivity, total nitrogen, and total phosphorus remained
elevated. Changes in algal biomass followed changes in water chemistry with the high magnitude floods
producing conditions unfavorable for algal growth. Algal biomass decreased in the wetland coinciding
with the two high magnitude floods. Increases in algal biomass coincided with the short/low magnitude
flood. Wetland and river water column diatom assemblages were dominated by periphytic taxa. The
diatom assemblage in the river and wetland were distinct, except during the short/high magnitude flood.
During this period, floodwaters brought diatoms into the wetland and both systems were dominated by
planktonic centric taxa. Similar diatom taxa were observed in the wetland water column assemblage and
the assemblage collected in settling chambers, although their relative abundances varied. Shifts in the
settling diatom assemblage coincided with periods of flooding, indicating that river floodwaters leave a
discernable signal within this assemblage. Our findings indicate that caution should be exercised when
using diatom-based bioassessment in frequently flooded wetlands as the wetland diatom assemblage is
influenced by river floodwaters and changes may depend on the duration and magnitude of flooding.
Key Words: algae, floodplain, floodpulse, nitrogen, phosphorus, sediment trap
INTRODUCTION
and it is estimated that only 42 high quality, undammed river segments greater than 200 km in
length remain in the lower 48 states (Benke 1990).
Floodplain wetlands can be conceptualized as
existing in two phases: an isolated phase, when the
wetland and adjacent river are physically separated
with no surface water exchange, and a connected
phase, where floodwaters dissolve the ecotone
between the river and the wetland and they can be
viewed as one hydrologic unit. In turn, algal
communities in shallow, floodplain wetlands are
shaped by the interplay of internal forces and the
periodic external forcing of floodwaters. During the
connected phase, floodwaters dominate the wetland’s physical structure, raising water levels,
depositing sediments, and potentially scouring
substrates (e.g., Robinson et al. 1997a, 1997b,
Stromberg et al. 1997). The floodpulse can affect
Floodplain wetlands are intimately linked to
rivers and their watersheds through the pulsing of
floodwaters. The pulsing of river floodwaters is
thought to be the major driver of biota in the
floodplain (Junk et al. 1989). Several studies have
demonstrated the influence of the river floodpulse
on the biota of floodplain aquatic habitats. Primary
productivity and algal species composition and
diversity are closely related to hydrologic regime
(Brinson et al. 1981, Engle and Melack 1993, Garcia
de Emiliani 1993, Ibanez 1998). Anthropogenic
hydrologic modifications to rivers and their drainage
basins, as well as, upland development have changed
the physical, chemical, and biological character of
river-floodplain systems. Approximately 98% of the
rivers in the U.S. are regulated (Vitousek et al. 1997)
473
474
the floodplain wetland algal assemblage in two main
ways. First, floodwaters may influence the wetland
algal assemblage directly by serving as a source of
colonizers that develop within the wetland during
and after flooding. While several studies have found
riverine taxa deposited within the surface sediment
of the wetland after flood events (Hay et al. 1997,
Thoms et al. 1999, Hay et al. 2000, Gell et al. 2002,
Medioli and Brooks 2003), these studies lack the
temporal resolution to demonstrate if river algae
thrive in the wetland following flooding or are
merely deposited. Second, floodwaters can have
indirect impacts, including flushing out wetland
algae or changing environmental conditions (e.g.,
turbidity, nutrients, sediments; Engle and Melack
1993, Robertson et al. 1999) so that the pre-flood
assemblage is no longer favored and a different set
of species can dominate. The shear volume of
floodwater has been shown to dilute water chemistry
(Gell et al. 2002) and biota (Hamilton and Lewis
1987). Floodwaters can also bring in nutrients,
favoring certain algal assemblages (Engle and
Melack 1993). Squires and Lesack (2001) noted
increases in phytoplankton during peak floods
associated with increases in nutrients and decreases
in light. Engle and Melack (1993) observed increases
in periphyton biomass during early flooding when
turbidities were highest and shifts to phytoplankton
species as turbidity in floodwaters decreased. The
influence of flooding on wetland algae may vary with
the duration and magnitude of flooding. Short pulses
of flooding may cause minimal environmental
changes and flushing, and have only minor effects
on the wetland algal assemblages. Longer, higher
magnitude flooding may flush wetland algae and
nutrients and cause a reset that changes the trajectory
of algal development following flooding, thereby
leaving a discernable signal on the wetland biota.
After floodwaters recede, floodplain wetlands
return to the isolated phase, where internal forces
governing shallow systems (e.g., water depth, light,
nutrients, substrate, and trophic interactions) influence the algal assemblage (Scheffer 1990, Goldsborough and Robinson 1996). If water levels are
low and water clarity is high, a macrophyte
community can develop in the wetland, favoring
attached algae. Benthic algae accounted for 75% of
production in clear shallow lakes and 5% in turbid
shallow lakes (Liboriussen and Jeppesen 2003).
Nutrients and turbidity interact to influence the
type of algal assemblage that develops in the
wetland. McDougal et al. (1997) found that wetland
algae shifted from epiphyton to metaphyton as
nutrients were added. However, Burkholder and
Cuker (1991) found that benthic periphyton were
WETLANDS, Volume 28, No. 2, 2008
the major contributors to algal primary production
in shallow reservoirs with elevated nutrients.
Assessment of condition in floodplain wetlands
may be confounded by the changes brought on by
floodwaters. The use of biological indicators in
hydrologically variable wetlands has been cautioned
against due to the influence of water-level history in
shaping biologic assemblages (Wilcox et al. 2002).
While the utility of diatoms in the bioassessment of
lakes, streams, and isolated wetlands has been welldocumented (Dixit et al. 1992, Pan and Stevenson
1996, Stevenson and Pan 1999, Weilhoefer and Pan
2007), the extension of diatom-based bioassessment
to frequently flooded wetlands warrants an examination of the signal floodwaters leave on the wetland
diatom assemblage.
The objectives of this study were: 1) to document the
changes river flooding brings to floodplain wetlands in
terms of habitat, water quality, and diatom assemblages, 2) to examine the effect of flood duration and
magnitude on the changes induced in the floodplain
wetland, and 3) to determine the river floodwaters’
signal on the floodplain wetland settling diatom
assemblage. While studies have correlated seasonal
changes in floodplain wetland sediment diatom assemblages with periods of flooding and drawdown (e.g.,
Gell et al. 2002), we directly examine the river influence
by concurrently sampling river and wetland water
column assemblages. These data will allow us to
determine if a unique flood diatom assemblage exists.
MATERIALS AND METHODS
Site Description
River-floodplain wetland interactions were studied
within the Jackson Bottom Wetlands Preserve,
Hillsboro, OR, USA (Figure 1). The preserve is a
107 ha complex of wetlands and uplands located
within the annual floodplain of the Tualatin River.
The Tualatin River is approximately 130 km in length
and drains a 1,837 km2 watershed. The watershed is
bounded by the Tualatin Mountains to the north and
the Chehalem Mountains to the southwest (Hawksworth 2001). The basin is underlain by a mixed
geology, including basalts, sedimentary formations,
and alluvial depositions. As a consequence of this
geology, some soils in the basin are naturally rich in
phosphorus. Land use in the basin is 50% forested,
35% agricultural, and 15% urban. The region is
characterized by a moderate climate, with warm, dry
summers and cool, wet winters and 72% of precipitation falls between November and March (Hawksworth 2001). Stream flows peak in the winter and
spring due to precipitation and snowmelt.
Weilhoefer et al., RIVER-WETLAND INTERACTIONS
475
aimed at improving urban habitat for wildlife. For
this study, data were collected within a single
wetland within the complex, the ‘‘Gene Pool’’, and
at one location in the Tualatin River upstream of the
Gene Pool wetland (Figure 1). Treated wastewater is
released occasionally into various areas of the
Preserve, including the Gene Pool wetland. During
the course of this study, treated wastewater was
released into the Gene Pool on November 20, 2003
and December 8, 2004, resulting in short pulses of
increased nitrogen and phosphorus concentrations
within the wetland. The Gene Pool wetland was
selected because there is a continuous water quality
monitoring station, providing conductivity, dissolved oxygen, pH, and water level data, located in
this wetland. Summer wetland vegetation is dominated by pondweed (Potamogeton spp.), spike rush
(Eleocharis spp.), and invasive reed canary grass
(Phalaris arundinacea). Submersed and emergent
aquatic vegetation die-back in the winter months.
The Tualatin River is a 5th order river at the
sampling location with dense riparian vegetation.
Field Sampling
Figure 1. Gene Pool wetland and Tualatin River
sampling locations within the Tualatin River Watershed,
Hillsboro, OR, USA.
The Tualatin River and its riparian area have
been extensively modified (Hawksworth 2001).
Water storage in the Upper Tualatin basin has
increased peak winter flows and surface runoff and
reduced summer flows. Agriculture has increased
erosion and subsequently sediment and nutrient
loads to the river. Loss of riparian vegetation and
replacement of native species (Oregon Ash (Fraxinus
latifola), black poplar (Populus balsamifera), big leaf
maple (Acer macrophyllum), and redoiser dogwood
(Cornus stolonifera)) by the invasive shrub Himalayan blackberry (Rubus discolor) is common. Subsequently, water quality in the Tualatin River has
declined, with increases in temperatures, turbidity,
nutrient levels, and bacteria and decreases in
dissolved oxygen levels.
The Jackson Bottom Wetlands Preserve was
chosen for this study because it receives an annual
floodpulse from the Tualatin River. Historically, the
area was bottomland hardwood forest. Most of the
area was drained for farming and sewage disposal.
Since the 1970s, the wetlands within the Preserve
have undergone a series of restoration projects
To assess the changes to wetland water chemistry,
habitat, and algal biomass (chlorophyll-a) brought
about by floodwaters, water samples were collected
weekly in the wetland and river from November
2003 to October 2004 and monthly from October
2004 through September 2005. Samples for diatom
species compositional analysis were collected monthly from the river and wetland water columns
throughout the study. Within the wetland and river,
samples for water chemistry analysis were collected
and water quality variables were measured near the
water’s edge where water depth was approximately
1.5 m. Conductivity and temperature were measured
at mid-depth in the water column with a YSI-85
oxygen, conductivity, salinity, and temperature
meter. Water pH was measured using an Orion
Model 210A meter. Turbidity was measured with a
HACH 2100P Turbidimeter. Water samples were
collected at mid-depth for nutrient analysis at the
same location where water quality measurements
were taken. Two water samples were collected at
each site, one filtered on site (47 mm Millipore type
HA filters, 0.45 mm pore size), the other left
unfiltered and stored on ice until returned to the
laboratory and frozen until nutrient analyses. A
YSI-multiparameter probe outfitted with a water
sensor recorded water levels in the wetland at 1 hour
intervals and was used to determine wetland water
depth associated with the time of sampling.
476
Algal assemblages within the wetland and river
were assessed using measures of biomass (chlorophyll-a) and diatom species composition. The water
column algal assemblage was sampled at the same
location where water quality samples were collected
by collecting the water below the surface in the river
and wetland in a 1 L bottle. Samples were preserved
with formalin (final concentration 4%) and allowed
to settle for 2 days to concentrate algae in the
bottom 50 ml. Algal biomass was sampled by
filtering a known volume of water collected 0.5 m
below the water surface in the wetland and river
(47 mm Millipore type HA filters, 0.45 mm pore
size). Filters were saturated with MgCO3 and frozen
for chlorophyll analysis.
To determine if river flooding leaves a discernable
signal in the floodplain wetland settling diatom
assemblage, sinking particles were collected from
November 2003 and November 2004 using floating
sediment traps. Traps collected settling material over
the course of a month to integrate temporal changes
in algal assemblages in the water column. The
collection device was held constant at 0.5 m below
water surface. A floating device was employed
because water level within the wetland can rise by
more than 3 m during flood periods, making it
difficult to retrieve sediment traps. The collection
device, containing two collection bottles, was
anchored at the deepest part of the wetland near
the location of the water level sensor. The sample
from one bottle was used for analysis of settling
diatoms and the sediment sample in the other bottle
was used to quantify sedimentation rate and
sediment composition (data not presented).
Laboratory Analyses
Nutrient analyses for all samples were performed
during a 2-week period 6 months after sample
collection. Total nitrogen was measured by alkaline
persulfate digestion followed by cadmium reduction
(Ameel et al. 1993). Total phosphorus was measured
by the alkaline persulfate digestion method followed
by the ascorbic acid method (Ameel et al. 1993).
Chlorophyll-a was measured using a spectrophotometer and calculated using the equations of
Lorenzen (Arar 1997). Approximately 25 ml of the
concentrated water column algal sample and settling
material sample were prepared for diatom identification. Samples were digested using concentrated
sulfuric acid and potassium dichromate for
12 hours. Samples were rinsed repeatedly with
deionized water until the pH was approximately
neutral and mounted on slides with Naphrax high
resolution mounting medium. Transects along the
WETLANDS, Volume 28, No. 2, 2008
slide were scanned until at least 600 diatom valves
were identified and enumerated to the species level
using a Nikon Eclipse E600 microscope at 10003
magnification. The primary references for diatom
taxonomy were Krammer and Lange-Bertalot (1986,
1988, 1991a,b, 2000) and Patrick and Reimer (1966,
1975). Taxon dominance was calculated as the
relative abundance of the most abundant taxa at
each site. Habitat metrics (% periphytic taxa, %
planktonic taxa) were calculated based on expert
opinion for the diatom assemblage at each site.
Data Analysis
Principal components analysis (PCA) was used to
examine major gradients in environmental data and
separate sites at each sampling date based on
environmental variables. All environmental data,
other than pH, were normalized by log-transformation prior to analysis. A correlation matrix was used
in the analysis to standardize variances. Time-series
plots were used to examine differences in environmental variables over the two-year study period.
Diatom species data were summarized as relative
abundance within a sample (taxon abundance/total
abundance * 100). Boxplots were used to compare
relative abundances of common diatom taxa and
diatom metrics between the river, wetland, and
wetland settling material during different hydrologic
states. Variables were considered to be different if
quartiles did not overlap. For hydrologic states with
only one sampling point (short/low magnitude and
short/high magnitude floods), variables were considered to be different if means did not overlap
quartiles. Bray-Curtis index was used to calculate
the similarity in the overall diatom assemblages
between the river, wetland, and wetland settling
material for each sampling date. Bray-Curtis similarity includes both species richness and abundance
in its calculation. Sorensen’s similarity index, which
is based solely on presence-absence data and not
taxa abundance, was calculated between the river,
wetland, and wetland settling material diatom
assemblages at each sampling date to determine if
similar taxa were present in both the samples. All
taxa were included in Bray-Curtis and Sorensen’s
indices calculations and diatom data were arcsine
square root transformed prior to analysis. Patterns
in diatom assemblages between river, wetland, and
wetland settling material were examined visually
using non-metric multidimensional scaling ordination (NMDS, PC-ORD v. 4, Bray-Curtis distance
measure, 40 real runs and 400 maximum iterations;
McCune and Mefford 1999). The Monte-Carlo
permutation procedure was used to determine if
Weilhoefer et al., RIVER-WETLAND INTERACTIONS
477
Figure 2. a) Total monthly rainfall at Jackson Bottom Wetlands Preserve and b) wetland water depth. Shaded areas
indicate periods of hydrologic connection.
the axes extracted by NMDS explained more
variation than by chance alone (PC-ORD v. 4, 50
randomized runs). Diatom data were arcsine square
root transformed prior to analysis and species with
relative abundances less than 1% in two or fewer
samples were eliminated from the dataset to reduce
the influence of taxa with only a few occurrences.
RESULTS
Characterizing Hydrologic Connectivity between the
River and Wetland
Rainfall patterns and subsequently hydrologic
connectivity at Jackson Bottom varied between the
two years of the study (Figure 2a). During the first
year of the study, rainfall was high from November
through February, peaking in December 2003
(16.0 cm). There was less rain in the winter months
during the second year of the study, with rainfall
peaking in October 2004 (7.5 cm) and December
2004 (7.7 cm). Heavy rainfall also occurred in
March 2005 and May 2005.
Hydrologic connectivity between the river and
wetland followed the rainfall patterns. The river and
wetland exhibited surface hydrological connectivity
in three periods during the study. The first flood
period was long in duration and high in magnitude,
with the river and wetland exhibiting continual
surface hydrological connections from January 11,
2004 through April 23, 2004. Water levels in the
wetland responded to this connection (Figure 2b).
Water depth ranged between 0.4 and 1.4 m prior to
flooding during the first year of the study. Water
depth gradually increased during the connected
phase, peaking at 3.3 m in February 2004. Water
levels gradually declined throughout the summer,
with occasional small peaks associated with rainfall
events. There were two brief periods of hydrologic
connectivity between the river and wetland during
the second year of the study, in February 2005 and
April 2005. For 5 days in February 2005, the river
just overtopped into the wetland and flooding was
of short duration and low magnitude. Water levels
in the wetland rose to 1.3 m (Figure 2b). For 7 days
in April 2005, the river flooded into the wetland.
This flooding period was short but of high
magnitude. Wetland water levels rose to 2.9 m
during this period (Figure 2b). Water levels in the
wetland declined throughout the summer 2005.
Effects of Flood Duration and Magnitude on
Physicochemical Variables
Floodwaters from the river altered the water
quality of the wetland. The first two PCA axes
explained 64% of the variability in the environmental data (Figure 3). In general, the first PCA axis
(46% of variability) separated wetland sites during
the isolated phase from wetland sites during the
connected phase and from river sites in both the
isolated and connected phases. Conductivity, TN,
TP, and chlorophyll were positively related to PCA
axis 1 (Table 1). An exception to this pattern was the
wetland site during the short/low magnitude flood of
February 2005, when water chemistry variables were
similar to those in the wetland during the isolated
phase.
478
WETLANDS, Volume 28, No. 2, 2008
Table 1. Principal components analysis eigenvector
values for first and second PCA axes.
Total phosphorus (mg/L)
Total nitrogen (mg/L)
Conductivity (mS/cm)
pH
Turbidity (NTU)
Chlorophyll-a (mg/L)
Figure 3. PCA (Principal Components Analysis) plot of
water quality parameters (total nitrogen, total phosphorus, conductivity, pH, turbidity, chlorophyll-a) measured
on each sampling date. Points are labeled by sampling
date and habitat category (habitats connected or isolated).
Turbidity in both the river and wetland was
variable, with changes coinciding with both flood
times and biotic factors (Figure 4a). Turbidity
increased in the wetland at the onset of flooding
and maximum water levels in both years (January
2004 5 22 NTU, February 2005 5 34 NTU, April
2005 5 19 NTU). Turbidity levels in the wetland
decreased rapidly after the initial floodpulse even as
the river and wetland were still in the connected
phase. River turbidity increased during and after
flooding, with peaks of 53 NTU in January 2004,
and 14 NTU in both February and April 2005. River
turbidity was low in the summer, ranging between 3
and 7 NTU. Wetland turbidity increased in July
2004 (25–40 NTU) and late September/early October 2004 (28–51 NTU), associated with floating
green algal mats (personal observation).
The effect of hydrologic connectivity on water
chemistry varied with the duration and magnitude of
flooding. Both the long/high magnitude flood during
the first year of the study and the short/high
magnitude flood during the second year of the study
mixed river and wetland water (Figure 4). Conductivity in the river was slightly lower during times of
connection brought on by high magnitude flooding
than the isolated phases (Figure 4b). Mean conductivity in the wetland was higher during the isolated
phases of both years (year 1 5 332 mS/cm, year 2 5
213 mS/cm) than the connected phase brought on by
high magnitude flooding (year 1 5 123 mS/cm, year
2 5 89 mS/cm). Total nitrogen (TN) and total
phosphorus (TP) in both the river and wetland were
PCA 1
PCA 2
0.49
0.44
0.51
0.29
0.20
0.41
0.18
20.26
0.17
0.57
20.71
20.20
high throughout the study period (river TN range
226 - 4517 mg/L, river TP range 37 - 415 mg/L,
wetland TN range 293 - 7910 mg/L, wetland TP
range 31 - 1975 mg/L; Figure 4c, d). During the
connected period brought on by high magnitude
flooding in both years, wetland TN concentrations
decreased and were similar to that of the river
(Figure 4c). Wetland TP concentrations displayed a
similar pattern to TN, decreasing during times of
high magnitude flooding to levels similar to the river
(Table 2, Figure 4d) TP in the river did not show
strong patterns with flooding.
In contrast to the two periods of high magnitude
flooding, water chemistry in the river and wetland
did not appear to mix during the short/low
magnitude flood (February 2005; Figure 4). Conductivity in the wetland dropped during this
period (211 mS/cm) but not to levels of the river
(118 mS/cm; Figure 4b). Both the river and wetland
exhibited peaks in TN associated with flooding
(Figure 4c). River and wetland TP concentrations
did not appear to be influenced by this flood
(Figure 4d).
Characterizing the Effects of Flood Duration and
Magnitude on Algae
The effects of hydrologic connectivity on algal
biomass and diatom assemblages differed. Patterns
in algal biomass followed those of water chemistry
with both the long/high magnitude flood of the first
year and short/high magnitude flood of the second
year lowering algal biomass in the wetland to levels
similar to that of the river (Table 2, Figure 4e).
Wetland chlorophyll-a concentrations increased
slightly with the onset of the long/high magnitude
flood of 2004, but decreased to levels similar to the
river by the end of the flood period. Wetland
chlorophyll-a concentrations ranged between 1 mg/L
prior to flooding in the first year and were lowest
during early February 2004, corresponding to peak
water levels (Figure 4e). During the short/high
magnitude flood of 2005, wetland chlorophyll-a
levels dropped to 7 mg/L, similar to those of the river
Weilhoefer et al., RIVER-WETLAND INTERACTIONS
479
Figure 4. a) Turbidity, b) conductivity, c) total nitrogen, d) total phosphorus, and e) chlorophyll-a in the wetland (closed
circle) and river (open circle). Shaded areas indicate periods of hydrologic connection.
(4 mg/L; Table 2). In contrast, during the short/low
magnitude flood of February 2005, chlorophyll-a
levels in the wetland peaked and did not decrease to
levels similar to the river (Figure 4e). Wetland
chlorophyll-a levels exhibited peaks in the winter
months of both years. Water column chlorophyll-a
levels in the river were low throughout the study
period, ranging between 0 and 18 mg/L (Figure 4e).
Overall, diatom assemblages in the river and
wetland differed, except for April 2005, coinciding
with the short/high magnitude flood. The wetland
was dominated by periphytic taxa throughout much
of the study. Cocconeis placentula was the most
abundant taxon in the wetland (relative abundance
in samples range 1%–57%) and present at each
sampling date. During the first year of the study, the
relative abundance of C. placentula in the wetland
decreased during the long/high magnitude flood
period. Cocconeis placentula relative abundance in
the wetland was lower during the second year of the
study and decreased during the short/high magnitude flood of April 2005 (Figure 5a). Gomphonema
parvulum (range 0%–12%) and G. cf. lippertii (range
0%–32%) were also common taxa in the wetland,
being present at 95% and 100% of the sampling
dates, respectively. These Gomphonema taxa decreased during the short/high magnitude flood of
April 2005 (Figure 5b). The river diatom assemblage
was also dominated by periphytic taxa (range 18%–
94%), including Achnanthidium minutissimum (range
0%–13%; Figure 5c), Planothidium lanceolatum
(range 1%–13%; Figure 5d), and Rhoicosphenia
curvata (range 1%–10%). Achnanthidium minutissimum was present throughout the year in the river
(Figure 5c). During the short/high magnitude flood
period (April 2005), river and wetland diatom
samples were dominated by similar species, including planktonic taxa such as Aulacoseira ambigua
(Grun.) Simonsen, Cyclotella pseudostelligera Hustedt, and Cyclostephanos invisitatus (Hohn & Hellerman) Theriot, Stoermer & Håkansson.
The responses of diatom assemblage metrics to
hydrologic connectivity were variable between the
river and wetland. Diatom taxa richness was higher
in the river than the wetland for all periods except
the short/high magnitude flood (Figure 6a). River
taxa richness was highest during the long/high
magnitude flood of 2004 (range 82–88). Taxa
richness in the wetland showed no strong trend with
hydrologic connectivity (Figure 6a). For all but the
89
19
133
3933
7
94
14
13
2624
4
211
34
306
7910
117
118
14
111
4517
3
(52)
(6)
(65)
(683)
(23)
123
10
119
1054
20
Conductivity (mS/cm)
98 (48)
239 (132)
109 (62)
332 (111)
105 (19)
213 (104)
78 (25)
Turbidity (NTU)
11 (9)
15 (12)
9 (5)
19 (14)
8 (6)
10 (5)
17 (1)
Total phosphorus (mg/L) 107 (63)
379 (389)
102 (46)
499 (368)
129 (30)
577 (590)
100 (95)
Total nitrogen (mg/L)
1294 (850) 2503 (1988) 1106 (708) 2628 (1540) 1790 (747) 3988 (2616) 1030 (499)
Chlorophyll-a (mg/L)
3 (3)
38 (55)
2 (1)
36 (51)
6 (6)
76 (91)
3 (3)
W
R
W
R
R
R
Overall
W
R
YEAR 1
W
Isolated
R
YEAR 2
W
Long/High
W
Short/Low
Connected
Short/High
WETLANDS, Volume 28, No. 2, 2008
Table 2. Environmental variables (mean and standard deviation in parenthesis) for the river (R) and wetland (W) summarized by hydrologic state. No standard
deviations were calculated for the short/low magnitude flood and short/high magnitude flood as only one data point was collected during these floods.
480
first sampling date, maximum relative abundance
(dominance) of any single taxon in river was less
than 20%. The wetland assemblage exhibited greater
dominance (maximum relative abundance) than the
river during both isolated periods and the long/high
magnitude flood period (Figure 6b). Dominance in
the wetland was lower during both short periods of
connection than at any other time. Shannon-Weaver
diversity index was higher in the river than the
wetland during all periods except the short/high
magnitude flood (Figure 6c). Relative abundance of
periphytic taxa was similar in the river and wetland
during the short/high magnitude flood (Figure 6d).
The response of the diatom assemblage to
hydrologic connectivity differed from the algal
biomass and water chemistry responses. A twodimensional NMDS solution explained 89% of the
variance in the diatom assemblage distance matrix
(Figure 7). River and wetland assemblages were
separated by the first ordination axis, with river
sampling points clustering on the right side of the
ordination diagram and wetland sampling points
clustering on the left side. Except for the short/high
magnitude flood (April 2005), diatom assemblages
in the river and wetland were distinct from one
another. The April 2005 wetland diatom assemblage
was the only sampling point that clustered near the
river sites on the right side of the NMDS ordination.
Bray-Curtis similarity between the wetland and river
diatom assemblage was highest (63%) during this
sampling period (Table 3). Nitzschia dissipata (r 5
0.88), R. curvata (r 5 0.81), P. lanceolatum (r 5
0.81), Achnanthidium pyreniacum (r 5 0.76), and
Asterionella formosa (r 5 0.71) were positively
correlated with the first ordination axis and
abundant in river and wetland samples during
flooding (Table 4). Cocconeis placentula (r 5
20.66) and several Gomphonema taxa were negatively correlated with the first axis, indicating taxa
abundant in the wetland.
Characterizing the Flood Signal on Wetland
Settling Diatoms
The wetland settling diatom assemblage was more
similar to the wetland water column assemblage
than the river assemblage regardless of hydrologic
connectivity, clustering on the far left of the
ordination diagram (Figure 7). Settling assemblages
formed three clusters corresponding to whether they
were collected prior to (left-center of ordination
plot), during (bottom-left of ordination plot), or
post-flooding (top-left of ordination plot), with
assemblages being most unique during the long/high
magnitude flood of 2004. Similar to the wetland
Weilhoefer et al., RIVER-WETLAND INTERACTIONS
481
Figure 5. Relative abundance of dominant diatom taxa in the wetland (W), settling material (S), and river (R)
summarized by hydrologic state. a) Cocconeis placentula, b) Gomphonema parvulum and Gomphonema cf. lippertii, c)
Achnanthidium minutissimum, and d) Planothidium lanceolatum.
water column assemblage, the wetland settling
diatom assemblage was dominated by periphytic
taxa. Cocconeis placentula was the most abundant
taxa (range 3%–63%) and was present at all
sampling dates (Figure 5a). Cocconeis placentula in
the wetland settling assemblage showed a similar
response to flooding as that in the wetland water
column, with lower relative abundances during the
long/high magnitude flood. Gomphonema parvulum
was also a common taxon (range 0%–18%) and was
present during 91% of the sampling dates. Achnanthidium minutissimum was common during
flooding periods in the wetland settling material
(Figure 5c). While similar taxa were found in the
wetland water column and settling assemblage, their
relative abundance varied. Bray-Curtis similarity
between these two assemblages ranged between 25
and 49% during the isolated phases and 36%–41%
during the connected phase (Table 3). Similarity
between the wetland water column and settling
diatom assemblages based on presence absence data
(Sorensen’s similarity measure) were higher (ranges
33%–57%; Table 3) than those based on relative
abundance. Diatom taxa richness, dominance,
Shannon-Weaver diversity, and relative abundance
of periphytic taxa did not differ between the wetland
water column and settling material during periods of
isolation (Figure 6). Wetland settling diatom assemblage taxa richness and Shannon-Weaver diversity
were lower than the wetland water column assemblage and dominance was higher during the long/
high magnitude flood (Figure 6). Similarity between
the wetland settling assemblage and river diatom
assemblages was low throughout the study (Bray
482
WETLANDS, Volume 28, No. 2, 2008
Figure 6. a) Taxa richness (number of taxa), b) dominance (maximum relative abundance of a diatom taxa within a
sample), c) Shannon-Weaver diversity, and d) percent periphytic taxa in the wetland (W), settling material (S), and river
(R) summarized by hydrologic state.
Curtis range: 11%–40%; Sorensen’s range: 16%–
40%; Table 3).
DISCUSSION
Both external forces (e.g., floodwaters) and localized, within wetland forces are instrumental in
shaping the water quality and diatom assemblage of
floodplain wetlands. Flood duration, magnitude, and
frequency may influence the type of environmental
changes brought about in wetlands. In our study,
floodwaters altered wetland habitat and water chemistry and the resulting changes to the diatom
assemblage appear to be related to the magnitude of
the connected state. Both the long/high magnitude
flood (year 1) and short/high magnitude flood (April
2005) mixed river and wetland water chemistry.
Conductivity, TN, and TP in the wetland decreased
to levels similar to the river, possibly due to the large
volume of floodwater brought in by the river and
consequent flushing of the wetland. Agostinho et al.
(2000) also observed homogenization of river and
floodplain water chemistry during high floods. In
contrast, the short/low magnitude flood in February
2005 did not bring in enough water from the river to
thoroughly mix water in the two systems. The effects
of river floodwater on wetland water quality may vary
depending on the nutrient status of the river and the
wetland. While floodwaters were a source of increased
nitrogen and phosphorus to three of five European
floodplain lakes (Van den Brink et al. 1993), Gell et al.
(2002) observed reductions in conductivity and
nutrients in three Australian floodplain wetlands
receiving floodwaters. The Gene Pool wetland is
Weilhoefer et al., RIVER-WETLAND INTERACTIONS
483
Figure 7. Non-metric multidimensional scaling (NMDS) ordination of diatom assemblages in the river water column,
wetland water column, and wetland settling material.
eutrophic, in part due to periodic treated wastewater
releases to parts of the Jackson Bottom Preserve, and
therefore large volumes of river waters served to dilute
its water chemistry. Water levels increased by over
two meters and turbidity increased during these
floods, reducing light levels in the wetland. Lower
algal biomass in the wetland coincided with these high
magnitude flood periods. The two floods of high
magnitude observed in our study appear to have
produced conditions unfavorable for algal growth.
Shifts in wetland diatom assemblages were also
observed during the time of flooding, although the
Table 3. Bray-Curtis (BC) and Sorensen’s (S) similarity indices (mean and standard deviation in parenthesis) summarized
by hydrologic state. Similarity indices were calculated for diatom relative abundance data between the wetland water
column assemblage and the river water column assemblage, the wetland water column assemblage and the wetland settling
material assemblage, and the river water column assemblage and the wetland settling material assemblage. ‘‘..’’ indicates no
data collected. No standard deviations were calculated for the short/low magnitude flood and short/high magnitude flood
as only one data point was collected during these floods.
Isolated
BC: Wetland - River
BC: Wetland - Settling
BC: River - Settling
S: Wetland - River
S: Wetland - Settling
S: River - Settling
Connected
Overall
Year 1
Year 2
37
43
26
47
44
31
31
45
24
41
46
30
35 (6)
..
..
46 (9)
..
..
(10)
(10)
(8)
(11)
(8)
(7)
(7)
(12)
(9)
(10)
(9)
(8)
Long/High
47
38
32
59
41
33
(6)
(2)
(4)
(6)
(3)
(4)
Short/Low
Short/High
48
..
..
58
..
..
63
..
..
62
..
..
484
WETLANDS, Volume 28, No. 2, 2008
Table 4. Correlations of diatom species relative
abundance with non-metric multidimensional scaling
ordination axes.
Diatom Taxa
Cocconeis placentula Ehrenberg
Gomphonema gracile Ehrenberg
Gomphonema clavatum Ehrenberg
Gomphonema cf. lippertii Reichardt &
Lange-Bertalot
Gomphonema parvulum Kützing
Asterionella formosa Hass
Navicula gregaria Donkin
Nitzschia linearis (C. Agardh) W. Smith
Achnanthidium pyreniacum (Hustedt)
Kobayasi
Aulacoseira alpigena (Grunow) Simonsen
Planothidium lanceolatum (Breb. ex Kütz)
Rnd & Bukh
Rhoicosphenia curvata (Kütz) Grunow
Nitzschia dissipata (Kütz) Grunow
Gomphonema kobayasii Kingston and
Kociolek
Axis 1 Axis 2
20.66
20.61
20.59
20.55
0.58
0.47
0.59
0.66
20.55
0.71
0.72
0.73
0.76
0.43
20.21
20.12
20.14
20.12
0.78 20.08
0.81 20.03
0.81 20.01
0.88 20.11
0.84 20.08
specifics of these shifts varied between years.
Cocconeis placentula was more abundant during
the isolated period and decreased during the long/
high magnitude flood of 2004. Relative abundances
of several planktonic species (A. granulata, C.
invisitatus, C. pseudostelligera) increased in the
wetland during both high magnitude floods. The
floodwaters of April 2005 brought river taxa into the
wetland as evidence by highest Bray-Curtis similarity between the wetland and river in April 2005
(63%). At this date, the wetland assemblage became
more similar to the river assemblage than wetland
assemblages at other sampling dates. Changes in the
diatom assemblage of floodplain aquatic habitats
brought on by flooding has varied among studies.
Experimental increases in water depth and a
reduction of light reaching the sediments shifted
algal assemblages from periphytic to planktonic
species in a prairie lakeshore wetland (Robinson et
al. 1997a) and planktonic assemblages were observed in the sediments of several floodplain aquatic
habitats following flooding (Osborne et al. 1993,
Gell et al. 2002). Other studies reported decreases in
phytoplankton biomass during flooding due to
flushing of the system (Hamilton and Lewis 1987,
Agostinho et al. 2000). Engle and Melack (1993)
hypothesized that epiphytic algae may dominate in
wetlands after flooding because they may remain
attached to substrates as taxa in the water column
get flushed out due to water movement. We suspect
that the short/high magnitude flood of April 2005
increased water levels, favoring a planktonic community, but was of short enough duration (,
1 week) to only partially flush this diatom community from the wetland.
Observed shifts in the wetland diatom assemblage
during the connected phase may also have resulted
from shifts in water chemistry. Both floods of high
magnitude flushed nutrients and ions from the
wetland. Gell et al. (2002) found differences in the
types of periphytic diatoms in floodplain wetlands
before and after flooding in response to changes in
salinity and nutrient levels in a subtropic Australia
river, and Garcia de Emiliani (1993) observed a
successional reset of South American floodplain lake
phytoplankton following flooding. Engle and Melack (1993) found that the type of assemblage that
developed in an Amazon floodplain wetland following flooding depended on the nutrient and sediment
characteristics of floodwaters. In their study, planktonic diatom assemblages developed after floods
that lowered nutrient levels, similar to the changes
we observed in April 2005. The diatom assemblages
in the wetland returned rapidly to pre-flood
conditions after the April 2005 flood. Both total
nitrogen and total phosphorus levels returned
rapidly to pre-flood levels in the month following
flooding, possibly caused by the mixing processes of
floodwaters mobilizing nutrients sequestered in the
sediments (Kadlec 1986) and allowing taxa with high
nutrient preferences to again flourish after floodwaters receded.
In contrast to the high magnitude floods, the
short/low magnitude flood of February 2005 did not
appear to leave a discernable signal on either
wetland water chemistry or the algal assemblage.
Water chemistry remained distinct between the river
and wetland and algal biomass was not reduced by
the influence of floodwaters. In addition, no changes
were detected in the wetland diatom assemblage,
with assemblages prior, during, and after the flood
exhibiting high similarity (Figure 7). We believe that
this flood, which only brought enough water to
overtop the wetland for a few days, barely raising
water levels, did not cause significant habitat
alterations to the wetland or bring in enough water
to outweigh local wetland forces.
The diatom assemblage captured in the sediment
traps reflected the water column assemblage in the
wetland. While Bray-Curtis similarity was low
overall between the water column and settling
assemblages (23%–78%), similar taxa were found
in both assemblages. The settling assemblage may
integrate species growing in different wetland
habitats, potentially leading to a discrepancy between this and the wetland water column assem-
Weilhoefer et al., RIVER-WETLAND INTERACTIONS
blage. Medioli and Brooks (2003) found low
correspondence between water column and sediment
diatom assemblages in Canadian lakes, with Fragilaria capucina making up 20%–70% of the lake
water column assemblage and less than 5% of the
sediment assemblage. However, in an English lake,
the relative abundance of common species in the top
1 cm of a sediment core reflected the relative
abundance in weekly plankton samples (Haworth
1980). The shallow, well-mixed nature of the
wetland may allow diatoms from all habitats to
reach the sediment traps prior to dissolution.
Floodwaters from the river produced a discernable signal in the wetland settling diatom assemblage, with taxa composition shifting during the
connected phase (Figure 7). Similar taxa were found
in the wetland settling, river, and wetland assemblages during this time, although their relative
abundances varied. Approximately 93% of the taxa
found in the wetland settling assemblage during the
months of flooding were also observed in the river
during flooding. The wetland settling assemblage
was dominated by A. minutissimum, a taxon washed
in from the river and Bray-Curtis similarity between
the river and the wetland sediment assemblage were
elevated during the connected phase. Based on our
results, floodplain wetland habitats do contain a
record of river flooding.
While we detected changes in wetland algal
assemblages coinciding with flooding, we are unable
to fully attribute our findings to the impact of
floodwaters as seasonal shifts may be occurring
concurrently. Seasonal shifts in algal assemblages
have been observed in many aquatic habitats.
However, since high magnitude floods produced
dramatic changes in wetland physical habitat and
water chemistry, we feel that there is strong evidence
that observed shifts in algal assemblages were due, at
least in part, to the influence of river floodwaters.
In conclusion, the signal of the river floodpulse on
floodplain wetlands appears to depend on the
magnitude and duration of flooding. Floods of high
magnitude appeared to mix water chemistry between
the wetland and river and flush algae from the
wetland, while the short/low magnitude flood had
little effect on the wetland. While the short floodpulse of high magnitude in April 2005 deposited
river taxa within the wetland, longer duration and
higher magnitude flooding in 2004 produced shifts
in species but did not deposit distinct river taxa.
Caution should be used when extending diatombased bioassessment to frequently flooded wetlands
as the diatom assemblage within the wetland may
reflect both wetland condition and the influence of
river floodwaters.
485
ACKNOWLEDGMENTS
We thank Jackson Bottom Wetlands Preserve.
Particular gratitude goes to Frank Opila for
assistance with continuous water quality monitoring
data. We also thank Carlo Pearson, Leah Bower,
and John Van Voorhies for assistance with sample
collection and water chemistry analysis. The manuscript was improved by the comments of Joseph
Maser and two anonymous reviewers. This publication was developed under a STAR Research
Assistance Agreement (No. U915886) awarded by
the U.S. Environmental Protection Agency. It has
not been formally reviewed by the EPA. The views
expressed in this document are solely those of the
authors and the EPA does not endorse any products
or commercial services mentioned in this publication.
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Manuscript received 19 June 2007; accepted 18 February 2008.