-------
Eggert, Burton and Mullen
additional statistical analysis of diversity data
using the non-parametric randomized
intervention analysis was redundant. A check of
all data sets indicated that all assumptions of
BACI had been satisfied, thus making RIA
unnecessary.
The abundance of a dominant diatom species
Achnanthes minutissima has followed a
predictable pattern of high dominance during
summer periods and low dominance during the
winter (Figure 2). All diatom abundance data
were transformed using the arcsin square root
of the mean transformation and tested for
additivity. Seasonal transformed data for the
before period were marginally additive (summer:
p < 0.06 and winter: p < 0.08).
Ourbin-Watson independence tests of summer
data found no significant autocorrelation
problem (d = 1.98), while winter data were
significantly autocorrelated (d - 0.90, p <
0.05). Results of unpaired t-tests for A.
minutissima indicated that there were no
significant inter-site changes in mean
differences for either seasonal "before" and
"after" periods, or year-to-year comparisons.
Since the winter abundance data were found to
be significantly autocorrelated, BACI results
were verified using RIA. RIA reflected the
results obtained in the BACI comparisons of
both the summer and winter abundance data
(Table 1).
In an attempt to detect even more subtle
changes in diatom abundances, we ran BACI
analyses on monthly data at peak A.
minutissima abundances (Figure 2). All May and
June data for the years 1983-1985 were
pooled to represent the "before" period and all
May and June data for 1986-1990 as the
"after" period so that mean differences beween
sites could be examined. The data appeared to
be significantly negatively serial correlated (d =
3.33, p < 0.05). Since negative
autocorrelations are conservative with regard to
probability levels, an unpaired t-test was run on
the monthly data. There was no significant
change in the inter-site relationship after
antenna operation according to the BACI
analysis (Table 1). This comparison could not
be verified with RIA due to the limited number
of observations available.
For the chlorophyll a. standing crop and species
diversity parameters where significant
differences were found using either BACI or
RIA, the data were scrutinized further to
determine whether ELF electromagnetic
radiation or another factor had caused the
observed differences. Significant differences
found by BACI or RIA do not imply that a
suspected perturbation has caused a change,
nor do these tests reveal at what point in time
the change occurred. Ecological and procedural
considerations should be examined in all cases.
Analysis of covariance (ANCOVA) of
chlorophyll a standing crop with ELF exposure
included as a covariant indicated that a variable
other than ELF electromagnetic radiation caused
the change in relationship between sites.
Significant positive correlations between water
temperature and chlorophyll a. during drought
periods from 1986 to 1990 suggest that the
observed differences were related to weather
variables. ANCOVA of species diversity data
using ELF exposure as a covariate also
indicated that a factor other than antenna
exposure was responsible for the significant
BACI and RIA results.
Along with the need for careful interpretations
regarding the possible sources of variation, data
set sample sizes should be considered when
deciding on the appropriate statistical analysis.
Generally, the statistical power of
non-parametric tests such as RIA is smaller
than that of a similar parametric test
(Welkowitz et at. 1976). When populations are
normally distributed, the number of
observations required by RIA should be larger
than the sample size required by the BACI
analysis in order to obtain the same amount of
power. Carpenter et al. (1989) reported that
RIA could consistently detect manipulation
effects with sample sizes of 40 or more.
Stewert-Oaten et al. (1986) did not suggest a
32
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Comparison of RIA and BACI Analysis
minimum sample size for the BACI analysis.
Although the authors did present an example in
which only 23 data points were used to detect
a manipulation effect, Monte Carlo simulations
are required to definitively determine the
effective sample size required by the BACI
method. Thus, year-to-year BACI comparisons
presented in this study should be interpreted
with some caution, since BACI's ability to
detect perturbation effects with sample sizes of
nine to eleven (Table 1) data points remains
unknown. The sample size problem also
emphasizes the need for long term monitoring
of potential pollutant effects.
In summary, an environmental impact study
should provide quantitative evidence to support
regulatory decisions regarding potential
environmental pollutants. Both the BACI
analysis and RIA offer a means of quantitatively
detecting whether perturbations such as toxic
effluents, pipeline construction, or power plant
discharges may be impacting an ecosystem.
The BACI and RIA results should be interpreted
with some care and caution however, as
significant findings do not imply that the
suspected pollutant has caused the observed
differences. Our analysis of ELF effects on a
riverine algal community using BACI and RIA
suggests that the following statistical protocol
will accurately and quantitatively allow the
detection of environmental perturbations. First,
the parametric BACI analysis should be used for
data sets satisfying plausible assumptions of
independence and additivity. If the relationship
between control and impacted sites has
changed significantly over time, or if the
independence, normality or additivity
assumptions appear to be questionable, then
the non-parametric randomized intervention
analysis may be used to examine the data.
Finally, if the inter-site relationship is found to
change over time using RIA, final conclusions
of the perturbations effects should be based on
ANCOVA results (using the magnitude of the
perturbation as the covariate) and/or other
ecological considerations. When used in this
manner, BACI and RIA represent complimentary
and practical tools with which to make sound
ecological decisions regarding potential
environmental impacts.
Acknowledgments
We thank Stephen R. Carpenter for copies of
the programs used to run the randomized
intervention analyses. Jennifer Molloy and Mark
P. Oemke assisted with the diatom
identification and enumeration. Support for this
research was provided by the Naval Electronic
Systems Command through a subcontract to IIT
Research Institute under contract numbers
N00039-81, N00039-84-C-0070, and
N00039-88-C-0065.
Literature Cited
Burton, T. M., D. M. Mullen, and S. L. Eggert.
1991. Effects of extremely low frequency (ELF)
electromagnetic fields on the diatom community
of the Ford River, Michigan, pp. 17-25 In: T.P.
Simon and W.S. Davis (editors). Proceedings of
the 1991 Midwest Pollution Control Biologists
Meeting. U.S. EPA Region V, Environmental
Sciences Division, Chicago, IL. EPA-905/R-
92/003.
Carpenter, S. R. 1990. Large-scale
perturbations: opportunities for innovation.
Ecology 71:2038-2043.
Carpenter, S. R., T. M. Frost, D. Heisey, and T.
K. Kratz. 1989. Randomized intervention
analysis and the interpretation of
whole-ecosystem experiments. Ecology
70:1142-1152.
Durbin, J. and G. S. Watson. 1951. Testing for
serial correlation in least squares regression II.
Biometrika 38:159-178.
Hulbert, S. H. 1984. Psuedoreplication and the
design of ecological field experiments.
Ecological Monographs 54:187-211.
Jassby, A. D. and T. M. Powell. 1990.
Detecting changes in ecological time series.
Ecology 71:2044-2052.
33
-------
Eggert, Burton and Mullen
Reckhow, K. H. 1990. Bayesian inference in
non-replicated ecological studies. Ecology
71:2053-2059.
Stewart-Oaten, A., W. W. Murdoch, and K. R.
Parker. 1986. Environmental impact
assessment: "Pseudoreplication" in time?
Ecology 67:929-940.
Steel, R. G. and J. H. Torrie. 1960. Principles
and procedures of statistics. McGraw-Hill, New
York, 481 pp.
Tukey, J. W. 1949. One degree of freedom for
non-additivity.Biometrics 5:232-242.
Welkowitz, J., R. B. Ewen and J. Cohen. 1976.
Introductory statistics of the behavioral
sciences. 2nd edition. Academic Press, New
York, 316 pp.
34
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The Freshwater Annelida (Polychaeta, Naidid and Tubificid Oligochaeta,
and Hirudinea) of the Great Lakes Region-an Overview
Donald J. Klemm
Bioassessment and Ecotoxicology Branch
Environmental Monitoring Systems Laboratory
U.S. Environmental Protection Agency
3411 Church Street, Cincinnati, Ohio 45244
Jarl K. Hiltunen
Sugar Island
Sault Ste. Marie, Michigan 49783
Abstract
The segmented worms are important components of benthic communities in nearly every freshwater
biotope. They are widely distributed, and some groups are found in great abundance. Several of the
annelid groups have been used for monitoring and detecting changes in water quality and physical
habitats. The habitat and water quality requirements as well as the pollution tolerance of many
species of freshwater annelids have been documented in the literature by a few investigators.
Practical taxonomic keys are now available to species, but many benthic water quality assessment
studies still do not treat the annelid groups adequately because the investigators lack the knowledge
and experience in using these keys. Furthermore, most bioassessment monitoring studies do not use
adequate sampling and processing (preservation) techniques for aquatic annelids. The inadequate
treatment by some investigators represents a loss of valuable ecological information for use in
biological assessment of the quality of water resources, water pollution, or other changes in aquatic
ecosystems resulting from natural causes or anthropogenic activities. The current aspects of
morphology, taxonomy, distribution, and organic pollution to polychaetes, naidid and tubificid
oligochaetes, and leeches of the Great Lakes Region species are presented and discussed.
Kev Words: macroinvertebrates, polychaetes, oligochaetes, leeches, pollution, water quality, organic
enrichment.
Introduction
Benthic animals, including the segmented
worms, are commonly used to demonstrate the
effects of pollution on the biological integrity of
surface waters and changes in the biotic
community (species composition, presence or
absence, and relative abundance of tolerant and
intolerant species) resulting from natural causes
and destructive activities by man (Aston 1984;
Brinkhurst 1974a,b; Carr and Hiltunen 1965;
Goodnight and Whitley 1960; Hiltunen 1967,
1969a-c, 1971; Hiltunen and Manny 1982;
Howmiller and Scott 1977; Milbrink 1983;
Sawyer 1974; Klemm 1991 and papers cited
therein). This paper is a taxonomic overview of
the freshwater polychaetes, naidid and tubificid
oligochaetes, and leeches of the Great Lakes
region with emphasis on their use to
demonstrate pollution effects and changes in
biotic community. A checklist of the species is
found in Table 1.
Annelida is an important and major phylum in
the animal kingdom. The body of annelids is
divided into rings (somites) or segments with
serially arranged organs. The phylum includes
three major classes, Polychaeta, Oligochaeta,
and Hirudinea. The distribution of aquatic
annelids is usually determined by the physical,
chemical, and biological characteristics of the
environment. Published accounts are relatively
sketchy for understanding the roles that some
of these characteristics play in the distribution
of annelids (gross chemical pollution not
35
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Klemm and Hiltunen
Table 1. Checklist of Polychaetes, Naidid and Tubificid Oligochaetes, and Leeches in the Great Lakes
Region.
Family Sabellidae
Manavunkia soeciosa
Family Naididae
Allonais pectinata
Amphicaeta america
Amphicaeta levdioii
Arcteonais lomondi
Bratislavia unidentata
Chaetogaster diaphanus
Chaetoqaster diastroohus
Chaetooaster limnaei
Chaetogaster setosus
Dero digitata
Dero f urcata
Dero nivea
Dero obtusa
Dero vaga
Haemonais waldvoqeli
Nais alpina
Nais bretspheri
Nais barbata
Nais behninoi
Nais communis
Nais elinouis
Nais pardalis
Nais pseudobtusa
Nais simplex
Nais variabilis
Class Polychaeta: Order Sabellida
Class Oligochaeta: Order Tubificida
Qphidonais seroentina
Paranais frici
Piguetiella michiaanensis
Piouetiella blanci
Pristina aequiseta
Pristina breviseta
Pristina leidyi
Pristina lonqiseta bidentata
Pristina lonaiseta lonoiseta
Pristina plumaseta
Pristina svnclites
Pristinella acuminata
Pristinella ienkinae
Pristinella osborni
Ripistes oarasita
Slavina appendiculata
Soecaria iosinae
Steohensoniana trivandrana
Uncinais uncinata
Veidovskvella comata
Veidovskvella intermedia
36
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Great Lakes Region Oligochaeta, Polychaeta, and Hirudinea
Table 1. Checklist of Polychaetes, Naidid and Tubificid Oligochaetes, and Leeches in the Great Lakes
Region (continued).
Class Oligochaeta: Order Tubificida
Family Tubificidae
Aulodrilus americanus
Aulodrilus limnobius
Aulodrilus piaueti
Autodrilus pluriseta
Bothrioneurum veidovskvanum
Branchiura sowerbvi
Haber cf. soeciosus
llvodrilus temoletoni
Isochaetides f revi
Isochaetides curvisetosus
Limnodrilus cervix
Limnodrilus cervix (variant form)
Limnodrilus claoaredianus
Limnodrilus hoffmeisteri
Limnodrilus hoffmeisteri (soiralis form)
Limnodrilus hoffmeisteri (variant form)
Limnodrilus maumeensis
Limnodrilus profundicola
Limnodrilus udekemianus
Phallodrilus hallae
Potamothrix bavaricus
Potamothrix bedoti
Potamothrix hammoniensis
Potamothrix moldaviensis
Potamothrix veidovskvi
Psammorvctides californianus
Quistadrilus multisetosus
Rhyacodrilus coccineus
Rhyacodrilus montana
Rhvacodrilus punctatus
Rhvacodrilus sodalis
Soirosperma ferox
Spirosperma nikolskvi
Tasserkidrilus harmani
Tasserkidrilus kessleri
Tasserkidrilus superiorensis
Teneridrilus flexus
Tubifex ianotus
Tubifex tubifex
Varichaetadrilus auoustipenis
Class Hirudinea: Order Arhynchobdellida
Family Haemopidae
Haemopis arandis
Haemopis lateromaculata
Haemopis marmorata
Haemopis plumbea
Haemopis terrestris
Family Hirudinidae
Macrobdella decora
Philobdella gracilis
37
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Klemm and Hiltunen
Table 1. Checklist of Polychaetes, Naidid and Tubificid Oligochaetes, and Leeches in the Great Lakes
Region (continued).
Family Erpobdellidae
Eroobdella dubia
Erpobdella parva
Erpobdella punctata
Mooreobdella bucera
Mooreobdella fervida
Mooreobdella microstoma
Nepheloosis obscura
Class Hirudinea: Order Rhynchobdellida
Family Glossiphoniidae
Actinobdella annectens Helobdella fusca
Actinobdella ineouiannulata Helobdella paoillata
Actinobdella pediculata Helobdella staanalis
Helobdella transverse
Alboqlossiphonia heteroclita Helobdella triserialis
Desserobdella michiaanensis
Desserobdella phalera
Desserobdella Dicta
Gloiobdella elonqata
Glossiphonia comolanata
Marvinmeveria lucida
Placobdella hollensis
Placobdella montifera
Placobdella ornata
Placobdella papillifera
Placobdella parasitica
Family Piscicolidae
Cvstobranchus meveri
Cvstobranchus verrilli
Mvzobdella luoubris
Theromvzon biannulatum
Theromvzon rude
Piscicola geometra
Piscicola milneri
Piscicola punctata
Piscicolaria reducta
withstanding). Despite the fact that annelids
may occur in all aquatic habitats and in great
numbers, especially certain oligochaete groups,
it must be stressed that much work remains to
be done on the ecology and pollution biology of
the annelids. The improper or inadequate
treatment of the segmented worms is attribut-
able, in part, to investigators that lack
appropriate experience or do not understand the
morphological terms and characters used in
practical keys. To interpret the quality of water
resources, the water quality requirements and
pollution tolerances, the animals should be
identified to the species level (Resh and
Unzicker 1975).
Class Polychaeta: Order Sabellida: Family
Sabellidae
General Morphology and Taxonomy
Polychaetes are segmented annelids typically
with parapodia associated with each body
38
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Great Lakes Region Oligochaeta, Polychaeta, and Hirudinea
segment. A high degree of modification in the
basic plan of many polychaetes has resulted in
different modes of existence, ranging from
sedentary (tubicolous) forms to highly free-
moving (errant) forms. The group is very
diverse, most forms are mainly found free
living; some are commensal with other inverte-
brates, and only a few are parasitic. The class
contains about 85 families and many species,
most of which are marine forms. World-wide,
only 10 families are represented in freshwater.
In North America only four families and 11
species are represented (Klemm 1985a,c):
Nereididae with six species, Ampharetidae with
one species, Sabellidae with one species, and
Serpulidae with two species. In the Great Lakes
Region only the family, Sabellidae is
represented by one species, Manavunkia
soeciosa; it is sedentary and inhabits a tube
built of mud or sand and mucus. The size of
this species is usully 2 to 5 mm long. The body
is divided into distinct regions, the thorax and
abdomen, with reduced or vestigal parapodia,
with simple capillary chaetae and hooks or
uncini. The prostomium is small or indistinct,
without appendages. In Sabellidae, the anterior
end is modified to form a branchial (tentacular)
plume or crown surrounding the mouth which
is used for food getting (filter feeding) and
respiration. For more information on the
taxonomy of the North America freshwater
forms, see Klemm (1985a,c).
General Distribution and Ecology
This species is widely distributed in the Neartic
region, from Duluth Harbor in western lake
Superior to St. Marys River, Lake St. Clair, the
Ottawa River in Ontario to Western Lake Erie,
Lake Ontario, and the upper St. Lawrence River;
eastward to the Finger Lakes and Hudson River,
New York, Schuylkill River and Delaware River,
Pennsylvania, Egg Harbor River, New Jersey,
and Lake Champlain, Vermont, south to North
Carolina, South Carolina, and Georgia. It has
also been reported from the Pacific northwest,
California, and Oregon to Alaska. One reason
for not detecting this species more often is its
small size; the sieve used to process the sample
may have mesh openings too large to retain the
specimens. Most specimens may pass through
a Standard No. 30 sieve, and a Standard No.
60 should be used.
Mackie and Qadri (1971) reported, during a
limnological survey of the Ottawa River, that
specimens of M- soeciosa occurred only in
substrates composed of silt and sand and in
moderately moving waters. Hiltunen (1965)
also found a large number of M- speciosa at the
mouth of the Detroit River, suggesting some
relationship between water movement and the
frequency of occurrence. Specimens in the
Mackie and Qadri study did not occur in
polluted water where BOD value exceeded 4
ppm nor where the DO content was less than 5
ppm. M- speciosa has also been found in lentic
habitats, many locations in Lake Erie (Hiltunen
1965, Krieger 1990), and in several lakes in
Alaska (Holmquist 1973). Spencer (1976)
found it in Cayuga Lake, New York at depths of
20 m or less where densities of more than
I000/sq. m were found occasionally. Poe and
Stefan (1974) reported this polychaete from the
Schuylkill River, Pennsylvania, near the type-
locality, and they reported that it appears to
have a wide range of tolerances for environ-
mental parameters such as DO (1.8 to 14.0
ppm), depth (0.3 to 16.0 m), pH (6.8 to 8.8),
and water temperature (2.8 to 28.3°C)
(obviously as low as 0°C because it survives
ice-cover seasons), and concluded that the only
environmental factor which may limit its
distribution is the requirement for fine
paniculate material in the substrate for the
construction of the tube in which it lives (gross
chemical pollution notwithstanding).
Class Oligochaeta
Introduction
Naidids and tubificids are predominantly found
in freshwater but some are strictly marine
forms. Most oligochaetes have chaetae, with a
few exceptions, and have no parapodia as in
the polychaetes. The body is segmented into
39
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Klemm and Hiltunen
somites or compartments separated by septum,
and each segment by convention is indicated by
a Roman numeral, progressing from anterior to
posterior. Segment I (including mouth and
prostomium) is devoid of chaetae, hence
numerical orientation of segments is achieved
by counting posteriad of the chaetophorous
segments, beginning with II. Normally each
segment bears four fascicles "bundles" of
chaetae, two dorso-lateral and two ventro-
lateral. There are two basic types of chaetae
(crotchets and capillif orms) whose numbers and
morphology in the various body regions are
taxonomically important. A crotchet can be
straight or curved (sigmoid), and usually
possess a more or less median thickening (the
node or nodulus), and may be simple-pointed or
have a bifid (cleft) distal end. Crochets are
found in all oligochaetes. Capilliform chaetae
which are elongate and simple-pointed, may be
smooth or finely serrated. Capilliform chaetae
when present, are found only in the dorsum of
the Naididae and Tubif icidae. For more detailed
information on aquatic oligochaete biology, see
Brinkhurst and Jamieson (1971), Brinkhurst and
Cook (1980), Bonomi and Erseus (1984), and
Brinkhurst and Diaz (1987).
Order Tubificida: Family Naididae
General Morphology and Taxonomy
Naidids are relatively small, commonly I mm to
10 mm and more or less transparent when alive.
All or nearly all can be identified to species by
the external morphology, particularly the shape
and arrangement of the chaetae. There are 20
genera and more or less 48 species known or
likely to occur in the Great Lakes region. In
North America 21 genera and 75 species are
reported. Keys work well for most species, but
some species descriptions are incomplete for
North American material. The kinds of chaetae
are much like those in the Tubificidae, except
the naidids have dorsal acicular (short needle-
like) chaetae that accompany the long capilli-
form (hair) chaetae. Some species of naidids
also bear pectinate chaetae like the tubificids.
Dorsal chaetae can begin in segment II or
posteriad to it; dorsal chaetae that accompany
capilliform (hair) chaetae are often very
different from ventral chaetae. Dorsal fascicles
often contain 1-2 capilliform chaetae and I-2
acicular chaetae. In summary, some naidids
have ventral chaetae only (Chaetooaster spp.);
other species have dorsals and ventrals with
bifids only, while still other species have
ventrals bifid and dorsals with capilliform plus
simple, bifid, pectinate, or palmate acicular
chaetae. The dorsal chaetae may begin in II, III,
or further back, usually V or VI, rarely beyond.
Some species may have eyes; may be found
budding (a form of asexual reproduction), and
when sexually mature, may bear genital
chaetae in segments V or VI; spermathecae in
segments IV, V, or VII; male pores on segments
V, VI, or VIII. However, these features are not
used in species identification. For more
information on taxonomy of the naidids, see
Hiltunen and Klemm (1980, 1985), Brinkhurst
(1986), and Brinkhurst and Kathman (1983).
General Distribution and Ecology
The naidids are an ecologically diverse group
(Leaner 1979) and are found in both lotic and
lentic waters. The naidids are widely distributed
and commonly inhabit the littoral zones of lakes
or other shallow waters in streams, ditches,
and ponds. Some species are sediment dwellers
(like tubificids) while other species are
characteristically found among the aquatic
plants. Naidid populations are usually reduced
where siltation and mud occur. Plants with a
thick growth habit and well-developed peri-
phyton community can support sizeable naidid
populations. Riffles and similar areas where the
substrate is primarily sand and gravel often
contain substantial naidid populations. Longi-
tudinal zonation of naidids along rivers has been
demonstrated. Learner et al. (1978) concluded
that factors associated with changes in altitude
and slope of a river (water velocity, substrate
type, presence and type of vegetation, and the
influence of municipal and industrial wastes)
can be important in influencing the distribution
of naidids. They are generally less significant in
lakes where they are confined primarily to the
40
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Great Lakes Region Oligochaeta, Polychaeta, and Hirudinea
littoral zone. The behavior of most naidids is
unlike that of most other oligochaetes because
some naidids can swim as well as crawl, and
others are small enough to be passively carried
by strong water movements.
Feeding habits of most species are unknown,
but some have been observed to feed on
detritus, grazing upon bacteria, protozoans, and
algae. Probably most oligochaetes are
herbivorous, but some Chaetooaster species are
primarily or perhaps entirely predaceous.
Reproduction occurs by paratomy (architomy,
asexually budding), where the posterior
segments of the naidid develop into daughter
zooids that break free after development is
complete, or by fragmentation. Sometimes the
worms are found consisting of two or more
individuals that have not yet separated from the
parent (anterior) section. Sexual reproduction is
considered uncommon in many species.
Order Tubificida: Family Tubificidae
General Morphology and Taxonomy
Tubificids are medium-sized to large worms,
commonly more than 20 mm long, that never
have eyes, never reproduce by asexual budding,
but occasionally regeneration of the posterior
section in some species suggests fragmen-
tation. A variety of chaetae are found among
the species. Crotchets are always present but
capilliform (hair) chaetae may or may not be
present depending on the species. Most species
are red when alive and coil or loop when
disturbed. In the Great Lakes region there are
presently 17 genera and 37 species. There are
21 genera and 64 species reported to occur in
North America. Tubificids are identified by the
characteristic shape of the somatic chaetae and
their genital chaetae (spermathecal or penial
chaetae) if present, or by mature male genitalia.
In some species penis sheaths in segment XI
are especially helpful in species identification.
Spermathecae are located in segment X, and
males pores are in Segment XI. Dorsal chaetae
always begin on segment II, dorsal chaetae are
often broadly similar in form to ventral chaetae;
dorsal fascicles often bear a complement of
more than 2 capilliform (hair) chaetae and 2 or
more crotchets. Therefore, some species of
tubificids have dorsal and ventral chaetae bifid;
other species have dorsal capilliform and
pectinate chaetae and ventrals mostly bifid.
Pectinate chaetae may be narrow and hairlike
distally in appearance. Some species have
dorsal capilliform and bifid chaetae and ventral
bifid chaetae. For more information on the
taxonomy of the tubificids, see Stimpson,
Klemm, and Hiltunen (1982, 1985) and
Brinkhurst (1986, 1989).
General Distribution and Ecology
Tubificids are most commonly found in soft
sediments rich in organic matter; several
tubificid species characteristically live in large
numbers in habitats that receive organic
pollution (Aston 1984, Brinkhurst 1974a,b,
Carr and Hiltunen 1965, Goodnight and Whitley
1960. Hiltunen 1967,1969a-c, 1971, Hiltunen
and Manny 1982, Howmiller and Scott 1977,
Krieger 1990, Milbrink 1983). Tubificids respire
cutaneously, but some species can tolerate
anoxic conditions and environmental stresses
(e.g., Limnodrilus hoffmeisteri and Tubifex
tubifex). A number of species in the family are
very stress-sensitive (Hiltunen 1967, Howmiller
and Scott 1977, Milbrink 1983). Tubificids
burrow in soft sediments, often in tubes of mud
and mucus secretions, as the classic name
implies. A few species occur in fine gravel or
sand. The quantity and quality of organic
matter reaching the sediment may be more
important in determining which tubificid species
will occur in a locality (gross chemical pollution
notwithstanding) than the physical-chemical
variables of water or sediment (Brinkhurst and
Cook 1974).
Tubificids are mostly deposit feeders living on
organic detritus and its associated bacteria,
microflora, and fauna. Tubificids typically feed
with their heads buried below the sediment
surface with their tails protruding above it. The
feeding activities of tubificids play an important
role in mixing the physical and chemical
41
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Klemm and Hiltunen
characteristics of sediments (Brinkhurst 1974).
Most tubificids reproduce by sexual repro-
duction even though they are hermaphrodites.
Tubificids enclose their eggs in cocoons and
deposit them on sediments.
Class Hirudinea (Not Hirudinoidea)
Leeches are serially segmented worms and are
considered closely related to the oligochaetes.
They are also hermaphroditic, i.e., they contain
both male and female organs in each individual.
Muscles and a hydrostatic skeleton are used in
locomotion. The nervous, excretory and
vascular systems are segmentally arranged.
Leeches have well-developed anterior and
posterior sucker, 34 segments (indicated by
Roman numerals I-XXXIV)) which are sub-
divided into annul!, a reduced coelom and
intestinal caeca, and usually two separate male
and female gonopores with male gonopore
anterior to the female gonopore. Leeches are
devoid of chaetae, except Acanthobdella
oeledina. a leech which has chaetae in the
anterior segments of the cephalic region
(Klemm 1985b,c). Although some leeches are
well adapted to a sanguivorous existence, the
group is also well-represented by species which
are both predatory and can engulf small animals
whole or parasitic fluid-feeders. Leeches are
found on all the continents, in terrestrial,
freshwater, estuarine, and marine
environments.
General Taxonomy and Morphology
Five families in the orders Arhynchobdellida and
Rhynchobdellida are represented in the Great
Lakes region: Haemopidae, Hirudinidae,
Erpobdellidae, Glossiphoniidae, andPiscicolidae.
Nineteen genera and 43 species have been
recorded from that region. Twenty-one genera
and 66 nominal species presently are reported
to occur in North America. For more infor-
mation on the taxonomy of leeches, see Klemm
(1985b,c, 1991) and Sawyer (1972,1986).
Leeches are found in most freshwater habitats,
but are often ignored by biologists because they
are thought to be difficult to identify to genus
and species. Also, investigators neglect them
because they lack an understanding of the
diagnostic (morphological) terms and characters
used in keys to identify specimens to the
lowest taxonomic level.
The external morphological characters are
usually sufficient for the identification of most
leeches. Internal characters used to identify
certain species of Haemopis are discussed in
Klemm (1985b,c). The general external
diagnostic features that are important for
identifying the leeches to species are: size of
mouth, general shape of body, form of suckers,
form of cephalic region, number and
arrangement of eyes, jaws and teeth, eyespots
(ocelli), papillae, pulsatile vesicles, digitate
processes on rim of caudal sucker, caudal-
sucker separation from body on narrow pedicle,
copulatory gland pores, the number of annul!
between gonopores, and pigmentation patterns.
Typically, the mouth opening of the haemopids
and hirudinids is medium to large, occupying
the entire sucker cavity, and the body is large,
linear, elongate and well-muscled, length 75-
300 mm. They are good swimmers. Haemopids
and hirudinids always have 5 pairs of eyes.
The mouth opening of erpobdellids is medium,
occupying the entire sucker cavity; the body of
erpobdellids is moderate size, linear, elongate,
length to 100 mm, and they are also good
swimmers. They usually have 3, or 4 pairs of
eyes (or eyes absent). The mouth of
glossiphoniids is a small pore on the rim or
within the oral sucker cavity; the body of this
group is dorso-ventrally flattened with the
posterior half usually much wider than the
tapering cephalic end, length to 40 mm. They
have I, 2, 3, or 4 pairs of eyes. The mouth of
piscicolids is a small pore within the oral sucker
cavity, and the body of the piscicolids is
cylindrical, narrow, posterior half can be slightly
flattened, length to 30 mm. The body may be
divided into a narrow neck (trachelosome) and
wider body (urosome) regions; caudal sucker
with or without eyespots, and body with or
without pulsatile vesicles. Piscicolids can have
42
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Great Lakes Region Oligochaeta, Polychaeta, and Hirudinea
one or two pairs of eyes (or eyes absent).
General Distribution and Ecology
In North America, freshwater leeches reach
their greatest species diversity in lakes,
permanent and temporary ponds, woodland
pools, bogs, wetlands, rivers, and streams
(Klemm 1972, Sawyer 1972). Klemm (1977,
1991) summarized the distribution of leeches in
the Great Lakes states (Illinois, Indiana,
Michigan, Minnesota, New York, Ohio, and
Wisconsin) bordering the Great Lakes, including
Ontario, Canada. Additional collecting in all
habitats, substrate types, and host organisms
will undoubtedly extend the regional distribution
of some taxa.
The leeches of the Great Lakes region are a
significant part of the continental freshwater
fauna. Leeches are biologically important in
food webs, and at trophic levels they function
mainly as ectoparasites, predators, or both. An
important ecological factor in the distribution of
leeches is the availability of prey. Other
environmental factors, such as composition of
substratum, lentic or lotic waters, depth, size
and type of water body, hardness and pH,
dissolved solids, water temperature, dissolved
oxygen, siltation and turbidity, and salinity are
characteristics of aquatic habitats that also
influence leech distribution and abundance
(Klemm 1972, 1991; Sawyer 1974), not
withstanding toxic substances in the aquatic
environment. We have a relatively sketchy
understanding of the role that these and other
environmental factors play in the distribution of
leeches. It must be stressed that much work
remains to be done before we can have a clear
picture of the problem.
The little we know of the feeding habits of
leeches indicates that they are far more diverse
than most people realize; many are not
sanguivorous (blood feeders). The Haemopidae
have no teeth or have varied and poorly
developed jaws that are armed with small
numbers of blunt teeth for masticating food,
and are able to swallow prey whole. They are
mainly predators of macroinvertebrates, but the
ones with teeth are perhaps capable of sucking
blood. In the Hirudinidae, the jaws are well
developed, armed with numerous small, sharp,
saw-like teeth suitable for making cuts in the
epidermis of prey, such as reptiles, amphibians,
and mammals. These leeches are blood
sucking ectoparasites. Surprisingly, studies on
the North American haemopids and hirudinids
indicate that these leeches are predominantly
predatory and extremely opportunistic, and
consume larvae and eggs of amphibians and
small invertebrates. They dwell mostly in
freshwater, but some species can travel
overland, and a few species are terrestrial. In
Erpobdellidae, the mouth is large and adapted
to predation. It contains a muscular pharynx for
crushing and swallowing macroinvertebrate
prey whole. They are highly mobile and are
good swimmers. They live exclusively in
freshwater. The Glossiphoniidae are without
teeth or jaws and have a very small oral
opening (pore). This name refers to the
mechanism by which these leeches feed. They
insert a tube-like proboscis into their prey and
suck out the body fluids. The glossiphoniids
parasitize turtles, mollusks, waterfowl, fishes,
amphibians, mammals, including man, and even
other leeches. They are ectoparasites or
predators. Most travel slowly with a looping
movement, but a few species are active
swimmers. Brooding behavior is well developed,
and cocoons are brooded over substrate or
directly on the venter of the parent. They are
found exclusively in freshwater. The
Piscicolidae are primarily ectoparasites on fish.
Some are permanent parasites on specific
hosts, but most are opportunistic and feed on
a variety of host fishes. A few feed on
invertebrate groups, such a Decapoda and
Cephalopoda. Piscicolids generally have a large
and well-developed anterior sucker surrounding
the mouth. As in the glossiphoniids, the oral
opening is small. To feed, piscicolids insert a
protruding sucking proboscis into their host.
Parent leeches lay hard-shelled cocoons on a
substrate, but they do not brood their cocoons
or young. Only a few piscicolids are active
43
-------
Klemm and Hiltunen
swimmers. Muscles are generally poorly
developed and locomotion is usually by looping
movements over a substrate or host. Most
species are marine, but some are found in
brackish or fresh waters.
Tolerance to Organic Pollution
The species of annelids in the Great Lakes
region that are commonly or occasionally
associated with organically enriched waters are
indicated in Table 2A-C. The tolerance values in
Table 2A-C can be used with the Trophic
Condition Index (Howmiller and Scott 1977;
Milbrink 1983) and modified Hilsenhoff Biotic
Index (Klemm et al. 1990; Plafkin et al. 1989).
However, more pollutional studies are needed
for these annelid groups because little is known
about their tolerances and the biological effects
of various contaminants. For more information
on the water quality requirements and pollution
tolerance of freshwater naidids, tubificids, and
leeches, see Brinkhurst (1974a,b), Carr and
Hiltunen (1965), Hiltunen (1967, 1969a-c),
Howmiller and Scott (1977), Klemm (1972,
1991), Klemm et al. (1990), Krieger (1990),
Milbrink (1983). and Sawyer (1974).
General Collection and Preservation
Aquatic worms are usually collected using
dredges, grabs, cores and other sampling
devices that provide bulk collections of bottom
subtrate. This material is then sieved or hand-
picked so that the organisms are separated
from the accompanying silt and debris. This
must be done carefully, especially if a sieve is
used. The abrasion of the soft-bodied worms
against a sieve surface may break specimens or
damage the specimens by breaking or
displacing chaetae, particularly capillif orm (hair)
chaetae, for example. Although a US Standard
No, 30 mesh sieve (28 meshes per inch, 0.595
mm openings) is usually used, it should be
noted that many small individuals may be lost
during the sieving process and that the use of
a finer sieve (for example, No. 60 mesh, 0.25
mm opening) or no sieving at all may be
required to ensure collection of all individuals.
Even when sieving has been accomplished care-
Table 2A. Pollution Tolerance of Selected
Freshwater Annelids
Taxa
Tolerance to Organic Wastes*
T F I
ANNELIDA - POLYCHAETA
SABELLIDAE
Manavunkia soeciosa
ANNELIDA - OLIGOCHAETA
NAIDIDAE
Amohichaeta americana
Chaetogaster diaohanus
C. diastroohus
Dero dioitata
D. nivea
D. obtusa
D. pectinata
Naj§ barbata
N. behninqi
N. bretscheri
N. communis
N. elinauis
N. oardalis
N. simplex
N. variabilis
Qphidonais serpentina
Pristina aeauiseta
Slaving aopendiculata
Specaria josinae
Stvlaria fossularis
S. lacustris
Veidovskvella comata
V. intermedia
2
2
2
2
3
3
2
3
3
4
5
4
5
4
3
2
2
3
3
* Ranking from 0 to 5 with 0 being the least
tolerant. T = tolerant; F = facultative; I =
intolerant
fully, some individuals will nevertheless
fragment. Only head-end sections and whole
worms should be enumerated. The initial sorting
of specimens from sediment residue in the
laboratory should be done at a 5-10 X
magnification using a dissection microscope or
44
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Great Lakes Region Oligochaeta, Polychaeta, and Hirudinea
Table 2B. Pollution Tolerance of Selected
Freshwater Annelids.
Tolerance to Organic Wastes*
Taxa T F I
ANNELIDA • OLIGOCHAETA
TUBIFICIDAE
Aulodrilus americanus
A. limnobius
A. DJoueti
4, pluriseta
Bothrioneurum veidovskvanum
Branchiura sowerbvi 4
llvodrilus tenrmletoni
Isochaetides curvisetosus
Limnodrilus cervix
L. claoaredianus
L. hoffmeisteri
L. maumeensis
L. udekemianus
Potamothrix moldaviensis
P. vejdovskvi
Quistadrilus multisetosus
Spirosperma carolinensis
S. ferox
S. nikolskvi
Tubifex tubifex
3
3
3
3
2
3
2
4
4
5
5
5
3
3
3
3
2
'Ranking from 0 to 5 with 0 being the least
tolerant. T = tolerant; F = facultative; I =
Intolerant.
lens. They can also be selectively hand-picked,
fixed, and preserved in the field.
Leeches are also found attached to various
substrates such as rocks, boards, logs, or
almost any inanimate object littering both lentic
and lotic environments or collected from prey
organisms. Annelid specimens should be fixed
in 5010% formalin, and transferred after 48
hours to 70% ethanol or 5-10% buffered
formalin for storage. Undesirable shrinkage is
kept to a minimum with this process. The use
of alcohol as a fixative should be avoided
Table 2C. Pollution Tolerance of Selected
Freshwater Annelids
Tolerance to Organic Wastes*
Taxa T F I
ANNELIDA - HIRUDINEA
ERPOBDELLIDAE
Eroobdella parva 4
E. ounctata 4
Mooreobdella microstoma 4
HAEMOPIDAE
Haemoois grandis 3
H. marmorata 3
GLOSSIPHONIIDAE
Alboolossiphonia heteroclita 3
Gloiobdella elonoata 4
Helobdella stagnalis 4
H. triserialis 3
Glossiohonia complanata 4
Placobdella multilineata 2
P. ornata 3
P. papillifera 3
P. parasitica 3
PISCICOLIDAE
MvzobdeHa luoubris 3
Piscicola punctata 3
* Ranking from 0 to 5 with 0 being the least
tolerant. T = tolerant; F = facultative.
because polychaetes, oligochaetes, and leeches
initially preserved in alcohol without first being
fixed in formalin tend to deteriorate and
disintegrate. If the specimens of oligochaetes
are to be cleared and they have been preserved
in 70% alcohol, they should be placed in 30%
alcohol and then in water for a short time to
leach out the alcohol to enable placement into
a tissue-clearing solution (e.g., Amman's
lactophenol). Alcohol retards the clearing
process of Amman's lactophenol.
45
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Klemm and Hiltunen
To allow internal structures to be seen,
oligochaete specimens should be cleared before
specific examination. Temporary mounting
media, Amman's lactophenol (lOOg phenol, 100
mL lactic acid, 200 mL glycerine, and 100 mL
water) or CMCP-9, or CMCP-10, can be used
for rapid processing of specimens. Oligochaete
specimens must be cleared and mounted on
glass slides for examination under a compound
light microscope capable of magnification up to
1000X (oil immersion). An 18 mm diameter,
No. 0 or 1 round cover glass is appropriate
because it will adequately accommodate nearly
the size range of naidids and tubificids and the
shape allows for maneuvering the specimens
into the most desired position by gentle
pressure and rotation of the coverglass. When
preparing a temporary or permanent slide
mount, an attempt should be made to place the
specimen on its side, revealing both dorsal and
ventral fascicles of chaetae. Permanent mounts
of oligochaetes can be made following alcohol
dehydration of specimens and clearing, using
methyl salicylate or xylene, and mounting the
specimens in a synthetic resin, such as
Harleco's Coverbond or Canada balsam.
Permanent mounts of oligochaetes are suitable
for systematic study and may last over 20
years. Most leech specimens can be identified
to species by examining the external features
using a dissecting microscope (450X).
Additional instructions for sorting, processing,
and identifying polychaetes, naidid and tubificid
oligochaetes, and leeches specimens can be
found in a number of taxonomic guides
(Brinkhurst 1986; Hiltunen and Klemm 1980;
Stimpson, Klemm, and Hiltunen 1982; Klemm
1985a-c; Klemm at al. 1990).
Literature Cited
Aston, R. J. 1984. Tubificids and water
quality. A review. Environmental Pollution
(5)1-10.
Bonomi, G. and C. Erseus (eds.). 1984.
Aquatic Oligochaeta. Hydrobiologia 115:1-
240.
Brinkhurst, R.O. 1974a. The Benthos of lakes.
St. Martin's Press, New York. 190 pp.
Brinkhurst, R.O. 1974b. Aquatic earthworms
(Annelida: Oligochaeta). ]n: C.W. Hart, Jr. and
S.L.H. Fuler (eds.). Pollution ecology of
freshwater Invertebrates. Academic Press,
Inc., New York, pp. 3-156.
Brinkhurst, R.O. 1986. Guide to the
freshwater aquatic microdrile oligochaetes of
North America. Canadian special publication
of fisheries and aquatic sciences 84,
Canadian Government Publishing Centre,
Supply and Services Canada, Ottawa,
Ontario, Canada K1A OS9. 259 pp.
Brinkhurst, R.O. 1989. Varichaetadrilus
auoustipenis (Brinkhurst and Cook 1966),
new combination for Limnodrilus auaustipenis
(Oligochaeta; Tubificidae). Proc. Biol. Soc.
Wash. 102(2):311-312.
Brinkhurst, R.O. and D.G. Cook (eds.). 1980.
Aquatic oligochaete biology. Plenum, New
York. 529 pp.
Brinkhurst, R.O. and R.J. Diaz (eds.). 1987.
Aquatic Oligochaeta. Hydrobiol. 155:1-323.
Brinkhurst, R.O. and B.G.M. Jamieson. 1971.
Aquatic Oligochaeta of the world. Toronto:
Univ. Toronto Press. 860 pp.
Brinkhurst, R.O. and R.D. Kathman. 1983. A
contribution to the taxonomy of the Naididae
(Oligochaeta) of North America. Can. J. Zool.
61:2307-2312.
Carr, J.F. and J.K. Hiltunen. 1965. Changes
in the bottom fauna of western Lake Erie from
1930 to 1961. Limnol. Oceanogr. 16:551-
569.
Goodnight, C.J. and L.S. Whitley. 1960.
Oligochaetes as indicators of pollution. Proc.
15th Ind. Waste Conf., Purdue Univ., Indiana,
pp. 139-142.
46
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Great Lakes Region Oligochaeta, Polychaeta, and Hirudinea
Hiltunen, J.K. 1965. Distribution and
abundance of the polychaete, Manavunkia
soeciosa Leidy, in western Lake Erie. Ohio J.
Sci. 65:183-185.
Hiltunen, J.K. 1967. Some oligochaetes from
Lake Michigan. Trans. Am. Microsc. Soc.
86(4):433-454.
Hiltunen, J.K. 1969a. Distribution of
oligochaetes in western Lake Erie. 196I.
Limnol. Oceanogr. 14(21:260-264.
Hiltunen, J.K. 1969b. Invertebrate
macrobenthos of western Lake Superior.
Mich. Academician 1(3-4):123-133.
Hiltunen, J.K. 1969c. The benthic
macrofauna of Lake Ontario. In: Limnological
Survey of Lake Ontario. 1964, pages 39-50.
Great Lakes Fish. Comm. Tech. Rep., No. 14.
Hiltunen, J.K. 1971. Limnological data from
Lake St.Clair, I963 and 1965. Dept.
Commer., NOAA/NMFS, Data Rept. No.
54(CON-71-00644). 54 pp.
Hiltunen, J.K. and D.J. Klemm. 1980. A guide
to the Naididae (Annelida: Clitellata:
Oligochaeta) of North America. U.S.
Environmental Protection Agency, Office of
Research and Development, Environmental
Monitoring and Support Laboratory,
Cincinnati, Ohio 45268. EPA-600/4-80-031.
Hiltunen, J.K. and D.J. Klemm. 1985.
Freshwater Naididae (Annelida: Oligochaeta).
In: D.J. Klemm (ed.). A guide to the
freshwater annelida (Polychaeta, Naidid and
Tubificid Oligochaeta, and Hirudinea) of North
America. Kendall/Hunt Publ. Co. Dubuque,
Iowa. pp. 24-43.
Hiltunen, J.K. and B.A. Manny. 1982.
Distribution and abundance of
macrozoobenthos in the Detroit River and
Lake St. Clair, 1977. Great Lakes Fishery
Laboratory Administrative Report No. 82-2,
U.S. Fish & Wildlife Service, Ann Arbor,
Michigan, pp. 87.
Holmquist, C. 1973. Fresh-water polychaete
worms of Alaska with notes on the anatomy
of Manavunkia speciosa Leidy. Zool. Zb. Syst.
Bd. 100,5.497-516.
Howmiller, R.P. and M.A. Scott. 1977. An
environmental index based on relative
abundance of oligochaete species. JWPCF
49(5):809-815.
Klemm, D.J. 1972. The leeches (Annelida:
Hirudinea) of Michigan. Mich. Academician
4(4):405-444.
Klemm, D.J. 1977. A review of the leeches
(Annelida: Hirudinea) in the Great Lakes
region. Mich. Academician 9(4):397-418.
Klemm, D.J. 1985a. Freshwater Polychaeta.
]n: D.J. Klemm (ed.). A guide to the
freshwater Annelida (Polychaeta, Naidid and
Tubificid Oligochaeta, and Hirudinea) of North
America. Kendall/Hunt Publ. Co., Dubuque,
Iowa. pp. 14-23.
Klemm, D.J. 1985b. Freshwater leeches. ]n:
D.J. Klemm (ed.). A guide to the freshwater
Annelida (Polychaeta, Naidid and Tubificids
Oligochaeta, and Hirudinea) of North America.
Kendall/Hunt Publ. Co., Dubuque, Iowa. pp.
70-173.
Klemm, D.J. (ed.). 1985c. A guide to the
freshwater Annelida (Polychaeta, Naidid and
tubificid Oligochaeta, and Hirudinea) of North
America. Kendall/Hunt Publ. Co., Dubuque,
Iowa. 198 pp.
Klemm, D.J. 1991. Taxonomy and pollution
ecology of the Great Lakes region leeches
(Annelida: Hirudinea). Mich. Academician,
24:37-103.
Klemm, D.J., P.A. Lewis, F. Fulk, and J.M.
Lazorchak. 1990. Macroinvertebrate field and
47
-------
Klemm and Hiltunen
laboratory methods for evaluating the
biological integrity of surface waters.
Environmental Monitoring Systems
Laboratory, U.S. Environmental Protection
Agency, Cincinnati, Ohio 45268.
Krieger, K.A. 1990. Changes in the benthic
macrotnvertebrate community of the
Cleveland Harbor area of Lake Erie from 1978
to 1989. Ohio EPA, Division of Water Quality
Planning and Assessment, Columbus, Ohio
43212.
Leaner, M.A. 1979. The distribution and
ecology of the Naididae (Oligochaeta) which
inhabit the filter-beds of sewage-works in
Britain. Water Res. 13:1291-1299.
Leaner, M.A., G. Lochhead, and B.D. Hughes.
1978. A review of the biology of British
Naididae (Oligochaeta) with emphasis on the
lotic environment. Freshwater Biol. 8:357-
375.
Mackie, G.L. and S.U. Qadri. 1971. A
polychaete, Manavunkia soecuisa. from the
Ottawa River, and its North American
distribution. Can. J. Zoology 49:780-782.
Milbrink, G. 1983. An improved
environmental index based on the relative
abundance of oligochaete species.
Hydrobiologia 102:89-97.
Plafkin, J.L., M.T. Barbour, K.D. Porter, S.K.
Gross, and R.M. Hughs. 1989. Rapid
bioassessment protocols for use in streams
and rivers: benthic macroinvertebrates and
fish. EPA/440/4-89-001 (printed erroneously
as EPA/444/4-89-001). Assessment and
Watershed Protection Division, USEPA,
Washington, D.C. 20460
Poe, T.P. and D.C. Stefan 1974. Several
environmental factors influencing the
distribution of the fresh-water polychaete,
Manavunkia soeciosa Leidy. Chesapeake Sci.
15:235-237.
Resh, V.H. and J.D. Unzicker. 1975. Water
Quality monitoring and aquatic organisms: the
importance of species identification. J. Wat.
Pollut. Control Fed. 47:9-19.
Sawyer, R.T. 1972. North American
freshwater leeches, exclusive of the
Piscicolidae with a key to all species.
Monogr. 46(1 ):1-154.
I. Biol.
Sawyer, R.T. 1974. Leeches (Annelida:
Hirudinea). in: C.W. Hart, Jr. and S.L.H. Fuller
(eds.). Pollution ecology of freshwater
invertebrates. Academic Press, Inc., New
York. pp. 81-142.
Sawyer, R.T. 1986. Leech biology and
behavior. Volume II: Feeding biology, ecology,
and systematics. Clarendon Press, Oxford.
pp. 418-793.
Spencer, D.G. 1976. Occurrence of
Manavunkia soeciosa (Polychaeta: Sabellidae)
in Cayuga Lake, New York, with additional
notes on its North American distribution.
Trans. Am. Microsc. Soc. 95(1 ):127-128.
Stimpson, K.S., D.J. Klemm, and J.K.
Hiltunen. 1982. A guide to the freshwater
tubificidae (Annelida: Clitellata: Oligochaeta)
of North America. U.S. EPA, Environmental
Monitoring and Support Laboratory,
Cincinnati, Ohio 45268, EPA-600/3-82-033.
61 pp.
Stimpson, K.S., D.J. Klemm, and J.K.
Hiltunen. 1985. Freshwater Tubificidae
(Annelida: Oligochaeta). ]n: D.J. Klemm (ed.).
A guide to the freshwater Annelida
(Polychaeta, Naidid and Tubificid Oligochaeta,
and Hirudinea) of North America. Kendall/
Hunt Publ. Co., Dubuque, Iowa. pp. 44-69.
48
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A Comparison of Macroinvertebrates Collected from Bottom Sediments
in Three Lake Erie Estuaries
Philip A. Lewis
Bioassessment and Ecotoxicology Branch
Environmental Monitoring Systems Laboratory
U.S. Environmental Protection Agency
3411 Church Street, Cincinnati, Ohio 45244
Mark E. Smith
Technology Applications Inc.,
C\0 U.S. Environmental Protection Agency,
3411 Church Street, Cincinnati, Ohio 45244
Abstract
Macroinvertebrates were collected from bottom sediments from three Lake Erie tributaries as a part
of EMSL's Biomarker Project. The objective of this paper is to compare the macroinvertebrate
populations collected from the three water bodies and relate these populations to possible pollutional
stresses and/or habitat characteristics. Three grab samples were collected with either a petite Ponar
or a standard Ekman-on-a-stick at three different stations at each site. The sampling stations were
chosen randomly from among the nine stations used for collecting fish at each site. In the Black River
above a Coking Plant, 60-80% of the organisms were tolerant oligochaete worms but some pollution
sensitive organisms were also present indicating organic enrichment but not toxic pollution. All of
the individuals collected from below the plant were oligochaete worms (90%) and other organisms
tolerant of both organic and toxic pollution. In Old Woman Creek, over 80% of the individuals
collected were oligochaete worms and blood worms (midges) characteristic of organically enriched
sediments associated with high oxygen demand. Toussaint Creek samples were characterized by a
variety of midge larvae and many empty mollusk shells but few live mollusks. Less than 50% of the
individuals were oligochaete worms. This may be a reflection of the sediment characteristics which
consisted of gravel and clay with little of the muck substrate characteristic of the other two sites.
The data indicate that all three sites are effected by organic enrichment and/or agricultural runoff,
but the Black River macroinvertebrate community below the Coking Plant appears to be stressed by
something in addition to organic enrichment.
Key Words: macroinvertebrates, bottom sediments, pollution, water quality, organic enrichment,
biotic index.
Introduction
As part of EMSL-Cincinnati's Biomarkers
Research Program, fish and macroinvertebrates
were collected from the Black River, Old
Woman Creek and Toussaint Creek during April
23-25, 1990. The project objective is to
determine if biomarkers can be used to detect
a wide range of pollutants. The Black River site
was chosen as one study site because of
polyaromatic hydrocarbons that have been
detected in the sediment downstream from a
coking plant and the resulting high incidence of
liver cancer in brown bullhead that inhabit the
site. Old Woman Creek and Toussaint Creek
were sampled as possible control sites.
The purpose of this paper is to compare the
macroinvertebrate populations collected from
the three water bodies and attempt to relate
these populations to pollutional stresses and/or
habitat characteristics.
Methods
Nine fyke fish nets were set at each site along
both east and west banks over a distance of
approximately one mile. Each net was defined
49
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Lewis and Smith
as a sampling station for fish. Three of these
stations were randomly chosen as benthic
macroinvertebrate collection stations on each
water body. Three six inch Ekman or Ponar grab
samples were collected at each station and
preserved in 70% ethanol. The samples were
returned to the laboratory, sorted and the
organisms identified to the lowest taxonomic
level possible. Pollution tolerance value for each
taxa were taken from EMSL-Cincinnati's
Biological Methods Manual (USEPA 1990) or
Hilsenhoff (1987). Sediment and water samples
were also collected at each station for chemical
analysis.
Stations were analyzed using Hilsenhoff's
(1977) modification (HBI) of Chutter's (1972)
Imperial Biotic Index, Shannon-Weaver's mean
diversity (d), and equitability (e). Although the
HBI was designed to give a measure of the
effects of organic pollution on the macro-
invertebrate communities inhabiting stream
riffles, it should be useful in analyzing grab
samples collected from soft river substrates if
care is used in interpreting the data. An HBI
score of < 1.75 would indicate excellent water
quality, 1.76 - 2.50 good water quality, 2.51 -
3.75 fair water quality, 3.76 - 4.00 poor water
quality, and >4.00 would indicate grossly
polluted conditions. Mean diversity values of
less than 1.0 are characteristic of gross
pollution, values between 1.0 and 3.0 indicate
fair to poor water quality, and values above 3.0
are common for clean water stations.
Equitability values above 0.5 are indicative of
good water quality, 0.3 • 0.5 fair, and values
below 0.3 indicate degradation of water quality.
Stations within each site and the three water
bodies (between sites) were compared using
the Community Loss Index and Jaccard's
Coefficient of Similarity (Plafkin et a). 1989).
Trophic condition index (Howmiller and Scott
1977; Milbrink 1983) was also determined for
each station but the results were not helpful in
interpreting the data, possibly because most of
the oligochaete worms could not be identified
to species.
Station descriptions
Old Woman Creek - Station 1 was located 200
feet north of the RR bridge about 25 feet from
the west shore. Substrate was muck and well
rotted organic material; water depth was about
18 inches. Station 2 was located 100 feet west
of the observation deck about 50 feet from
shore. Substrate was muck with some clay;
water depth was about 18 inches. Station 3
was located 500 feet south of the Highway 2
bridge and about 150 feet west of a high gravel
bank near the east shore of the estuary.
Substrate was muck and rotting leaves; water
depth was about 16 inches.
Toussaint Creek - Station 1 was located 200
feet north of Highway 2 bridge about 50 feet
from the east shore. Water depth was about six
feet and the substrate was sandy and silty.
Station 2 was located about one half mile south
of Highway 2 bridge about 50 feet from the
east shore just south of a small island. Water
depth was about two feet and substrate was
hard packed gravel with some clay and silt.
Station 3 was located across the bay from
Station 2 about 50 feet east of a very small
island. Water depth was about three feet and
substrate was muck, clay and hard packed
gravel. A large ditch entered the bay about 50
feet north of the station.
Black River - Station 1 was located about four
miles upstream from the mouth of the river
about 30 feet from the west bank across from,
and about 100 feet upstream from, the upper
end of a large island. Water depth was about
seven feet and the sediment was mostly fine
silt. Station 2 was located about 100 feet
downstream from the lower end of the large
island about 20 feet from the bank. Water
depth was about two feet and the sediment
was muck, clay and leaves. Station 3 was
about one half mile downstream from a major
discharge from the coking plant at the bend of
the river about 20 feet from the west bank at
about river mile three. Water depth was 21/2
feet and the sediment was mud and silt:
50
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Comparison of Macroinvertebratesin Lake Erie Estuaries
Results and Discussion
Toussaint Creek. The Ponar samples collected
at Station 1 (Table 1) yielded 13 taxa, two of
which (see below) are not generally found in
polluted waters and six of which are highly
tolerant of organic pollution. Just over half
(56%) of the individuals were oligochaete
worms which is an indication of non-polluted
conditions (Goodnight and Whitley 1960). The
HBI of 3.76, the diversity index (2.3) and
equitability (0.5) all indicate fair to marginal
water quality. The large number of empty
mollusk shells (123 individuals representing 8
species) indicate that conditions have been
present, at least some time recently, for a
diverse population of these organisms to
develop. The midge larvae Ablabesmvia
mallochi. which is very sensitive to metals
contamination, and the gastropod
Somatoovrus. which is not generally found in
organically polluted waters indicate that this
station is probably not, or only slightly,
impaired by pollutants. Probably this is the least
impacted station sampled during this study;
therefore it was used as the reference station
for this study.
The three Ekman samples collected at Station 2
(Table 1) contained only five living taxa,
consisting of the midge Polvpedilum scalaenum.
which is generally restricted to unimpaired
waters, two pollution tolerant oligochaete
worms, and two facultative (wide range of
tolerance) midge species. Less than half the
individuals (36%) collected were oligochaete
worms which indicates non impaired conditions
(Goodnight and Whitley, 1960). The HBI of
3.33, the diversity index (1.8) and equitability
(1.0) all indicate fair to good water quality
conditions at this station. The presence of the
midge Polvpedilum scalaenum also indicates
that toxic substances are probably not present
in concentrations high enough to effect the
biological integrity of the aquatic community.
The lack of a diverse fauna is probably due to
the difficulty of obtaining a sample because of
the hard packed gravel and clay substrate. The
Ekman sampler used was one with a handle so
that it could be pushed with some force into
the bottom or the samples could not have been
taken here.
The three Ponar samples taken at Station 3
(Table 1) contained only seven living taxa but
14 taxa of empty mollusk shells. It seems odd
that so many shells were collected without any
live animals. Perhaps something had recently
occurred that killed all the mollusks, but it is
obvious that conditions in the recent past must
have been conducive to the establishment of a
diverse population. Only two taxa present (the
midge Cryptochironomus sp. and the
oligochaete worm Limnodrilus hoffmeisteri) are
tolerant of pollution, however, these two taxa
make up about 80% of the individuals
collected. This station was very near a ditch
that enters the bay here and it is possible that
periodically toxic and/or organic substances
may flow into the bay from this ditch causing
stress on the benthic community. Our sampling
may have occurred during the beginning of the
recovery period. The samples were collected in
the spring soon after planting and storm runoff
from nearby agricultural lands may have entered
the bay by way of this ditch. Krieger (1989)
reported that agricultural herbicides used with
corn and soybeans reach higher concentrations
in rivers of northwest Ohio than in rivers
anywhere else in North America. The Biotic
Index of 3.73 indicates only a slight impact as
do the diversity index (1.5) and equitability
(0.5). The high percentage of oligochaete
worms at this station as compared to the other
two may be an indication that whatever has
effected the benthic community was not
limiting to the oligochaete worms and probably
was not widespread throughout the bay.
Community Loss Index and Jaccard's
Coefficient of Similarity indicate that the three
stations are quite dissimilar with the greatest
difference between stations 1 and 2. If station
1 is considered the control, station 2 shows a
loss of 2.2 and station 3 shows a loss of 1.4.
These differences would appear to be signifi-
cant and one might expect that stations 2 and
51
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Lewis and Smith
Table 1. Macroinvertebrates Collected and Pollution Tolerance Values for Toussaint Creek. Intolerant
taxa denoted by *.
Number of Individuals
Taxa
Chironomidae
Coelotanypus concinnus
Cryptochironomus sp.
Cryptochironomus fulvus gr.
Procladius nr. bellus
Chironomus plumosus gr.
Ablabesmyia mallochi*
Cladotanytarsus sp.
Polypedilum scalaenum*
Tanytarsus guerlus gr.
Other Diptera
Ceratopogonidae
Nr. Probezzia sp.
Crustacea
Isopoda
Lirceus lineatus
Bryozoa
Urnatella gracilis
Bivalvia
Sphaeriidae
Gastropoda
Somatogyrus sp."
Oligochaeta
Limnodrilus hoffmeisteri
L. maumeensis
Branchiura sowerbyi
Unidentified Oligochaeta
Total Individuals
Total Taxa
% Oligochaeta
Biotic Index (HB[)
Mean Diversity (d)
Equitability (e)
Station 1
11
4
30
1
1
2
2
P
3
1
7
1
1
62
126
13
56
3.76
2.3
0.5
Station 2
5
15
1
2
10
33
5
36
3.33
1.8
1.0
Pollution Tolerance
Station 3
1
2
2
2
1
1
1
31
41
7
78
3.73
1.5
0.5
Value
4
4
3
3
5
2
3
2
3
3
3
3
4
2
5
5
4
4
3 would be quite similar because the
Community Loss Index between them is small,
but Jaccard's Coefficient (<0.50) indicates
otherwise. HBI scores for stations 1 and 2 and
for 2 and 3 are statistically different from each
other but HBI scores for stations 1 and 3 are
similar. Probably most of the differences
observed are ecological and not caused by
pollution.
52
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Comparison of Macroinvertebratesin Lake Erie Estuaries
The data suggest that Toussaint Creek Bay is
not affected to any great extent by pollution,
however, the presence of many species of
empty mollusk shells at stations 1 and 3 leads
me to suspect that occasional instances may
occur, either natural or man caused, that stress
the aquatic community and temporarily affect
the biological integrity of the bay in the vicinity
of the canal that enters the bay on the west
shore. Runoff from agricultural lands during
storms may be a factor here as in most other
northwest Ohio rivers (Krieger 1989). A total of
17 taxa consisting of 200 individuals
(11.8/taxa) were collected in the nine grab
samples taken from Toussaint Creek.
Old Woman Creek. The three Ekman samples
collected at Station 1 (Table 2) yielded nine
taxa, including the bryozoan Pectinatella
maanifica which is known to be sensitive to
toxic contaminants and five taxa which are
highly tolerant to organic pollution. Of the 96
individuals collected 83 (86%) wereoligochaete
worms characteristic of organically enriched
sediments. The HBI of 4.01 and the high
percentage of oligochaete worms (Goodnight
and Whitley 1960) indicate organic enrichment
but the diversity index (2.3) and equitability
(0.7) indicate fair to good water quality. The
presence of the bryozoan Pectinatella maanifica
at this station would indicate good water
quality.
The three Ekman samples collected at Station 2
(Table 2) yielded ten taxa, none of which are
known to be sensitive to pollution and seven
which are highly tolerant of organic pollution.
Of the 132 individuals collected, 107 (81%)
were oligochaete worms characteristic of
organically enriched sediment. The blood worm
Chironomus Dlumosus. which is highly tolerant
of sediments with high oxygen demand, was
also common at this station. The HBI of 4.03
and the high percentage of oligochaete worms
indicate organic enrichment but the diversity
index (1.6) and equitability (0.4) would indicate
fair water quality.
The three Ekman samples collected at Station 3
(Table 2) contained seven taxa, including the
amphipod Gammarus pseudolimnaeus which is
generally not found in waters containing toxic
substances other than organic enrichment and
four of which are highly tolerant to organic
pollution. Of the 31 individuals collected, 23
(74%) were oligochaete worms characteristic
of organically enriched sediments. The HBI of
3.84 and the presence of Gammarus
pseudolimnaeus would indicate that toxic
substances are probably not major limiting
factors at this station. The percentage of
oligochaete worms was less at this station than
the other two and fell in the moderately
polluted category of Goodnight and Whitley
(1960). The diversity index (1.8) and
equitability (0.6) indicate fair to good water
quality.
Community Loss Index and Jaccard's
Coefficient of Similarity show that stations 1
and 2 are quite similar and that station 3 is
significantly different from the other two
stations. The Community Loss Index values also
indicate that station 3 has fewer taxa in
common with either of the other two stations.
HBI score for station 3 was also lower than for
the other two stations, but the difference was
not significant (P = 0.051. However, HBI for Old
Woman Creek stations 1 and 2 were
significantly higher than for the reference
station while station 3 was not significantly
different (P = 0.05).
The data suggest that Old Woman Creek
Estuary may have been temporarily affected by
storm runoff from agricultural lands (Klarer and
Millie, 1989) and organic enrichment from
decaying vegetation and that station 3 is
slightly less effected than the other two. The
limiting factor at this site would most likely be
the highly enriched substrate consisting of
muck and decaying vegetation probably
accompanied by periods of low DO and high
levels of nitrites resulting from storm runoff
from nearby agricultural lands (Klarer and Millie,
1989). Using Goodnight and Whitley's (1960)
53
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Lewis and Smith
Table 2. Macroinvertebrates Collected and Pollution Tolerance Values for Old Woman
Number of Individuals
Taxa Station 1 Station 2
Chironomidae
Coelotanypus concinnus
Cryptochironomus fulvus
Procladius nr bellus
Chironomus plumosus gr
Oicrotendipes sp.
Other Diptera
Ceratopogonidae
Nr Probezzia sp.
Coleoptera
Dubiraphia sp
Coieoptera
Donacia sp.
Cruatacea
4
gr 1
4
1
1
2
1
6
1
2
9
7
Gammarus pseudolimnaeus*
Oligochaeta
Limnodrilus maumeensis
L. hoffmeisteri
L. cervix
Branchiura sowerbyi
llyodrilus templetoni
Unidentified Oligochaeta
Bryozoa
Pectinatella magnifies*
Total Individuals
Total Taxa
% Oligochaeta
Biotic Index (HBi)
Mean Diversity (d)
Equitability (e)
metric based on percent
present, all three stations
2
5
76
P
96
9
86
4.01
2.3
0.7
oligochaete worms
would be considered
polluted. A total of 15 taxa consisting of 259
2
3
2
2
1
97
S
132
10
81
4.03
1.6
0.4
Creek.
Pollution Tolerance
Station 3 Value
5
1
1
1
2
1
19
S
31
7
74
3.84
1.8
0.6
or organically polluted waters.
species,
Harnischia
4
3
3
5
4
3
3
3
2
5
5
4
4
4
4
1
Two midge
Dicrotendioes neomodestus and
curtilamellata. present
at this station
individuals (17.3/taxa) were collected in the
nine grabs taken from Old Woman Creek.
Black River. The three Ponar samples collected
at Station 1 (Table 3) contained 18 taxa most
of which are characteristic of slightly impaired
are usually not found under contaminated
conditions. Of the 137 individuals collected 106
(77%) were oligochaete worms which would
indicate moderately polluted or organically
enriched sediments. The HBI of 3.92, the
diversity index (2.4) and equitability (0.4) all
54
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Comparison of Macroinvertebrates in Lake Erie Estuaries
Table 3. Macroinvertebrates Collected and
Pollution Tolerance Values for Black River.
Number of Individuals
Taxa Station 1 Station 2
Chironomidae
Harnischia curtilamellata* 1
Ablabesmyia mallochi*
Tanytarsus guerlus gr
Phaenopsectra prob dyari
Eukiefferiella claripennis
Hydrobaenus pilipes gr*
Cryptochironomus f ulvus gr 7
Cryptochironomus sp 1
Diplocladius cultriger 1
Dicrotendipes neomodestus 1
Polypedilum ophioides 1
Glyptotendipes lobiferus 1
Orthocladius sp.
Cricotopus tremulus gr
Corichapelopia sp.
Procladius nr. bellus
Nanocladius distinctus
Other Diptera
Hemerodromia sp 2
Chaoborus punctipennis 1 0
Nr. Probezzia sp. 3
Ephydridae 1
Unidentified Diptera
Coleoptera
Elmidae
Stenelmis sp. *
Dubiraphia sp.
Staphylinidae
Helodidae
Scirtes sp
Noteridae
Berosus sp
Ephemeroptera
Caenis sp 2
Lepidoptera
Bactra sp (?)
Hydracarina
Trombidiformes
Odonata
Argia apicalis
3
2
2
2
1
1
1
2
2
2
1
1
3
1
1
21
4
1
1
1
2
1
1
1
Pollution Tolerance
Station 3 Value
2
2
3
3
4
2
11 3
1 4
4
2
3
1 3
1 3
3
1 3
9 3
3
3
1 3
3
3
3
2
3
2
3
3
3
2
2
3
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Lewis and Smith
Table 3. Macroinvertebrates Collected
(Continued)
Taxa
Crustacea
Isopoda
Asellus communis
Bivalvia
Pisidium casertanum
Gastropoda
Physella vinosa
Bryozoa
Lophopodella carted*
Oligochaeta
Limnodrilus hoffmeisteri
L. udekemianus
L. cervix
Potamothrix vejdovskyi*
Tubifex tubifex
llyodrilus templetoni
Quistadrilus multisetosus
Oero sp.
Nais elinguis
Nais communis
Enchytraeidae
Unidentified Oligochaeta
Total Individuals
Total Taxa
% Oligochaeta
Biotic Index
Mean Diversity (d)
Equitability (e)
Station 1
17
7
1
1
1
1
78
137
18
77
3.92
2.4
0.4
and Pollution Tolerance
Number of Individuals
Station 2
6
1
P
23
5
6
2
2
2
3
1
55
164
35
60
3.62
3.7
0.5
Values for Black River.
Station 3
1
1
24
1
4
1
172
229
13
88
4.00
1.5
0.3
Pollution
Tolerance
Value
3
4
4
1
5
5
4
3
5
4
4
5
4
4
4
4
* Intolerant taxa
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Comparison of Macroin vertebrates in Lake Erie Estuaries
indicate fair to poor water quality at this station
which is located upstream from the major
effluent from the coking plant. The mayfly
Caenis sp., present at this station, is the most
tolerant of the mayflies and is not a good water
quality indicator.
The three Ponar samples collected at Station 2
(Table 3) yielded a surprising 35 taxa, many of
which would not be expected in mud substrate
where tittle current is present. Nine of the taxa
are tolerant of pollution, one (Lophopodella
carter!) is very sensitive to pollution (except in
the statoblast stage) and seven others are not
characteristically found in impaired waters.
Almost exactly 60% of the individuals are
oligochaete worms characteristic of organically
polluted conditions which would indicate some
organic enrichment, possibly due to decaying
vegetation. It is not likely that toxic pollution is
present in the sediment at this station because
Potamothrix vejdovskvi. an intolerant
oligochaete, was among the diverse worm
fauna present. The diversity of midge and other
insect groups would indicate that conditions are
conducive to the development of a balanced
benthic community. This may be due, however,
to the possibility that these organisms are
drifting into this station from some stream that
may enter the river behind the island just
upstream or from a spring entering the river
from under the stream bank. The HBI of 3.62,
diversity index (3.7) and equitability (0.5) all
indicate fair to good water quality.
The three Ponar samples collected at Station 3
(Table 3) contained 13 taxa, eight of which are
characteristic of polluted conditions. The other
five species are all facultative and could be
present in moderately polluted waters. Of the
229 individuals present, 202 (88%) were
oligochaete worms indicating grossly polluted
waters. It is interesting that the only deformed
Procladius midge (Warwick, 1989) found during
this study was collected from this station which
is located one half mile downstream of the main
effluent from the coking plant. The HBI of 4.00
indicates poor water quality as does the
equitability (0.3) while the diversity index (1.5)
is borderline between fair and poor conditions.
Community Loss Index and Jaccard's
Coefficient of Similarity show that the three
stations are considerably different from one
another. Both stations 1 and 3 show significant
community loss when compared with station 2.
The HBI scores for the three stations are similar
(P = 0.05), however the HBI scores for
stations 1 and 3 are both significantly higher
than for the control reference station.
The data suggest that the benthic community
at Station 1, located upstream from most of the
effects of the coking plant, does show some
stress on the biota. Station 2 samples contain
intolerant organisms (including riffle beetles)
which are not characteristic of muddy
substrates and some tolerant forms that are
characteristic of organic pollution. Station 3 is
noticeably degraded as compared to the other
stations sampled during this study. A total of
47 taxa consisting of 530 individuals (7.2/taxa)
were collected in the nine grab samples taken
from the Black River.
Summary and Conclusions
The Community Loss Index and Jaccard's
Coefficient of Similarity show that Old Woman
Creek and Toussaint Creek macroinvertebrate
communities are more similar to each other
than either one is to the Black River but these
similarities are not very great. There is no
significant community loss between Old
Woman Creek or Toussaint Creek and the Black
River when composite data are compared.
Based on the Community Loss Index, it would
appear that Old Woman Creek and Toussaint
Creek are both slightly less polluted than the
Black River, but this is likely due to the diverse
fauna collected from Black River station 2 as
discussed above.
Toussaint Creek station 1 is considered the
best control station because of its
representative substrate and overall quality
based on the combination of metrics used in
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Lewis and Smith
this analysis. Using this as the reference
station, Community Loss, Jaccard's Coefficient
of Similarity and t-values based on a
comparison of HBI scores for the other stations
sampled are as follows:
Community
Stations Loss Jaccard's HBI
Compared Index Coeff. t-values
Toussaint2 2.2 0.13 2.597
Toussaint 3 1.4 0.19 0.040 ns
Old Woman 1 0.3 0.44 4.590
Old Woman 2 0.6 0.47 4.531
Old Woman 3 1.6 0.13 0.241 ns
Black River! 0.6 0.11 2.531
Black River 2 0.2 0.12 0.146 ns
Black River 3 0.7 0.18 3.983
Old Woman Creek
Composite 0.4 0.37 3.073 *
Black River
Composite 0.1 0.10 1.108 ns
•Significant at p < 0.05, df = 4.
These metrics indicate that Toussaint Creek
stations 2 and 3, Old Woman Creek Stations 1,
2, and 3 and Black River stations 1 and 3 are
different from the reference station. The
differences between the Toussaint Creek
stations can be explained by substrate
differences but the others could be related to
pollution, including agricultural runoff (Krieger
1989). As might be expected Black River
station 2 did not show a community loss (see
discussion of the individual stations above).
Based on Jaccard's Coefficient all the stations
are significantly different from the reference
station and all but Old Woman Creek stations 1
and 2 are vastly different. All of the HBI scores
differ significantly (P = 0.05) from the reference
station except Toussaint Creek station 3, Old
Woman Creek station 3, Black River station 2,
and the Black River composite. The reason the
Black River composite HBI scores did not differ
from the reference station is probably because
of the high variability due to station 2 Black
River samples. As mentioned above, the
dissimilarities between the reference station
and stations 2 and 3 at Toussaint Creek and
Black River station 2 may be attributable to
environmental and/or physical factors. The
other dissimilarities could be due to organic or
toxic stresses. The Community Loss Index and
Jaccard's Coefficient scores for the composite
data indicate that Old Woman Creek is more
like the reference station than is the Black
River. The low Community Loss Index score
and the low t-value for the composite Black
River samples as compared to the reference
station are mostly due to the diverse fauna
collected at station 2 as discussed above.
Because the sediment samples that were
collected for chemical characterization have not
yet been analyzed, it is impossible to reach any
real definitive conclusions based on the
macroinvertebrate collections alone. However,
the benthic macroinvertebrate grab collections
seem to indicate that all three Old Woman
Creek stations were organically enriched, with
oligochaete worms making up over 60% of the
individuals and the remaining taxa characteristic
of waters with high oxygen demand. In
Toussaint Creek only Station 3 located near a
drainage ditch showed signs of stress. That
may be due to periodic discharge of toxic
and/or organic pollutants into the bay from this
ditch, possibly during storms. Oligochaete
worms made up 60% or more of the individuals
collected from the Black River at all three
Stations indicating organic enrichment or toxic
substances in the sediment (Krieger 1990).
However, a few sensitive taxa were found at
Stations 1 and 2 indicating reasonably good
water quality at these two stations. Station 3
samples contained about 90% worms and no
sensitive taxa indicating a stressed community
of benthic macroinvertebrates.
The complete absence of Hexagenia mayflies,
the limited number of chironomids and
gastropods, and the increase in oligochaetes in
Lake Erie bays has been correlated with
increased pollution in the lake (Edwards 1990,
Reynoldson et al. 1989). Because no Hexagenia
mayflies and only a few live gastropods were
collected at any of the sites sampled during this
58
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Comparison of Macroinvertebratesin Lake Erie Estuaries
study, it might be reasonable to assume that all
of the sites were polluted.
Neither Old Woman Creek or Toussaint Creek
appear to be good control sites. Grab samples
collected from the Grand and Chagrin Rivers
will be analyzed and the data compared with
the Black and Cuyahoga Rivers to determine if
either of them might be better control sites for
the benthic phase of this study.
Acknowledgments
We would like to thank Larry Linns, Eastern
District Office, U.S. EPA Region 5 for providing
a boat and assisting with collection of the
samples. This work was a part of Susan
Cormier and Tim Neiheisel's Biomarker
Research Project.
Literature Cited
Chutter, P.M. 1972. An empirical biotic index
of the quality of water in South Africa streams
and rivers. Water Research 6:19-30.
Edwards, C.J. 1990. Biological surrogates of
mesotrophic ecosystem health in the Laurentian
Great Lakes. Great Lakes Science Advisory
Board, Windsor, Ontario.
Hilsenhoff, W.L. 1977. Use of arthropods to
evaluate water quality of streams. Technical
Bulletin 100, Department of Natural Resources,
Madison, Wl.
Hilsenhoff, W. L. 1987. An improved biotic
index of organic stream pollution. Great Lakes
Entomologist. 20:31-39.
Howmiller, R.P. and M.A. Scott. 1977. An
environmental index based on relative
abundance of oligochaete species. Journal of
the Water Pollution Control Federation
49(51:809-815.
Goodnight, C.J. and L.S. Whitley. 1960.
Oligochaetes as indicators of pollution.
Proceedings of the 15th Industrial Waste
Conference, Purdue University Engineering Bull.
100:139-149.
Klarer, D.M. and D.F. Millie. 1989. Amelioration
of storm-water quality by a freshwater estuary.
Archives of Hydrobiology 116(3):375-389.
Krieger, K.A. 1989. Chemical limnology and
contaminants. Pages 149-175 in K.A. Krieger
editor). Lake Erie estuarine systems: Issues,
resources, status and management. Estuary of
the Month Seminar Series No. 14, NOAA
Estuarine Programs Office.
Krieger, K.A. 1990. Assessing lake quality
improvement using trends in benthic
macroinvertebrate communities: A case study
in Lake Erie. Presented at the 1990 Midwest
Pollution Control Biologists Meeting, Chicago,
Illinois.
Milbrink, G. 1983. An improved index on the
relative abundance of oligochaete species.
Hydrobiologia 102:89-97.
Plafkin, J.L.; M.T. Barbour: K.D. Porter, S.K.
Gross and R.M. Hughes. 1989. Rapid
bioassessment protocols for use in streams and
rivers: Benthic macroinvertebrates and fish.
U.S. Environmental Protection Agency, Office
of Water Regulations and Standards,
Washington, D.C. 20460. EPA/440/4-89/001.
Reynoldson, T.B.; D.W. Schloesser and B.A.
Manny. 1989. Development of a benthic
invertebrate objective for mesotrophic Great
Lakes waters. Journal of Great Lakes Research
15(4):669-686.
USEPA. 1990. Macroinvertebrate field and
laboratory methods for evaluating the biological
integrity of surface waters. Klemm, D.J., P.A.
Lewis, F. Fulk, and J.M. Lazorchak. U.S.
Environmental Protection Agency,
Environmental Monitoring System Laboratory,
Office of Research and Development,
Cincinnati, OH 45268. EPA/600/4090/030
59
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Lewis and Smith
Warwick, W.F. 1989. Morphological deformities
in larvae of Procladius Skuse (Diptera:
Chironomidae) and their biomonitoring potential.
Canadian Journal of Fisheries and Aquatic
Sciences 46{7):1255-1271,
60
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A Comparison of the Results of a Volunteer Stream Quality Monitoring
Program and the Ohio EPA's Biological Indices
Mark A. Dilley"
School of Natural Resources
The Ohio State University
2021 Coffey Road
Columbus, OH 43210
Abstract
Volunteer stream quality monitoring is increasing in popularity around the country, and
organizations involved with the administration of volunteer stream quality monitoring programs
are becoming interested in the effectiveness of their monitoring techniques. This research
compares the results of the Ohio Department of Natural Resources (ODNR) volunteer-oriented
Scenic Rivers Stream Quality Monitoring Program and the Ohio Environmental Protection
Agency's (OEPA) biological assessments. The volunteer biological monitoring ("kick-seining")
technique was performed on 12 Ohio rivers and tributaries, at 47 different sites, to coincide with
the OEPA's monitoring agenda for the summer of 1989. Comparisons were made between the
volunteer stream quality monitoring ratings and the OEPA's Index of Biotic Integrity (IB)) and
Invertebrate Community Index (ICI). Sites which were rated "excellent" using the ODNR volunteer
method tended to meet the OEPA's criteria for attainment of aquatic life uses for both the IBI and
ICI. Sites which were determined to be "fair" or "poor" with the volunteer method corresponded
to IBI and ICI scores falling in the non-attainment of aquatic life uses range. Although revisions in
the sampling and rating system for the volunteer program could improve the predictive value of
these results as compared to OEPA's indices, the volunteer technique assessments currently
appear to have merit when interpreted in terms of aquatic life use attainment or non-attainment.
Key Words: volunteer monitoring, biological indices, stream quality, kick-seining, Ohio, Scenic
Rivers.
Introduction
In 1983, the Ohio Department of Natural
Resources (ODNR) developed the Ohio Scenic
Rivers Stream Quality Monitoring Program
with assistance from the Ohio Environmental
Protection Agency (OEPA). This program uses
volunteers to conduct simple stream quality
assessments at designated monitoring
stations on the state's ten Scenic Rivers.
ODNR's stream quality monitoring technique
involves assessments based on the presence
or absence of 20 taxa of macroinvertebrates
which are divided into three categories,
according to each groups pollution tolerance
level (Fig. 1).
Group One, the pollution intolerant organisms,
includes mayfly and stonefly nymphs,
dobsonfly, caddisfly and water penny beetle
larvae, riffle beetles, and gill-breathing snails.
Group Two macroinvertebrates, with inter-
mediate pollution tolerances, include dragon-
fly and damselfly nymphs, beetle and cranefly
larvae, scuds, crayfish, sowbugs, and clams.
Group Three, the pollution tolerant organisms,
consists of aquatic worms, pouch snails,
black fly and midge larvae, and leeches.
Many of the taxa used in the program encom-
pass entire orders (i.e. mayflies - Order
Ephemeroptera, caddisflies - Order Trichop-
tera) so identification is not refined.
Current address: Metcalf & Eddy, Inc., 2800 Corporate Exchange Drive,
Suite 250, Columbus, Ohio 43231
61
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Dilley
GROUP 1 IThe&e oigonum* HAS. gznvuiUy poLtu£u>n-±itCotuian£. TheAJi. dominance
gtnvuULLu Ug>u.fc
-------
Volunteer Monitoring Program Comparison
In this Stream Quality Monitoring Program,
volunteers collect macroinvertebrates from
riffles using the "kick-seine" technique.
Riffles, with little or no vegetation and stones
up to 15 inches in diameter, are the type of
habitat best suited to this method of sampling
(Frost et al. 1971). The "kick-seine"
technique involves disturbing the substrate
upstream of the seine to dislodge the
macroinvertebrates which cling to, and hide
under the rocks and debris in the riffle. Once
freed of the substrate, the macroinvertebrates
are carried by the current into the seine. After
a sample has been collected, the seine is
taken to the stream bank where the
organisms are hand-picked from the net and
identified on site. Macroinvertebrates often
exhibit patchy distributions in streams (Rabeni
and Minshall 1977, Schwenneker and
Hellenthal 1984). Therefore, volunteers are
encouraged to take samples from a variety of
habitats until they feel that no new taxa are
represented in their samples. No set number
of samples has ever been established for the
program, however.
After all the samples have been collected,
volunteers fill out an assessment form
indicating the station sampled (according to
OEPA river miles), water conditions such as
clarity, algal bedgrowths, and odor, and the
macroinvertebrate groups found (Fig. 2). For
each macroinvertebrate taxon group located
at the station, an estimated count letter code
is entered on the assessment form. The letter
codes A, B, and C represent 1 to 9, 10 to 99,
and 100 or more individuals, respectively.
Using estimated counts allows the ODNR
staff to get an idea of population sizes while
not placing the burden of counting the
organisms on the program volunteers.
The final assessment score, referred to as the
Cumulative Index Value, or CIV, is based only
on the diversity of macroinvertebrates found
and not the quantity. In the scoring system,
each Group One taxon in the sample receives
a point value of three, each Group Two taxon.
a point value of two, and each Group Three
taxon, a point value of one. The CIV is the
sum of the points given to each category. The
final step in the stream quality assessment is
determining a qualitative rating for the station
based on the CIV. Cumulative Index Values of
over 22 are given an "excellent" rating.
Scores between 17 and 22 are rated "good."
"Fair" is 11-16, and a "poor" rating is given
to scores of less than 11.
Once completed, assessments are sent to
ODNR where they are entered into a
computer database. The database allows for a
quick review of the history of a given
monitoring station to determine if the site has
experienced any significant impacts over the
years it has been monitored. It is believed that
this may allow for early detection of
degradation on the Scenic Rivers. The Ohio
Department of Natural Resources stresses
that the procedure "is not intended to
pinpoint subtle changes in water quality, but
rather the general condition of the river," and
that "information which indicates potential
decreases in water quality will be coordinated
with the Ohio Environmental Protection
Agency." (ODNR n.d.). The program is not
intended to completely assess the source or
degree of degradation, but rather provide an
inexpensive and enjoyable way for the public
to flag the attention of responsible
enforcement agencies in the event that
further study may be warranted.
The simplicity and accessibility of the
program has made it popular among schools,
conservation groups, scouts, and families.
Since its beginning, the Stream Quality
Monitoring Program has grown quite rapidly.
During the 1990 monitoring season,
approximately 3,000 volunteers monitored
Ohio's Scenic Rivers. In addition to this
considerable volunteer force, many Soil and
Water Conservation Districts in Ohio have
expressed interest in ODNR's method to
develop volunteer stream quality monitoring
programs for streams within their counties
63
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Dilley
STREAM QUALITY ASSESSMENT FORM
STATION
LOCATION
COUNTY
GROUP OR
INDIVIDUALS
STREAM
SAMPLE i
TOWNSHIP/CITY
DATE
TIME
NO. OF PARTICIPANTS
DESCRIBE WATER CONDITIONS (COLOR, ODOR, BEDGRQWTHS,
SURFACE SCUM. ETC.)
HACH KIT RESULTS (If used) AND
OTHER OBSERVATIONS
USE BACK OF FORM IF NECESSARY
WIDTH OF RIFFLE
WATER DEPTH
WATER TEMP. (°F)
BED COMPOSITION OF RIFFLE (X)
SILT L_I SAND LJ
COBBLES (2"- 10")
LJ
GRAVEL (V- 2"
BOULDERS (> 10") f~]
MACnOINVERTEBRATE
TALLY
ESTIMATED COUNT
LETTER CODE
B ^0 to 99
c > 100 or more
NUMBER OF TAXA
(tiMU)
INDEX VALUE 3
NUMBER OF TAXA
(tiMtA)
INDEX VALUE 2
NUMBER OF TAXA
(•tine*)
INDEX VALUE 1
CUMULATIVE
INDEX VALUE
STREAM QUALITY ASSESSMENT
EXCELLENT (> 22) O SOOD (17-Z2) f""l
FAIR (11-16) Q POOR « 11) O
PLEASE SEND THIS FORM TO:
Mr. John S. Kopec, Planning Supervisor
Division of Natural Areas and Preserves
Ohio Scenic Rivers Program
1889 Fountain Square Court
Columbus. Ohio 4322* Phone: (614) 265-6458
Figure 2. Ohio Department of Natural Resources Scenic Rivers Stream Quality Monitoring Program
assessment form.
64
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Volunteer Monitoring Program Comparison
(Kopec 1989). Other states and private
organizations are also patterning programs
after ODNR's technique.
Generally, identification of invertebrates to
only the order level of classification is
considered to have limited ecological
meaning. Species level identification is
necessary for a more sensitive measure of
water quality. (Resh and Unzicker 1975). This
fact, and the increasing interest in ODNR's
Stream Quality Monitoring Program, caused
OEPA and ODNR staff to question quality
assurance and quality control for the program.
To examine the accuracy of ODNR's stream
quality monitoring technique, a source of
reliable stream health information was needed
for comparison. James Karr (1981) stated
that it would be impossible, because of the
complexity of stream ecosystems, to ever
recognize all the potential factors that may
impact biological communities. Although no
techniques exist which can fully acknowledge
all the processes at work in an aquatic
ecosystem, biological monitoring integrates
the effects of many processes that occur in
streams. To assess stream health, the OEPA
uses biological indices which have been
closely studied and tested, making the
OEPA's methods the best available source of
stream health and biological integrity
information in Ohio.
The OEPA monitors rivers and streams using.
three primary indices as criteria for
assessment. The Index of Biotic Integrity, or
IBI, originated by Karr (1981), is based on fish
populations. Invertebrate samples are used to
compile the Invertebrate Community Index, or
ICI. The third index, the Index of Well-being,
or Iwb, was not examined in this study.
These indices are used to rate the relative
quality of Ohio's rivers and are translated into
ratings of "exceptional, good, fair, poor, and
very poor." The reason for the use of more
than one organism group (fish and
invertebrates) is explained in the OEPA
publication Biological Criteria for the
Protection of Aquatic Life: Volume I. which
states "The need to use both groups is
apparent in the ecological differences
between them, differences that tend to be
complementary in an environmental
evaluation" (Ohio EPA 1988).
In order to address the quality
assurance/quality control issue for ODNR's
Stream Quality Monitoring Program, this
research examined the correlation between
the OEPA's indices (IBI and ICI) and ODNR's
CIV and also the agreement between ODNR
staff- and volunteer-generated stream quality
assessments. The general objective of this
paper is to illustrate how accurately the
results of ODNR's simple approach to
biological monitoring can reflect stream health
assessments based on more sophisticated
approaches.
Methods and Materials
Over the summer of 1989 (late June to mid-
September), the standard ODNR stream
quality monitoring technique (as described
above) was performed on 12 of Ohio's rivers
and tributaries which were also being
monitored by the OEPA (Fig. 3). The sites on
these rivers represented a variety of habitat
and impact types. With the help of the
OEPA's staff, a sampling schedule was
arranged which closely adhered to their
agenda. This was done to help reduce the
effects of seasonal or temporary variations in
stream quality. All ODNR stream quality
assessments were made within 0.8 river miles
of the area sampled by the OEPA and all
ODNR assessments were made within two
weeks of OEPA's testing.
A 9 in. high, 18 in. wide rectangular frame
1732 in. mesh dip net was substituted for the
seine to allow for solo collections. This type
of net and the standard 1/16 in. mesh seine
are used interchangeably in ODNR's program
to allow stream quality monitoring
coordinators to make collections alone. At
65
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Dilley
Figure 3. Ohio rivers monitored in study.
each site sampled, four regular samples were
collected from areas approximately 9 ft.
square and a search was conducted along the
stream's edge for macroinvertebrates such as
dragonfly naiads, which may prefer slower
water velocity or vegetation. An index value
was calculated for each sample and a CIV
was calculated for the riffle. The CIV was
then translated into a qualitative rating.
Assessments were made on 37 different sites
for comparison with the IBI, and many of
those sites were monitored twice, resulting in
56 assessment records. For the comparison
with the ICI, 30 assessment records from 30
sites were collected. The data were entered
into a FoxBase Mac database.
In the spring of 1990, the OEPA finished
processing all of its 1989 data, and their
assessments were merged into a master
database. The study sites were then
examined for correlations between ODNR's
stream quality monitoring results and the
indices of OEPA, the IBI and ICI.
To examine volunteer accuracy, ODNR's
volunteer monitoring database was searched
for sites which were monitored both by staff
members and volunteers within a three month
period of time. Matched records in which one
sample was taken in the early months of
spring and the other in the summer were
discarded, due to the notable changes in
benthic community composition between
these time periods. Spring CIVs are typically
higher and are usually not comparable to
summer CIVs. Over 200 usable matched
records were located in the database.
Results and Discussion
Comparison of the IBI and ODNR's CIV
For sites rated "excellent" by ODNR's
method, corresponding IBIs ranged from 30 to
57 (Fig. 4). This range includes IBIs which the
OEPA would consider indicative of "fair" to
"exceptional" conditions. The CIV ratings did
not match exactly those of the OEPA.
However, 86% of the corresponding IBIs did
fall at a value of 40 or above, indicating
attainment of aquatic life uses, as designated
by the OEPA. For sites rated "good," the
corresponding IBIs again showed a wide
range, with approximately half the values
indicating that sites did attain aquatic life
uses, while the other half indicated that sites
did not attain life uses. All "fair" and "poor"
ratings were observed at sites where IBIs
were less than 40, indicating non-attainment
of life uses.
A primary reason for lack of complete
agreement between the CIV and IBI
qualitative ratings is the inherent differences
between the indices. The IBI is an index
based on fish collected from a 200 meter
reach of stream and ODNR's CIV is based on
macroinvertebrates sampled from a riffle only.
However, another factor, drainage area, was
found to affect the correlation. The OEPA
designates sampling sites as headwater,
wading, and boat sites, based on the drainage
area. When the boat sites were eliminated
66
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Volunteer Monitoring Program Comparison
Excellent
Good Fair
ODNR CIV RATING
Poor
Figure 4. Notched box plots of Cumulative Index Value (CIV) qualitative ratings versus Index of
Biotic Integrity (IBI) scores, 25th and 75th percentiles, IBI range, and IBI outliers (>2 interquartile
ranges from median). IBI qualitative ratings (exceptional, good, fair, poor, and very poor) appear
on the right vertical axis. Shading indicates approximate boundaries between ratings and the
variability of the index.
from the comparison, the definition between
ODNR's qualitative ratings and the
corresponding IBIs increased (Fig. 5). The
median IBI score for each corresponding
ODNR rating fell in the correct qualitative
range for the IBI, and the IBI ranges for the
"excellent" and "good" ratings were
shortened and more defined. The IBI range for
sites rated "good" was still considerably
large, but it was centered in the correct IBI
qualitative range. For "fair" sites, all IBIs fell
in the non-attainment range of less than 40.
A box plot of those sites with drainage areas
greater than 200 square miles further
illustrates the impact of drainage area. (Fig.
6). Notice that, for sites of larger drainage,
there is no detectable definition between sites
rated "excellent" and "good" and the
corresponding IBIs. The IBI ranges for the CIV
ratings are notably similar.
Comparison of the ICI and ODNR's CIV
For sites rated "excellent" by ODNR's
method, ICIs ranged from 41 to 57 (Fig. 7).
This range includes ICIs which the OEPA
would consider "good" to "exceptional." As
in the IBI comparison, the CIV ratings did not
exactly match the ICI ratings of the OEPA.
However, all of the ICIs corresponding to the
"excellent" rating did fall at a value of 35 or
above, indicating attainment of aquatic life
uses, as designated by the OEPA for the ICI.
For sites rated "good," the corresponding ICIs
showed a wide range, with approximately
62% of the values indicating attainment of
aquatic life uses, while the other 38% of the
values indicated non-attainment. ICI values
were less than 35 at sites where "fair" ODNR
results were observed, indicating non-
attainment of life uses. No "poor" sites for
67
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Dilley
Excellent
Good Fair
ODNR CIV RATING
Poor
Figure 5. Notched box plots of Cumulative Index Value (CIV) qualitative ratings versus Index of
Biotic Integrity (IBI) scores, 25th and 75th percentiles, IBI range, and IBI outliers (>2 interquartile
ranges from median) for sites with drainage area <. 200 sq. mi. IBI qualitative ratings
(exceptional, good, fair, poor, and very poor) appear on the right vertical axis. Shading indicates
approximate boundaries between ratings and the variability of the index.
Excellent
Good Fair
ODNR CIV RATING
Poor
Figure 6. Notched box plots of Cumulative Index Value (CIV) qualitative ratings versus Index of
Biotic Integrity (IBI) scores, 25th and 75th percentiles, IBI range, and IBI outliers (>2 interquartile
ranges from median) for sites with drainage area > 200 sq. mi. IBI qualitative ratings
(exceptional, good, fair, poor, and very poor) appear on the right vertical axis. Shading indicates
approximate boundaries between ratings and the variability of the index.
68
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Volunteer Monitoring Program Comparison
Excellent
Good
ODNR CIV RATING
Fair
Figure 7. Notched box plots of Cumulative Index Value (CIV) qualitative ratings versus
Invertebrate Community Index (ICI) scores, 25th and 75th percentiles, ICI range, and ICI outliers
(>2 interquartile ranges from median). ICI qualitative ratings (exceptional, good, fair, poor, and
very poor) appear on the right vertical axis. Shading indicates approximate boundaries between
ratings and the variability of the index.
comparison with the ICI were present in the
data set. Overall, there was a closer
correlation (the ICI ranges for the CIV ratings
were more defined) between ODNR's CIV
ratings and the ICI than ODNR's ratings and
the IBI. The "good" CIV rating still
encompassed a large range of Ids, however,
and the actual CIV and ICI ratings did not
always match.
The differences between ODNR and OEPA
macroinvertebrate assessments may be due,
in part, to the fact that the OEPA retains its
collections for microscopic investigation and
they are better able to locate and identify
small early instar forms of these organisms.
Another factor is that the OEPA researchers
always make an attempt to sample a riffle, a
run, and a pool area when performing their
qualitative collection procedure. This could
result in a higher diversity of organisms in
their samples as compared to ODNR's
samples, which are taken only from riffle
areas. In addition, both the IBI and ICI
incorporate a correction factor to adjust for
drainage area impacts, while ODNR's
technique does not. For the ICI/CIV
comparison, drainage area impacts were not
found to noticeably affect the correlation.
Comparison of Volunteer and Staff
Assessments
Volunteer ratings tended to be higher than
assessments made by ODNR staff members
(Fig. 8). For sites rated "excellent" by staff,
approximately 80% of volunteer CIVs fell in
69
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Oilley
e
Z
40
33.
30.
25.
20.
15.
10
5.
Excellent Good Fair
STAFF-COLLECTED CIV RATING
Poor
Figure 8. Notched box plots of OONR staff ratings versus volunteer-generated Cumulative Index
Values (CIVs), 25th and 75th percentiles, CIV range, and CIV outliers (>2 interquartile ranges
from median). CIV qualitative ratings (excellent, good, fair, and poor) appear on the right vertical
axis. Dashed lines indicate boundaries between ratings.
the excellent range, showing agreement. For
sites rated "good" or "fair" by staff, the range
of corresponding volunteer CIVs was wide,
including CIVs which would be rated "fair" to
"excellent." Differences between staff and
volunteer ratings may be due to
misidentification of organisms by volunteers,
a misconception among program volunteers
that water quality is always "excellent" in
Ohio's Scenic Rivers (potential bias), or a
greater sampling effort by volunteers as
compared to staff members, who may rush
through many reference sites in a day. In the
Central Ohio area (ODNR's headquarters),
stream quality monitoring coordinators have
received better instruction on sampling
strategy through frequent contact with
program administrators and, as a result, the
volunteer and staff assessments for this
region showed closer agreement. This
suggests that part of the reason for the
general lack of agreement may be due to
insufficient sampling by the ODNR stream
quality monitoring coordinators, although all
of the aforementioned factors probably
contribute to the high variability of these
results.
Summary
The qualitative ratings of ODNR's volunteer
monitoring technique do not necessarily agree
with the qualitative ratings of the OEPA.
However, ODNR's CIV ratings do tend to
reflect the attainment ("excellent" CIVs) or
non-attainment ("fair" and "poor" CIVs) of
aquatic life uses, as designated by the Ohio
EPA, for both the IBI and the ICI. Hence, the
assessments may be useful in screening sites
at a basic level.
CIV ratings tend to reflect IBI ratings more
accurately in streams and rivers with smaller
70
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Volunteer Monitoring Program Comparison
drainage areas. Drainage area did not appear
to have a marked effect on the correlation
between the CIV and ICI, but further
collection of data could amplify an otherwise
undetectable effect. Adequacy of sampling
with the use of ODNR's technique may also
affect the correlation. The results of this
research suggest that larger drainage areas
may require a modified approach, although
determining exactly what that approach
should entail is beyond the scope of this
project.
A review of ODNR's database revealed that
program volunteers tend to overrate the
health of Ohio's Scenic Rivers as compared to
staff assessments. This is probably due to a
lack of standardization in the number of
samples collected and misidentification of the
organisms. These problems could be solved
through more thorough training and better
communication between ODNR, the regional
coordinators, and the volunteers. To improve
on the program, a measure of sampling effort
and better quantitative estimates could be
incorporated.
It should be noted that the range of observed
CIV ratings used for the comparisons in this
paper is constricted. There were relatively
few sites which were rated "fair" or "poor"
using ODNR's "kick-seine" method. Further
collection of data will be necessary before
suggestions of revisions to the scoring criteria
or rating system can be made.
Acknowledgements
This research was supported by the School of
Natural Resources, The Ohio State University,
through an Undergraduate Honors Research
Scholarship. This research could not have
been completed without the input and
assistance of my field assistant/secretary (and
fiancee) Chris McKinney; Ed Rankin, Chris
Yoder, Dennis Mishne, Jeff DeShon, Mike
Bolton, and many others at the Ohio
Environmental Protection Agency; Stuart
Lewis and John Kopec of the Ohio
Department of Natural Resources; and my
faculty advisor from The Ohio State
University School of Natural Resources, Dr.
David L. Johnson. My sincere thanks goes out
to each of these individuals.
Literature Cited
Frost, S., A. Huni, and W.E. Kershaw. 1971.
Evaluation of a kicking technique for sampling
stream bottom fauna. Canadian Journal of
Zoology 49:167-173.
Karr, J.R. 1981. Assessment of biotic
integrity using fish communities. Fisheries
6:21-27.
Kopec, J.S. 1989. The Ohio Scenic Rivers
Stream Quality Monitoring Program: Citizens
in action, pp. 123-127. ]n W.S. Davis and
T.P. Simon (eds). Proceedings of the 1989
Midwest Pollution Control Biologists Meeting,
Chicago, IL. USEPA Region V, EPA 905/9-
89/007.
Ohio Department of Natural Resources, n.d.
Ohio's Scenic River Stream Quality Monitoring
Program - A citizen action program.
Columbus: Ohio Department of Natural
Resources.
Ohio Environmental Protection Agency. 1988.
Biological criteria for the protection of aquatic
life: Volume I. The role of biological data in
water quality assessment. Columbus, Ohio.
Rabeni, C.F. and G.W. Minshall. 1977.
Factors affecting microdistributton of stream
benthic insects. Oikos 29(1):33-43.
Resh, V.H. and J.D. Unzicker. 1975. Water
quality monitoring and aquatic organisms: The
importance of species identification. Journal
of the Water Pollution Control Federation
47:9-19.
Schwenneker, B.W. and R.A. Hellenthal.
1984. Sampling considerations in using
71
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Dilley
stream insects for monitoring water quality.
Environmental Entomology 13:741-750.
72
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The Effects of Sediment Deposition on Insect Populations
and Production in a Northern Indiana Stream
Gary W. Kohlhepp and Ronald A. Hellenthal
Department of Biological Sciences
University of Notre Dame
Notre Dame, Indiana 46556
Abstract
In 1986 the St. Joseph County, Indiana, Drainage Board began conducting routine maintenance
operations in and along Juday Creek, a third-order tributary of the St. Joseph River. These activities,
which include debris and snag removal from stream channels, have led to a large increase in sediment
deposition into the lower reaches of the stream. Monthly benthic invertebrate samples were collected
from June 1989 to June 1990 from a riffle area in Juday Creek and insect densities and secondary
production rates during this time were compared to those from a previous study at the same site in
1981-82. Invertebrate density and production rate responses varied based on functional feeding
group. Among filter-feeders, two species showed significantly lower mean annual densities in 1989-
90 compared to 1981-82, two species showed significantly lower densities during several months
in 1989-90 versus corresponding months in 1981-82, and only one species (Hvdropsvche morosa)
showed significantly greater density in 1981-82. Among collector-gatherers, mean annual densities
were significantly higher for five of six species collected in 1989-90. Shredders showed mixed
responses, with two species having significantly higher mean annual densities in 1981-82, and one
species, Taeniootervx nivalis. having higher densities in 1989-90. While production rates of three of
the five species for which production rates were calculated increased in 1989-90, the net effect of
the increased sediment deposition was a reduction in the combined production rates of the five
species from 2765.1 mg/m2/year in 1981-82 to 653.8 mg/m2/year in 1989-90.
Key Words: sediment, water quality, secondary production, functional feeding group, filter-feeders,
collector-gatherers, shredders, benthic macroinvertebrates
Introduction
The effects of anthropogenic disturbances such
as clear-cut logging, modification of riparian
vegetation, and changes in land-use practices
on streams and aquatic organisms are well
documented (Dance and Hynes 1980, Swanson
et al. 1982, Ward 1984). Physical and chemical
attributes of streams affected by these
activities include light penetration (Mclntire and
Colby 1978), water temperature (Hall and Lantz
1969, Holtby 1988), nutrient concentrations
(Chauvet and Decamps 1989), and inputs of
woody debris and sediments (Bryant 1983,
Swanson et al. 1987). Such changes have been
shown to affect algae (Minshall 1978),
macrophytes (Hynes 1970), invertebrates
(Newbold et al. 1980), fish (Thedinga et al.
1989, Ward and Stanford 1989) and other
vertebrates (Hawkins et al. 1983, Spencer et
al. 1991). However, impacts associated with
more subtle or routine activities such as
removal of snags and woody debris from
streams are less understood. Because such
operations likely represent common and
widespread management practices in the
midwestern United States, quantifying their
impact on the biota is essential.
In 1986, the St. Joseph County (Indiana)
Drainage Board began to perform maintenance
operations in and along Juday Creek, a third-
order tributary of the St. Joseph River. Included
in these activities are removal of snags, debris,
and trees that block water flow and that might
result in flooding or pooling along the stream.
Coincident with these practices in Juday Creek
has been a large increase in sediment transport
and subsequent deposition in downstream
73
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Kohlhepp and Hellenthal
locations. Because data on insect production
rates and populations were collected from this
stream in 1981-82 (Schwenneker 1985) and
1985-86 (M.B. Berg pers. comm.), prior to
drainage board activities, we had an
opportunity to document changes in population
densities and production rates of stream
benthos that may have resulted from stream
maintenance operations. Our objectives were to
consider 1) the effects of increased sediment
deposition on benthic macroinvertebrate
populations; 2) whether changes in population
densities were reflected by changes in
invertebrate secondary production rates; and 3)
the implications of these results for predicting
invertebrate responses to sediment deposition.
Study Site
This study was conducted in a riffle area of the
creek (41°43'N, 86°16'W, elevation - 206m)
on land owned and maintained by the St.
Joseph Co. Chapter of the Izaak Walton League
of America. Juday Creek is important from a
conservation perspective because it is one of
only a few streams in Indiana known to support
breeding trout populations. The upper segment
flows through flat, agricultural land, and then
makes its way through primarily residential
areas. The lower segment, which includes the
study location, flows through natural,
deciduous woodlands and has a gradient of
1.3%. The site is heavily shaded from the late
spring through the early fall, and the substrate
is a mixture of sand, gravel, and cobble. Some
of the physical and chemical data collected at
this site in 1981-82 and 1989-90 are
summarized in Table 1.
A silt trap is located approximately 100 m
upstream from the site. It was built to protect
the lower segment of Juday Creek from
excessive sediment deposition. Prior to drainage
board operations, approximately 18 yd3 of
sediment were removed from the silt trap every
1.5-2 years. Since 1988 the silt trap has been
dredged once a year. In 1989, approximately
90 yd3 of sediment were removed from the trap
(J. Moore, pers. comm.). However, during the
1989-90 sampling period, the trap was
completely filled after 4 months. Therefore, it
provided little or no protection to the lower part
of the stream for 8 months. During March and
early April 1990, a large pulse of sediment
entered the study area, resulting in extensive
coverage of the gravel and cobble substrate
with sand. This pulse was apparently a result of
an unusually high number of stream
maintenance activities during this period
compared to previous years.
Materials and Methods
Benthic samples were collected monthly for a
period of thirteen months during the course of
two separate studies. The first was from
September 1981 to September 1982
(Schwenneker 1985) and the second from June
1989 to June 1990. In both studies, ten
random benthic bottom samples were collected
each month from the riffle using a 0.09 m2
Hess sampler with a mesh size of 333 fjm.
Because of extremely high water levels,
samples could not be collected in January and
March of 1982, and these two months were
omitted from comparisons between years.
Samples were preserved in 80% ethanol and
transported back to the laboratory for
processing and analysis.
In the laboratory, invertebrates were sorted
from the substrate using sugar flotation
(Anderson 1959) and then identified to species
and instar or size class. Instar determinations
were based on head capsule width, except for
stoneflies and mayflies. These organisms were
divided into size classes based on body length
from the front of the head to the base of the
cerci. Population densities were recorded for
the most abundant species (excluding
chironomids), and each was assigned to a
functional feeding group (Merritt and Cummins
1984). Mean annual densities were compared
between years using a repeated measures
analysis of variance on log transformed data
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Sediment Deposition Effects on Insects
Table 1. Comparison of physical and chemical
characteristics of Juday Creek in 1981-82 and
1989-90.
Temperature (°C)
Current Velocity (m/s)
Depth (cm)
Alkalinity (mg/l CaC03)
Nitrate (mg/l-N)
Orthophosphate (mg/l)
Conductivity (t/mho/cm)
1981-82
2.5-17.0
0.3-0.5
20-30
150
1.0
0.13
600
1989-90
1.5-20.5
0.4-0.65
20-40
182
1.5-1.7
0.16
610-680
In addition, densities of each species were
compared month by month between years (i.e.,
January 1982 versus January 1990) using a
one-way ANOVA. Production rates were
calculated for all species in the 1981-82 study,
and for five species in the 1989-90 study,
including Hydropsvche soarna Ross,
Hydropsvche betteni Ross, Qptioservus
fastiditus (LeConte), Baetis vaoans
McDunnough, and Taenioptervx nivalis (Fitch).
Dry weights for each instar or size class of
these organisms were obtained by drying at
70°C for 48 hours. Production rates were
measured using either the instantaneous growth
method for those species with distinguishable
cohorts, or the size-frequency method for those
with cohorts that could not be distinguished.
Instantaneous growth calculations were made
from computer programs developed by
Schwenneker (1985) and Berg (1989). Size-
frequency production calculations were made
using the Aquatic Ecology-PC software package
(Ekblad 1986). Because early instars often are
inefficiently sampled, apparent increases in
population densities are sometimes seen during
cohort development. For production rate
calculations, densities were back-calculated
using the catch-curve method of Waters and
Crawford (1973). The result of this correction
was that if densities were lower at sampling
time T-1 than at time T, densities at time T-1
were set equal to those at time T. Given the
typical log-type decline exhibited by stream
invertebrates, this correction represents a
conservative estimate of densities
(Schwenneker 1985).
Results
Population responses of individual species to
sedimentation varied depending on functional
feeding group. Among the six species of filter-
feeders, four had lower mean annual densities
in 1989-90, although only two of these
differences were statistically significant (Table
2). Hvdropsvche soarna and Chimarra obscura
(Walker) showed the most dramatic density
reductions in 1989-90 (over 80% and 95%
respectively). Mean annual densities of
Hvdropsvche betteni and Cheumatopsvche
petiti (Banks) were not significantly different
between years, but each did show significant
density differences in eight of the ten monthly
comparisons (Fig. 1). During six of these
months, ]H- betteni showed higher densities in
1981-82 compared to 1989-90, and £. petiti
had significantly higher densities in 1981-82
during five months. Hvdropsvche morosa (Ross)
mean annual density was significantly higher in
1989-90 compared to 1981-82, and Simulium
sp. had similar densities between years. All six
filter-feeders had lower densities, four of which
were significant (p<.05), in April and May of
1990 versus 1982.
The reduction in secondary production rates of
filter-feeders in 1989-90 was much more
pronounced than changes in population
densities (Fig. 2). The production rate of ]H.
sparna dropped more than 90% in 1989-90
from the 1981-82 rate. From 1981-82 to 1989-
90, the production rate of JH. betteni declined
by more than 50%, while mean annual density
was reduced about 30% in 1989-90.
Population densities of collector-gatherers
showed different responses to the increased
sediment deposition than the of filter-feeders.
Five of the six species of collector-gatherers
showed significantly greater mean annual
densities in 1989-90 compared to 1981-82
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Kohlhepp and Hellenthal
Table 2. Mean annual densities (N/M2) and p values of individual taxa in
= p < 0.05; •• = p < 0.01.
1981-82 1989-90
Mean (1 SE) Mean (1 SE)
Filter-feeders
Cheumatopsvche petiti
Chimarra obscura
Hvdropsyche betteni
Hvdropsvche morosa
Hvdropsyche sparna
Simulium sp.
Collector-gatherers
Antocha so.
Baetis vaoans
Macronvchus olabratus
Qptioservus fastiditus
Stenelmis crenata
Stenonema spp.
Shredders
Arnphinemura delosa
Taenioptervx nivalis
Tipula abdornjnalis
(Table 2). The elmid beetle
223.2(18.2)
44.2 (4.0)
258.8 (23.0)
26.3 (2.4)
1648.7 (136.9)
45.9 (5.4)
43.9 (3.3)
17.3 (2.4)
22.1 (1.7)
3.9 (0.7)
56.5(5.1)
66.6 (8.6)
6.5(1.4)
3.3 (0.5)
larvae Qptioservus greater
194.3(30.6)
2.2 (0.4)
187.9(24.2)
113.6(18.9)
293.7 (35.6)
47.3(13.4)
57.6 (5.2)
32.7(12.0)
34.0 (4.5)
144.0(12.8)
155.1 (13.0)
91.3 (10.7)
3.4 (0.7)
38.8(8.1)
1.3 (0.3)
in 1 989-90.
1981 -82 and 1989-90. «
P
0.270
••0.001
0.097
••0.001
••0.001
0.376
••0.001
••0.001
••0.001
••0.001
••0.001
••0.001
••0.001
••0.001
Densities of some species
fastiditus and Stenelmis crenata (Say) had the
largest increases in mean annual densities
between years (650% and almost 4000%,
respectively I. Macronvchus glabratus (Say) was
collected so rarely in 1981-82 that its density
was not reported by Schwenneker (1985).
Therefore, we could not compare densities
between years, although densities were clearly
of collector-gatherers dropped in April and May
of 1990, although some declines could be
explained partly by life history characteristics.
We compared secondary production rates
between years for two species of collector-
gatherers, Qptioservus fastiditus and Baetis
vagans. Production rates showed changes
76
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Sediment Deposition Effects on Insects
200
** •
LLL
Apr May Jun Jul Aug 8«P Oel Nov Dec
•I Wat-62 ^ 19M-BO
Combined H. iparna H. bttltnl O tutldllui B. v«gtrn T. nlvtlli
•11981-82 £211989-80
ftt Apr M«y Jun Jul Aug S*p Oet Nm Dec
Figure 1. (A) Mean densities (N/M2) by month
for Hvdropsvche betteni in 1981-82 and 1989-
90. (B) Mean densities (N/M2) by month of
Cheumatopsvche oetiti in 1981-82 and 1989-
90. * = p<0.05; ** = p<0.01).
similar to those of the population densities for
these species. Production rates of Q. fastiditus
and B. vaaans were more than 300% and
200% greater, respectively, in 1989-90 than in
1981-82 (Fig. 2).
Figure 2. Combined (all 5 species) and
individual species secondary production rates
(mg/m2/yr) in 1981-82 and 1989-90.
Population densities of shredders showed mixed
responses between years (Table 2). Mean
annual densities of Amohinemura delosa
(Ricker) and Tioula abdominalis (Say) were
significantly lower in 1989-90 than in 1981-82,
Mean annual densities of Taeniootervx nivalis.
on the other hand, increased significantly in
1989-90. However, the secondary production
rate of T. nivalis was only 20% higher in 1989-
90 (Fig. 2), because of a greater mean
individual biomass in 1981-82.
The predator functional feeding group is not
represented because the only predaceous
invertebrates at this location besides f latworms
were the filter-feeding hydropsychid
caddisflies. The only obligate scraper,
Glossosoma intermedium (Klapalek), had greater
densities in 1981-82 than in 1989-90 (8.91 /m2
vs. 6.6/m2), but differences were not significant
(p = 0.389).
Three of the five species for which we
77
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Kohlhepp and Hellenthal
Collector-gatherers
10.96
Shredders
0.56
B
Collector -gatherers
394
Shredders
383
Filter-feeders
88.48
Filter-feeders
5677
Shredders
Collector-gatherers 4 25
1.29 ^
Filter-feeders
94.46
Filter-feeders
561
Collector-gatherers
2136
Shredders
22.54
Figure 3. Percent contribution by functional feeding group to (A) 1981-1982 total densities, (B)
1989-90 total densities, (C) 1981-1982 secondary production rates, (D) 1989-90 secondary
production rates. Relative values are calculated using only five species in Figure 2.
calculated secondary production rates had
higher rates in 1989-90 compared to 1981-82.
However, summing the production rates for all
five species resulted in a substantial decline in
production rates in 1989-90 compared to
1981-82 (Fig. 2). Of these five species, the
two filter-feeders in 1981-82 contributed
88.48% of the numbers and 94.46% of the
combined secondary production rate (Fig. 3). In
1989-90, these filter-feeders declined in both
densities and production rates, resulting in a
large decline in total secondary production
rates. The three species that increased in
production rate in 1989-90 versus 1981-82
were those that contributed little to total
production rate. The result is a shift from a
community dominated by filter-feeders in both
numbers and production rate in 1981-82 to a
community in 1989-90 in which collector-
gatherers and shredders increased in
importance in terms of relative contribution to
both numbers and production (Fig. 3).
Discussion
It seems likely that the shift in invertebrate
community structure and decreased secondary
production rates observed in 1989-90 was the
result of increased sediment deposition into the
study area. Differences cannot be explained by
changes in the physicochemical characteristics
of Juday Creek between years (Table 1). Other
authors have noted that reduction or elimination
of the overhead canopy can result in abrupt
changes in abundances of stream invertebrates
(Behmer and Hawkins 1986, Newbold et al.
1980). However, the riparian vegetation in this
part of the stream has remained undisturbed.
Attributing differences between years to
increased fish predation also is unsupported.
Furthermore, data collected in 1985-86 showed
78
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Sediment Deposition Effects on Insects
that insect species abundances were similar to
those from 1981-82 in this section of the
stream (M.B. Berg, pers. comm.). The county
drainage board began maintenance operations
along Juday Creek soon after the 1985-86
study. Almost immediately, sediment deposition
into the lower portion of Juday Creek increased
substantially. The result has been severely
reduced population densities for some species
and higher densities for others.
Functional feeding group classification proved
to be a good predictor of a species' population
response to the sediment deposition. Filter-
feeder population densities were the most
severely reduced (Table 2, Figure 1). Reduction
of filter-feeder densities due to greater sediment
deposition is consistent with their biology and
habitat requirements. These organisms require
a solid and somewhat stable substrate on
which to spin their nets (Hynes 1970, Minshall
1984), and are considered intolerant of heavily
silted and sandy areas where they lose their
attachment to the substrate (Marsh and Waters
1980, Tebo 1955). In addition, large amounts
of suspended inorganic sediments can clog nets
and interfere with the feeding mechanisms of
the net-spinners (Nuttall and Bielby 1973).
These results generally are consistent with
those found in other studies. The net-spinning
caddisflv Arctoosvche orandis (Banks) is highly
sensitive to sedimentation due to the filling of
interstitial spaces of rocks (Cline et al. 1982,
McClelland and Brusven 1980). Barton (1977)
found reduced densities of Hvdropsvche
slossonae Banks (48%) and Cheumatopsvche
sp. (66%) in an Ontario stream immediately
downstream from a highway construction site
compared to upstream locations. He suggested
that damage to benthos only occurs when
stones are buried by sediment, which was the
case in our study. Gurtz and Wallace (1986)
and Benke et al. (1984) found lower production
rates of net-spinning caddisflies on sand when
compared with more stable substrates. Other
studies also have found that filter-feeders are
intolerant of sediment additions (Cherry et al.
1979, Nuttall and Bielby 1973, White and
Gammon 1976).
In contrast to the filter-feeders, mean annual
densities of all collector-gatherers increased
significantly in 1989-90 from the 1981-82
levels (Table 2). Although most of the
deposited sediment was inorganic sand during
the 1989-90 study, the amount of organic
material deposited in the riffle area probably
increased as well. This would result in a greater
food supply for the collector-gatherers, which
forage along the substrate for food particles
(Berkman et al. 1986, Merritt and Cummins
1984). Elmid beetles showed the greatest rise
in numbers (Table 2). Although studies have
shown that Qptioservus and Stenelmis prefer
larger, solid substrates (Cummins and Lauf
1969, Marsh and Waters 1980, Rabeni and
Minshall 1977), these organisms are known to
tolerate fine sediments to some extent (Brown
1987, White and Gammon 1976). The reason
for the increased density of Macronvchus
glabratus in 1989-90 is unclear, since this
species is almost always associated with wood
substrate (Brown 1987, Hynes 1970). The
amount of woody debris in the stream was not
quantified in 1981-82 or in 1989-90.
Stenonema also is somewhat tolerant of silt
(Dance and Hynes 1980, Jones and Clark
1987), and could benefit from an increase in
deposition of organic matter. Investigators have
found that Baetis nymphs generally prefer larger
substrates and drift in the presence of large
quantities of sediment (Culp et al. 1986,
Wagner 1989, White and Gammon 1976).
However, numbers of the European species
Baetis rhodani (Pictet) increase as sediment
deposition increases (Nuttall and Bielby 1973,
Scullion and Edwards 1980). Wallace and Gurtz
(1986) found that although production rate of
Baetis sp. was greatest on cobble and gravel,
individuals of this genus were found in
moderate numbers in sandy areas. Culp et al.
(1986) showed that while sediment transport
reduced B. tricaudatus Dodds densities by
67%, sediment deposition actually decreased
drift rates of this species. As with the other
collector-gatherers in this study, the potential
79
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Kohlhepp and Hellenthal
increase in food material transported into the
study area may have offset any negative
effects on B. vaaans associated with the loss of
stable substrate. However, densities of Q.
fastiditus. Antocha sp.f and Stenonema spp. all
decreased substantially in April and May of
1990, coinciding with the sediment pulse that
covered the study site. It is possible that
continued heavy sediment deposition could
eventually cause population declines in many of
these collector-gatherers.
The population responses of the shredder
functional feeding group varied depending on
the species (Table 2). Taeniootervx nivalis
appeared to benefit from the increased
sediment deposition. T. nivalis nymphs are
generally found in leaf packs or other debris
along the stream margins (Sephton and Hynes
1984, Stewart and Stark 1988). This species
may not be affected by sediment that covers
the cobble and gravel substrate as long as there
are sufficient leaf packs in the stream. During
this study, there were many leaf pack
accumulations both in the main channel and
along the margins and backwater areas. Like
the collector-gatherers, T. nivalis may have
benefirted from a potential increase in available
detritus in the study area. Because adults
emerge in February and the nymphs diapause
deep in the substrate until September (Stewart
and Stark 1988), this species would not have
been affected by the sediment pulse that
occurred in the spring. In contrast, Tioula
abdominalis had significantly lower densities in
1989-90 compared to 1981-82 (Table 2).
Cummins and Lauf (1969) found that Tioula
calootera Loew preferred coarse substrates, but
showed a wide tolerance for finer sediments,
including silt. They suggest that these larvae
are probably found in microhabitats of finer
sediments and organic material between and
behind coarse sediments. If J_. abdominalis has
similar tolerances, then this species should be
moderately affected by the increased
deposition. Amphinemura delosa densities also
were significantly lower in 1989-90 (Table 2).
Another congeneric species, A. sulcicollis
(Stephens), is known to move from leaves to
stone as it develops (Hynes 1976). If A., delosa
exhibits a similar habitat shift, then a reduction
in numbers would be expected with increasing
sediment deposition. However, Scullion and
Edwards (1980) found that A. sulcicollis was
tolerant to mine discharge siltation. The lack of
tolerance of A. delosa in our study may
represent species-specific differences in
tolerance or may be related to the amount of
material deposited.
Most studies that examine the effects of
sediment deposition on benthic invertebrates
look only at changes in population densities or
relative changes in species composition (Barton
1977, Nuttall and Bielby 1973, Scullion and
Edwards 1980). The few studies that have
looked at changes in biomass found it to be a
better measure of benthos response to
sedimentation than density and diversity
(Letterman and Mitsch 1978, Marsh and
Waters 1980). We have found no studies that
examine the effects of sediment deposition on
the secondary production rates of stream
invertebrates. Studies have compared
invertebrate secondary production rates
between logged and unlogged areas (eg.
Wallace and Gurtz 1986), but have difficulty
separating the relative effects of sediments,
increased algal growth, and water
temperatures.
There are many advantages to calculating
secondary production rates rather than looking
only at changes in abundances. Secondary
production incorporates a measure of individual
growth as well as population density, and
provides a measure of the functional
importance of a species to stream energy flow
and nutrient processing (Short et al. 1987).
From a management perspective, invertebrates
are an important component of fish diets and
may limit fish production (Benke 1984, Tebo
1955). Our results suggest that changes in
secondary production rates between years give
a clearer picture of what is happening to the
benthos in Juday Creek than changes in
80
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Sediment Deposition Effects on Insects
abundance. Because eight species show greater
densities in 1989-90 compared to six species
with greater densities in 1981-82, examination
of numbers atone suggests that the only impact
of the sediment on the invertebrate community
was a change in relative abundance of species.
However, the species that had reduced
densities in 1989-90 are those that contributed
the most to invertebrate secondary production
rates in 1981-82. Those that had increased
densities in 1989-90 contributed little to
secondary production in 1981-82 (Figure 2).
The result was that the combined production
rates for five species decreased from 2765.1
mg/mz/year in 1981-82 to 653.8 mg/mz/year in
1989-90. This 78% reduction in secondary
production rate represents a substantial decline
in the amount of food available to support
higher trophic level organisms such as fish. The
response of Taeniootervx nivalis populations
also represents a good example of the value of
calculating production rates. While densities of
this species increased approximately 6 times in
1989-90 from 1981-82, production rates
increased only 20% (Fig. 2). The large increase
in mean densities of this species was mostly
offset by lower individual growth rates.
Our results suggest that a functional feeding
group classification of organisms offers a
method for predicting the downstream impacts
of stream maintenance activities. If a
community is dominated by filter-feeders, then
substantial impacts associated with an increase
in sediment deposition may be expected. If
collector-gatherers contribute the majority of
invertebrate production, then moderate
increases in sediment transport and deposition
may actually enhance population densities and
production rates through increased import of
organic material. However, heavy, sustained
sediment deposition will probably have a
negative impact on many of these species as
well. In addition, secondary production rates
are a better measure of invertebrate community
response than diversity and population density.
Finally, many studies have documented that the
benthos recovers rapidly when the source of
sediment is eliminated and high flows are
allowed to wash the sediment downstream
(Barton 1977, Cherry et al. 1979, Tebo 1955).
If silt traps are properly maintained and cleaned
out before filling with sediment, then it is
possible that negative impacts associated with
instream maintenance operations can be
reduced.
Acknowledgements
We would like to express our appreciation to
the St. Joseph Co. Chapter of the Izaak Walton
League of America for allowing access to the
stretch of Juday Creek that runs through their
property, and especially the caretaker, John
Moore, for all of his help during the study. We
also are grateful to Barb Hellenthal for her
comments on an earlier version of this
manuscript. This research was supported by a
grant from the Indiana Academy of Science.
Literature Cited
Anderson, R.O. 1959. A modified flotation
technique for sorting bottom fauna samples.
Limnology and Oceanography 4: 223-225.
Barton, B.A. 1977. Short-term effects of
highway constuction on the limnology of a
small stream in southern Ontario. Freshwater
Biology 7: 99-108.
Behmer, D.J. and C.P. Hawkins. 1986. Effects
of overhead canopy on macroinvertebrate
production in a Utah stream. Freshwater
Biology 16: 287-300.
Benke, A.C. 1984. Secondary production of
aquatic insects. Pages 289-322 in V.H. Resh
and D.M. Rosenberg (editors). The ecology of
aquatic insects. Praeger Press, New York.
Benke, A.C., T.C. Van Arsdall, Jr., D.M.
Gillespie, and F.K. Parrish. 1984. Invertebrate
productivity in a subtropical blackwater river:
the importance of habitat and life history.
Ecological Monographs 54(1): 25-63.
81
-------
Kohlhepp and Hellenthal
Berg, M.B. 1989. The role of Chironomidae
(Diptera) in stream insect secondary production.
Ph.D. Dissertation, University of Notre Dame,
Notre Dame, Indiana. 239 pp.
Berkman, H.E., C.F. Rabeni, and T.P. Boyle.
1986. Biomonitors of stream quality in
agricultural areas: fish versus invertebrates.
Environmental Management 10(3): 413-420.
Brown, H.P. 1987. Biology of riffle beetles.
Annual Review of Entomology 32: 253-273.
Bryant, M.D. 1983. The role and management
of woody debris in west coast salmonid nursery
streams. North American Journal of Fisheries
Management 3: 322-330.
Chauvet, E. and H. Decamps. 1989. Lateral
interactions in a fluvial landscape: the River
Garonne, France. Journal of the North American
Benthological Society 8(1): 9-17.
Cherry, D.S., S.R. Larrick, R.K. Guthrie, E.M.
Davis, and F.F. Sherberger. 1979. Recovery of
invertebrate and vertebrate populations in a
coal ash stressed drainage system. Journal of
the Fisheries Research Board of Canada 36:
1089-1096.
Cline, L.D., R.A. Short, and J.V. Ward. 1982.
The influence of highway construction on the
macroinvertebrates and epilithic algae of a high
mountain stream. Hydrobiologia 96: 149-159.
Gulp, J.M., F.J. Wrona, and R.W. Davies.
1986. Response of stream benthos and drift to
fine sediment deposition versus transportation.
Canadian Journal of Zoology 64:1345-1351.
Cummins, K.W. and G.H. Lauff. 1969. The
influence of substrate particle size on the
microdistribution of stream macrobenthos.
Hydrobiologia 34: 145-181.
Dance, K.W. and H.B.N. Hynes. 1980. Some
effects of agricultural land use on stream insect
communities. Environmental Pollution (Series A)
22: 19-28.
Ekblad, J. 1986. Aquatic Ecology-PC. Oakleaf
Systems, Decorah, Iowa.
Gurtz, M.E. and J.B. Wallace. 1986.
Substratum-production relationships in net-
spinning caddisflies (Trichoptera) in disturbed
and undisturbed hardwood catchments. Journal
of the North American Benthological Society
5(3): 230-236.
Hall, J.D. and R.L. Lantz. 1969. Effects of
logging on the habitat of coho salmon and
cutthroat trout in coastal streams. Pages 355-
376 in T.G. Northcote (editor). Symposium on
salmon and trout in streams. February 22-24,
1968. University of British Columbia.
Hawkins, C.P., M.L. Murphy, N.H. Anderson,
and M.A. Wilzbach. 1983. Density of fish and
salamanders in relation to riparian canopy and
physical habitat in streams of the northwestern
United States. Canadian Journal of Fisheries
and Aquatic Sciences 40(8): 1173-1184.
Holtby, L.B. 1988, Effects of logging on stream
temperatures in Carnation Creek, British
Columbia, and associated impacts on the coho
salmon (Oncorhvnchus kisutch). Canadian
Journal of Fisheries and Aquatic Sciences 45:
502-515.
Hynes, H.B.N. 1970. The ecology of running
waters. Liverpool University Press, Liverpool,
Great Britain.
Hynes, H.B.N. 1976. Biology of Plecoptera.
Annual Review of Entomology 21: 135-154.
Jones, R.C. and C.C. Clark. 1987. Impact of
watershed urbanization on stream insect
communities. Water Resources Bulletin 23(6):
1047-1056.
Letterman, R.D. and W.J. Mitsch. 1978. Impact
of mine drainage on a mountain stream in
Pennsylvania. Environmental Pollution 17: 53-
82
-------
Sediment Deposition Effects on Insects
73.
Marsh, P.C. and T.F. Waters. 1980. Effects of
agricultural drainage development on benthic
invertebrates in undisturbed downstream
reaches. Transactions of the American Fisheries
Society 109: 213-223.
McClelland, W.T., and M.A. Brusven. 1980.
Effects of sedimentation on the behavior and
distribution of riffle insects in a laboratory
stream. Aquatic Insects 2:161-169.
Mclntire, C.D. and J.A. Colby. 1978. A
hierarchical model of lotic ecosystems.
Ecological Monographs 48:167-190.
Merritt, R.W. and K.W. Cummins. 1984. An
introduction to the aquatic insects. Second ed.,
Kendall/Hunt Publishing Company, Dubuque,
Iowa.
Minshall, G.W. 1978. Autotrophy in stream
ecosystems. BioScience 28(12): 767-771.
Minshall, G.W. 1984. Aquatic insect-
substratum relationships. Pages 358-400 in
V.H. Resh and D.M. Rosenberg (editors). The
ecology of aquatic insects. Praeger Press, New
York.
Newbold, J.D., D.C. Erman, and K.B. Roby.
1980. Effects of logging on macroinvertebrates
in streams with and without buffer strips.
Canadian Journal of Fisheries and Aquatic
Sciences 37: 1076-1085.
Nuttall, P.M. and G.H. Bielby. 1973. The effect
of china-clay wastes on stream invertebrates.
Environmental Pollution 5:77-86.
Rabeni, C.F. and G.W. Minshall. 1977. Factors
affecting microdistribution of stream benthic
insects. Oikos 29: 33-43.
Schwenneker, B.W. 1985. The contribution of
allochthonous and autochthonous organic
material to aquatic insect secondary production
rates in a north temperate stream. Ph.D.
Dissertation, University of Notre Dame, Notre
Dame, Indiana. 363 pp.
Scullion, J. and R.W. Edwards. 1980. The
effects of coal industry pollutants on the
macroinvertebrate fauna of a small river in the
South Wales coalfield. Freshwater Biology 10:
141-162.
Sephton, D.H. and H.B.N. Hynes. 1984. The
ecology of Taeniootervx nivalis (Fitch)
(Taeniopterygidae; Plecoptera) in a small stream
in southern Ontario. Canadian Journal of
Zoology 62: 637-642.
Short, R.A., E.H. Stanley, J.W. Harrison, and
C.R. Epperson. 1987. Production of Corvdalus
cornutus (Megaloptera) in four streams differing
in size, flow, and temperature. Journal of the
North American Benthological Society 6(2):
105-114.
Spencer, C.N., B.R. McClelland, and J.A.
Stanford. 1991. Shrimp stocking, salmon
collapse, and eagle displacement. BioScience
41(1): 14-21.
Stewart, K.W. and B.P. Stark. 1988. Nymphs
of North American stonefly genera (Plecoptera).
The Thomas Say Foundation, Vol. 12. 460 p.
Swanson, F.J., S.V. Gregory, J.R. Sedell, and
A.G. Campbell. 1982. Land-water interactions:
the riparian zone. Pages 267-291 in R.L.
Edmonds (editor) Analysis of coniferous forest
ecosystems in the western United States.
US/IBP Synthesis Series 14. Hutchinson Ross
Publishing, Stroudsburg, PA.
Swanson, F.J., L.E. Benda, S.H. Duncan, G.E.
Grant, W.F. Megahan, L.M. Reid, and R.R.
Ziemer. 1987. Mass failures and other
processes of sediment production in Pacific
Northwest forest landscapes. Pages 9-38 in
E.O. Salo and T.W. Cundy (editors). Streamside
management: forestry and fishery inter- actions.
University of Washington, Institute of Forest
83
-------
Kohlhepp and Hellenthal
Resources, Seattle.
Tebo, L.B., Jr. 1955. Effects of siltation,
resulting from improper logging on the bottom
fauna of a small trout stream in the southern
Appalachians. The Progressive Fish-Culturist
17:64-70.
Thedinga, J.F., M.L. Murphy, J. Heifetz, K.V.
Koski, and S.W. Johnson. 1989. Effects of
logging on size and age composition of juvenile
coho salmon (Oncorhvnchus kisutchl and
density of presmolts in southeast Alaska
streams. Canadian Journal of Fisheries and
Aquatic Sciences 46: 1383-1391.
Wagner, R. 1989. The influence of artificial
stream bottom siltation on Ephemeroptera in
emergence traps. Archive fur Hydrobiologie
115(1): 71-80.
Wallace, J.B. and M.E. Gurtz. 1986. Responses
of Baetis mayflies (Ephemerop- tera) to
catchment logging. American Midland Naturalist
115(1): 25-41.
Ward, J.V. 1984. Ecological perspectives in the
management of aquatic insect habitat. Pages
558-577 in V.H. Resh and O.M. Rosenberg
(editors) The ecology of aquatic insects.
Praeger Press, New York.
Ward, J.V. and J.A. Stanford. 1989. Riverine
ecosystems: the influence of man on catchment
dynamics and fish ecology. Pages 56-64 in D.P.
Dodge (editor). Proceedings of the International
Large River Symposium. Can. Spec. Publ. Fish.
Aquat. Sci. 106.
Waters, T.F. and G.W. Crawford. 1973. Annual
production of a stream mayfly population.
Limnology and Oceanography 18: 286-296.
White, D.S. and J.R. Gammon. 1976. The
effect of suspended solids on macroinvertebrate
drift in an Indiana creek. Proceedings of the
Indiana Academy of Science 86: 182-188.
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Agricultural Impacts on the Fishes of the Eel River, Indiana
James R. Gammon and Clifford W. Gammon
Department of Biological Sciences
DePauw University,
Greencastle, Indiana 46135
Abstract
The Eel River of northern Indiana is a major tributary of the Wabash River. It is approximately 177
km (110 mi) in length with an average rate of descent of 0.457 m/km (2.41 ft/mi). Approximately
79% of its 210,800 ha (814 m2) drainage basin is devoted to row-crop agriculture. The fish
communities and habitat were studied during the summer of 1990. Fish were collected from 25 sites
located throughout the Eel River and some of its tributaries. A 3/16 inch mesh, 30 ft by 4 ft seine
was effective in collecting small fish including darters. Backpack electrofishing was also used at most
stations on two separate dates. Historic records of the fish communities were examined and, when
possible, converted into Indexes of Biotic Integrity values so that changes over time could be
estimated. Habitat evaluation included a mainstem reconnaisance, a habitat survey (HEP) conducted
at all collecting stations, and a synoptic turbidity survey on July 16 and 17,1990. Estimates of the
amount of woodland were made from conventional analysis of enlarged infrared photographs.
Analyses of existing suspended sediment data were used to evaluate possible impacts of nonpoint-
source influence from agricultural fields as well as historic records of fish kills and chemical spills
within the Eel River watershed. The 1990 fish community was generally better than the community
found in 1982. However, this improvement is probably temporary and the result of a series of recent
years when both river discharge and suspended sediment concentrations were lower than normal.
From a longer time perspective the fish community is degraded, with many species which were
common 50 years ago now either absent or very severely reduced. Rainbow darter, orangethroat
darter, bluebreast darter, and stonecat were not collected at all. Sculpin, greenside darter, blackside
darter, silver shiner, rosyface shiner, longear sunfish, and smallmouth bass were very restricted in
distribution.
Key Words: non-point source, habitat, suspended sediment, fish community, Indiana
Introduction
In 1982 populations of smallmouth bass
(Micropterus dolomieui) were found to be
virtually lacking by Braun and Robertson
(1982) who collected from the same sites as
Taylor (1972). Exerting roughly equivalent
effort and similar methods, Taylor (1972)
collected 98 smallmouth bass while Braun and
Robertson (1982) found only 3. During the
1980's smallmouth bass populations in the
lower part of the Eel River were augmented by
stocking fin-clipped fingerlings (5130 fish on
10-28-83; 5000 on 9-17-85; and 6960 on 4-
17-86). A limited number of stations were
more intensively sampled and additional
tributaries were also investigated.
The current study was planned to provide
information about the fish communities at all
of Taylor's sites and a few additional sites. It
included an evaluation of instream and near-
stream habitat from the standpoint of
agricultural nonpoint sources of pollution and
their possible influence on those fish
communities. This report is a condensation
and extension of a larger report to the Indiana
Department of Environmental Management
(Gammon and Gammon 1990).
The Study Area
The Eel River is a major tributary of the
Wabash River in northern Indiana. Originating
in northwest Allen county near Ft. Wayne, it
flows southwest for approximately 177 km
(110 miles) through Kosciusko, Whitley,
Wabash and Miami counties into the Wabash
85
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Gammon and Gammon
River at Logansport in Cass county. Its rate of
descent is approximately 0.457 m/km (2.41
ft/mile) with a lower rate in the upper third and
a slightly higher rate in the lower 20 km.
This area originally contained glacial lakes and
swampy wetlands, but it was extensively
ditched and drained prior to 1900 for
agricultural use. Approximately 79% of its
2,148 km2 (814 m2) drainage basin area
(Hoggatt 1975) is devoted to rowcrop agricul-
ture, primarily corn and soybeans. Most of the
smaller tributaries and the upper river have
been channelized to facilitate drainage. Low
mill dams have been constructed at various
locations, many of which are currently in a
state of disrepair except in Logansport. That
dam severely restricted the movement of
Wabash River fishes into the Eel River and
facilitated evaluations of impacts.
Materials and Methods
The study included a) a reconnaisance float
trip of the entire river, b) sampling each
station twice by electrofishing, c) sampling
most of these same stations once by seining,
and d) a habitat survey (HEP) at each station.
Secchi transparency and temperature were
routinely measured on each occasion. In
addition, synoptic short-term profiles of
turbidity, temperature, and dissolved oxygen
concentration were determined on three
separate dates.
Single stations were located on lower Twelve
Mile, Paw Paw, Squirrel, Beargrass, and Sugar
Creeks, and also upstream and downstream of
Columbia City on Blue River. The remaining 16
stations were located on the matnstem of the
Eel River. A few mainstem stations (Taylor's
2B, 2, and 3) and Squirrel Creek were not
seined because of inappropriate seining
habitat.
Seining was conducted with a 30-foot by
4-foot seine having 3/16 inch mesh weighed
down by a heavy steel chain tied to the
bottom. This method was very effective at
capturing darters and minnows. Three seining
passes along 20 meters of shoreline
constituted each seine sample.
Electrofishing utilized a Safari Bushman 300
backpack shocker carried in a canoe or while
wading, depending on place and depth. Each
electrofishing sample was about 20 minutes in
duration along approximately 400 meters of
shoreline. This method was effective in
capturing larger fish such as redhorse and
suckers and species which prefer nearshore
cover such as sunfish and bass.
All captured fish were identified to species,
weighed and measured, then released back
into the river. Those fish not easily identified in
the field were preserved in formalin and
returned to the laboratory for identification
(Trautman 1981).
Fish data were analyzed using the Iwb and the
IBI. The 1990 Iwb values were based upon the
average of two electrofishing catches at each
station. The rationale of this community
parameter is presented by Gammon (1980),
who recommended multiple collections at each
site.
The Iwb was calculated as:
Iwb = 0.5 In N
Div.wt.
+ 0.5 In W + Div.no. +
Where, N = number of fish captured per km;
W = weight in kg of fish captured per km;
Div.no. = Shannon diversity based on
numbers; Div.wt. = Shannon diversity based
on weight.
The IBI methodology has been thoroughly
discussed by Karr (1981 and 1987), Karr et al.
(1986 and 1987), and Angermeier and Karr
(1986). Regional applications are summarized
by Miller et al. (1988).
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Agricultural Impacts in the Eel River
Table 1. Scoring criteria used to determine IBI
for Eel River fish collections.
Metric
1 (worst)
Score
3 5(best)
^20
Fish species (total) 0-9 10-19
Darter species 0-1 2-3
Sunfish Species 0-1 2-3
Sucker Species 0-1 2-3
Intolerant Species 0-1 2-3
No. Individuals 0-100 101 -200 & 201
Percent individuals as:
Green sunfish 11-100
Omnivores 45-100
Insect, cyprinids 0-20
Top carnivores 0-2
Hybrids 4-10
Diseased 6-10
6-10 0-5
21-44 0-20
21-4445-100
3-10 £11
2-3 0-1
2-5 0-1
The original criteria for determining IBI (Karr,
et. al., 1987) were modified slightly for the Eel
River (Table 1). The scaled metrics are those
used in studies of the Sugar Creek system
(Gammon et al. 1990a) and an agricultural
analysis of several streams in west-central
Indiana (Gammon et al. 1990b). They differ in
some details from the criteria used in other
studies. The 1990 IBI values were based upon
the combined catches from electrofishing and
seining. The IBIs calculated on data from earlier
Eel River studies may be influenced to an
unknown degree by the somewhat different
methodologies used to collect fish. Taylor
(1972) used a combination of electrofishing
and rotenone, while Braun and Robertson
(1982) used more intensive electrofishing. We
have elected to use the same criteria
regardless of stream order.
Habitat was quantitatively evaluated at each
mainstem collecting site, except for the most
downstream site near the Logansport dam and
Taylor's site 1, using a habitat evaluation pro-
Table 2. Habitat assessment scoring criteria
(HEP).
Condition
Parameter Excellent Good Fair Poor
PRIMARY INFLUENCE
Substrate and Instream Cover
1. Substrate/cover 16-20 11-156-10 0-5
2. Embeddedness 16-20 11-156-10 0-5
3. Water velocity 16-20 11-156-10 0-5
SECONDARY INFLUENCE
Channel Morphology
4. Channel Alter 12-158-11 4-7 0-3
5. Scour/Deposition 12-15 8-11 4-7 0-3
6. Pool/Riffle Ratio 12-158-11 4-7 0-3
TERTIARY INFLUENCE
Riparian and Bank Structure
7. Bank stability 9-10 6-8 3-5 0-2
8. Bank vegetation 9-10 6-8 3-5 0-2
9. Bank cover 9-10 6-8 3-5 0-2
cedure (HEP; Plafkin et al. 1989) adapted from
Platts et al. (1987). HEP quantifies 9 habitat
characteristics summarized in Table 2. The
total score for each site was based upon data
from 10 transects at each site spaced 25, 50,
or 100 feet apart.
In addition, several other physical
measurements were taken whenever fish
collections were made during special
longitudinal surveys. Stream turbidity was
measured with a secchi disc or a Minispec20
nephelometer. Water temperature and
dissolved oxygen were measured using a YSI
meter. Water velocity was measured with a
Gurley pygmy meter. ALI distances were
measured optically using a Leitz rangefinder.
Estimates of the amount of woodland were
based on conventional analyses of enlarged
LandSat infrared photographs taken on May 2,
1981. These were obtained from the U.S.
87
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Gammon and Gammon
Geological Survey (ESIC), EROS Data Center,
Sioux Falls, SD.
The drainage area perimeter was determined
using topographic maps of tributaries. This
scaled map was superimposed over the
infrared photographs on a light table. Plots of
land with permanent tree cover were outlined
on the topographic map.
Using a light table, the marked topographic
map was traced onto a fine grid. Individual
grids with more than 50% woodland was
calculated. Land use in a few tributaries was
not determined because of insufficient
coverage of LandSat infrared photographs.
Results
A total of 6,635 fish comprising 46 species
were captured by electrofishing and seining.
Forty species and 4154 individuals (63%) were
taken by seining alone. Electrofishing catches
also yielded 40 species, but only 2481
individuals or 37% of the total.
Bluntnose minnow (Pimeohales notatus) was
very common comprising 40.9% of the total
number seined, while sand shiner (Notroois
stramineusl. spotfin shiner (N. soilopterus).
striped shiner (N. chrvsocephalus). silverjaw
minnow (Ericvmba buccata). and creek chub
(Semotilus atromaculatusl together contributed
another 37%.
The electrofishing catch was more evenly
distributed with common shiner (N. cornutus)
and white sucker (Catostomus commersoni)
each contributing about 15% to the catch.
Substantial numbers of the following were also
collected: creek chub (9.3%), bluntnose
minnow (9%), rock bass (Ambloolites
ruoestris: 7.4%), and northern hog sucker
(Hvpentilium nioricans: 7.1%).
Smallmouth bass (Microoterus dolomeiui)
adults and subadults were mostly found in the
lower 50 miles of the Eel River and only in Paw
Paw and Twelve Mile Creeks. Catch rates
were higher in the lower 30 miles of river and
attenuated from RM 30 to RM 51.7. Three of
12 smallmouth bass 250 mm and longer were
fin clipped, indicating that they were stocked
fish. Two of these were collected by
electrofishing at RM 37.8(1) near Roann and
the other at RM 27.3(68) near Chili. Young-of-
the-year smallmouth bass were taken only in
the extreme lower part of the Eel River and in
Paw Paw and Twelve Mile Creeks.
Largemouth bass (M. salmoides) formed a
minor component of the catch. Fair numbers of
small spotted bass (M- punctulatusl were
scattered throughout the mainstem Eel and in
Paw Paw and Twelve Mile Creeks. This species
had not been present since they could easily
have been misidentified as small largemouth
bass. Spotted bass young-of-the-year were
found even in the otherwise poorer habitat of
the upper 30 miles above South Whitley. This
species has been shown to be tolerant of high
turbidity and sedimentation (Gammon 1970).
Rock bass was taken at all stations except
Squirrel Creek. Longear sunfish (Lepomis
meaalotis) were most common at the upper
mainstem stations and in the Blue River and
were sporadic in the lower river. Green sunfish
(L- cvanellus) also occurred at most sites, but
was more abundant in the upper mainstem and
in the Blue River. Substantial numbers of
bluegill (L. macrochirusl were also taken more
regularly in the upper mainstem Eel from RM
63.5 to RM 80.
The most abundant catostomid was white
sucker with greatest numbers in the upper
mainstem from RM 63.5 to RM 80 and in the
Blue River, Sugar Creek, and Beargrass Creek.
They were uncommon in the lower 60 miles of
the mainstem. Northern hogsucker was widely
distributed throughout the mainstem and most
tributaries. Spotted sucker (Minvtrema
melanops) was found in good numbers only in
the pool above the Logansport dam.
88
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Agricultural Impacts in the Eel River
ao
75
70
65
60
55
SO
45
40
35
30
25
IBI and Iwbxtt
a a
0 10 20 30 40 50 60 70 80
River Mil*
-&-1990IBI
-1990lwb
Modlwb
IBI baaad upon Sugar Craak erllarla and
Inoludn aakilng data.
Mod. Iwb daMaa 'lotoraitt' apaatoa
55
50
45
40
35
30
25
IBI
D O
0 10 20 30 40 50 60 70 80
River Mil*
••°- 1972IBI -*- 1982IBI -&- 1990IBI
Baa«d upon Sugar Craak erltarla.
Fig. 1. IBI, Iwb, and modified Iwb values for
1990 Eel River fish communities.
Fig. 2. IBI profiles of mainstem Eel River fish
communities for 1972, 1982, and 1990.
Golden redhorse (Moxostoma ervthrurum, was
the most common of the three redhorse
species, but it was not all that abundant. It
was absent between RM 56.5-80, as well as,
from all tributaries including Blue River. Black
redhorse (Moxostoma duauesnei) was almost
as common as golden redhorse, but was
mostly restricted to the lower 30 miles of the
mainstem. Greater redhorse (Moxostoma
valenciennesil is a rare species throughout
Indiana and most of its range, but a healthy
population thrives in the Eel River system. It
was particularly abundant in the lower 20
miles of river, but was also found in Paw Paw
and Squirrel Creeks.
The distribution of smaller species of minnows
and darters is best illustrated by the seining
catches. Bluntnose minnow (Pimeohales
notatus) was the dominant species, occurring
throughout the mainstem and tributaries.
Common shiner was even more frequently
encountered by electrofishing and was also
widely distributed throughout the Eel River
system.
Spotf in shiner and sand shiner mostly occurred
in the lower 50 miles of the mainstem. Creek
chub was common only in the tributaries.
Redfin shiner (Notropis umbratilus) and
rosyface shiner (Notroois rubellus) were most
common in the lower river, but also occurred in
Sugar and Twelve Mile Creeks. River chub
(Nocomis micropoaonl was regularly taken by
seine and electrofishing mostly downriver from
RM 65. A few bigeye chub (Hyboosis amblopsl
were also present in the lower river.
Among the darters, only johnny darter
(Etheostoma nigrum) was common and
widespread. Blackside darter (Percina
maculata). greenside darter (E. blennioides) and
eastern sand darter (Ammocrvpta pellucida)
were found only in the lower river. Dusky
darter (P. sciera) was taken only from upper
Eel River (RM 88.0) and Beargrass Creek.
89
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Gammon and Gammon
130
110 •
HEP tcor*
0 B 12 18 24 30 36 42 48 54 BO 66 72 78
70
SO
30
- IBI IcorM * TributariM
Fig. 3. IBI and HEP values for the mainstem Eel
River and tributaries.
NTU
10 20 30 40 SO 60 70 10 90 100
M«in»t*m
CIS Tributaries
Fantail darter (E. flabellare) was collected only
at RM 63.5). Mottled sculpin (Cottus bairdi)
was taken only at RM 56.5 and RM 63.5).
Important community index values are
summarized in Table 3. IBI values were also
calculated on less extensive data sets provided
by Braun et al. 1984,1986) on five collections
of fish from each of three stations; 2B (RM
3.3), 3B (RM 8.3) and 3 (RM 46.4) during the
years 1984 and 1985. The mean IBI values at
stations 2B, 3B, and 3 were 39.6, 42.0, and
43.6 in 1984 and 43.2, 41.2, and 42.9 in
1985, respectively.
The IBI and Iwb profiles for the Eel River
mainstem are shown in Figure 1. An additional
modified Iwb is also shown, wherein four
tolerant species were deleted prior to
calculation, carp, bluntnose minnow, creek
chub, and green sunfish.
All three profiles indicate somewhat depressed
fish communities in the lower river, probably
because of the ponding effect of the dam,
followed by relatively good communities from
RM 8 to RM 25. From RM 30 to RM 80 there
is considerable variation from place to place,
but the communities are generally depressed,
especially at RM 70.
In Figure 2, the 1990 IBI profile is repeated and
compared to IBI profiles based on Taylor's
(1972) and Braun et al.'s (1982) series of
collections. The 1990 fish communites are
clearly much better than they were in 1982.
However, both profiles indicate better
communities in the lower river than in the
upper river. In 1972 there was less difference
in the lower mainstem but equal variation
between stations.
Fig. 4. Turbidity (NTU) for the mainstem Eel
River and its tributaries on July 16 and 17,
1990.
Habitat Evaluatipn
Habitat scores were generally lower in the
upper part of the watershed and higher
90
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Agricultural Impacts in the Eel River
NTU
70
60
so
40
30
20
10
0 5 10 15 20 25 30 35 40
% woodland
Turbidity deteratkiee' en 7-M ( IT. 1MO.
Figure 5. Turbidity (NTU) of Eel River
tributaries in relation to percent woodland in
their drainage basins.
Speciea
Rainbow darter r-
Roayface ahmer IE
Silver ahiner i
Blackaide darter Si
Qreenaide darter
Rock baaa
Smallmouth baaa
Longear aunfiah ai
Sculpin •
Stonecat F
Orangetnreat darter p
bluebreaat darter p
20
40 60
% frequency
80
10O
EH3 1940-41
11990
1940-41: 6 aialnetaei • 4 trlbutarlee
19*0: 12 aialnalen • e tributaries
Figure 6. Frequency of occurrence of some
species collected by seining in 1990 compared
to 1940-41.
IBI
downstream. Upstream from South Whitley,
habitat features were uniformly low in quality
and homogeneous because of past
channelization and recent deforestation of both
banks. Habitat scores of tributaries were
generally higher than the mainstem reaches
into which they flowed (Figure 3). An
excpetion was Paw Paw Creek which was
somewhat lower.
Turbidity and Landuses
Turbidity determination during the synoptic
surveys on July 16-17, 1990 are portrayed
graphically in Figure 4. Scattered showers fell
throughout northern Indiana during the week
previous to the turbidity determinations. It is
not known to what extent these results may
be affected by differential rainfall. Tributaries
which were distinctly more turbid than others
included Squirrel Creek, Otter Creek, Simonton
Creek, Hurricane Creek, Blue River, Solon
Ditch, and Johnson Ditch. A bridge was under
construction in the Simonton watershed, but
55
50
45
40
35
301 I I I I I I I I I I I I I I I I I I I I I I I
1975 1980 1985 1990 1995
Year
•If HM0**n Crfe
0 Eel River
+ Sugar Creek
O Big Raccoon crk.
IBI Meed uvea Sneer Craek criteria.
Figure 7. Changes in the IBI values for the Eel
River compared to Big Raccoon Creek and
Sugar Creek.
91
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Gammon and Gammon
Table 3. Fish community indicies for Eel River stations.
Station 1 990
Taylor
1B
2B
3B
4B
5B
6B
7B
1
2
3
4
5
6
7
8
11
RM
1.0
3.3
8.3
12.0
19.0
27.3
32.0
37.8
41.4
46.4
51.7
56.5
63.5
66.0
70.3
79.8
Iwb
5.8
6.4
6.7
7.5
5.3
7.5
5.4
6.7
6.1
7.2
5.4
7.2
6.6
6.6
6.6
6.7
No.
1972
12
14
10
17
14
11
17
14
18
16
10
18
13
15
12
19
. Soec. elec.
1982
13'
14
11
14
17
12
10
8
9
11
12
13
8
9
6
9
1990
15
14
12
17
18
16
—
10
17
13
16
10
13
13
16
17
1972
38
44
40
44
38
40
44
38
42
46
40
44
36
36
39
42
IBI
1982
44
42
38
40
40
44
36
32
34
36
40
36
36
32
28
32
1990
44
44
50
50
46
42
—
42
44
44
46
40
42
40
32
40
Tributary Stations
Twelvemile Creek
Paw Paw Creek
Squirrel Creek
Beargrass Creek
Sugar Creek
Blue River - upstream from Columbia City
Blue River - downstream from Columbia City
44
40
40
40
40
40
44
animals were also pastured in the stream and
some corn fields extended to stream banks.
The turbidity of mainstem water was high in
the upper river mainly because of highly turbid
Johnson ditch. The water cleared considerably
after passing through two mainstem gravel pits
at RM 84 then again became progressively
more turbid as it flowed downstream.
During this same period the turbidity gradually
increased in the mainstem from the upper river
to the lower river, although there were
localized sharp increases in turbidity
downstream from both Johnson ditch and
South Whitley. Earlier in the summer (June 12,
1990) when water levels were higher the
turbidity (NTU) was 45 in the lower 60 km (40
mi) of river and between 46-48 in the upper
river. In some streams lateral erosion can be a
major source of sediment and turbidity, but
scoured banks were a very limited component
of the lower portions of the Eel River
mainstem. However, they were highly evident
in the channelized upper portions. The entire
upper 50 km (33 mi) of the Eel River has been
stripped of its trees and bushes along both
banks. During this study the trimmings had
92
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Agricultural Impactsinthe Eel River
been removed from the river and were piled
along the shore for burning.
Woodlands were readily determined from the
infrared photographs, but other kinds of
permanent vegetation such as brushlands,
pastures, and winter wheat were
indistinguishable from one another, estimates
of woodland ranged from only 7.0% in the
Beargrass Creek watershed to 40.9% in the
Weesaw Creek watershed. There was a greater
percentage of landuse in agriculture south of
the mainstem and in the upper two-thirds of
the Eel River watershed than north of the
mainstem and in the lower third. There was an
inverse relationship between the percentage of
tributary watersheds in woodland and the
measured turbidity (Figure 5).
Discussion
The Eel River in 1990 was found to- support
fairly diverse fish communities throughout
most of the watershed, although the upper
reaches had depressed populations and
reduced numbers of species. Many species of
juvenile fish were caught, with larger numbers
at stations 1B and Twelve Mile Creek. This is
an indication that reproduction for many
species was successful during the past couple
of years.
Several usually common species which Braun
and Robertson (1982) did not collect were
found in good numbers in 1990: river chub,
bigeye chub, several species of shiners
including silver shiner, spotfin shiner, rosyface
shiner, redf in shiner, and blackside and johnny
darters.
Some species present in 1972 were found only
rarely or not at all in 1990. These species
included mottled sculpin, blacknose dace,
unidentified madtom species, suckermouth
minnow, largemouth bass, and carp.
It is difficult to evaluate long-term changes in
abundance of any single species of fish
because of the different collecting
methodologies. The comprehensive study of
Gerking (1945) used the seine as the primary
collecting gear and our effort in 1990 was
comparable. Gerking collected from five
mainstem sites and four tributaries. We
collected from 12 mainstem sites and six
tributaries.
A comparison of percent frequency of
occurrence from these studies indicates rather
drastic reductions for many species
populations of sediment sensitive fish (Figure
6). Rock bass, johnny darter, and eastern sand
darter appear to be distributed much as they
were 50 years ago. However, many species
have suffered drastic declines including
rainbow darter (E. caeruleum), orangethroat
darter (E. soectabile). and bluebreast darter (E.
camuruml which may have been totally
eliminated from the river.
Changes over time in populations of clams and
mussels parallel those of fish. Henschen
(1988) concluded that while the Eel River once
supported a diversity of mussel species
throughout its length, its currently reduced
population is mostly confined to the lower river
in Cass and Miami Counties.
Changes in the Fish Community over Time
The IBI offers one way of addressing questions
about how the overall fish community has
changed over time and how it compares to fish
communities in other streams. The mean IBI
values for the Eel River mainstem stations
declined from 40.7 in 1972 to 36.9 in 1982.
This increased substantially to 43.1 in 1990.
The IBI values estimated from data of the
studies of Braun et al (1984,1986) and Braun
(1990) generally corresponded to improving
trend noted in the 1980's.
The overall Eel River fish community appears
to have improved rapidly from the degraded
community found in 1982.
93
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Gammon and Gammon
60
50
0340
30
20
1 1 1 — 1 — 1 — 1 — 1 — 1 — 1 — 1 — 1 — 1 — 1 — 1— 1 — 1 — 1 — 1 — 1 —
o "*"
O D 1990 O
0 + 1988 ^ '
++1.°
H" Toflfi
-------
Agricultural Impacts in the Eel River
Nevertheless, a series of years with high runoff
and increased non-point source pollution can
depress fish communities to equally low levels.
Other factors which may modify the recovery
of fish populations in the Eel River system
include the absence of high quality tributaries
to serve as refugia for sensitive species during
unfavorable years and the dam blockage at
Logansport which reduces recolonization by
species from the Wabash River.
The Potential Influence of Habitat and Turbiditv
Much of the upper Eel River is characterized by
low HEP values, e.g. channelized stream beds,
poor riffle/pool development, and a lack of
instream structure. In addition, riparian trees
have been removed recently from many older
previously channelized sections of the river.
The bottom substrate usually included much
fine sediment, as indicated by the low
embeddedness scores for almost all of the
mainstem stations and most tributaries.
Turbidity was high for virtually the entire
summer. At Roann we saw a layer of mud two
centimeters deep on top of a flat boulder after
higher water had subsided in a pool.
The lower 48 km (30 mi) of Eel River contained
much better habitat than the upstream
reaches. Beds of water willow (Dianthera)
were mostly limited to the lower 64 km (40
mi) of the mainstem Eel. This section also had
fairly good riparian protection and good
instream habitat.
Habitat in the tributaries generally scored
higher than the mainstem. Twelve Mile Creek,
with 26.5% of its watershed in forest,
contained the best habitat, followed by Squirrel
Creek. The Blue River is approximately the
same size as the Eel River where the two
streams converge. With only 11 % permanent
vegetation cover, its turbidity readings were
among the highest recorded. Fish from this
stream, and Paw Paw Creek, were commonly
infected with blackspot disease (Simon 1989).
Potential Negative Effects from Point-Source
Pollution
Agricultural point-source pollution in Indiana
often occurs because of accidents or careless
handling of animal wastes and farm chemicals.
Spilled materials, animal wastes applied to
fields, and the contents of waste holding
lagoons may be flushed into ditches and
streams following rain storms. Fish kills
reported to the Indiana Department of
Environmental Management (IDEM) since 1969
include five incidents on Paw Paw Creek and
single kills on Twelve Mile, Pony, Beargrass,
and Clear Creeks.
There were 39 additional reports of spilled
materials which are not known to have
resulted in fish kills, but which may have
exerted sublethal damage. Most of the
materials were fertilizer and animal wastes,
which include wastes generated by chickem,
turkey, veal, and swine rearing operations.
All of the known causes of ifsh kills and most
of the spills reported within the Eel River
watershed are agriculturally based. The actual
number of fish kills and spills is unknown, but
would certainly far exceed the number of
reported cases.
In the decade following passage of the Clean
Water Act of 1972, it was estimated that
municipal BOD loads decreased by 46% and
industrial BOD loads decreased at least 71%
(U.S. Environmental Protection Agency, 1982).
Most of the communities in the area have
improved waste treatment and reduced BOD
concentrations by at least 50%. Some
previously unsewered communities now have
a central treatment system. It is likely that any
negative influences from these point sources
are masked by the magnitude of non-point
source impacts.
Weather and Nonooint Source Pollution
Unlike point source pollution, nonpoint sources
of influence such as occurs from plowed fields
are most severe during storm events. The
95
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Gammon and Gammon
discharge of rivers is roughly proportional to
the amount of rainfall, hence, non-point
sources are most severe when rainfall is great
and river discharge is high. Conversely, non-
point sources are reduced during periods of dry
weather. Fish populations are negatively
affected by non-point sources during the
reproductive period and in the months
immediately after hatching, spring and
summer.
From October 1974 through September 1980
the U.S. Geological Survey Water Resources
Division determined daily sediment loads for
the Eel River near Logansport (Anonymous
1974 through 1980). Data from the Eel River
and rivers throughout Indiana is analyzed and
discussed by Crawford and Mansue (1988).
They estimated that for the Eel River the mean
annual suspended sediment yield was 178
tons/square mile/year and the flow-weighted
mean annual suspended sediment
concentration was 89 mg/l (median = 53
mg/l). These values are high for the northern
moraine/lake portion of Indiana which
Crawford and Mansue found to have the
lowest sediment yield. Only that part of the Eel
River watershed north of the mainstem resides
within the moraine area. The portion of the
watershed situated south of the mainstem is
located in the Tipton Till Plain where both
parameters were generally much larger.
Monthly data from May through August for the
years 1974 through 1980 was used for
regression analysis of suspended solids
concentration on discharge. The regression
equation obtained was then used to estimate
the suspended solids concentration for the
months May through August for the years
following 1980 (Figure 9). Suspended solids
concentrations were highest during May and
June when relatively high discharges occurred
during half of the years since 1974. "Wet"
summers of relatively high suspended solids
concentration include the years 1974, 1975,
1980,1981,1982, and 1986. "Dry" summers
when Eef River water was relatively clear
include the period from 1976 through 1979,
1983 through 1985, and 1987 through 1988.
During "dry" summers the effects of point
sources of pollution such as from population
centers would theoretically increase, but
nonpoint source pollution should be less than
normal. For streams influenced mostly by NFS
the fish communities following a sequence of
"dry" summers should improve. The Eel River
fish communities did improve somewhat, but
less than might have been expected compared
to fish communities in Big Raccoon Creek.
The 1990 fish communities may be as good as
the Eel River is able to support considering
present land use. The summers of 1989 and
1990 were relatively "wet". Therefore,
reproductive success and survivorship through
the first year of life would be expected to be
lower than normal. It is likely that the 1991
fish communities will be poorer than they were
in 1990 and the prognosis for improvement in
the future is bleak unless changes in land-use
are implemented.
Summary
The Eel River is essentially a linear stream. Its
drainage basin is long and narrow and its
tributaries are generally small first and second
order streams. Improving landuse in these
tributaries will be necessary in order to improve
the mainstem of the Eel River. Thorough
surveys of all tributary watersheds should be
conducted using both Geographic Information
System (GIS) technology and ground study.
Twelve Mile Creek, Paw Paw Creek, and,
possibly, Squirrel Creek appear to be less
influenced by agriculture than other tributaries.
These tributaries may act as refugia for
sensitive species during periods of stress and
serve as species reservoirs to replenish the
mainstem during more benevolent times. They
should receive special attention to ensure that:
a) the streamside riparian buffer zone is
96
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Agricultural Impacts in the Eel River
maintained, b) tilled fields do not impinge on
the stream itself, c) hogs and cattle are not
pastured directly in the streams, d) appropriate
forms of conservation tillage are encouraged,
e) animal wastes are properly disposed.
Several other tributaries appear to be more
environmentally degraded than others. Otter
Creek, Simonton Creek, Hurricane Creek, Blue
River, Solon Ditch, and Johnson Ditch
delivered higher than average sediment loads
to the Eel River during the survey of July 16
and 17, 1990. While this survey is only a brief
"snapshot" in time, it nevertheless suggests
that these streams may have greater than
average negative impacts on the Eel River
system. They should also receive the same
items of attention listed above.
Streams in the upper watershed are referred to
and used as drainage ditches. Nevertheless,
these streams are permanent "creeks" and
should support normal aquatic life. Their
rehabilitation could contribute positively
toward the improvement of the lower
mainstem. The creation of a "green belt"
riparian corridor would also contribute toward
a greater ecological diversification.
Acknowledgements
The Eel River study was funded by the Indiana
Department of Environmental Management.
Undergraduate students D. Wallace and M.
Giesecke acted as the primary field crew.
Assistance in collecting fish was provided by
J. Riggs, C. Hansen, J. Hecko, and N. Masten,
who were supported by grants from Eli Lilly
and Company and PSI Energy. The Big
Raccoon Creek data was collected through a
grant from Heritage Environmental Services,
Indiana.
The study benefitted significantly through
consultation with J. Ray, C.L. Bridges, and
R.J. Gammon of the Indiana Department of
Environmental Management (IDEM) and E.
Braun and T. Stefanavage of the Indiana
Department of Natural Resources (IDNR). This
contribution is dedicated to the successful
completion of the first year of life of Robert
Wayne Pitman-Gammon, who is considerably
more alert, lively, and curious than he was a
year ago and to his parents.
Literature Cited
Anonymous. 1975-1981. Water resources
data for Indiana. U.S. Geological Survey Water-
Data Reports for Water Years 1974 through
1980.
Braun, E.R. 1990. A survey of the fishes of the
Eel River in Wabash and Miami Counties,
Indiana 1989. Indiana Department of Natural
Resources, Division of Fish and Wildlife, 607
State Office Building, Indianapolis, Indiana
46204. 30 pp. mimeo.
Braun, E.R. and R. Robertson. 1982. Eel River
watershed fisheries investigation 1982. Indiana
Department of Natural Resources, Division of
Fish and Wildlife, 607 State Office Building,
Indianapolis, Indiana 46204. 60 pp. mimeo.
Braun, E.R., R. Robertson, and T. Stefanavage.
1984. Evaluation of smallmouth bass stocked
in the Eel River 1984 progress report. Indiana
Department of Natural Resources, Division of
Fish and Wildlife, 607 State Office Building,
Indianapolis, Indiana 46204. 47 pp. mimeo.
Braun, E.R., R. Robertson, and T. Stefanavage.
1986. Evaluation of smallmouth bass stocked
in the Eel River 1985 progress report. Indiana
Department of Natural Resources, Division of
Fish and Wildlife, 607 State Office Building,
Indianapolis, Indiana 46204. 84 pp. mimeo.
Crawford, C.G. and L.J. Mansue. 1988.
Suspended sediment characteristics of Indiana
streams, 1952-84. U.S. Geological Survey,
Open File Report 87-527. 79 pp.
Gammon, C.W. and J.R. Gammon. 1990. Fish
communities and habitat of the Eel River in
relation to agriculture. A report for the Indiana
97
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Gammon and Gammon
Department of Environmental Management,
Office of Water Management, Indianapolis, IN
74pp.
Gammon, J.R. 1970. The effect of inorganic
sediment on stream biota. Water Pollution
Control Research Series 18050DWC 12/70:1-
141.
Gammon, J.R. 1980. The use of community
parameters derived from electrofishing catches
of river fish as indicators of environemntal
quality, pp. 335-363 in Seminar on Water
Quality Management Tradeoffs. U.S.
Environmental Protection Agency, Washington,
D.C. EPA 905/9-80-009.
Gammon, J.R. and J.R. Riggs. 1983. The fish
communities of Big Vermilion River and Sugar
Creek. Proceedings Indiana Academy of
Science 92: 183-190.
Gammon, J.R., C.W. Gammon, and M.K.
Schmid. 1990. Land use influence on fish
communities in central Indiana streams, pp.
111-120. in W.S. Davis (editor). Proceedings
1990 Midwest Pollution Control Biologists
Meeting. U.S. Environmental Protection
Agency, Region V, Environmental Sciences
Division, Chicago, IL. EPA 905/9-90-005.
Gammon, J.R., C.W. Gammon, and C.E.
Tucker. 1990. The fish communities of SUgar
Creek. Proceedings Indiana Academy of
Science 99: in press.
Gammon, J.R. 1990. The fish communities of
Big Raccoon Creek 1981 -1989. A report for
Heritage Environmental Services, One
Environmental Plaza, 7901 West Morris Street,
Indianapolis, IN 46231. 120 pp.
Henschen, M. 1988. The freshwater mussels
(Unioinidae) of the Eel River of northern
Indiana. Indiana DNR, Division of Fish and
Wildlife, 607 State Office Building,
Indianapolis, IN 46204. 73 pp. mimeo.
Hoggart, R.E. 1975. Drainage areas of Indiana
streams. U.S. Geological Survey, Water
Resources Division, Indianapolis, IN. 231 pp.
Karr, J.R. 1981. Assessment of biotic integrity
using fish communities. Fisheries 6: 21-27.
Karr, J.R. 1987. Biological monitoring and
environmental assessment: a conceptual
framework. Env. Management 11: 249-256.
Karr, J.R., K.D. Fausch, P.L. Angermeier, P.R.
Yant, and I.J. Schlosser. 1986. Assessing
biological integrity in running waters: a method
and its rationale. Illinois Natural History Survey
Special Publications 5, Urbana.
Karr, J.R., P.R. Yant, K.D. Fausch, and I.J.
Schlosser. 1987. Spatial and temporal
variability of the index of biotic integrity in
three midwestern streams. Transactions
American Fisheries Society 116: 1-11.
Miller, D.L., P.M. Leonard, R.M. Hughes, J.R.
Karr, P.B. Moyle, L.H. Schrader, B.A.
Thompson, R.A. Daniels, K.D. Fausch, A.
Fitzhugh, J.R. Gammon, D.B. Halliwell, P.L.
Angermeier, and DJ. Orth. 1988. Regional
applications of an index of biotic integrity for
use in water resource management. Fisheries
13: 12-20.
Plafkin, J.L., M.T. Barbour, K.D. Porter, S.K.
Gross, and R.M. Hughes. 1989. Rapid
Bioassessment Protocols for use in streams
and rivers: benthic macroinvertebrates and
fish. EPA 444/4-89-001.
Platts, W.S., C. Armour, G.D. Booth, M.
Bryant, J.L. Bufford, P. Cuplin, S. Jensen,
G.W. Lienkaemper, G.W. Minshall, S.B.
Monsen, R.L. Nelson, J.R. Sedell, and J.S.
Tuhy. 1987. Methods for evaluating riparian
habitats with applications to management.
U.S. Department of Agriculture, Forest Service,
Intermountain Research Station, General
Technical Report INT-221.177 pp.
98
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Agricultural Impacts in the Eel River
Simon, T.P. 1989. Biological Survey of the
instream fish and water quality evaluation of
Wayne Reclamation and Recycling, Whitley
County, Indiana. U.S. Environmental Protection
Agency, Central Regional Laboratory, Chicago,
IL. 60605. 19 pp. mimeo.
Taylor, M. 1972. Eel River watershed fisheries
investigations report 1972. Indiana Department
of Natural Resources, Division of Fish and
Wildlife, 607 State Office Building,
Indianapolis, Indiana 46204. 65 pp. mimeo.
Trautman, M.B. 1981. The fishes of Ohio.
Ohio State University Press, Columbus. 782
PP.
U.S. Environmental Protection Agency. 1982.
National water quality inventory: 1982 report
to Congress. Washington, D.C. 63 pp.
99
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Selenastrum Algal Growth Test: Culturing and Test Protocol at the Illinois EPA
Michael J. Charles and Greg Searle
Illinois EPA
Office of Ecotoxicology, if 31
2200 Churchill Road
P.O.Box 19276
Springfield. IL 62794-9276
Abstract
The availability of quality test organisms is of fundamental concern in conducting any regularly
scheduled biomonitoring activities. The Selenastrum algal assay requires a continuous supply of pure
log-phase algae. The most convenient means to meet this demand is through the establishment of
in-house cultures. Upon receipt of a pure Selenastrum "starter" culture from an outside source,
laboratory stock cultures are initiated. The algae is aseptically transferred to a series of culture flasks
containing synthetically prepared algal medium. Once pure algal cultures have been established, a
regime of routine cell transfers will provide the laboratory with a steady supply of log-phase algal
cells suitable for testing purposes. Back-up reserve cultures are stored on agar slants and plates.
Testing of municipal and industrial effluents at the Illinois EPA using Selenastrum algae follows
USEPA test protocol. Through experience running the test and repeated attempts to get confident
results, the testing has been refined and the integrity of the analysis is ensured. Various techniques
are employed in the testing that serve to tighten the USEPA protocol and may be of interest to other
regulatory bioassay personnel.
Keywords: Selenastrum algae, log-phase, aseptically, agar slant, agar plate.
Introduction
At the present time, the Illinois EPA (IEPA) is
the only state run bioassay laboratory in USEPA
Region V conducting the Selenastrum caori-
cornutum algal growth test. The algal growth
test is conducted in the Toxicity Testing Unit
(TTU) which is one of two units in the Office of
Ecotoxicology (OE). OE serves as a support
laboratory for the various control divisions
within the IEPA (air, land, water, and public
water supplies). IEPA uses USEPA protocol.
Short Term Methods for Estimating the Chronic
Toxicity of Effluents and Receiving Waters to
Freshwater Organisms (USEPA 1989) as a
guideline for culturing and conducting tests.
TTU maintains a continuous supply of in-house
stock cultures for use in bioassays. Various
techniques are employed in culturing and
testing that serve to tighten USEPA protocol.
Establishing and Maintaining Selenastrum Stock
Cultures
The Selenastrum algal assay calls for healthy
log-phase-growth cells which are harvested
from resident in-house stock cultures. The TTU
laboratory maintains a continuous supply of
these log-phase Selenastrum cultures in
quantities sufficient to meet all testing
demands. The culturing process is relatively
straight-forward and consists of six basic
components: 1) general culture setup/
conditions; 2) algal nutrient culture medium; 3)
aseptic technique; 4) routine cell transfers; 5)
back-up/reserve cultures; and, 6) quality
assurance/quality control considerations.
1. General Culture Setup/Conditions.
The Selenastrum cultures are maintained in an
environmental chamber at 25 ± 1°C under a
continuous "cool-white" fluorescent illumination
of 400 ± 40 ft-c (4306 ± 431 lux). The algal
cells are kept in a constant state of suspension
through the use of a mechanical shaker at
approximately 100 cpm (cycles per minute).
The culture flasks are arranged on the shaker
table and allowed to incubate anywhere from
four to seven days, depending upon current
testing schedules. This incubation interval
100
-------
IEPA Selenastrum Culturing and Test Protocol
Table 1. Nutrient Stock Solutions For
Maintaining Algal Stock Cultures (adapted from
USEPA 1989).
Nutrient
Stock Compound
Solution
Amount dissolved
in 500 ml
Distilled H20
1
2
3.
4
MgCI2'6H20
CaCI2'2H20
H3BO,
MnCI2'4H20
ZnCI2
FeCI3'6H20
CoCI2-6H20
Na2Mo042H20
CuCI22H20
Na2EDTA2H20
NaN03
MgSCy7H20
K2HP04
NaHC03
6.08 g
2.20 g
92.8 mg
208.0 mg
1.64mg'
79.9 mg
0.714 mgb
3.63 mg°
0.006 mgd
1 50.0 mg
12.75 g
7.35 g
0.522 g
7.50 g
•ZnCI2 - Weigh out 164 mg and dilute to 100
ml. Add 1 ml of this solution to Stock #1.
bCoCI2-6H2O - Weigh out 71.4 mg and dilute to
100 mL. Add 1 mL of this solution to Stock #1.
cNa2MoCy2H20 - Weigh out 36.6 mg and dilute
to 10 mL. Add 1 mL of this solution to Stock
#1.
dCuCI2'2H2O - Weigh out 60.0 mg and dilute to
1000 mL. Take 1 mL of this solution and dilute
to 10 mL. Take 1 mL of the second dilution and
add to Stock #1.
provides plenty of viable, log-phase
suitable for bioassay purposes.
cells
2. Algal Nutrient Culture Medium.
Upon receipt of a Selenastrum "starter" culture
from an established outside source, in-house
stock cultures are initiated by aseptically
transferring a portion of the cells to freshly
prepared algal nutrient medium. The culture
medium consists of a mixture of various macro-
and micronutrients prepared in four separate
stock nutrient solutions using the reagent grade
chemicals listed in Table 1.
The nutrient medium is prepared by adding 1
mL of each of the four stock solutions, in order
as listed in Table 1, per liter of distilled water.
The solution is mixed well and then pH-
adjusted to 7.5 ± 0.1 by dropwise addition of
0.1 N NaOH or HCL, as appropriate. The
medium is then immediately filtered through a
pre-washed 6.2 /jm pore diameter membrane at
a vacuum pressure of approximately 8 psi. The
algal nutrient medium is then ready to be
dispensed into the various culture flasks and
inoculated as needed. Any leftover portions of
the sterile medium may be stored in a
refrigerator at 4°C until needed. Care should be
taken, however, to seal off the storage vessel
well so as to prevent loss of water by
evaporation. Evaporation losses will alter the
concentration of macro-micronutrients in the
final medium, thus compromising its quality for
use in culturing purposes.
3. Aseptic Technique.
Extreme care is exercised to prevent
contamination of the cultures by other
microorganisms. All glassware products used in
the culturing process are thoroughly cleaned,
sealed with aluminum foil, and sterilized at
121°C in an autoclave. All pipet tips used in
handling the algal cells during routine cell
transfer procedures are of the disposable type,
and they too are autoclaved at 121°C. The algal
nutrient culture medium is cold-sterilized before
use by passing it through a 0.2 jjm pore
diameter membrane filter, as described above.
Despite these efforts, contamination problems
do occur from time to time. Contaminated
cultures are either discarded or used as food for
Ceriodaphnia cultures.
4. Routine Cell Transfers.
To meet scheduled testing demands, the TTU
laboratory maintains a continuous source of
101
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Charles and Searle
log-phase Selenastrum cells. This is achieved
through a series of routine cell transfers from
existing stock cultures to various afiquots of
fresh algal nutrient medium. An inoculum is
prepared from a four - to seven - day stock
culture by concentrating the cells of the culture
through a centrif ugation process. The algal cell
concentrate is then diluted with distilled water
to provide an initial density of approximately
10,000 cells/mL in the culture flasks. A Coulter
Counter" (model ZM) is utilized in cell density
determinations for both the stock cultures and
the final inoculum.
Once prepared, 1 mL of inoculum is aseptically
transferred to each of three 500 mL Erlenmeyer
culture flasks containing 250 mLs of fresh algal
medium each. After inoculation, the flasks are
situated in the environmental chamber on a
mechanical shaker for incubation purposes. An
incubation period of four to seven days renders
plenty of healthy, log-phase Selenastrum cells
ready for harvest and use in testing and/or
other purposes. Routine cell transfers are
carried out twice per week, with each transfer
staggered 3*4 days apart. This arrangement will
provide a continuous supply of log-phase cells
suitable for biomonitoring purposes.
The volume of stock cultures required depends
on the test loads involved and any other
targeted uses for the algae (ie., food source for
Ceriodaphnia cultures, etc.). The TTU
laboratory meets all of its current algae
demands by inoculating 1.5 - 2 L of fresh
culture medium weekly.
5. Back-Up/Reserve Cultures.
As mentioned above, contamination of the
stock cultures seems to be inevitable from time
to time. It is therefore essential to have in place
some type of a back-up/reserve system for
storing clean, pure Selenastrum stocks that
may be called upon to rejuvenate "dirty" or
"fouled" cultures. The TTU laboratory meets
this objective through the use of a system of
agar slants and plates. The agar medium is
prepared with the same stock nutrients, in the
same amounts, as the standard liquid algal
medium. The only difference is that the stock
nutrients are dissolved in a 1-2% BactoRAgar
solution. The agar nutrient medium is mixed up
in an AgarMatic^bench top agar sterilizer, which
in turn is linked up to a PourMatic™automatic
plate dispensal system. Thus, the agar medium
is mixed, sterilized, and poured into plate form
all in one process. Any excess medium is then
hand-poured into test tubes for use as slants. A
large batch of plates and slants are poured all at
once, the bulk of which is then stored in a
refrigerator at 4°C until needed.
At scheduled intervals of approximately once a
month, several fresh agar plates are "streaked"
with Selenastrum cells from existing stock
cultures. A 10 /j\ inoculating loop is used to
transfer the cells from the liquid stock cultures
to the agar, where they are streaked out into
quadrants on the plated medium. The plates are
then arranged on a rack situated in a partially
enclosed glass box shelter in the environmental
chamber for incubation purposes. The glass
box, along with rubber bands used to seal the
lids on the petri dishes, serves to break up the
airflow patterns of the chamber around the
immediate vicinity of the plates, thereby
minimizing dessication problems of the media.
The plated cultures need air exchange for
proper growth, but too much airflow will only
serve to dry out the plates completely,
rendering them useless for storage purposes.
An incubation period of 1-2 weeks yields
several distinct Selenastrum colonies that may
then be targeted for transfer to fresh liquid
nutrient medium, thereby rejuvenating active
stock cultures.
Agar slants are also streaked up from time to
time as needed, but serve primarily in a
secondary backup role. After incubation, the
slants displaying healthy Selenastrum colonies
are pulled from the environmental chamber and
stored in a refrigerator at 48C for up to several
months. In this way, they may serve as a
"backup" to the backup cultures.
102
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IEPA Selenastrum Culturing and Test Protocol
6. Quality Assurance/Quality Control
Considerations.
At each cell transfer, the stock cultures are
examined microscopically for species
verification purposes and to look for any signs
of microbial contamination. This information,
along with general observations on the overall
condition of the cells themselves, is then
recorded in an algal culture logbook. This
enables the TTU laboratory to keep a running
history of culture activities and any special
problems/ solutions encountered. Stock cultures
are also subjected to monthly NaCI reference
toxicant tests. EC 50 point estimates are
calculated for each reference test, and these
values are then plotted on a standard reference
toxicant control chart for quality control
purposes (Figure 1). These steps are taken to
ensure the quality and suitability of the
Selenastrum stock cultures for use in
biomonitoring activities.
Selenastrum Algal Growth Test Protocol
Samples received by TTU for algal bioassays
consist of municipal and industrial effluents and
their ambient receiving waters (upstream of the
effluent outfall). For the purposes of the IEPA,
Illinois is divided into seven regions. All regions,
except Region 4 (Field Operations Services in
Champaign, IL) and Region 5 (Field Operations
Services in Springfield, IL), ship samples to OE
via bonded courier (e.g., Emery Worldwide).
Regions 4 and 5 hand deliver samples to OE. All
samples are received in the laboratory and
testing started within 28 hours of sampling. A
chain of custody is maintained by field and
laboratory personnel to ensure the samples are
not tampered with.
When samples arrive in TTU they are logged in,
warmed to the proper temperature, aerated,
and initial water chemistries are performed.
Initial water chemistries consist of alkalinity,
hardness, chlorine, and ammonia
determinations. Measurement of these
parameters helps resolve the cause of toxicity.
When the samples have been warmed and
aerated, dilutions are poured. A 0.5 dilution
series is used. Temperature, pH, conductivity,
and dissolved oxygen are measured on each
dilution to determine if these parameters are
within the range for normal growth of
Selenastrum. A 250 ml portion of each dilution
is then poured off for the algal bioassay.
Each 250 mL dilution is enriched with 250 fjL
of each of the four nutrient stock solutions
(with EDTA). To reduce the possibility of
contamination in the algal bioassay, aseptic
techniques are employed throughout the test.
All glassware is washed with non-phosphate
detergent and rinsed with tap water, acetone,
hydrochloric acid, tap water, and distilled
water. Glassware and pipet tips are autoclaved
at 121°C.
Each dilution (with nutrients) is filtered through
a 0.2 fjm membrane filter. This filtration
removes any indigenous algae from the
dilutions. Following filtration, 150 mL of each
dilution is measured in each of three 125 mL
Erlenmeyer test flasks (three flasks per dilution
with 50 mL of diluent per flask). The entire test
consists of three test flasks in each of the
following concentrations; control, 0%, 6.25%,
12.5%, 25%, 50%, 100%.
The test flasks are inoculated with 1 mL of
log-phase-growth Selenastrum (4 to 7 days old)
to provide an initial cell density of 10,000
cells/mL (±10%). At IEPA the algal cells are
not "washed" prior to inoculation (there is no
need to remove EDTA from the test cells since
the test nutrients contain EDTA). The required
volume of stock culture needed to inoculate the
test flasks is calculated as follows:
number of x volume of test x 10,000
test flasks solution/flask cells/mL
cell density (cells/mL) in the stock culture
= volume (mL) of stock culture required
Test flasks are covered with aluminum foil for
autoclaving. After the flasks are inoculated, the
103
-------
Charles and Searle
s
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z
z
o
§~
• •
zl
S8
§1
^^
i
£
u
IL
8
U
U
2.9
2.8
2.7
2.6
2.5
9 A
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2.3
2.2
2.1
2
1.9
1.8
1.7
1.6
I.S
L D
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a
-
a Da
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a °
a
-
-
-
1 U III IV V VI VM VHI IX X XI XII
MONTHLY DATA
Figure 1. Illinois EPA Reference Toxicity Test (NaCI).
Front
Front row moves to back
Figure 2. Rotational pattern.
104
-------
IEPA Selenastrum Culturing and Test Protocol
aluminum foil is removed and replaced with
100 mL plastic beakers for incubation. Cell
density is checked in at least three flasks within
two hours of inoculation. The test flasks are
randomized on the holding tray prior to
incubation. The flasks are rotated at 24, 48,
and 72 hours using a standard rotation pattern
(Figure 2) to help ensure that all test containers
receive equal amounts and intensity of light and
even temperature throughout the 96 hour
incubation.
The algal incubator is kept at a constant
temperature of 25 ± 1°C. Lighting is turned on
approximately four hours before the start of the
test to allow the lights to reach equilibrium.
During the test the flasks are rotated
mechanically at 100 cycles per minute
continuous rotation. Light intensity during the
test is 400 ± 40 ft-c (4306 ± 431 lux). Light
lux is measured at the beginning of the test and
at the end of the test. The pH of the 0% and
the 100% is also measured at the beginning
and at the end of the test.
Test termination is at 96 ± 2 hours. Algal cell
density in each flask is1 measured using a
Coulter Counter" Model ZM. Test cultures are
diluted with Isoton" (a sodium chloride
electrolyte solution) and counted directly on the
Coulter Counter". Three cell counts are taken
for each aliquot and the mean cell volume is
averaged for the three counts. Each test flask
is mixed thoroughly following US EPA
procedure. For IEPA purposes, the counts from
each test culture must have less than 10%
variability.
Test results are considered acceptable if the
average cell counts in the control flasks are
greater than 2 x. 106 cells/mL and control
variability does not exceed 20%. When using
stock nutrient solutions without EDTA,
obtaining average cell counts greater than
200,000 cells/mL was not a problem, however
keeping control variability below 20% was
difficult. Without EDTA, cell counts in the
controls ranged from 400,000 to 800,000
cells/mL. Variability in the three control flasks
was as high as 79%. variability in control flasks
inoculated with stock nutrient solutions contain-
ing EDTA consistently remained below 20%.
EDTA can lower toxicity of a sample by
complexing heavy metals. EDTA facilitates algal
growth by increasing the availability of
micronutrients. Based on the control flask
variability (when EDTA is not used) the decision
was made at IEPA to conduct the Selenastrum
algal bioassay with EDTA in the stock nutrient
solutions. Adverse effects on Selenastrum cell
growth, expressed in LOEC and NOEC values,
are obtained using Dunnett's Procedure.
Statistics are analyzed using an in-house
written computer program.
Conclusion
The Selenastrum algal bioassay is a useful
aquatic toxicity test, and is an important
component of lEPA's testing program. In
addition to detecting phytotoxic contaminants,
the bioassay could identify wastewaters which
are nutrient rich and biostimulatory. By
incorporating a freshwater primary producer
(Selenastrum) into the bioassay regime, toxicity
could be detected which is not detected by
tests using primary consumers (Ceriodaohnia
dubia. at IEPA) or secondary consumers
(Pimeohales promelas. at IEPA).
Literature Cited
USEPA. 1989. Short-Term Methods for
Estimating the Chronic Toxicity of Effluents and
Receiving Waters to Freshwater Organisms.
Environmental Monitoring Systems Laboratory,
U. S. Environmental Protection Agency,
Cincinatti, Ohio, EPA/600/4-89/001.
105
-------
Effects of Acute Sublethal Levels of pH on the Feeding Behavior of Juvenile
Fathead Minnows
Robert D. Hoyt1 and Hanan Abdul-Rahim2
Graduate Center for Toxicology
The University of Kentucky
Lexington, KY 40506
Abstract
This study was conducted to determine the impact of acute sublethal pH levels on the feeding
behavior of juvenile fathead minnows. Eighteen to 24 day-old juveniles were fed live or dead brine
shrimp under light or dark conditions in order to identify the role of the senses of vision,
chemoreception, and mechanoreception in feeding at different pH's. Feeding trials were conducted
at various pH combinations; 5.0, 7.0, 10.0; 4.5, 7.0, 11.0; and 3.5, 7.0, 11.5. Total mortality was
observed at pH 3.5 and 11.5. Appetitive behavior was present at all pH levels as evidenced by
frequencies of occurrence of feeding ranging from 93-100%. No relationship was observed between
pH and the number of fish feeding. The fathead minnow is chiefly a visual feeder and vision was not
affected at any pH level as 99.9% of all brine shrimp, live and dead, were consumed in the light.
Significantly fewer brine shrimp were consumed in the dark than in the light and significantly fewer
brine shrimp were consumed in the dark at the lower pH's than in the dark at pH's 7.0+ . No
measurable impact on feeding behavior was observed at pH 7.0 and 10.0 + . Chemoreception was
stressed at low pH levels. The ability of chemoreception and mechanoreception to successfully
function in consort in capturing living prey in the dark at low pH levels was noticeably impacted. The
effect of low pH on mechanoreception was not determined.
Keywords: pH, fathead minnow, feeding behavior
Introduction
The science of behavioral toxicology is a
recently developed diagnostic approach to
measuring and recording observations of
behavior that reflect biochemical and
ecological responses of organisms to
environmental contamination (Little 1990).
Introduction
The science of behavioral toxicology is a
recently developed diagnostic approach to
measuring and recording observations of
behavior that reflect biochemical and
ecological responses of organisms to
environmental contamination (Little 1990).
Behavioral activities are rapidly becoming
recognized as highly sensitive indicators of
sublethal toxicity (Diamond et al. 1990, Little
and Finger 1990). A variety of behaviors has
been used to study sublethal toxicities
including ventilation and cough frequencies,
feeding activities, temperature preference,
predator avoidance, swimming performance,
schooling behavior, and pH detection and
avoidance (Hill 1989). However, while it is
readily acknowledged by investigators that
differing behavior activities involve a diversity
of sensory-motor pathways and physiological
processes (Sandheinrich and Atchison 1990),
little attention has been given to the impacts
that toxicants selectively impart to specific
senses or sensory pathways. Although the
1 Present Address: Department of Biology, Western Kentucky University, Bowling Green, KY
42101
2 Present Address: Salem College, Salem, NC 27108
106
-------
Sublethal pH Effects
embryo-larval-juvenile life cycle stages are
accepted as being the most sensitive for
toxicity tests (McKim 1977), little
consideration has been given to the impacts of
sublethal toxicants upon the sensory systems
of these life cycle stages, many of which
exhibit a gradient of sense organ development
from the time of hatching until the completion
of successful behavior formation (Noakes and
Godin 1988).
Although acid stress and depressed pH
conditions have been studied at length from
many different perspectives (Zischke et al.
1983, Leino et al. 1987, Mills et al. 1987,
Jansen and Gee 1988, among others), and are
receiving much local press in relation to acid
precipitation, little attention has been given to
the effects of acid stress on fish behavior
(Jones et al. 1985). Lemly and Smith (1985)
summarized the literature supporting fathead
minnows as being among the most acid sensi-
tive fishes. Jones et al. (1985), in pursuing the
effects of sublethal pH levels on the behavior
of arctic char, reported acid stress to suppress
chemoreception. Lemly and Smith (1985,
1987) found acidification to significantly
affect the ability of fathead minnows to detect
or respond to chemical stimuli. Jones et al.
(1985) described this chemo-suppression to
likely result from the reduced stimulatory
nature of amino acids at reduced pH's and the
damage of epithelial tissues (olfactory epi-
thelium) by acidic conditions. Lemly and Smith
(1987) suggested that increased olfactory
mucous thickness in response to lowered pH
prevented normal stimulus-receptor interaction
and/or that chemical interaction at the sub-
cellular level was impaired because of stearic/
charge changes at the receptor cells.
Whatever the explanation, these observations
are of potentially profound importance to
environmental biologists in that they represent
avenues for unrecognized massive larval
mortalities among those fish species that are
dependent upon chemoreceptors of
chemoreceptors/mechanoreceptors in the
formation of exogenous feeding behavior.
The purpose of this study was to determine
the effects of different acute pH levels on the
senses of vision, mechanoreception and
chemoreception in the feeding behavior of
juvenile fathead minnows, Pimeohales
promelas.
Methods and Materials
Test Fish
Fathead minnows used in the project were
obtained within 12 hours of hatching from the
U.S. EPA Newtown Fish facility, Newtown,
Ohio, on 3 July, 1990. Fish were maintained
in 2 1 finger bowl in ASTM water at 24 + /-1C
and were fed freshly hatched brine shrimp
twice daily, 0800 and 1700 h. At the onset of
the pH trials, the minnows were 18 days old
and averaged 11.58 mm in total length (range
10.1 mm- 13.0mm). All experimentation was
conducted at the Graduate Center for
Toxicology at the University of Kentucky,
Lexington, KY.
pH Test Solutions
pH solutions in 3.0 1 aliquots were prepared
using ASTM water with Nitric acid to produce
low pH levels and Sodium Hydroxide to
produce high pH levels. A pH of 7.0 was
achieved by adding either Nitric acid or
Sodium Hydroxide as required. An Orion pH
meter was used to determine pH levels in
producing the desired pH concentrations. Acid
pH's tested were 5.0, 4.5, 4.0, and 3.5.
Basic pH's were 10,10.5,11.0, and 11.5. An
acidic, a basic, and a pH 7.0 concentration,
was used daily for four consecutive days in
the following sequence: Day 1 - pH's 5.0, 7.0,
10.0; Day 2 - pH's 4.5, 7.0, 10.5; Day 3 -
pH's 4.0, 7.0, 11.0; Day 4 - pH's 3.5. 7.0,
11.5.
Sense Organ Isolation
Each daily combination of pH test solutions
was applied to four different feeding regimes,
living brine shrimp fed in the light, living brine
shrimp fed in the dark, dead brine shrimp fed
in the light, and dead brine shrimp fed in the
' dark (see Table 1 for design).
107
-------
Hoyt and Abdul-Rahim
Table 1. Frequency of occurrence of juvenile fathead minnows eating at least one brine shrimp during
live-dead, light-dark, feeding trials at varying pH levels.
pH Level
Light-Live
Light-Dead
Dark-Live
Dark-Dead
pH Level
Light-Live
Light-Dead
Dark-Live
Dark-Dead
pH Level
Light-Live
Light-Dead
Dark-Live
Dark-Dead
pH Level
Light-Live
Light-Dead
Dark-Live
Dark-Dead
Test Procedure
55
35
4/5
5/5
5/5
5/5
5/5
5/5
55
5ft
5/5
55
5.0
5/5 4/5
5/5 5/5
5/5 5/5
5/5 5/5
(93%)
4.5
5/5 5/6
4/5 4/5
5/5 5/5
5/5 5/5
(95%)
4.0
5/5 55
5/5 5/5
5/5 5/5
5/5 4/4
(100%)
3.5
55
5/5
5/5
5/5
55
55
55
55
55
45
55
6/6
Total
Mortality
7.0
5/5 45
55 55
5/5 55
5/5 3/3
(98%)
7.0
4/5 55
55 55
5/5 55
5/5 45
(95%)
7.0
55 55
55 55
55 55
5/5 55
(98%)
7.0
No
Teat
The fish
55
55
55
55
55
55
55
55
55
55
45
4/4
10.0
4/5 45
5/5 55
5/5 55
5/5 55
(95%)
10.5
5/5 55
5/5 55
5/5 55
5/5 55
(100%)
11.0
5/5 55
5/5 55
4/5 55
4/4 4/4
(96%)
11.5
Total
Mortality
were allowed to acclimate in the
A total of 180 minnows were selected at
approximately 1600 h the day before a trial.
The fish were separated into 12 groups of 15,
each group of which was placed in a 500 ml
beaker containing water of a specific pH
concentration (daily test combinations
described above). The fish were then placed in
an environmental chamber at 25.0 C with an
8 hour dark period (2200 to 0600 h) for
acclimation until approximately 1300 h the
following day. Five fish from each pH
concentration were then placed in each of
three 150 ml finger bowls containing fresh
mixtures of the test pH's for replicate trials.
finger bowls for ten minutes in either light or
dark before food was added. Light feeding
trials with live and dead brine shrimp were
conducted prior to similar dark feeding trials.
Feeding
Brine shrimp (Salt Lake City variety) were
raised in the laboratory and fed immediately
following 24 hours incubation. The brine
shrimp used in the feeding trials were the
same variety and size used to raise and
maintain the minnows. Average brine shrimp
length was 0.7 mm. Fifty live or dead brine
shrimp for each fish subsample were selected
108
-------
Sublethal pH Effects
Table 2. Number and Percent of brine shrimp remaining following juvenile fathead minnow live-dead,
light-dark feeding trials at varying pH levels.
pH Level
Light-Live
Light-Dead
Dark-Live
Dark-Dead
0
0
4
9
5.0
0
0
0
14
0
0
0 (2.7%)
8 (20.7%)
0
0
0
8
(5.8%)
pH Level
Light-Live
Light-Dead
Dark-Live
Dark-Dead
0
0
17
15
4.5
1
0
22
5
2
0
23 (41.3%)
21 (27.3%)
0
0
1
9
(17.2%)
pH Level
Light-Live
Light-Dead
Dark-Live
Dark-Dead
0
0
30
1
4.0
0
0
8
3
0
0
36(49.3%)
7 (7.3%)
0
0
0
1
(14.2%)
pH Level
Light-Live
Light-Dead
Dark-Live
Dark-Dead
3.5
Total
Mortality
7.0
0
0
0
3
0
0
0
0 (7.3%)
0
0
5
2
(1.8%)
7.0
0
0
1
0
0
0
3 (3.3%)
0 (6.0%)
0
0
0
2
(2.3%)
7.0
0
0
0
0
0
0
0
1 (1.3%)
0
0
2
3
(0.3%)
7.5
No
Test
10.0
0
0
5
0
0
0
3 (8.7%)
0
(1.3%)
(2.5%)
10.5
0
0
0
0
0
0
0
0
(1.3%)
(0.3%)
11.0
0
0
0
2
0
0
0
2
(1.3%)
(4.7%)
(1.5%)
1L5
Total
Mortality
with a 10 cc syringe and counted using a
dissecting microscope. Each of nine syringes
was loaded immediately prior to the feeding
exercise and the fifty brine shrimp added to
each group of five fish following the ten
minutes acclimation to the test pH's. Feeding
time for all tests was ten minutes. Brine
shrimp for the dead feeding trials were killed
by treatment in an ultrasonic bath for two to
four minutes. Fish were aspirated from the
test dishes immediately following the feeding
trial and isolated in holding dishes. While the
number of brine shrimp ingested by each
individual fish could not be determined, the
number of fish having consumed at least one
brine shrimp was recorded using a dissecting
microscope. Each feeding test dish was
examined with the aid of a dissecting
microscope and the number of brine shrimp
remaining following the feeding trial was
counted. All fish used in a feeding exercise
were excluded from further feeding.
109
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Hoyt and Abdul-Rahim
Results
Survival
All fish survived every trial except those at pH
3.5 and 11.5 in which 100 percent mortality
was observed.
Frequency of Occurrence of Feeding
The number of fish ingesting at least one brine
shrimp during the feeding trials ranged from
93 to 100 percent for all pH levels (Table 1).
No relationship was observed between number
of fish feeding and light or dark or live or dead
food conditions. No relationship between
frequency of occurrence of fish feeding and
decreasing or increasing pH levels was
observed. The average frequency of
occurrence of feeding by fishes in pH 7.0
water, 97.7%, was the same as that (97.7%)
for fish in the increasing pH level trials and
only slightly greater than that (96.1 %) for fish
in the decreasing pH level trials (Table 1).
Light-Dark Feeding versus pH Level
With the exception of three brine shrimp at pH
4.5, all food organisms (99.9%). live or dead,
were consumed during light feeding trials, at
all pH levels (Table 2). In the dark, however,
live or dead brine shrimp remained after
feeding in 15 of 18 trials. The number of brine
shrimp remaining in each dark trial ranged
from 1.3% to 49.3% (Table 2). At pH 7.0,
slightly more dead brine shrimp remained in
the dark than live brine shrimp, 5,0% and
1.1%, respectively. At the lower pH's 5.0,
4.5, and 4.0 combined, more live brine shrimp
(31.1%) remained in the dark than dead
shrimp (18.4%). The number of live and dead
brine shrimp remaining at the higher pH's was
generally similar to that of pH 7.0 (Table 2).
Significantly fewer live brine shrimp were
consumed at pH 4.5 and 4.0 than at the
higher levels while significantly fewer dead
brine shrimp were consumed at pH 5.0 and
4.5 (Table 2).
Discussion
Based upon the high frequency of occurrence
of feeding (93 + %) by juvenile fathead
minnows at all pH levels during all test
regimens, appetitive behavior was concluded
to be present to the lethal pH levels of 3.5 and
11.5. Hill (1989), in a chronic study of low pH
effects on feeding behavior of smallmouth
bass, also observed no loss of appetitive
behavior at pH 4.2. Mortality at pH 3.5 in this
study was consistent with the report by
Mount (1973) that most lethal pH values
recorded from laboratory data occur below
4.0. No similar data were found regarding high
pH mortality, although Carlender (1969)
reported the fathead minnow to have a broad
tolerance for pH. The persistence of appetitive
behavior through all pH levels was further
supported by the minnows eating all except
three (99.9%) live and dead brine shrimp fed
during the light feeding trials. Consequently,
since appetitive behavior persisted throughout
the study, any observed reductions in feeding
activity were considered to be the result of the
selective impairment of sense organs by the
different pH levels, or the behavioral
inactivation of the feeding response during
certain environmental conditions of the test (
i.e..dark), or possibly a combination of both
these features.
According to the feeding patterns observed in
this study, the fathead minnow is
predominantly a visual, daylight feeder. The
removal of all except three brine shrimp, live
and dead, in the light at all pH levels identified
the eyes, or the eyes in conjunction with the
senses of chemoreception and mechano-
reception, as the major sense organs involved
in early life stage feeding. Klemm (1985)
reported the fathead minnow to be primarily
omnivorous and to possess large black
eyes,presumably functional if heavily
pigmented, at the time of hatching. The eye
must play a strategically greater role in early
life stage feeding as indicated by the
recommendations by Birch et al. (1975) and
Klemm (1985) that fathead minnow fry at
least 6 days to 28 days old be fed live, freshly
hatched (small) brine shrimp during the day,
while older individuals may be fed frozen brine
110
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Sublethal pH Effects
shrimp or varying forms of dry chow. This
recommendation suggested a greater visual
feeding success on living prey by larval-
juvenile individuals and more successful visual-
chemoreceptive feeding behavior in older
individuals.
The eyes did not appear to be functionally
impacted by different pH levels in this study.
All brine shrimp, living and dead except for
three individuals, were consumed in the light
at all pH levels. However, the small test
chambers (150 ml) with concomitant short
reaction distance and number of brine shrimp
per trial (50) might have alleviated any eye
stress that was present or developing. Hill
(1989) reported low pH levels to impair visual
acuity, coordination, and agility, subsequently
resulting in lower growth in chronic trials with
smallmouth bass. The question of acute
versus chronic study interpretations is brought
into focus at this point. Hill's (1989)
suggestion that acute bioassays may be too
short to detect certain biological parameters,
such as growth and survival, is not contested.
However, in meeting the objectives of studies
such as this, acute sublethal tests, especially
feeding, may be more sensitive than chronic
growth studies (Sandheinrich and Atchison
1990). That the fathead minnow is not an
adept nocturnal feeder was supported by the
number of uneaten live and dead brine shrimp
in dark feeding trials at all pH's. The presence
of more dead (5.0%) than live (1.1%) brine
shrimp following dark feeding trials at pH 7.0
indicated a slightly greater ability or behavioral
preference by the fathead minnow to select
living food over dead food in the dark.
However, at the lower pH's during the dark,
feeding was greatly reduced and more dead
brine shrimp were consumed than live shrimp.
Chemoreception was considered to be
impaired at pH 5.0 when 20.7% of the dead
brine shrimp remained, worsened at pH 4.5
when 27.3% remained, and then inexplicably
improved at pH 4.0 when 7.3% of the shrimp
remained. The marked improvement in dark-
dead feeding at the lowest pH could not be
explained. Feeding on dead food in the dark
was considered to be entirely a function of
chemoreception since stimuli for visual and
mechanoreceptive senses were not present.
Consequently, since appetitive behavior was
known to exist, this decreased feeding on
dead food in the dark strongly suggested
chemoreceptive inhibition. Yoshii and Kurihara
(1983) reported that bluegill without
functional lateral lines did not produce
successful feeding strikes in the dark based on
chemoreception alone. The omnivorous
feeding capability of certain species such as
bluegill sunfish and fathead minnow might
employ the sense of chemoreception at levels
not yet described and different than other
species. However, the findings by Jones et al.
(1985) that pH 5.0 suppressed chemo-
orientation in the Arctic char, and Lemly and
Smith (1985, 1987) that pH 6.0 eliminated
fathead minnow responses to chemical stimuli
supported the initial conclusions drawn in this
study that chemoreception was impaired by
the lower pH levels.
Living prey in the dark seemed to represent
the maximum sensory challenge in feeding,
especially for the visual feeding fathead
minnow. Consequently, low pH exhibited its
greatest impact on feeding on live brine shrimp
in the dark. Although live brine shrimp were
successfully preyed upon in the dark at pH 7.0
and higher, only slightly more than 50% were
captured at the lowest pH's. Live food under
dark conditions would suggest the
involvement of the combined senses of
chemoreception and mechanoreception in
successful feeding. Enger et al. (1989) and
Montgomery (1989) presented evidence
substantiating the role of mechanoreceptors in
detecting moving prey. Montgomery (1989)
further described the role of mechanoreceptors
as operating synergistically with vision in
daylight planktivory and singly in total
darkness. No mention was made by
Montgomery of any receptor involvement with
mechanoreceptors in dark feeding. Hara
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Hoyt and Abdul-Rahim
(1986) summarized the extensive literature
reviews on the role of chemoreception in
feeding behavior and identified the first step in
the feeding sequence as arousal to the
presence of food which is primarily mediated
by olfaction. Atema (1980) excepted the most
visual fish species, i.e., anosmic sticklebacks,
from the above behavior and suggested that
accompanying senses such as
mechanoreception and vision may also be
involved in initial prey recognition. Should
these sensory assumptions be correct and
chemoreception was impaired to a reduced
level of effectiveness in establishing the
presence of living food, then several
explanations regarding the role of
mechanoreception in dark feeding become
likely. First, mechanoreceptors act
synergistically with chemoreception in dark-
live feeding and the impairment of one sense
automatically reduced the success of the
other; secondly, mechanoreceptors also were
impaired by the lowered pH rendering them
incapable of successfully detecting and/or
locating the moving prey; or thirdly,
mechanoreceptors alone are incapable of
successfully detecting and locating living prey
in complete darkness in strongly visually
directed species.
Summary
Fathead minnow feeding behavior was not
observed to be affected by any non-lethal pH
in the light. The interaction of vision with
chemoreception, or mechanoreception, or
both, produced successful feeding on live
brine shrimp at all pH levels. Likewise, in the
dark at pH's of 7.0 and higher, the senses of
chemoreception and mechanoreception
operated successfully in feeding on live food
(98.9%) in the dark. However, at low pH's in
the dark, chemoreception and
mechanoreception were impaired in detecting
and locating living prey. While chemoreception
was observed to be stressed at low pH's, no
evidence of such an effect on
mechanoreception was detected. Future
studies using streptomycin sulfate
(Montgomery 1989) or cobalt (Karlsen and
Sand 1987) to ablate mechanoreceptors might
provide valuable insights into the role of
mechanoreceptors in initiating as well as
concluding the feeding response.
Acknowledgements
This study was supported by funding from the
National Science Foundation Kentucky EPSCoR
Program to the first author and the 1990
Undergraduate Summer Research Experience
for Minorities, Graduate Center for Toxicology,
University of Kentucky, the second author. We
thank W.J. Birge of the Departments of
Biological Sciences and Toxicology, the
University of Kentucky, for serving as project
host and providing the physical facilities
necessary for this investigation. Special thanks
go to T. Short and W.A. Robison for their
technical expertise and assistance. Additional
thanks go to the numerous other members of
Dr. Birge's laboratory who contributed to the
project.
Literature Cited
Atema, J. 1980. Chemical senses, chemical
signals, and feeding behavior in fishes, pp. 57-
94. In: Bardach, J.E. et al. (eds.), Fish
behavior and its use in the capture and culture
of fishes. International Center for Living
Aquatic Resources Management, Manila.
Birch, T.J., J.P. Abrams, and G.L. Martin.
1975. Design and operation of a static rearing
unit for fathead minnows. Ohio Environmental
Protection Agency, Division of Surveillance
and Laboratory Services, Columbus, 8 pp.
Carlender, K.D. 1969. Handbook of freshwater
fishery biology. Vol. 1. Iowa State University
Press, Ames, I A.
Diamond, J.M., M.J. Parson, and D. Gruber.
1990. Rapid detection of sublethal toxicity
using fish ventilatory behavior. Environ. Tox.
Chem. 9:3-11.
112
-------
Sublethal pH Effects
Enger, P.S.. AJ. Kalmijn, and 0. Sand. 1989.
Behavioral investigations on the functions of
the lateral line and inner ear in predation, pp.
575-587. In: Coombs, S. et al. (eds), The
Mechanosensory Lateral Line. Springer-Verlag,
New York.
Hara, T.J. 1986. Role of olfaction in fish
behavior, pp. 152-176. In: Pitcher, T.J. (ed.).
The behavior of teleost fishes. Croom Helm,
London.
Hill, J. 1989. Analysis of six foraging
behaviors as toxicity indicators, using juvenile
smallmouth bass exposed to low
environmental pH. Arch. Env. Contamination
Tox. 18:895-899.
Jansen, W.A., and J.H. 1988. Effects of
water acidity on swimbladder function and
swimming in the fathead minnow, Pimephales
promelas. Can. J. Fish. Aquatic Sci. 45:65-77.
Jones, K.A., T.J. Hara, and E. Scherer. 1985.
Behavioral modifications in arctic char,
(Salvelinus aloinus) chronically exposed to
sublethal pH. Phys. Zool. 58(4):400-412.
Karlsen, H.E., and 0. Sand. 1987. Selective
and reversible blocking of the lateral line in
freshwater fish. J. Exp. Biol. 133:249-262.
Klemm, D.J. 1985. Distribution, life cycle,
taxonomy, and culture methods. Fathead
minnow (Pimephales promelas). pp. 112-125.
In: United States Environmental Protection
Agency, Methods for Measuring the Acute
Toxicity of Effluents to Freshwater and Marine
Organisms. U.S. Environmental Protection
Agency, EMSL, Cincinnati. EPA/600/4-
85/013.
Leino, R.L., P. Wilkinson, and J.G. Anderson.
1987. Histopathological changes in the gills of
pearl dace, Semotilus maroarita. and fathead
minnows, Pimephales promelas. from
experimentally acidified Canadian lakes. Can.
J. Fish. Aquatic Sci. 44 (Suppl. 11:126-134.
Lemly, A.D., and R.J.F. Smith. 1985. Effects
of acute exposure to acidified water on the
behavioral response of fathead minnows,
Pimephales promelas. to chemical feeding
stimuli. Aquatic Tox. 6:25-36.
Lemly, A.D., and R.J.F. Smith. 1987. Effects
of chronic exposure to acidified water on
chemoreception of feeding stimuli in fathead
minnows (Pimephales promelas): mechanisms
and ecological implications. Env. Tox. Chem.
6:225-238.
Little, E.E. 1990. Behavioral toxicology:
stimulating challenges for a growing discipline.
Env. Tox. Chem. 9:1-2.
Little, E.E. and S.E. Finger. 1990. Swimming
behavior as an indicator of sublethal toxicity in
fish. Env. Tox. Chem. 9:13-19.
McKim, J.M. 1977. Evaluation of tests with
early life stages of fish for predicting long-term
toxicity. J. Fish. Res. Board Can. 34:1148-
1154.
Mills, K.H., S.M. Chalanchuk, L.C. Mohr, and
I.J. Davies. 1987. Responses of fish
populations in Lake 223 to 8 years of
experimental acidification. Can. J. Fish.
Aquatic Sci. 44 (Suppl. 1) 114-125.
Montgomery, J.C. 1989. Lateral line detection
of planktonic prey, pp. 561-574. In: Coombs,
S. et al. (eds.). The Mechanosensory Lateral
Line. Springer-Verlag, New York.
Mount, D.I. 1973. Chronic effect of low pH on
fathead minnow survival, growth and
reproduction. Water Res. 7:987-993.
Noakes, D.L.G. and J.G.J. Godin. 1988.
Ontogeny of behavior and concurrent
developmental changes in sensory systems in
teleost fishes, pp. 345-395. In: physiology of
developing fish, Part B: Viviparity and
posthatching juveniles. Academic Press.
113
-------
Hoyt and Abdul-Rahim
Sandheinrich, M.B. and G.J. Atchison. 1990.
Sublethal toxicant effects on fish foraging
behavior: empirical vs. mechanistic
approaches. Env. Tox. Chem. 9;107-119.
Yoshii, K. and K. Kurihara. 1983. Role of
cations in olfactory reception. Brain Res.
274:239-248.
Zischke, J.A., J.W. Arthur, K.J. Norlie, R.O.
Hermanutz, D.A. Standen, and T.P. Henry.
1983. Acidification effects on
macroinvertebrates and fathead minnows
(Pimeohales oromelas) in outdoor experimental
channels. Water Res. 17:47-63.
114
-------