AQUATIC CONSERVATION: MARINE AND FRESHWATER ECOSYSTEMS
Aquatic Conserv: Mar. Freshw. Ecosyst. 22: 588–597 (2012)
Published online 6 June 2012 in Wiley Online Library
(wileyonlinelibrary.com). DOI: 10.1002/aqc.2251
Harvesting an invasive bivalve in a large natural lake: species
recovery and impacts on native benthic macroinvertebrate
community structure in Lake Tahoe, USA
MARION E. WITTMANNa,*, SUDEEP CHANDRAb, JOHN E. REUTERc, ANDREA CAIRESb,
S. GEOFFREY SCHLADOWa and MARIANNE DENTONb
a
Tahoe Environmental Research Center, University of California Davis, Incline Village, NV 89451, USA
b
Department of Natural Resources and Environmental Science, University of Nevada, Reno, NV 89512, USA
c
Department of Environmental Science and Policy and Tahoe Environmental Research Center, University of
California Davis, CA 95616, USA
ABSTRACT
1. The increasing dispersal and establishment of aquatic invasive species in natural freshwater ecosystems has led
to efforts to remove non-native taxa and/or restore native species. An invasive bivalve, Asian clam (Corbicula
fluminea), recently (2002) became established in a large, natural subalpine lake (Lake Tahoe, USA). In 2009,
experimental efforts were undertaken to harvest C. fluminea from Lake Tahoe sediments using a manually
operated suction dredge apparatus.
2. Treatment and control plots were monitored for a 450 day period after dredging to observe target species
(C. fluminea) and non-target macroinvertebrate recovery rates. A paired Before-After-Control Impact analysis
was used to assess the short- and long-term impacts of suction dredging.
3. Physical harvest resulted in short-term reductions of C. fluminea (1500 individuals m-2 before treatment to 60
individuals m-2 14 days after treatment) with significant disruption to benthic macroinvertebrate community
structure. The impact to the target invasive species (C. fluminea) was present 450 days after treatment and
community diversity (as represented by Simpson diversity index) did not recover after 1 year (365 days) in
dredged sites. Certain non-target macroinvertebrate taxa (Chironomidae and native clam (Pisidium spp.))
increased in suction dredge plots to levels greater than before treatment or in control plot conditions at the end
of the study period.
4. Harvesting C. fluminea significantly reduced population densities for a period of 450 days after the removal.
Recolonization rates of C. fluminea and non-target species over multiple reproductive seasons will determine the
feasibility for this method as a long-term control strategy.
Copyright # 2012 John Wiley & Sons, Ltd.
Received 04 December 2011; Revised 20 February 2012; Accepted 01 April 2012
KEY WORDS:
lake; littoral; biological control; recolonization; monitoring; benthos; invertebrates; alien species; dredging
INTRODUCTION
The establishment of aquatic invasive species
continues to affect freshwater ecosystems and
present costly challenges to natural resource
managers. While prevention of invasive species
introductions is considered to be the most effective
means to reduce invasive species impacts (Leung
*Correspondence to: Marion E. Wittmann, Department of Biological Sciences, University of Notre Dame, Notre Dame, IN 46556.
E-mail: Marion.E.Wittmann.3@nd.edu
Copyright # 2012 John Wiley & Sons, Ltd.
HARVESTING INVASIVE CLAMS: IMPACT TO THE MACROINVERTEBRATE COMMUNITY
et al., 2002; Finnoff et al., 2007; Keller et al., 2008), it
is a complex and resource-intensive endeavour that is
often complicated by undetected propagules,
illegal releases, or accidental introductions that can
confound prevention goals. As a result, natural
resource managers are often tasked with controlling
or removing an introduced species after it has
become established.
Harvesting invasive species has promise as a
non-chemical means to reduce negative economic
and ecological impacts and increase water quality
and commercial and recreational use of ecosystems
(Simberloff, 1999; Mack et al., 2000). Efforts to
physically remove invasive species have been
attempted for a number of taxa including rusty
crayfish (Hein et al., 2007), dreissenid mussels
(Wimbush et al., 2009), aquatic macrophytes
(Tobiessen et al., 1992; Eichler et al., 1993), and
smallmouth bass (Weidel et al., 2007) with varied
levels of success. Unintended effects of invasive
species removal include shifts to native community
structure (Rinella et al., 2009) or increases in
population growth rates of the management
target (Zipkin et al., 2009). Although rare,
invasive species management goals have been
accomplished through long-term programmes of
physical removal (Wimbush et al., 2009) or
combining physical removal with other treatment
methods (Madsen, 1997).
Many methods exist to remove species from
aquatic systems such as hydraulic dredging, hand
removal, trapping, or electroshocking and each
has impacts on the surrounding environments
and biological communities. Dredging or suction
removal is a widely used method of species
extraction from sediments, and has been used for
the removal both of desirable (i.e. commercial)
species such as the razor clam (Ensis spp.) and
nuisance benthic macrophytes (Nichols and
Cottam, 1972; Tobiessen et al., 1992; Eichler
et al., 1993; Hauton et al., 2007). In general,
dredging in aquatic environments is considered a
major disturbance to benthic systems as it
reduces benthic macroinvertebrate populations
(Kenny and Rees 1996; Pranovi et al., 1998;
Lewis et al., 2001) and disrupts the population
structure of native communities (Grassle and
Sanders 1973; McCall, 1977; Dernie et al., 2003).
These types of disturbances have also been
observed with dredging activities specifically
targeted towards biotic removal (Tuck et al.,
2000; Morello et al., 2005).
Copyright # 2012 John Wiley & Sons, Ltd.
589
Corbicula fluminea is a sediment-dwelling bivalve,
introduced and invasive to North America and
native to temperate and tropical regions of Asia,
Africa, and Australia (Counts, 1986). The impacts
of C. fluminea both on natural and on man-made
systems are known (Isom, 1986) and have been
observed to affect native invertebrate communities
(Karatayev et al., 2003), phytoplankton assemblages
(Lopez et al., 2006), benthic habitats (Hakenkamp
and Palmer, 1999) and nutrient cycling (Lauritsen
and Mozley, 1989). This species is considered an
economic nuisance because of its ability to biofoul
water intakes (Eng, 1979), particularly in nuclear
and hydropower production (Isom et al., 1986;
Williams and McMahon, 1986) where damage
caused by C. fluminea has been estimated at $1
billion annually (Pimentel, 2005).
Established populations of C. fluminea were first
observed in Lake Tahoe, CA-NV in 2002 and in
2008 high density populations (up to 6000 clams m-2)
were observed in nearshore habitats by scientists,
natural resource managers, and community
stakeholders who responded by creating a
science-based
rapid
response
management
programme. To explore the feasibility of reducing
C. fluminea abundance with physical harvesting,
a dredging experiment was carried out on
sub-populations of C. fluminea. Benthic substrate
removal through diver assisted suction dredging
was applied to treatment plots with established
populations of C. fluminea and monitored for
450 days. The objectives of this study were to
assess the recolonization rates of C. fluminea and
co-occurring benthic macroinvertebrate taxa after
suction removal to understand community recovery.
METHODS
Study site
Lake Tahoe (Figure 1) is a large, deep (surface area:
495 km2, maximum depth: 501 m) oligotrophic lake
located at a subalpine elevation of 1898 m in the
Sierra Nevada Mountain Range. The Tahoe
Basin’s largely granitic geology, the lake’s large
volume (150 km3) and relatively small drainage
(800 km2) basin explain its low nutrient
concentrations and primary productivity rates
(Goldman, 1988). Annual water temperature
ranges from 5 to 28 C in the littoral zone, with
upper and lower temperature extremes occurring
in marina locations. The lake is oligomictic,
Aquatic Conserv: Mar. Freshw. Ecosyst. 22: 588–597 (2012)
590
M. E. WITTMANN ET AL.
Figure 1. Location of C. fluminea treatment plots in Lake Tahoe, CA-NV.
Two treatment plots (surface area: 36 m2) at Marla Bay, NV (A) and three
at Lakeside, CA (B) were dredged to remove C. fluminea containing
sediments in April 2009. Both plot locations were at 5 m water depth
in the south-eastern portion of Lake Tahoe where C. fluminea populations
have been established since 2002.
mixing completely only in years of intense spring
storms (Goldman et al., 1989; Wetzel, 2001). The
photic zone extends to an approximate depth of
100 m, and the entire water column is oxygenated
throughout the year (Coats et al., 2006). Lake Tahoe
supports an assemblage of benthic invertebrates
dominated by oligochaetes, amphipods, ostracods,
and dipteran larvae (Frantz and Cordone, 1996).
Methodology
Diver-assisted suction dredging was applied in 5 m
water depth at two sites, Marla Bay and Lakeside,
which both have established C. fluminea
populations – with Marla Bay representing a high
density C. fluminea population (average abundance
~2000 m-2) and Lakeside representing a region with
lower densities (average abundance ~500 m-2). At
each site, three suction removal plots and one
control plot (surface area: 36 m2) were delineated
on the lake bottom to guide the dredging diver and
to demarcate sampling areas for monitoring
purposes (Figure 1). Owing to field resource
limitations, only two plots were suctioned at Marla
Bay and three at Lakeside. In March 2009 a
suction dredge apparatus (4 cm diameter hose,
engine specifications: 5.5 HP (4.1 kW) @3600 rpm
Copyright # 2012 John Wiley & Sons, Ltd.
net power output, 196 cm3 displacement, 12.4 Nm
@2500 rpm net torque) was used to remove
treatment plot sediments to a depth of 13 cm at
Marla Bay (9.5 m3 removed) and a depth of 8 cm
(depth of sediments above a clay hard pan
boundary) at Lakeside (8.5 m3 removed).
To collect benthic macroinvertebrates, sediment
grab samples (N = 3) were collected in each of the
treatment and control plots using a petite Ponar
grab sampler (2.4 L volume, 231 cm2 sample area,
Wildlife Supply Company, Yulee, FL, USA) at
7 days before and 14, 90, 240, 365, and 450 days
after dredging. Upon collection all samples were
screened (500 mm mesh) and the retained sediment
was then placed in a super-saturated sugar solution
to float invertebrates (Anderson, 1959). Samples
were then picked manually to remove all
macroinvertebrates. All organisms were preserved
in 70% ethanol until identification (Merritt and
Cummins, 1996; Thorp and Covich, 2001) Average
particle size distribution of Marla Bay and
Lakeside sediment types was determined using a
wet sieve method (Gordon et al., 1992) and
described using a Wentworth scale. Temperature
samples were collected using in situ loggers (4 h
sampling interval) (iButtonW #DS1922L, 0.5 C
accuracy, Embedded Data Systems, Lawrenceburg,
KY, USA).
Total macroinvertebrate abundance, abundance
by taxonomic grouping and Simpson Diversity
Index (Simpson, 1949) was calculated for each
sample. Abundance was calculated by dividing the
number of species individuals per sample by the
volume of the petite ponar grab sampler. Impact
of sediment removal on the invertebrate
community was assessed using Paired Before-After
Control-Impact (BACIP) analysis (Stewart-Oaten
et al., 1986; Underwood, 1991, 1994; GuerraGarcía et al., 2003). Effect size was calculated by
forming differences between site-specific pairs:
Dik ¼ XiCj XiIk ¼ m þ ’i þ eik
(1)
where X represents taxonomic abundance or
diversity index values, m is the mean difference
between control and impact, ’i the change in
difference from before to after, and eik the error
associated with the differences (Stewart-Oaten
et al., 1986). Differences (Dik), were then compared
for before and after periods using a two-sample
t-test. Both sites (Marla Bay and Lakeside) were
combined to calculate the overall mean difference
in effect size while maintaining site-specific pairing
Aquatic Conserv: Mar. Freshw. Ecosyst. 22: 588–597 (2012)
HARVESTING INVASIVE CLAMS: IMPACT TO THE MACROINVERTEBRATE COMMUNITY
between control and treatment plots. A significant
change in the mean difference (m) after the onset of
the perturbation was considered strong evidence of
an environmental impact. All statistical analysis was
carried out using program R 2.12.1.
RESULTS
Before treatment, average total macroinvertebrate
abundance (P < 0.001, t-test, n = 12, df = 22) was
lower at Lakeside (mean abundance standard
error: 4014 280 m-2) compared with Marla Bay
(6995 684 m-2). Common taxonomic groupings
were observed at both sites: Amphipoda (Hyalella
sp.), Chironomidae, Oligochaeta, Gastropoda
(Planorbidae and Physidae), and non-native and
native bivalves, Corbiculidae (C. fluminea) and
Sphaeriidae (Pisidium casertanum and compressum;
hereafter referred to as Pisidium spp.), respectively.
Other less common taxonomic groups observed at
both sites were Trichoptera (Leptoceridae,
Lepidostomatidae), Ceratopogonidae (Palpomyia
sp.), Ostracoda, Copepoda, Hydracarinidae,
Cladocera, Hirudinea and Nematoda, and were
not included in the analysis because of rare
occurrence. C. fluminea was dominant in Marla
Bay, with average relative abundance of 35%
(2450 343 m-2), compared with Lakeside where
C.
fluminea
abundances
represented
8%
(485 95 m-2) of the community (Figure 2).
Chironomids were the other dominant taxonomic
group at Marla Bay (2527 145 m-2, 31%), and
Amphipoda (669 76 m-2; 10%), Oligochaeta
(676 126 m-2; 10%), and Gastropoda (779 214 m-2;
11%) represented secondary groups. Pisidium spp.
Figure 2. Proportion of macroinvertebrate taxa at site A (Marla Bay)
and site B (Lakeside) in Lake Tahoe before suction removal treatment.
Proportions are based on average abundance of taxonomic grouping
based on benthic sampling carried out in March 2009.
Copyright # 2012 John Wiley & Sons, Ltd.
591
(78 19 m-2) comprised 1% of the population at
Marla Bay compared with more than 13%
(536 58 m-2) of the population at Lakeside. The
dominant groups at Lakeside were Amphipoda
(1354 188 m-2; 34%) and Chironomidae (559 74 m-2;
14%) (Figure 2).
Sediment substrate at the Lakeside site was
characterized as coarse to medium sand with a
median sediment particle size, Me = 0.375 mm and
very coarse sand at Marla Bay, Me = 1.180 mm.
Water temperature at the sediment–water interface
ranged from 5.6 to 17.9 C during the course of
the 450 day monitoring period.
Immediate (14 days after treatment) impacts of
suction removal were similar among most
taxonomic groupings (Table 1, Figure 3). There
was no significant difference in taxonomic
abundances and diversity indices between control
and treatment plots before suction removal
(Table 1). At 14 days (April 09) after suction
removal treatment, average total invertebrate
abundance (combined between two sites) was 410
(140 S.E.) individuals m-2; a 93% reduction
compared with pre-treatment condition (5892 594).
The initial impact on C. fluminea was significant
with a 96% reduction from 1484 (472) to 59 (34)
individuals m-2 in the treatment site. Amphipoda
also showed a dramatic reduction in abundance,
decreasing 96% from 1022 (143) in the
pretreatment condition to 44 (23) individuals m-2
at 14 days. While average abundances of both
chironomids and gastropods on day 14 in treatment
plots did significantly decline after dredging
(Figure 3), BACIP results showed that there was no
significant impact on mean effect size (Table 1).
Simpson diversity index values were significantly
affected immediately after treatment, with mean
effect size increasing more than two orders of
magnitude compared with pre-treatment and control
conditions (Table 1).
Mean effect size varied during the monitoring
period with differences observed among the
taxonomic groups. Abundance of the management
target C. fluminea in dredged plots was
significantly less than in control plots at 450 days
(July 2010) after treatment (Figure 3). However, at
365 days (April 2010) after treatment BACIP
results showed that effect size in dredged plots was
not significantly different from control or before
conditions. Total invertebrate abundance effect
size was not significant at 240 days (December
2010) after treatment. However, Simpson diversity
Aquatic Conserv: Mar. Freshw. Ecosyst. 22: 588–597 (2012)
592
M. E. WITTMANN ET AL.
Table 1. CI and BACI analysis results for taxonomic groups collected before suction dredge treatment (Before) and after treatment (days 14–450). The
first value for each taxon indicates the mean effect size: between control and impact sites (CI) in the Before column, and in subsequent columns (days 14–450)
before/after and control/impact (BACI). The second entry shows the t-statistic (H0: EffectSizeBefore = EffectSizeAfter) with significance level indicated:
*** = P < 0.001, ** = P < 0.01, * = P < 0.05
Before (Mar 09)
Total invertebrates
Amphipoda
Chironomidae
Oligochaeta
Gastropoda
Corbicula fluminea
Pisidium spp.
Simpson Index (diversity)
1230
0.06
34
0.33
651
0.13
69
0.99
138
0.87
453
0.22
31
1.26
0.007
0.02
14d (April 09)
4372
5.30***
551
2.11**
742
0.31
808
4.69***
310
1.11
1471
1.82*
369
3.22**
0.405
3.77***
90d (Jul 09)
3255
4.11*
322
1.37
332
1.06
754
4.75***
147
0.09
1343
2.01*
286
3.07**
0.355
5.16***
150d (Sep 09)
240d (Dec 09)
365d (Apr 10)
2674
3.29*
180
0.83
63
1.97*
682
4.42***
98
0.45
1285
2.00*
316
3.64***
0.336
6.00***
2248
2.54
218
0.98
134
2.43*
652
4.15***
63
0.89
1206
1.88*
236
2.94**
0.310
6.69***
1850
1.94
185
0.86
268
2.95**
706
4.38***
34
1.28
1033
1.48
147
1.67
0.256
6.26***
450d (Jul 10)
1565
1.54
177
0.84
442
3.59***
677
4.33***
29
1.39
1013
1.45
111
1.24
0.258
7.13***
Figure 3. Temporal changes of the abundance (number of individuals m-2) of each taxonomic group in suction dredge treatment (dashed line) and
control (solid line) plots from March 2009 (before treatment) to July 2010 (after treatment). Error bars represent one standard error. Time of sampling
is indicated by days after treatment and month/year in which the sampling event occurred.
index remained significantly different from the
pre-impact condition throughout the entire
sampling period (450 days; July 2010) suggesting
that community dynamics were altered by
harvesting. Chironomids had a significant positive
effect size at day 240 (December 2009), indicating
that abundances in treatment plots at this time
surpassed abundances in control and impact plots
before treatment. All other taxonomic groups had
negative, but reducing, effect sizes at the end of
the sampling period – indicating a trend towards
a return to background and pre-treatment
abundances. Figure 4 shows the proportional
Copyright # 2012 John Wiley & Sons, Ltd.
change in effect size for total invertebrate
abundance, C. fluminea abundance, native clam
(Pisidium spp.) abundance, and Simpson diversity
index. By the end of the sampling period, both
Pisidium spp. and total invertebrate abundance
show the greatest reductions in effect size,
declining from 0.92 at 14 days to 0.28 at 450 days
and 0.78 at 14 days to 0.28 at 450 days,
respectively. C. fluminea effect size was no longer
significant at the end of the monitoring period,
with a decrease from 0.76 to 0.53, while species
diversity remained affected, with a proportional
reduction to 0.63 at 450 days.
Aquatic Conserv: Mar. Freshw. Ecosyst. 22: 588–597 (2012)
HARVESTING INVASIVE CLAMS: IMPACT TO THE MACROINVERTEBRATE COMMUNITY
Figure 4. Temporal changes (days after treatment, calendar date) in
proportional effect size for three taxonomic groups and one diversity
index: total invertebrate abundance, native bivalve species (Pisidium
spp.) abundance, invasive clam species (Corbicula fluminea, i.e. treatment
target) abundance, and the Simpson Diversity Index. Changes in effect
size are standardized by differences in effect size before treatment.
DISCUSSION
While physical harvesting can decrease the
abundance of a target species, it can also disrupt
soft-sediment benthic communities by reducing
diversity and abundance of some taxa. Dredging to
remove C. fluminea populations in Lake Tahoe
reduced benthic macroinvertebrate abundances and
produced variable recolonization patterns that were
dependent on environmental parameters and
individual taxon responses. At the end of this
study’s monitoring period, total invertebrate
abundances in treatment plots were not significantly
different from control or pre-treatment plot
conditions, whereas diversity indices remained
significantly decreased, suggesting that community
dynamics were still altered 450 days after treatment.
In general, dominance of a few groups is a
common feature of macroinvertebrate communities
in the early stages of the recolonization process
(Ladle et al., 1980; Otermin et al., 2002), in part
because they are either able to persist in sediments
or can migrate into disturbed areas (Yount and
Niemi, 1990). In this study, chironomids were
among the first recolonizers in treatment plots, with
a 50% recovery rate in abundances at 90 days (July
2009). In contrast, the abundances of some taxa
such as amphipods and oligochaetes did not begin
to recover until 150 days (September 2009) after
treatment. At 150 days after treatment amphipod
abundances increased more rapidly than those of
oligochaetes. Chironomids are ubiquitous in Lake
Tahoe, and have been similarly observed as early
colonizers in other systems because of their
r-selected traits (Gray, 1981; Malmqvist et al., 1991;
Copyright # 2012 John Wiley & Sons, Ltd.
593
Otermin et al., 2002). In addition, chironomids were
not completely removed after suction treatment,
leaving behind 10% of the initial population
abundance as primary recolonizers. In comparison,
other taxa, such as amphipods and gastropods, were
more completely removed, with 4% of their original
population abundance remaining in treatment plots.
The rapid chironomid recolonization could also be
attributed to potential migration from adjacent
untreated plots given its high abundance compared
with other taxa observed in Lake Tahoe. Amphipods
had rapid recovery relative to control plot
conditions, which was probably attributable to their
mobility, and similar to chironomids, a high
potential for migration from adjacent, untreated
areas. Temporal patterns of amphipod abundance
mimicked control plot conditions, and by 150 days
(September 2009) effect sizes were insignificantly
different from controls. Similar to amphipods,
oligochaete abundances in treatment plots showed
increases correlated with control plots, but with
many fewer individuals throughout the entire
450 day monitoring period, suggesting that migration
of new individuals from adjacent areas was not as
frequent for oligochaetes as it was for other more
mobile taxa. Timing of reproductive cycles for these
taxa is unknown, but it was also possible that early
recolonization rates during the spring and early
summer periods for chironomids and amphipoda
were optimal in comparison with oligochaetes,
Pisidium spp., or C. fluminea, all of which did not
show significant increases in abundance until at
least 240 days (December 2009) after treatment.
Slight increases in the abundance of the target
invasive species C. fluminea in treatment plots did
not begin until 240 days (December 2009) after
treatment, and treatment plot densities were
statistically different from control plots at 450 days
(July 2010). However, during the April 2009
sampling (365 days after treatment), C. fluminea
treatment and control plots did not significantly
differ. This reduction of control plot abundances
could have been attributed to sampling effects
given the naturally heterogeneous distribution of
C. fluminea in space, or potentially from a
winter-time decrease in the abundances as a
result of a low temperature mortality event (Werner
and Rothhaupt, 2008). Subsequent sampling on
450 days after treatment (July 2010) showed
continued increases in both the control and
treatment plot abundances of C. fluminea, once
again returned to a significant difference between
Aquatic Conserv: Mar. Freshw. Ecosyst. 22: 588–597 (2012)
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M. E. WITTMANN ET AL.
control and treatment plots densities, but an
insignificant effect size as represented in the BACIP
results. While BACIP analyses have frequently been
used to assess the impact of dredging on
macrobenthic communities (Morello et al., 2005),
natural changes in control populations and timing
of monitoring events can affect results and
interpretations of recovery. This is particularly
important in relation to management scenarios,
where the determination of invasive species control
programmes are often based on monitoring results
that are limited in extent.
It is possible that recolonization by C. fluminea
was influenced by the following: (1) minimum
thermal thresholds for reproduction/juvenile
release, and (2) the coarse sediment type at Marla
Bay, compared with the clay substrate at Lakeside
remaining after removal. C. fluminea population
densities remained low at both sites until the
minimum temperatures required for reproduction (i.
e. gametogenesis, juvenile release) were approached
during the spring and summer periods. Through
either juvenile release, or movement of adults from
adjacent plots, recolonization of C. fluminea
occurred in the summer months, but with higher
abundances observed in Marla Bay as a result of
sediment type (coarse sediments in Marla, and clay
pan remaining after suction removal at Lakeside)
that provided more favourable habitat in which
to recolonize. In contrast, through site surveys
using SCUBA, Pisidium spp. were observed
inhabiting the upper clay and re-sedimented
layers at Lakeside after dredging.
Unlike the invasive target species, C. fluminea,
abundances of native bivalves Pisidium spp.
increased to levels greater than or equal to those
observed in the treatment plots by day 365 (April
2010) compared with control plot conditions. C.
fluminea is well known for its temperature-dependent
reproductive rates (Williams and McMahon, 1989),
generalist feeding preference (Way et al., 1990;
Hakenkamp and Palmer, 1999), and tolerance to a
wide variety of environmental conditions (Williams
and McMahon, 1989). The relative recolonization
success of Pisidium spp. compared with the slower
recolonization rates of opportunist C. fluminea,
suggests that C. fluminea is perhaps not as well
adapted to low water temperatures, limited food
availability, and other environmental variables not
measured here. Further study of inter-specific
competition in Lake Tahoe is needed to understand
this relationship.
Copyright # 2012 John Wiley & Sons, Ltd.
The 450 day monitoring period of the treatment
plots revealed several important factors of
recolonization dynamics in this system. First,
while total invertebrate abundance recovered to
pre-disturbance levels, species diversity (as
represented by the Simpson index) was significantly
different from pre-treatment and control plot value
throughout the entire post-dredging monitoring
period (14–450 days). This suggests a modified
community structure, which was most strikingly
highlighted through shifts in chironomid
abundances in treatment plots. Second, multiple
linear regression results suggest that, similar to
other pulse-disturbed freshwater soft-sediment
benthic communities (Lake, 2000), time since
treatment was a consistent predictor of C. fluminea
and total invertebrate abundances, as well as
Simpson diversity index effect sizes. Third, at
the end of this study period (450 days; July 2010),
the Simpson diversity index in the treatment
plots remained significantly different from both
pre-treatment and control conditions. This has
several implications: (i) continued monitoring is
required to understand whether the macrobenthic
invertebrate community can fully recover to
pre-treatment or control conditions, and (ii) further
studies relating to inter- and intra-specific dynamics
in treatment plots that can have longer-term
impacts on community recovery are warranted.
Recovery times of macroinvertebrate taxa after
disturbance in lentic habitats have been observed to
range from days to years and are dependent on
environmental condition, community composition,
and magnitude of the perturbation (Cowell, 1984;
Van de Meutter et al., 2006; Sychra and Adamek,
2011). Understanding benthic community dynamics
under a pulse disturbance scenario can further
indicate whether suction dredging is an effective
method for invasive species control of molluscs
in lentic habitats.
Several control strategies for invasive molluscs,
namely dreissenid mussels, have been developed
and in some cases, have been successful (Nalepa
and Schloesser, 1993; Claudi and Mackie, 1994;
D’Itri, 1997). Removal of adult C. fluminea and
other invasive molluscs has been used primarily in
power generation settings where biofouling of
important structures incurs significant costs and
removal was necessary for proper plant
functioning (Connelly et al., 2007). However, for
open water bodies these approaches are generally
not suitable as they can be ecologically damaging,
Aquatic Conserv: Mar. Freshw. Ecosyst. 22: 588–597 (2012)
HARVESTING INVASIVE CLAMS: IMPACT TO THE MACROINVERTEBRATE COMMUNITY
expensive, and ineffective (Wimbush et al., 2009).
There have only been two reports of bivalve
eradication (using chemical treatment) in open
water bodies (Bax, 1999; Virginia Department
of Game and Inland Fisheries, 2009) and one
non-chemical treatment that reduced but did
not eradicate a dreissenid mussel population
(Wimbush et al., 2009). In general, the only accepted
practical approach for managing molluscan
invasions in open waters is thought to be by
prevention-oriented management (Frischer et al.,
2005) with potential for control during early
invasion stages when population densities are
typically low (Hobbs and Humphries, 1995). While
C. fluminea abundances remained significantly low
at the end of the monitoring period (450 days after
treatment, and one full reproductive period),
continued monitoring of these populations is needed
to determine the effectiveness of the sediment
suction dredging method as a viable option for
population control. The cost of labour and materials
associated with the implementation of diver-assisted
suction removal were $265 m-2 and included dredge
apparatus equipment purchase, high altitude
commercial diver labour, sediment disposal, and
permitting fees.
Control of a biological invasion is most effective
when it uses a long-term, ecosystem-wide strategy
rather than a directed approach focusing on
removing individual invaders, in part because of
its impact on associated native taxa (Mack et al.,
2000). High rates of fecundity, competitive
resource utilization and juvenile dispersal allow
invasive species such as C. fluminea to invade
areas rapidly, which can limit management
options and reduce restoration efforts (Sakai et al.,
2001). While physical removal was economically
costly, it effectively reduced the target invasive
species abundance over a 450 day period. At the
same time, it also reduced associated invertebrate
abundances and caused disruptions to the
community structure of the macrobenthos.
ACKNOWLEDGEMENTS
The authors would like to thank Brant Allen, Steve
Sesma, Joe Sullivan, Katie Webb and Raph
Townsend for their dedicated efforts in the field
and laboratory. We thank the Tahoe Regional
Planning Agency, Tahoe Resource Conservation
District and the US Fish and Wildlife Service for
their support in this programme. Funding for this
Copyright # 2012 John Wiley & Sons, Ltd.
595
work was provided by the Southern Nevada
Public Lands Management Act, the Lahontan
Regional Water Quality Control Board and the
Nevada Division of State Lands.
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