Vol. 4
No 2, 103-117
2004
Assessment of anthropogenic influences
on littoral-zone aquatic communities
of Lake Texoma, Oklahoma-Texas, USA
Michael A. Eggleton1, Keith B. Gido2,
William J. Matthews3, Gary D. Schnell4
1Sam Noble Oklahoma Museum of Natural History, University of Oklahoma, 2401
Chautauqua, Norman 73072, Oklahoma USA. Current address: Aquaculture/Fisheries
Center, University of Arkansas-Pine Bluff, Box 4912, Pine Bluff 71601, Arkansas USA.
e-mail: meggleton@ uaex.edu.
2Division of Biology, Kansas State University, Ackert Hall, Manhattan, Kansas 66506, USA.
e-mail: kgido@ksu.edu.
3Sam Noble Oklahoma Museum of Natural History, University of Oklahoma, 2401
Chautauqua, Norman 73072, Oklahoma USA/University of Oklahoma Biological Station,
HC 71, Box 205, Kingston 73439, Oklahoma USA/Department of Zoology, University
of Oklahoma, Norman 73019, Oklahoma USA.
e-mail: wmatthews@ou.edu.
4Sam Noble Oklahoma Museum of Natural History, University of Oklahoma, 2401
Chautauqua, Norman 73072, Oklahoma USA/Department of Zoology, University of
Oklahoma, Norman 73019, Oklahoma USA.
e-mail: gschnell@ou.edu.
Abstract
From 1999-2001, we evaluated the effects of anthropogenic activities in and around Lake
Texoma, Oklahoma-Texas, USA, on the structure of littoral-zone fish and benthic invertebrate assemblages at 20 potentially impacted sites relative to paired reference sites. Spatial
structuring of both assemblages was strongly related to variables associated with water clarity, water-column chlorophyll-a levels, and degree of site exposure to wind and waves. Fish
assemblages at reference sites and impact sites exhibited minor differences, but none that
were considered indicative of severe anthropogenic stress. Conversely, benthic invertebrates exhibited greater differences between reference and impact sites. Procrustean analyses and Mantel tests indicated little concordance between reference site and impact site
benthic invertebrate assemblages. Greater abundances of oligochaetes at impact sites and
greater abundances of chironomids at reference sites contributed most to these differences,
with the largest assemblage differences found at sites influenced by agriculture and sanitary
dumps. Despite the fact that fishes and benthic invertebrates were structured along similar
environmental gradients, little concordance was observed between assemblages. Wide
annual fluctuations in the dominant taxa of each assemblage contributed most to the general discordance. Furthermore, discordant fish and invertebrate assemblages likely resulted
because responses of each assemblage to anthropogenic impacts occurred at different scales
of space and time.
Key words: reservoirs, fishes, benthic macroinvertebrates, multitaxon assessment,
assemblage structure, environmental impact, Procrustean analysis, Mantel test
104
M. A. Eggleton et al.
1. Introduction
Concern over the condition or "ecological
health" of reservoirs has increased over the past
quarter century with a recognition that chemical
standards have not protected aquatic resources
(Adler et al. 1993). However, when studying environmental impact or biotic integrity (sensu Angermeier, Karr 1984), reservoirs provide a more
difficult challenge than streams, rivers, or lakes
because they are artificial systems that lack natural reference sites (Hickman, McDonough 1996).
Furthermore, the temporary nature of reservoirs
on an evolutionary time scale precludes expectations of natural communities (Jennings et al.
1995). These attributes have led some researchers
to suggest the term "biotic integrity" may be inappropriate for reservoirs, recommending instead
that effort be devoted to developing multimetric
indices of reservoir health (e.g., "reservoir fish
assemblage index"; Hickman, McDonough 1996).
Despite the challenges presented in assessing
the ecological health of reservoirs, they comprise
a large proportion of freshwater resources in the
United States (Miranda 1996). Although most
contemporary methods focus on community-level
assessments, several recent studies (e.g., Jackson,
Harvey 1993; Allen et al. 1999a,b; O'Connor et al.
2000) have shown advantages in conducting community-level assessments on multiple assemblage
types within the same environment. Measures of
concordance between assemblage types implicit
with this approach provide added insight into
community-level responses to anthropogenic
activities in aquatic systems (Jackson, Harvey
1993). Few studies have examined environmental
impact in this manner, and none have been conducted in reservoirs.
Our objective was to assess effects of anthropogenic activities on aquatic assemblages in the
littoral-zone of a large reservoir. We assessed
effects at the community level, considering only
shifts in conventional assemblage measures
(species richness, organism abundance, and
assemblage structure) in response to anthropogenic influences. Most studies in reservoirs
have focused on off-shore fish assemblages consisting of large sport fish that can move readily,
whereas we focused on littoral-zone fish assemblages because they are an important functional
component of aquatic systems (Northcote 1988).
Littoral fishes tend to be small-bodied and have
little capacity to move great distances and so were
regarded as more likely to reflect local environmental stresses. Assemblages examined were
those most likely to show demonstrable effects
resulting from anthropogenic stressors that
impinge on reservoir ecosystems:
(1) juvenile/adult fishes (hereafter referred to as
fishes), which include the littoral-zone "watercolumn" community (e.g., juvenile temperate
basses, atherinids, shads, and minnows) and the
"benthic nesting" community (mainly centrarchids); and
(2) benthic invertebrates, which include substratedwelling macroinvertebrates such as chironomids,
oligochaetes, burrowing mayflies, and molluscs.
If anthropogenic stresses were prevalent in
Lake Texoma, our general expectation was that
strong differences would exist with fish and benthic invertebrate assemblages between reference
and impact sites. Further, we expected that assemblages might respond to anthropogenic influences
differently, with invertebrates responding to more
local impacts (e.g., substrate differences associated with anthropogenic activities) and fishes
responding to broader-scale influences (e.g., turbidity differences related to land use practices). In
addition, because human activities alter multiple
facets of the environment (Karr 1991), these
assemblage types represent very different taxonomic and functional groups and were expected to
provide a more complete view of how anthropogenic activities impinge on the Lake Texoma
ecosystem.
2. Materials and methods
Study area
Lake Texoma, Oklahoma-Texas, USA is a
35 200 ha impoundment in the southern Great
Plains region of the United States, located at the
confluence of the Red and Washita rivers on the
Oklahoma-Texas boarder. The reservoir was constructed in 1944 by the U.S. Army Corps of Engineers (USACE) for flood control and hydroelectric
power production. The watershed of the reservoir
encompasses about 103 000 km2; land use is predominantly agriculture, ranching, and forest with
relatively low human population densities. Highly
saline inflows from the Red River occur due to natural salt sources in the headwaters and tributaries.
As a result, conductance values in Lake Texoma are
high for freshwater (700 - 1200 µS cm-1; Gelwick,
Matthews 1990) and distinctly different between
river arms (Red River arm, usually >1000 µS cm-1;
Washita River arm, usually <1000 µS cm-1). Secchi
depths in the reservoir typically range from 50 to
125 cm (Matthews 1984) and are usually greater
downlake near the dam. The reservoir thermally
stratifies during the summer months and maximum
depth is around 25 m. Reservoir stages vary about
2 m per year on average, but fluctuations may
exceed 3 m in any given year (U.S. Army Corps of
Engineers, Tulsa District website). The ecological
health of Lake Texoma has significant repercussions on local economies because of its recreational
fisheries, which are valued at over 25 million dollars (U.S.) annually (Schorr et al. 1991).
Anthropogenic influences on aquatic communities
Experimental design
Statistically, the study was designed such that
potentially "impacted" sites in Lake Texoma were
matched or paired with physically similar "reference" sites for statistical validity. Twenty such
impacted sites were identified in Lake Texoma by
the U.S. Environmental Protection Agency
(USEPA) and used in this evaluation. Anthropogenic stressors of interest to USEPA included:
point-source chemical leachates (e.g., methyl tertbutyl ether [MTBE] and benzene-toluene-ethylbenzene-xylene [BTEX] compounds; An et al.
2002); nonpoint-source chemical leachates (e.g.,
herbicides and pesticides); nutrient enrichment
resulting from elevated runoff of inorganic agricultural fertilisers; sedimentation; and activities
associated with the alteration, destruction, or
removal of aquatic habitats. In Lake Texoma,
these major stressors were associated with one or
more of the following anthropogenic activities:
row-crop agriculture (mainly soy beans, corn, and
cotton); rural septic runoff; marinas and related
operations (e.g., boat repair/servicing and marine
gas facilities); sanitary dumping grounds; developed public beaches; and oil production facilities
(Schnell et al. 2002).
In classifying impacted sites to a particular
impact type, several factors were considered. For
marinas, sanitary dumps, public beaches, and oil
production facilities, the close proximity of the
activity to the site in question was the main characteristic considered. In most cases, the source
influence was located less than 200 m from the
site or was contained throughout the site. For rowcrop agriculture and rural septic runoff, the prevalence of the land use within reservoir sub-basins
and associated in situ characteristics were used to
classify sites. For instance, agricultural sites typically contained greater than 40% of their drainage
area in row crops and/or pasture lands, contained
elevated water-column chlorophyll-a levels (usually greater than 15 µg dm-1, presumably related to
nutrient enrichment), and had lower water transparency (presumably due to increased sedimentation). Sites exposed to rural septic runoff were
those with greater than 3% of their drainage area
constituting municipalities and/or moderate- to
high-density residential or commercial development. Septic-influenced sites frequently also had
elevated water-column and benthic chlorophyll-a
levels.
Principal components analysis (PCA) incorporating mainly site-specific physical variables
(e.g., slope, aspect, site exposure to wind and
waves) was used to pair each impact site with one
of 60 potential reference sites. Each impact site
was then paired with the reference site that was
closest in multivariate space. This approach kept
obvious physical differences within site pairs to a
105
minimum and yielded 20 site pairs that were used
for community-level assessments. Site pairs were
stratified within impact types as follows: agriculture (5 site pairs for evaluation), marinas (3 site
pairs), developed public beaches (2 site pairs),
sanitary dumping grounds (2 site pairs), rural
septic runoff (4 site pairs), and oil production
facilities (4 site pairs). The number of site pairs
evaluated for each impact type was proportional to
the prevalence of that influence in the reservoir. In
17 of 20 site pairings, impact sites were paired
with reference sites located less than 3 km away
(linear distance); between-site distances in the
remaining three pairs ranged 6.4-11.2 km (Schnell
et al. 2002).
Sample collections
Fish assemblages were assessed annually at
all 40 sites from 1999 through 2001. Assessments
were conducted during July of each year so that
community compositions would not be skewed
substantially by high numbers of young-of-theyear fishes (mostly fry). At each site, four adjacent
25 m reaches of shoreline were used for fish sampling. For each replicate reach, separate samples
were taken using a 7.62 m by 1.8 m bag seine (4.8
mm mesh, one offshore haul) and a 4.6 m by 1.2 m
(3.2 mm mesh, several inshore hauls) regular
seine. Fishes from all inshore hauls and the offshore haul were then pooled to depict the fish
assemblage for that 25 m reach. Gido et al. (2002)
demonstrated that this procedure was adequate to
depict fish assemblage structure in littoral-zone
areas of Lake Texoma, and provide additional
details on sampling and laboratory procedures.
Following processing, fishes were preserved in
50% isopropyl alcohol and archived in the Sam
Noble Oklahoma Museum of Natural History,
University of Oklahoma.
Benthic invertebrate assessments were made
annually at the same 40 sites during June and/or
July from 1999 through 2001. During each sampling period, three replicate samples were taken at
each site using a 15 cm by 15 cm Ponar dredge in
water 1.0-1.5 m deep. Following collection, each
dredge sample was rinsed through a 500 µm sieve
in the field to remove excess sediment and organic
debris. In 1999 and 2000, samples were preserved
in the field with 10% formalin containing rose
bengal dye; in 2001, samples were preserved with
70% ethyl alcohol. In the laboratory, invertebrates
were separated from the benthic detritus and other
extraneous material by trained technicians using a
dissecting microscope. Organisms were identified
to the lowest practical taxonomic level (usually
genus) using a compound microscope and standard taxonomic keys. Because most taxa collected
(primarily chironomid "midges" [Insecta] and
oligochaete "worms" [Annelida]) required that
106
M. A. Eggleton et al.
slide mounts be made for identification, a project
reference collection also was developed to aid in
identifications.
Data analysis
As a first-order assessment of reference and
impact sites, we used canonical correspondence
analysis (CCA; ter Braak 1986) to evaluate relationships between fish and invertebrate assemblages and environmental variables. CCA is a
direct gradient analysis that ordinates a
species/taxa by sample data matrix within the constraint that scores be linear combinations of environmental variables (Palmer 1993). Twenty-four
environmental variables, that included physical,
chemical, biotic, land use, and hydrological factors, were assessed for each of the 40 sites (Eggleton et al. in press). Significance of the species/taxa
-environment association was evaluated by comparing observed eigenvalues from the first three
ordination axes to those generated from randomisation of the data (10 000 iterations). Rejection of
this test meant that observed eigenvalues were
greater than expected by chance (i.e. eigenvalues
generated from randomised data), which indicated
a significant association existed between assemblages and environmental variables. All CCA calculations were done on species/taxa by sample
and environmental variable by sample data matrices using PC-ORD (McCune, Mefford 1999); all
species data and environmental data measured on
a continuous scale were square-root transformed
prior to analysis (ter Braak 1986).
Three main methods were used for assessment of assemblage differences between reference
and impact sites. First, simple paired analyses of
mean differences between reference and impact
site assemblages were done using Wilcoxon
signed-rank tests (Conover 1980). Assemblage
measures that were assessed included species or
taxa richness, organism abundance (as number of
fish seine haul-1 or number of invertebrates
dredge-1), and ordination site scores (CCA axis 1
and 2 only, as measures of assemblage structure).
Site-pair differences were derived by subtracting
values at reference sites from those of impact
sites; thus, negative differences imply that reference site values were greater than impact site
values. All paired analyses were done using Statistical Analysis Software (SAS Institute 2000).
Second, Procrustean analysis (Gower 1971)
was conducted using ordinations from correspondence analysis (CA; Legendre, Legendre 1998)
done separately for reference and impact sites.
Procrustean analysis superimposes ordinations
from reference and impact sites using a rotationalfit algorithm that minimises the sum-of-squared
residuals between the two matrices (Jackson 1995;
Peres-Neto, Jackson 2001). Residuals between the
original values and the best-fit solution are calculated for each site pair to identify outlying or
deviant site pairs. The m2 value is the test statistic
and also serves as a goodness-of-fit measure that
describes the degree of concordance between reference site and impact site matrices. Low m2
values and low residual sum-of-squares represent
high concordance between assemblage structures
of the different matrices; high m2 values and high
residual sum-of-squares represent low concordance between matrices. The statistical significance of the m2 statistic was evaluated using a
simple reshuffling algorithm (20 000 randomisations) as proposed by Jackson (1995). All correspondence analysis ordinations were done on
species/taxa by sample data matrices using PCORD (McCune, Mefford 1999); all species data
were square-root transformed prior to analysis (ter
Braak 1986). Procrustean analyses were performed using the first three CA axes with
PROTEST (Jackson 1995).
Third, we performed Mantel tests (Legendre,
Legendre 1998) to assess the spatial concordance
of reference site and impact site assemblages.
Mantel tests were performed between species/taxa
by sample matrices (square-root transformed raw
counts converted to relative abundances) using the
percent similarity distance measure. Mantel tests
represent a more holistic assessment method
because similarities among all reference sites and
all impact sites (not just similarities within preselected site pairs) are considered. The standardised Mantel statistic (Z) is the test statistic, with
higher values meaning greater concordance
between reference and impact sites. The statistical
significance of the Z-statistic was evaluated using
a simple reshuffling algorithm as with Procrustean
analysis (20 000 randomisations). All Mantel tests
were conducted using NTSYSpc (Rohlf 2002).
Significance for all analyses was declared at
an alpha level of 0.05. However, to account for
multiple comparisons in paired analyses separated
by impact type, we employed a sequential Bonferroni adjustment of alpha when appropriate following Rice (1989) and Gelwick, Matthews (1992).
3. Results
General patterns in assemblage structure
Canonical correspondence analysis results
indicated a significant association between fish
assemblages and environmental variables
(P=0.001). Fish assemblages were structured primarily along gradients of water transparency,
water-column chlorophyll-a, and annual discharge
(Fig. 1). These variables represented strong longitudinal gradients in Lake Texoma (i.e. varying
greatly from uplake to downlake), and were simi-
Anthropogenic influences on aquatic communities
107
2
Impact site
Reference site
Silt content
1
CCA axis 2
WC chlorophyll- a
Annual discharge
Residential area
0
Secchi reading
-1
Site exposure Sand content
-2
-2
-1
0
1
2
3
CCA axis 1
2
LABSIC
FUNNOT
LEPCYA
POMANU
LEPHUM
CCA axis 2
1
POMNIG
LEPMEG NOTCHR
LEPMAC
ICTBUB
CARCAR LEPOSS
APLGRU
GAMAFF
MACSTO
ICTPUN
PIMVIG
CYPLUT
HYBPLA
DORPET
DORCEP
CTEIDE
CYPCAR
MENBER
0
NOTATH
HIOALO
MICSAL
MICPUN
PERMAC
CYPVEN
MICDOL
MORCHR
MACEST
-1
MORSAX
Fig. 1. Scatterplot of site (top) and species
(bottom) scores from canonical correspondence analysis (CCA) of littoralzone fish assemblages at Lake Texoma,
1999-2001. Arrow lengths and directions
represent relative strengths of gradients.
First and second axes had eigenvalues of
0.306 and 0.233, respectively, and
together explained 28.9% of variation in
assemblage structure. Species codes:
APLGRU = Aplodinotus grunniens,
CARCAR = Carpiodes carpio, CTEIDE
= Ctenopharyngodon idella, CYPCAR =
Cyprinus carpio, CYPLUT = Cyprinella
lutrensis, CYPVEN = C. venusta,
DORCEP = Dorosoma cepedianum,
DORPET = D. petenense, FUNNOT =
Fundulus notatus, GAMAFF = Gambusia
affinis, HIOALO = Hiodon alosoides,
HYBPLA = Hybognathus placitus,
ICTBUB = Ictiobus bubalus, ICTPUN =
Ictalurus punctatus, LABSIC = Labidesthes sicculus, LEPCYA= Lepomis cyanellus, LEPHUM = L. humilis, LEPMAC =
L. macrochirus, LEPMEG = L. megalotis,
LEPOSS = Lepisosteus osseus, MACEST
= Macrhybopsis aestivalis, MACSTO =
M. storeriana, MENBER = Menidia
beryllina, MICDOL = Micropterus
dolomieu, MICPUN = M. punctatus,
MICSAL = M. salmoides, MORCHR =
Morone chrysops, MORSAX = M. saxatilis, NOTATH = Notropis atherinoides,
NOTCHR = Notemigonus chrysoleucas,
PERMAC = Percina macrolepida,
PIMVIG
=
Pimephales
vigilax,
POMANU = Pomoxis annularis,
POMNIG = P. nigromaculatus.
-2
-2
-1
0
1
2
3
CCA axis 1
lar to those reported by Gido et al. (2002). Sites
with high water transparency (i.e. Secchi readings,
greater axis-1 scores) were located almost exclusively in the main reservoir near the dam and associated with species such as striped bass Morone
saxatilis (MORSAX) and smallmouth bass
Micropterus dolomieu (MICDOL; Fig. 1). Higher
levels of water-column chlorophyll-a and annual
discharge (lower axis-1 scores) were characteristic
of the river arms and transitional areas and associated with species such as freshwater drum
Aplodinotus grunniens (APLGRU), white crappie
Pomoxis annularis (POMANU), bullhead minnow
Pimephales vigilax (PIMVIG), emerald shiner
Notropis athernoides (NOTATH), common carp
Cyprinus carpio (CYPCAR), and river carpsucker
Carpiodes carpio (CARCAR; Fig. 1).
Fish assemblages also were structured along
gradients related to the level of site exposure to
wind and waves and the sand content of the substrate (lower axis-2 scores), and levels of residential development and substrate silt content (higher
axis-2 scores). Species associated with high-exposure
sites and increased sand content included striped bass,
inland silverside Menidia beryllina (MENBER), and
several cyprinid species. Species typically found at
low-exposure sites with increased silt content included
brook silverside Labidesthes sicculus (LABSIC),
blackstripe topminnow Fundulus notatus (FUNNOT),
bluegill Lepomis macrochirus (LEPMAC), longear
sunfish L. megalotis (LEPMEG), and largemouth bass
M. salmoides (MICSAL; Fig. 1).
For benthic invertebrates, CCA also indicated
a significant association between assemblage
structure and environmental variables (P=0.001).
However, unlike fishes, invertebrate assemblages
exhibited a stronger spatial pattern within the
reservoir. Invertebrate assemblages were mainly
108
M. A. Eggleton et al.
2
Impact site
Reference site
Site exposure
Secchi reading
CCA axis 2
1
Conductivity
TKN
WC Chlorophyll- a
0
-1
Collection year
Wetland
-2
-2
-1
0
1
2
3
CCA axis 1
2
1
LIMNO
CRYP
PSEUPOLY
0
CCA axis 2
CRYPT
CLAD
COEL
PARAC
DICR
ENDO NEMA
PUPA
CRIC GASTR
EPHEM
TRICH
GLYP
ABLA
PROB
TANYT
-1
AULO
CHAO
CHIR TANYP
BRAN
MICRO
HIRUN
Fig. 2. Scatterplot of site (top) and taxa
(bottom) scores from canonical correspondence analysis (CCA) of littoral-zone
benthic invertebrate assemblages at Lake
Texoma, 1999-2001. Arrow lengths and
directions represent relative strengths of
gradients. First and second axes had
eigenvalues of 0.304 and 0.212, respectively, and together explained 18.9% of
variation in assemblage structure. Taxa
codes: ABLA = Ablabesmyia, AULO =
Aulodrilus, BRAN = Branchiura sowerbyi, CHAO = Chaoborus, CHIR = Chironomus, CLAD = Cladotanytarsus,
COEL = Coelotanypus, COLEO =
Coleoptera, CRIC = Crictopus, CRYP =
Cryptochironomus, CRYPT = Cryptotendipes, DERO = Dero, DICR = Dicrotendipes, ENDO = Endochironomus,
EPHEM = Ephemeroptera, GASTR =
Gastropoda, GLYP = Glyptotendipes,
HIRUN = Hirudinea, IMOLIG = Immature oligochaete, LARS = Larsia,
LIMNO = Limnodrilus, MICRO =
Microchironomus, NAIS = Nais, NEMA
= Nematoda, ODON = Odonata, PARA =
Parachironomus, PARAC = Paracladopelma, POLY = Polypedilum, PRIS
= Pristina, PROB = Probezzia, PROC =
Procladius, PSEU = Pseudochironomus,
PUPA = Chironomidae pupae, TANYP =
Tanypus, TANYT = Tanytarsus, TRICH =
Trichoptera, TUBIF = Tubifex.
PROC
PARA
TUBIF
IMOLIGNAIS
ODON
DERO
COLEO
-2
PRIS
LARS
-3
-2
-1
0
1
2
3
CCA axis 1
structured along gradients of water transparency,
water-column chlorophyll-a, and total Kjeldahl
nitrogen (TKN; Fig. 2). Sites with higher water
transparency (low axis-1 and high axis-2 scores)
were located mostly in the main reservoir near the
dam and associated with the chironomids
Polypedilum (POLY), Pseudochironomus (PSEU),
and Cryptochironomus (CRYP). Sites with low
transparency and high levels of TKN (greater axis1 and axis-2 scores near zero) were characteristic
of the turbid, more saline waters of the Red River
arm. Typical taxa from these assemblages were the
tubificid oligochaetes Limnodrilus (LIMNO) and
Branchiura sowerbyi (BRAN), and the chirono-
mids Tanypus (TANYP), Chironomus (CHIR),
Cladotanytarsus (CLAD), Coelotanypus (COEL),
Microchironomus (MICRO), and Paracladopelma
(PARAC, Fig. 2).
Secondarily, invertebrate assemblages were
structured by collection year (low axis-2 scores),
specific conductance (high axis-2 scores), and
proportion of wetlands in the site drainage subbasin (low axis-2 scores; Fig. 2). Sites with high
proportion of wetlands were located exclusively in
the Washita River arm, and typically contained the
oligochaetes Nais (NAIS) and Pristina (PRIS),
and the chironomids Larsia (LARS), Glyptotendipes (GLYP), and Tanytarsus (TANYT; Fig.
Anthropogenic influences on aquatic communities
109
measures were inadequate for
assessing anthropogenic stress
0.6
with fishes and benthic inverte*
brates in Lake Texoma.
*
In contrast, paired CCA
*
*
scores indicated differences in
assemblage structure between
0.4
reference and impact sites for
both fishes and benthic invertebrates. With fishes, axis-1 differences were near zero and not
0.2
significant whether data were
pooled across years (P=0.191)
or not (P=0.418-0.528; Fig. 3).
Conversely, CCA axis-2 scores
were significantly different
0.0
during all 3 years (1999,
Overall
1999
2000
2001
P=0.029; 2000, P=0.049; 2001,
P=0.012) and when data were
Fig. 3. Mean difference (difference = impact site - reference site) in CCA axis1 and axis-2 site scores of fish assemblages between paired reference and
pooled across years (P=0.001;
impact sites at Lake Texoma, 1999-2001. Asterisk denotes significant differFig. 3). This suggested that
ence (P < 0.05). Whisker represents standard error.
reference-impact structural
differences with fishes were
2). However, the wetlands gradient was dominated
more related to secondary environmental gradiby one site with a small drainage area, and was the
ents in Lake Texoma, such as level of residential
only site with a significant proportion of wetland
development and substrate content, which may
habitat in its sub-basin (35%, no other site with
have varied significantly within site pairs. Surprismore than 5%). Unlike fishes, invertebrate assemingly, site exposure also was important in discrimblages differed not only between the main reserinating reference and impact sites along CCA axis
voir and river arms on CCA axis 1, but also
2. Although small differences in site exposure did
between the two different river arms on CCA axis
exist within some site pairs, this observation was
2.
not expected because the site-pairing procedure
should have controlled for this effect. This observation was driven mostly by the increased abunReference-impact site comparisons
dances of sunfishes at impact sites in marinas
compared to their reference sites.
Paired analyses.
With benthic invertebrates, distribution of
Overall, neither fish species richness (P=0.3070.584) nor fish abundance
1.0
(P=0.452-0.985) were significantly different between reference
*
0.8
and impact sites. Similarly, no dif*
ferences in fish richness or abun*
0.6
*
dance were detected with any
particular impact types (P=0.0980.4
0.854, Bonferroni -adjusted).
Similar results were obtained for
0.2
benthic invertebrate assemblages
(richness, P=0.132-0.277; abundance, P=0.087-0.992); no differ0.0
ences in invertebrate richness or
abundance were detected with
-0.2
any particular impact types
CCA axis 1
CCA axis 2
(P=0.103-0.875,
Bonferroni-0.4
adjusted). Results with both
assemblages were identical
Overall
1999
2000
2001
regardless of whether data were
pooled across years or analysed
Fig. 4. Mean difference (difference = impact site - reference site) in CCA
by individual year. Furthermore,
axis-1 and axis-2 site scores of benthic invertebrate assemblages between
results indicated that species richpaired reference and impact sites at Lake Texoma, 1999-2001. Asterisk
denotes significant difference (P < 0.05). Whisker represents standard error.
ness and organism abundance
Mean difference
Mean difference
CCA axis 1
CCA axis 2
110
M. A. Eggleton et al.
0.4
m 2 = 0.92
P = 0.008
Reference/impact Procrustean residual
1999
0.3
0.2
0.1
0.0
AGRIC
BEACH
DUMP
MARINA
OIL
SEPTIC
0.4
m 2 = 0.47
P = 0.001
Reference/impact Procrustean residual
2000
0.3
0.2
0.1
0.0
AGRIC
BEACH
DUMP
MARINA
OIL
SEPTIC
0.4
m 2 = 0.64
P = 0.001
Reference/impact Procrustean residual
2001
0.3
0.2
0.1
0.0
AGRIC
BEACH
DUMP
MARINA
OIL
SEPTIC
Fig. 5. Procrustean analysis residuals from comparison of fish
assemblages at reference and impact sites in Lake Texoma, 19992001, averaged by impact type. Whisker represents standard error.
CCA axis scores also indicated structural differences between reference and impact sites (Fig. 4).
However, unlike fishes, differences were restricted
to axis-1 scores. Differences in axis-1 scores
between reference and impact sites were significant during all 3 years (1999, P=0.006; 2000,
P=0.001; 2001, P=0.009) and when data were
pooled across years (P=0.001; Fig. 4).
Conversely, differences in axis-2 scores
were not significant (individual years,
P=0.176-0.474; years pooled, P=0.227).
Thus, primary gradients such as watercolumn chlorophyll-a and Secchi reading (also identified from the fish CCA)
were important in discriminating benthic
invertebrate assemblages at reference
and impact sites.
Procrustean analyses.
Procrustean analyses gave different results for fishes and invertebrates.
With fishes, reference and impact site
assemblages were highly concordant
each year (P=0.001-0.008), meaning that
the general structure of fish assemblages
was consistently similar between reference and impact sites through time.
Mean Procrustes residuals indicated that
marinas usually contained the most different fish assemblages between reference and impact sites (Fig. 5), although
the difference was not large enough to
yield a nonsignificant test.
Conversely, benthic invertebrate assemblages were highly discordant all 3 years (P=0.14-0.88), which
indicated that reference-impact differences in assemblage structure approximated differences found between
random communities (Jackson 1995).
Mean Procrustes residual values indicated that sanitary dump (0.29±SE of
0.09) and agricultural (0.26±0.05) site
pairs contained the most different
assemblages (Fig. 6). However, mean
residuals exceeded 0.20 for all impact
types, which suggested consistent structural differences between reference and
impact sites. Regardless of impact type,
impact sites consistently contained
greater abundances of oligochaetes
(mean oligochaete-chironomid ratio
3.54±1.45), whereas greater abundances
of chironomids were found at reference
sites (0.42±0.10).
Mantel tests.
Mantel test results generally
supported Procrustean analysis results
for both fishes and benthic invertebrates.
With fishes, significant positive concordance
(P=0.001-0.027) was detected between the assemblage structures of reference and impact sites
during 2 of 3 years (Fig. 7). During the year for
which a nonsignificant result was obtained (2000),
the association between reference and impact sites
was still strongly concordant (P=0.079). With
111
Anthropogenic influences on aquatic communities
Procrustean analysis and Mantel
tests also were used to assess the concordance of fish and benthic invertebrate
assemblages. Both analyses were done
separately for references site and impact
site assemblages. Mantel tests indicated
no concordance between fish and invertebrate assemblages at reference
(P=0.667-0.984) or impact (P=0.3020.736) sites. Procrustean analyses also
indicated no concordance between
assemblages during 1999 and 2000 (reference sites, P=0.130-0.229; impact
sites, P=0.207-0.409). However, significant concordance between fish and invertebrates was detected during 2001 at both
reference (P=0.001) and impact
(P=0.023) sites, which indicated that the
arrangement of site pairs in ordination
space was closer than expected by
chance (Jackson 1995). This concordance also corresponded with the one of
three years that reference-impact invertebrate assemblages were judged concordant from Mantel tests. Although it is
possible that inclusion of additional axes
may have produced more significant
tests, three axes usually contain most of
the relevant biological information in
ordination (Legendre, Legendre 1998).
4. Discussion
Community responses to anthropogenic activities
m 2 = 1.18
P = 0.137
Reference/impact Procrustean residual
1999
0.5
0.4
0.3
0.2
0.1
0.0
AGRIC
BEACH
DUMP
MARINA
OIL
SEPTIC
0.6
2000
Reference/impact Procrustean residual
Concordance of fish and invertebrate assemblages
0.6
m 2 = 1.53
P = 0.882
0.5
0.4
0.3
0.2
0.1
0.0
AGRIC
BEACH
DUMP
MARINA
OIL
SEPTIC
0.5
2001
Reference/impact Procrustean residual
benthic invertebrates, no concordance
was detected between assemblage structures at reference and impact sites during
2 of 3 years (1999, P=0.401; 2000,
P=0.248; 2001, P=0.018; Fig. 8). Nonsignificant results in 1999-2000 were
largely due to high numbers of 0% similarities between reference and impact
sites, which indicated no taxa in common
within site pairs.
m 2 = 1.28
P = 0.281
0.4
0.3
0.2
0.1
0.0
AGRIC
BEACH
DUMP
MARINA
OIL
SEPTIC
Fig. 6. Procrustean analysis residuals from comparison of benthic
invertebrate assemblages at reference and impact sites in Lake
Texoma, 1999-2001, averaged by impact type. Whisker represents
standard error.
Fishes.
CCA axis scores identified some species/taxa
from each assemblage type that were mostly associated with either reference or impact sites. With
fishes, we did not observe a predominance of pollution-tolerant species (e.g., green sunfish L. cyanellus, golden shiner Notemigonus chrysoleucas, or
bullheads Ameiurus spp.; Ohio EPA 1987) at
impact sites in Lake Texoma. Overall, largemouth
bass and threadfin shads Dorosoma petenense on
average tended to be more abundant at reference
sites, whereas white crappies, bluegills, brook silversides, and blackstripe topminnows were consistently more abundant at impact sites. The
largest assemblage differences were observed at
marinas and, to a lesser extent, agricultural sites.
In marinas, reference-impact differences were due
112
M. A. Eggleton et al.
D. cepedianum, bluegills, white crappies,
and mosquitofishes Gambusia affinis
were found at impact sites; greater abunZ = 0.427
0.8
dances of threadfin shads and bullhead
P = 0.027
minnows were characteristic of reference
sites.
0.6
Fish species associated with
impact sites in Lake Texoma were not
0.4
characteristic of environmentally stressed
conditions, and are not classified as tolerant species indicative of point-source
0.2
organic or heavy-metal pollution (Ohio
EPA 1987; USEPA 1998). Rather, these
particular species are mostly trophic spe0.0
0.0
0.2
0.4
0.6
0.8
1.0
cialists and their abundances reflected a
wide array of human-induced influences
1.0
in Lake Texoma. Marina impact sites con2000
tain high levels of artificial structure (e.g.,
boat houses, docks, moorings) and shoreZ
=
0.265
0.8
P = 0.079
line modifications, which created a gradient of habitat heterogeneity between
impact sites and reference sites (e.g., Trial
0.6
et al. 2001). Observed assemblage shifts
at marina impact sites towards a topmin0.4
now-brook silverside-sunfish assemblage, which are mostly bottom- and
surface-feeding insectivores (Robison,
0.2
Buchanan 1988), were consistent with an
increase in habitat complexity associated
with proximity to marinas (Hosn, Down0.0
ing 1994).
0.0
0.2
0.4
0.6
0.8
1.0
Fish assemblage shifts at agricultural sites were less obvious. Differ1.0
ences in local habitat structure unrelated
2001
to agricultural practices may partially
0.8
explain assemblage differences. For
Z = 0.554
P = 0.001
instance, one impact site contained high
abundances of emergent vegetation not
0.6
found at other sites, which would correspond with high abundances of juvenile
bluegills and white crappies (Robison,
0.4
Buchanan 1988), whereas another impact
site was shallow (0.5 m) with low gradi0.2
ent (1%) and contained large areas of
flooded grass, consistent with the high
abundances of mosquitofishes (Robison,
0.0
Buchanan 1988). Given that fish assem0.0
0.2
0.4
0.6
0.8
1.0
blages respond rapidly and predictably to
Reference site PSI (%)
changes in habitat complexity (Gorman,
Karr 1978; Angermeier, Karr 1984; JenFig. 7. Scatterplots of percent similarity values (PSI) from Mantel
tests on fish assemblages at reference and impact sites in Lake nings et al. 1999), observed shifts at these
impact sites were, as in marinas, largely a
Texoma, 1999-2001.
function of predictable habitat changes
to increased abundances of blackstripe topminand not severe point- or nonpoint-source pollution
nows, brook silversides, and several Lepomis sunin Lake Texoma.
fishes at impact sites, and greater abundances of
largemouth bass, blacktail shiners Cyprinella
Benthic invertebrates.
venusta, inland silversides, and bigscale logperch
Benthic invertebrate assemblages at impact
Percina macrolepida at reference sites. At agriculsites in Lake Texoma exhibited three main charactural sites, greater abundances of gizzard shads
teristics. At impact sites, invertebrate richness was
1.0
Impact site PSI (%)
Impact site PSI (%)
Impact site PSI (%)
1999
Anthropogenic influences on aquatic communities
113
Impact site PSI (%)
Impact site PSI (%)
Impact site PSI (%)
1.0
on average about one taxon greater and
abundances 40% greater than reference
1999
sites. Abundances of invertebrates have
Z
=
0.017
0.8
been shown to increase in environmenP = 0.401
tally degraded conditions, especially
eutrophic conditions associated with
0.6
organic enrichment (Wiederholm 1984;
Cyr, Downing 1988). Impact sites also
showed a tendency towards an
0.4
oligochaete-dominated fauna, which is
consistent with previous pollution-based
0.2
studies with invertebrates (Howmiller,
Scott 1977; Wentsel et al. 1978; Winner
et al. 1980). Specifically, we observed
0.0
greater
abundances
of
tubificid
0.0
0.2
0.4
0.6
0.8
1.0
oligochaetes (Limnodrilus, Aulodrilus,
and Branchiura sowerbyi), which are
1.0
characteristic of polluted environments
2000
(Milbrink 1978; Lauritsen et al. 1985),
Z
=
0.066
and may have been indicators of anthro0.8
P = 0.248
pogenic impact at some sites.
Limnodrilus was consistently more abundant at all impact sites regardless of
0.6
impact type, whereas abundances of
Aulodrilus and Branchiura sowerbyi
0.4
were greater only at agricultural and oil
impact sites. All of these taxa are commonly associated with eutrophic condi0.2
tions or some form of nutrient
enrichment (Wiederholm 1984), a scenario likely at many of the impact sites in
0.0
Lake Texoma due to the extensive row0.0
0.2
0.4
0.6
0.8
1.0
crop agriculture and pasture lands in the
watershed.
1.0
Although the chironomid taxa col2001
lected were similar to those previously
Z = 0.264
0.8
found in Lake Texoma (Vaughn 1982),
P = 0.018
certain chironomids also may have indicated degraded environmental conditions
0.6
at some sites. The genus Cricotopus is
commonly associated with heavy-metal
pollution (Winner et al. 1980), but they
0.4
contributed only 0.4% of the total abundance of invertebrates, and heavy-metal
concentrations were always below detec0.2
tion limits at sites where the taxon was
collected. Abundances of Chironomus
were consistently greater at agricultural
0.0
0.0
0.2
0.4
0.6
0.8
1.0
impact sites, and Tanypus abundances
were consistently greater at agricultural
Reference site PSI (%)
and oil impact sites. Chironomus and
Fig. 8. Scatterplots of percent similarity values (PSI) from Mantel
Tanypus are common genera of which
tests on benthic invertebrate assemblages at reference and impact
some species have been associated with
sites in Lake Texoma, 1999-2001.
heavy sedimentation and nutrient enrichment (Wiederholm, Erickson 1979; Winnell,
cations sufficient for assessing many environmenWhite 1985; Dawson, Hellenthal 1986). The lack
tal impacts (Ferraro, Cole 1992).
of species-level identifications precludes a more
Overall, results suggest that observed shifts in
thorough assessment. However, pollution tolerinvertebrate assemblages between reference and
ances for many benthic invertebrates are fairly
impact sites were likely related to anthropogenic
similar across genera making genus-level identifiinfluences in Lake Texoma. However, the influ-
114
M. A. Eggleton et al.
ences were those (e.g., row-crop agriculture) with
the potential of producing widespread nonpointsource effects such as sedimentation and nutrient
enrichment. No evidence was found that suggested
assemblage shifts were in response to point-source
pollutants such as industrial effluents or runoff,
which is a priority concern for the USEPA.
Concordance of fish and invertebrate
assemblages
We observed that fish and invertebrate
assemblages in Lake Texoma were structured
around similar variables, yet two conventional
methods showed little concordance between fish
and invertebrate assemblages at reference or
impact sites. These results were opposite those
obtained by Jackson, Harvey (1993), who found
significant concordance between fish and invertebrate assemblages across 40 Canadian lakes,
despite assemblages being structured along different environmental gradients. Fish assemblages in
their study were structured around lake morphological measures (e.g., depth, area, volume),
whereas invertebrate assemblages were related
more to water chemistry measures (mainly pH, but
also dissolved organic carbon, nitrate, sulphate).
They speculated that observed patterns were the
result of biotic processes within and between
assemblages, most likely interactions between fish
predators and invertebrate prey.
So what factors might affect the concordance
or discordance between assemblages in a given
aquatic system? Multi-assemblage studies such as
Jackson, Harvey (1993) are rare; thus, it is not
widely established whether assemblages are likely
to be discordant or concordant in aquatic systems.
Furthermore, our study and theirs were conducted
at considerably different scales (i.e. 40 lakes vs. 40
sites in one reservoir), so potential causative
processes could well vary more among systems
than within systems.
In Lake Texoma, observed discordance
between fish and invertebrates was partly attributable to the extreme temporal variability exhibited
by the dominant taxa in each assemblage. With
fish assemblages, the three most abundant species
(88% of the total numbers) exhibited large fluctuations in abundance each year. Mean abundances
of striped bass increased approximately 90-fold
from 0.8±0.3 fish seine haul-1 in 1999 to 90±31
fish seine haul-1 in 2001. Threadfin shad abundance increased greater than 300% during 2000,
whereas inland silversides exhibited a 90%
increase between 1999 and 2000 followed by a
70% decrease in 2001 (Schnell et al. 2002). Similarly, annual fluctuations in dominant benthic
invertebrate taxa also were observed. Abundances
of two of the three most abundant oligochaetes
(22% of the total numbers) declined substantially
during the study. Abundances of Limnodrilus
declined from 7.1±1.4 dredge-1 in 1999 to 2.7±1.3
dredge-1 in 2001, whereas Branchiura sowerbyi
declined from 1.5±0.5 dredge-1 in 1999 to 0.5±0.2
dredge-1 in 2001 (Schnell et al. 2002). Additionally, three of the five most abundant chironomids
(30% of the total numbers) also exhibited substantial increases in abundance during the same
period: Polypedilum increased from 1.7±0.6
dredge-1 in 1999 to 6.9±1.4 dredge-1 in 2001;
Dicrotendipes increased from 0.4±0.2 dredge-1 in
1999 to 5.2±1.8 dredge-1 in 2001; and Glyptotendipes increased from 0.05±0.02 dredge-1 in
1999 to 1.3±0.7 dredge-1 in 2001 (Schnell et al.
2002). Declining abundances of dominant taxa
associated with impact sites (e.g., Limnodrilus and
Branchiura sowerbyi) were likely related to the
significant concordance of reference and impact
invertebrate assemblages suggested by Mantel
tests in 2001 (assemblages had been discordant in
1999 and 2000). In other words, declines in the
abundances of taxa that made reference and
impact assemblages different in 1999-2000 led to
greater concordance in assemblages in 2001.
Why are littoral-zone fish and macroinvertebrate assemblages in Lake Texoma so variable
from year to year in terms of species or taxa abundance? The cause of this phenomenon is not
entirely clear. Previous studies (e.g., Lienesch,
Matthews 2000; Gido et al. 2002; Eggleton et al.
in press) have suggested that littoral-zone areas in
Lake Texoma are highly unstable environments
for aquatic biota, especially for fishes and presumably (from this study) benthic invertebrates.
Lienesch, Matthews (2000) indicated that wind
velocity and wave height were strongly related to
the structure of littoral-zone fish assemblages.
Gido et al. (2002) showed significant relationships
between fish assemblages and degree of northern
exposure and silt/sand content, measures that are
related to site exposure. In addition, Eggleton et
al. (in press) reported that site exposure to wind
and waves was significantly related to both larval
and juvenile/adult fish assemblages in Lake
Texoma. Substrates along highly-exposed shorelines in Lake Texoma are mostly medium and
coarse sands, which are relatively unstable and
tend to support lower densities of invertebrates
(Minshall 1984). Although we can only speculate,
exposure to wind and waves and associated physical changes (e.g., sand-dominated substrates) at
many sites in Lake Texoma may be of such magnitude that common interactions purported to be
involved in the structuring of assemblages in other
types of systems (e.g., Lyons, Magnuson 1987;
Hinch et al. 1991; Benson, Magnuson 1994; those
suggested by Jackson, Harvey 1993) may be
absent or intermittent and difficult to detect. Thus,
the general absence of associations between fish
and macroinvertebrate assemblages may not be
Anthropogenic influences on aquatic communities
surprising in Lake Texoma given the environmental variability in shoreline areas and associated
variability in assemblages.
We did expect that responses to anthropogenic influences, if detected, might be different
between assemblage types, likely because fishes
and benthic invertebrates respond to impacts at
different scales of space and time (e.g., Allen et al.
1999b). For instance, fish and invertebrates were
structured along similar primary gradients, specifically water clarity and water-column chlorophylla. However, paired reference-impact differences
on CCA axis 1 were not significant for fishes but
significant for invertebrates. This suggested that
observed differences in these variables within reference-impact site pairs were not sufficient to
affect fish assemblages. In other words, although
fish assemblages may have may differed substantially between clear and turbid sites, water clarity
itself varied little between reference and impact
sites that usually were less than 3 km apart. However, because reference-impact invertebrate
assemblages differed significantly along this same
gradient (i.e., composed of mostly the same variables in ordination), differences in water clarity
and water-column chlorophyll-a may have been
substantial enough to influence invertebrate
assemblages. Small within-site pair differences in
water clarity, water-column chlorophyll-a, and/or
TKN (an inorganic nutrient measure) that were
negligible to fishes may have been significant to
benthic invertebrates. An example would be the
production of benthic algae, which provides food
and habitat for many invertebrates (Merritt, Cummins 1996), and might be affected by subtle differences in water clarity or nutrients. Benthic
invertebrates also have much less mobility than
fishes, and would likely reflect environmental
stress at the community level more readily. Thus,
benthic invertebrates may have reflected anthropogenic impacts at finer scales than fishes, which
might be indicators of larger-scale impacts in Lake
Texoma. This suggestion also is consistent with
rapid bioassessment protocols developed by the
USEPA (Barbour et al. 1999), which presuppose
such scale-related differences in assemblages in
their monitoring programmes.
5. Conclusions
Fish and benthic invertebrate assemblages in
Lake Texoma did not exhibit high levels of degradation attributable to anthropogenic activities.
Minor shifts in fish assemblage structure at some
impact sites may have been related to anthropogenic influences, but shifts were not indicative
of severe point-source organic or heavy-metal pollution. Rather, fish assemblage shifts appeared
related to human modification of shoreline habi-
115
tats, especially in marinas. Benthic invertebrates
exhibited greater differences in assemblage structure between reference and impact sites, presumably in response to nonpoint-source impacts
associated with row-crop agriculture. Furthermore, our analyses suggested that species or taxa
richness alone was inadequate for assessing
anthropogenic stress on fish and invertebrate
assemblages in Lake Texoma. This finding concurs with Allen et al. (1999a), who regarded richness as a somewhat ambiguous measure of biotic
integrity and called for more intensive analyses
that focused on other community structural attributes.
We found little concordance between fishes
and benthic invertebrates despite assemblages
being structured along similar environmental gradients. The high annual variation in the abundance
of several community-dominant fishes and invertebrates contributed most to this observation.
Overall results suggested that fishes and invertebrates likely responded to anthropogenic influences at different scales of space and time in Lake
Texoma. We do acknowledge that our study may
have been conducted at such a broad scale that
assemblage differences may have been difficult to
detect with conventional approaches unless environmental impacts were of large magnitude. Nevertheless, our community-based results provide a
thorough baseline assessment of present-day ecological conditions in Lake Texoma, and highlight
areas in need of future monitoring and research.
Acknowledgements
Financial support for this study was provided
by the U.S. Army Corps of Engineers and the U.S.
Environmental Protection Agency. We are grateful
to D. Cobb, R. Page, and L. Weider for the use and
maintenance of equipment and facilities provided
at the University of Oklahoma Biological Station,
and M. Cook of the U.S. Environmental Protection Agency. Field and laboratory assistance was
provided by R. Ramirez, C. Hargrave, F. March,
M. Furuyama, and E. Johnson.
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