Assessment of biocontamination of benthic macroinvertebrate communities in European inland waterways

Aquatic Invasions (2008) Volume 3, Issue 2: 211-230
DOI: 10.3391/ai.2008.3.2.12
© 2008 European Research Network on Aquatic Invasive Species


211
Research article
Assessment of biocontamination of benthic macroinvertebrate communities in
European inland waterways
Kęstutis Arbačiauskas1*, Vitaliy Semenchenko2, Michal Grabowski3, Rob S.E.W. Leuven4, Momir
Paunović5, Michail O. Son6, Bela Csányi7, Simona Gumuliauskaitė1, Alicja Konopacka3, Stefan
Nehring8, Gerard van der Velde9, Vasiliy Vezhnovetz2 and Vadim E. Panov10
1Institute of Ecology, Vilnius University, Vilnius, Lithuania
2Institute of Zoology, National Academy of Sciences, Minsk, Belarus
3University of Łódź, Łódź, Poland
4Department of Environmental Science, Institute for Wetland and Water Research, Radboud University Nijmegen, The Netherlands
5Institute for Biological Research “Sinisa Stankovic”, Belgrade, Serbia
6Institute of Biology of the Southern Seas, Odessa Branch, Odessa, Ukraine
7Environmental Protection and Water Management Research Institute (VITUKI Kht), Budapest, Hungary
8AeT umweltplanung, Koblenz, Germany
9Department of Animal Ecology and Ecophysiology, Institute for Wetland and Water Research, Radboud University Nijmegen,
The Netherlands / National Natural History Museum Naturalis, Leiden, The Netherlands
10St. Petersburg State University, St. Petersburg, Russia
*Corresponding author
E-mail: arbas@ekoi.lt
Received 27 June 2008; accepted in revised form 16 August 2008; published online 20 August 2008
Abstract
Introductions of alien species, regardless of their actual or potential impacts, can be considered as a biocontamination of the
ecosystem. A simple method to assess biocontamination is described and tested on benthic macroinvertebrate communities from
European inland waterways. This method includes calculations of abundance contamination and richness contamination at
ordinal taxonomic rank, from which integrated estimations of biocontamination are derived. Our method can be applied to data
collected during routine water quality monitoring, and allows estimation of biocontamination at specific study sites as well as
integrated assessment of ecosystems or assessment units. Results clearly show that the main European inland waterways are
highly biologically contaminated. They also indicate that richness contamination precedes abundance contamination, and that
severe abundance contamination may be caused even by a single ecologically aggressive alien species. Comparison of
biocontamination indices and ecological quality status by conventional methods suggests that these metrics are negatively
correlated, and richness contamination has a stronger negative affect than abundance contamination. Biocontamination warrants
inclusion within the development of holistic estimates of ecological quality status and should be considered in water management
policy.

Key words: biocontamination, alien species, ecological status, benthic macroinvertebrates, European inland waterways, aquatic
ecosystems

Introduction
Human-mediated introductions of invasive alien
species in European inland and coastal waters
are considered a serious environmental issue that
requires development of relevant management
approaches (Leppäkoski et al. 2002; Panov et al.
2002; Gherardi 2007). Alien species (AS) are
those that take up residence in a biogeographical
area, such as a river catchment, where they were
previously unknown. In the context of the EU
Water Framework Directive (European Commu-

K. Arbač i auskas e t a l . , Assessment o f b iocontaminat ion
212
nity 2000), invasive AS represent a significant
biological pressure. The assessment of such
pressure should therefore be considered within
an integrated catchment management strategy
and should receive special attention within the
context of ecological status assessments required
by the EU Water Framework Directive (Cardoso
and Free 2008).
Owing to adverse ecological and economic
consequences, AS invasions are perceived as
biological pollution of aquatic ecosystems.
According to Elliott (2003), “biological
pollution” is defined as the effects of introduced
invasive species sufficient to disturb an indivi-
dual, a population or a community; including the
causation of adverse economic consequences.
However, the quantitative assessment of negative
impacts of AS is difficult and requires compre-
hensive research and database efforts (Molnar et
al. 2008). Consequently, quantitative estimates
of “biological pollution” sensu Elliott (2003) in
aquatic ecosystems are lacking (Olenin et al.
2007). A more practical approach for assessing
the impact of AS on the ecological status of
water bodies, therefore, may be to assume that
their affect is proportional to their occurrence
and abundance within the invaded community. In
such a case, AS would be considered as
biological contaminants rather than biological
pollutants, and “biological contamination” (i.e.
biocontamination) means the presence of AS
regardless of their abilities to cause negative
ecological and/or socio-economic impacts (see
also Panov et al. 2008).
The purpose of the present study is to address
AS in ecological status assessments of water
bodies, including specific locations within water
bodies, considering the above “biocontami-
nation” concept. We describe a simple method to
measure the biocontamination of aquatic commu-
nities, which does not require sophisticated
research and can be applied to water bodies for
which routine monitoring data are available. This
method is tested on benthic macroinvertebrate
assemblages of European inland waterways.
Since invaders affect the structural organisation
of recipient communities (Simon and Townsend
2003), we hypothesize that the relative
abundance of aliens within a community and the
proportion of AS within a community at ordinal
taxonomic rank are sufficient quantitative
indicators to provide an integrated estimation of
biocontamination. The ratio in numbers of AS to
all species, i.e. abundance contamination,
measures the community dominance by aliens.
Whereas the proportion of alien orders within a
community, i.e. ordinal richness contamination,
can also be interpreted as a proxy of disparity
contamination. Different macroinvertebrate
orders represent particular ecomorphological
groups associated with specific feeding patterns,
therefore, the higher taxonomic richness
provides an index of disparity (Foote 1997), and
this concept is relevant to community structural
organization. In parallel, richness contamination
at familial and specific levels were investigated,
when data were available, to test their utility for
the assessment of alien contamination. More-
over, the relationship between biocontamination
and ecological status in the inland waterways of
Europe is analyzed and discussed.
Material and Methods
Assessment of biocontamination
Biocontamination of sampling sites was assessed
using a site-specific biocontamination index
derived from two metrics: abundance contami-
nation index (ACI) and richness contamination
index (RCI) at ordinal rank. These indices were
calculated as:
ACI = Na/Nt
where Na and Nt are numbers of specimens of
alien taxa and total specimens in a sample,
respectively, and
RCI = Ta/Tt (1)
where Ta is the total number of alien orders, and
Tt is the total number of identified orders (for
recorded AS and taxonomic resolution applied in
this study see Annexes 1 and 2, respectively).
With values of ACI and RCI, the site-specific
biocontamination index (SBCI) can then be
derived from matrix in Table 1. Five classes of
biocontamination ranging from 0 (“no” contami-
nation) to 4 (“severe” contamination) are
defined. Furthermore, these classes of SBCI
directly correspond to five ecological quality
classes sensu the Common Implementation
strategy for the EU Water Framework Directive
(2000/60/EC) (European Community 2000;
European Communities 2003), and allow status
ranking from “high” to “bad” (Table 1). The
threshold for the lowest quality limit (“bad”
class) is based on the assumption that when AS
represent more than half the detected orders or
when their abundance exceeds 50% of the

K. Arbač i auskas e t a l . , Assessment o f b iocontaminat ion

213
individuals, the community/assemblage has
developed as a consequence of the invasion by
non-native species.
In those cases when multiple estimates of ACI
and RCI are available for the same ecosystem or
assessment unit (i.e. samples were collected at
several sites), the integrated biocontamination
index (IBCI) can be derived by averaging ACI
and RCI of study sites, and ranking IBCI on
mean values following Table 1.
Table 1. Assessment of site-specific and integrated biocontamination indices (SBCI and IBCI, correspondingly) based on
abundance contamination index (ACI) and ordinal richness contamination index (RCI). SBCI and IBCI classes: 0 (no bio-
contamination, “high” ecological status, blue cell), 1 (low biocontamination, “good” ecological status, green cell), 2 (moderate
biocontamination, “moderate” ecological status, yellow cells), 3 (high biocontamination, “poor” ecological status, orange cells),
4 (severe biocontamination, “bad” ecological status, red cells).
ACI
RCI
none 0.01 – 0.10 0.11 – 0.20 0.21 – 0.50 >0.50
none 0
0.01 – 0.10 1 2 3 4
0.11 – 0.20 2 2 3 4
0.21 – 0.50 3 3 3 4
>0.50 4 4 4 4


Study sites and sampling
The method described above was tested on
several extensive data sets from the main inland
waterways of Europe. Samples of benthic macro-
invertebrates for evaluation of biocontamination
and ecological status of aquatic ecosystems were
collected in selected assessment units (AUs)
located within the three main European inland
invasion corridors (sensu Bij de Vaate et al.
2002; Galil et al. 2007; see Figure 1): 1) two
AUs within the Northern corridor (NC) - Neva
Bay (NC5, 8 study sites, 1999) and Lake Ladoga
(NC4, 8, 2000); 2) eight AUs within the Central
corridor (CC) - Lower Pripyat River (CC8, 5,
2007), middle Pripyat River (CC9, 3, 2007),
Pripyat-Bug canal (CC10, 5, 2007), middle
Nemunas River (CC11, 5, 2007), lower Nemunas
River (CC12, 7, 2007), Bug River (CC14a, 5,
2003), Vistula River (CC14b, 28, 2000), and
Oder River (CC16, 14, 2001); and 3) nine AUs
within the Southern corridor – lower Danube
River (SC2, 3, 2007), middle Danube River
(SC3, 3, 2007), Sava River (SC3a, 3, 2006), Tisa
River (SC3b, 3, 2001), upper Danube River
(SC4, 3, 2007), Main-Danube canal (SC5, 1,
1998), Main River (SC6, 2, 2001 and 2007),
Rhine River (SC7, 3, 2006), and Rhine River
Delta (SC8, 1, 1987-1999). In addition, the AU,
Sukhoy Liman, was selected within the Southern
Meridian corridor, as it links the southern parts
of all three main invasion corridors (SMC1, 12,
2008) (Annex 3).
In most cases, sampling was performed by
procedures comparable with AQEM methodo-
logy (AQEM 2002). In Neva Bay and Lake
Ladoga, samples were collected in the shallow
littoral (0.2-0.5 m depth) zone using a stovepipe
sampler designed for quantitative collection of
macroinvertebrates in reed beds (Panov 1996).
Samples from the Pripyat River and Pripyat-Bug
canal were collected using a dip net at depths
between 0.2 and 0.5 m by sweeping a 5 m
distance, in total. Samples from the Nemunas
River were collected from shoreline to 1.2 m
depth using a dip net over various substrates and
vegetation for 15 min by two persons. In the
Bug, Vistula and Oder rivers, a standard semi-
quantitative procedure using a benthic dip net
was applied (Grabowski et al. 2006). In the
southern part of the Southern invasion corridor,
sampling was also performed using a benthic dip
net during three surveys: International Tisa
Research (2001, organized by the International
Commission for the Protection of the Danube
River), Sava Survey (2006, supported by the
Serbian Government) and Joint Danube Survey 2
(2007, supported by the International Commi-
ssion for the Protection of the Danube River).
Data for the Main-Danube canal, the Main River
and the Rhine River were provided by the
German Federal Institute of Hydrology. The
sampling was performed from a ship by means of
an orange-peel grab. Data on macroinvertebrates
assemblages of the Rhine River Delta near

K. Arbač i auskas e t a l . , Assessment o f b iocontaminat ion
214
Lobith (rkm 882) in the Netherlands were
obtained from the Dutch Institute for Inland
Water Management and Waste Water Treatment
(RIZA). Over the period 1987-1999, macro-
invertebrates were collected using artificial
substrates (i.e. baskets filled with marbles). The
annual taxa richness and relative abundance of
macroinvertebrates were based on pooled data
(two baskets per sampling date; four to seven
sampling dates from spring to autumn).
Macroinvertebrate samples were preserved in
4% formaldehyde or 70% alcohol. Animals were
picked from whole samples or, when necessary,
sub-sampling was applied.


Figure 1. Selected assessment units (AUs) within European inland waterways (see Annex 3 for details). Dashed-dotted, dotted
and solid lines indicate the Northern, Central and Southern invasion corridors, respectively. Colour fillings of AUs indicate
integrated biocontamination estimates during the studied period (see Table 1 and Annex 3).

Calculations and statistical analysis
Site-specific indices of abundance and ordinal
richness contamination were calculated for all
study sites. From these estimates, the SBCI for
all study sites and the IBCI for all AUs were
derived. Numbers of AS within study sites and
AUs were counted for further analysis. In many
cases, the richness contamination at familial rank
was also evaluated following equation (1) and
considering Ta and Tt as the number of alien
families and the total number of families,
respectively. If monitoring data existed at
species-level resolution, the specific richness
contamination indices were analogically calcu-
lated. Estimates of ordinal, familial and specific
RCI were compared for their performance and
utility for evaluation of biocontamination.
The relationship between biocontamination
and ecological quality status was assessed using
the SBCI (and metrics of its derivation, when
appropriate) and BMWP (Biological monitoring
working party) scores at sampling sites. The
BMWP method is widely used in the EU and has
proved to be among the best indicators of
ecological quality of water bodies (Armitage et
al. 1993; Leeds-Harrison et al. 1996; Semen-
chenko and Moroz 2005).
Continuous variables were analysed by
ANCOVAs with studied waterways (which

K. Arbač i auskas e t a l . , Assessment o f b iocontaminat ion

215
actually corresponded to countries) as a catego-
rical predictor. This approach was taken because
differences resulting from biogeographical
factors and biocontamination level or applied
sampling methodologies might be responsible for
significant variation. Variation in continuous
variables within biocontamination classes
(categorical variable) was analysed by nested
ANOVAs, controlling for waterway effect. Sta-
tistical calculations were performed using Statis-
tica software (StatSoft, Inc., 2004, version 6).
Results
Annex 1 lists the AS recorded in various AUs.
This list includes only those species that were
identified in benthic samples from this study and
were used in the calculation of biocontamination
indices. Although the mysid Paramysis lacustris
is known to occur over the entire middle section
of the Nemunas River (CC11) (Arbačiauskas
2005), this species was not found in the dip net
samples (probably due to its low density).
Similarly, the polychaete Hypania invalida was
found only in samples collected by Ponar grab in
the port bay of Brest, the Pripyat-Bug Canal
(CC10), but not in the dip net samples. Conse-
quently, these two species were not indicated for
those AUs. Therefore, it should be noted that
Annex 1 does not represent a complete list of AS
known from the European inland waterways
included in this study.
Of the 43 AS listed in Annex 1, 24 are
crustaceans, including13 species of amphipods, 3
species of mysids, isopods and decapods, and 2
species of balanids. Molluscs are represented by
8 snails and 7 mussels, while only one species
each of flatworms, oligochaetes and polychaetes
are included. In rivers of the Baltic and North
Sea basins peracaridans prevail over other
invaders with respect to species richness,
whereas in the Black Sea basin rivers the
diversity of alien molluscs is higher than that in
more northern areas (Annex 1).
Site-specific and integrated estimates of abun-
dance contamination and ordinal richness conta-
mination (also familial and specific richness
contamination when data were sufficient) as well
as biocontamination assessments and BMWP
scores at study sites are given in Annex 3.
The time series data that existed for the Rhine
River Delta (SC8) allowed for the investigation
of a temporal trend in biocontamination (Figure
2A). These data indicate that richness contami-
nation preceded abundance contamination. When
comparing RCI at different taxonomic ranks, the
highest values were estimated at the ordinal level
while the lowest values were estimated at
specific level. In total, 17 non-indigenous macro-
invertebrate species were recorded on artificial
substrates in the River Rhine at Lobith (Annex
1). RCI estimates changed little during 1987-
1991 but increased between 1992 and 1998. This
increase in local fauna contamination could be
mainly attributed to the invasions of Ponto-
Caspian species after the opening of the Rhine-
Main-Danube Canal, or so-called Southern
invasion corridor, in 1992 (Van der Velde et al.
2000; Bij de Vaate et al. 2002).
In contrast to RCI estimates, the ACI
indicated a rapid increase in the abundance of
AS over the period 1987-1991, followed by a
state of dynamic equilibrium (mean value of
0.83) during 1992-1999 (Figure 2A). The high
ACI values could be mainly attributed to five AS
that dominated the macroinvertebrate assemb-
lages in those years: the isopod, Jaera istri,
amphipods, Chelicorophium curvispinum,
Dikerogammarus villosus, Chaetogammarus
ischnus and Gammarus tigrinus, and bivalve
molluscs, Dreissena polymorpha and Corbicula
fluminea.
Based on BMWP scores, the ecological
quality of the Rhine River Delta appeared to be
moderate over the period 1987-1999 (Annex 3).
However, the BMWP does not consider the
contribution of AS to macroinvertebrate
assemblages. The SBCI scores clearly indicate a
decrease of ecological quality from poor (SBCI =
3) to bad (SBCI = 4), due to severe biocontami-
nation of the river (Annex 3).
The surveys of the Nemunas River, Pripyat
River and Pripyat-Bug Canal allowed an analysis
of the spatial patterns of biocontamination. Eight
and five AS were recorded in the lower and
middle sections of the Nemunas River (CC11
and CC12), respectively (Annex 1). As indicated
by the ACI estimates, biocontamination of the
lower section of the river was severe (Figure
2B), resulting in bad IBCI-estimated ecological
status (Annex 3). This low status mainly resulted
from large numbers of the pontogammarid,
Pontogammarus robustoides, the snail, Litho-
glyphus naticoides, and mysids, P. lacustris and
Limnomysis benedeni. The high abundance of
these AS was expected because this section of
the Nemunas River is located downstream of the
Kaunas Water Reservoir into which Ponto-
Caspian peracaridan species were intentionally

K. Arbač i auskas e t a l . , Assessment o f b iocontaminat ion
216
introduced during the early 1960s (Arbačiauskas
2002). In contrast, the most biologically
contaminated part of the middle section of the
Nemunas River was at the most downstream
study site where the Kaunas Water Reservoir
begins, whereas the presence and abundance of
AS in macroinvertebrate communities decreased
upstream (Figure 2B). The two most upstream
study sites had SBCI scores of 2 (moderate
biocontamination), whereas the ecological status
of the entire middle section of the Nemunas
River must be considered as poor (IBCI = 3)
(Annex 3). Similarly to the Rhine River Delta,
the highest values of RCI were obtained using
ordinal taxonomic rank. However, at sites with
high ACI, estimates of richness contamination
were quite similar for the three different
taxonomic resolutions (Figure 2B).
In the Pripyat River (CC8 and CC9) and
Pripyat-Bug Canal (CC10), ten AS were
recorded (Annex 1). Generally, the number of
non-native species decreased with increasing
distance from the potential donor area, the Kiev
Water Reservoir located on the Dniepr River
downstream from the inflow of the Pripyat
River. Also there was an obvious increase in the
number of AS at river ports compared to other
study sites with the mean number being
significantly higher even if the most downstream
site was included (5.2 vs. 2.8 species, Kruskal-
Wallis test: H(1,n=13)=4.14, P<0.042). The most
common AS were the pontogammarids,
Dikerogammarus haemobaphes and D. villosus,
and the snail, L. naticoides. The latter species
was responsible for the dramatic abundance
contamination in the most downstream study site
of the Pripyat River. This snail accounted for
more than 1000 specimens per sample. Conse-
quently, the highest SBCI was observed at that
study site. The second highest biocontamination
was observed at the Mykashevichy River Port
(Site 2 on the Pripyat River, Figure 2C), which is
characterized by a high number of ship calls.
One more aspect of biocontamination revealed
itself in the Pripyat-Bug Canal. The highest
contamination here was found at Site 3.
Although only 2 AS were recorded, their share
among native taxa was comparatively high
because this part of the canal is artificial and the
diversity of native species is low compared to
other parts that are closer with respect to hydro-
morphology of natural rivers.
It should be noted that in most study sites of
the Pripyat River and Pripyat-Bug Canal, the

Figure 2. Temporal (A) and spatial (B, C) trends of
abundance contamination (ACI) and richness contamination
at specific (TCIs), familial (TCIf) and ordinal (TCIo) ranks
in the Rhine River Delta (A), the Nemunas River (B) and
the Pripyat River and Pripyat–Bug Canal (C), respectively.
Study sites in (B) and (C) are arranged in upstream-
downstream direction indicated by horizontal arrow (for
coordinates see Annex 3). Vertical arrows in (C) indicate
river ports.
richness contamination indices received higher
values than those for abundance contamination
(Figure 2C). Such a pattern suggests that bio-
contamination of this waterway is ongoing and
the decrease of ecological quality due to growth
of abundance contamination may be expected.
On the whole, two AUs within the Pripyat River

K. Arbač i auskas e t a l . , Assessment o f b iocontaminat ion

217
and the Pripyat-Bug Canal were assessed as
moderately biocontaminated systems (IBCI = 2)
(Annex 3).
Neva Bay (NC5) and Lake Ladoga (NC4)
were surveyed during 1999 and 2000,
respectively. The littoral zone of Neva Bay, the
freshwater part of the Neva River estuary, was
contaminated by two alien amphipods, the
Baikalian Gmelinoides fasciatus and Ponto-
Caspian P. robustoides. They were present in
most sampling locations. ACI and RCI at
familial and ordinal ranks ranged 0.082-0.470,
0.067-0.200 and 0.083-0.250, respectively,
resulting in estimation of IBCI of the littoral
zone of Neva Bay as moderate (Annex 3). In
Lake Ladoga, only one AS was recorded: the
Baikalian amphipod, G. fasciatus. Still this
single AS accounted for a dramatically large
proportion of the organisms at most study sites.
Consequently, SBCI estimates at most locations
were high, and IBCI for Lake Ladoga littoral
macroinvertebrate communities was estimated at
4, which indicates severe biological contami-
nation (Annex 3).
During the survey of the Bug River (CC14a)
in 2003, 4 alien macroinvertebrates, all of them
amphipods of Ponto-Caspian origin, C. curvi-
spinum, C. ischnus, D. haemobaphes and
D. villosus, were recorded (Annex 1). These
species occurred at two study sites located
downstream from the Pripyat-Bug Canal. Their
high abundances were responsible for high
values of SBCI at these locations, while the
remaining three sites were devoid of AS. The
IBCI value for the Bug River was 3 (Annex 3).
Altogether 7 AS were found during the survey
in the Vistula River (CC14b) (Annex 1). Among
them were the snail, Potamopyrgus antipodarum,
and the bivalve, D. polymorpha. All the other
species were crustaceans, including amphipods
of Ponto-Caspian origin, C. curvispinum,
C. ischnus, D. haemobaphes and D. villosus, and
the American crayfish, Orconectes limosus. AS
occurred at 22 of 28 sampled locations. Gene-
rally, the observed pattern shows that AS were
lacking in most upstream parts of the Vistula
River, while their presence in macroinvertebrate
communities increased downstream, resulting in
the increase of SBCI values from 1 to 3 in the
middle part of the river, and reaching 4 in the
lowest part. The high biocontamination indices
were mainly associated with a high abundance of
Dreissena and alien amphipods (Annex 3).
A total of 7 alien macroinvertebrates were
recorded from benthic samples in the Oder River
(CC16) during 2001. There were two Ponto-
Caspian molluscs, the gastropod, L. naticoides,
and the bivalve, D. polymorpha. The remaining
species were amphipods, including the North
American G. tigrinus, South European Gamma-
rus roeselii and Ponto-Caspian C. curvispinum,
D. haemobaphes and D. villosus (Annex 1). AS
occurred in 13 of 14 sampled locations. The only
site free of aliens was a study site in the most
upstream part of the river in Poland. At all the
other locations, the dominant AS, in terms of
abundance, were C. curvispinum and G. tigrinus,
with poor representation of native fauna. Thus,
with the exception of one and two sites with
SBCI values of 2 and 3, respectively, the
remaining 10 locations had high SBCI scores. As
a result, the IBCI for the Oder River was also
estimated at 4, indicating severe biocontami-
nation and bad ecological status (Annex 3).
In the Danube River (SC2, SC3 and SC4),
which belongs to the Southern invasion corridor,
a total of 19 non-native species were found
(Annex 1). All these species were recorded in the
main channel of the Danube River. Along the
investigated stretch of the Sava River (SC3a), 9
AS were identified, and 3 AS were recorded in
the Tisa River (SC3b). The most frequent and
abundant species was L. naticoides. This snail
was recorded at all sampling sites, with relative
abundance ranging from 0.8 to 54.4% of the total
benthic community. Considerable occurrence and
relative abundance were recorded for D. villosus.
This pontogammarid was found in 73% of all
samples, with relative abundance ranging from
0.4 to 43.8%. In addition to those species, the
mussel, C. fluminalis, was present in all
investigated AUs, except the Tisa River. The
tubificid worm, Branchiura sowerbyi, was
recorded in the middle Danube, along the entire
lower Danube and in surveyed tributaries. The
distribution of this species generally is asso-
ciated with hydromorphological modification of
rivers (Paunović et al. 2005). The Ponto-Caspian
mysid, L. benedeni, was found to be limited to
the lower and middle Danube River. Meanwhile
the polychaete, H.invalida, was found in samples
from the upper and middle Danube River. Its
relative abundance varied between 0.4 and 3.9%
of the total benthic community. This species was
also recorded in the Sava River. Both dreissenid
species were also detected among the non-native
fauna of the Danube River. The zebra mussel,
D. polymorpha, which is native for estuaries and
coastal waters of the Ponto-Caspian and the Aral
Sea basins and associated estuaries, was more

K. Arbač i auskas e t a l . , Assessment o f b iocontaminat ion
218
abundant and outspread in the Danube River,
while the quagga mussel, Dreissena rostriformis
bugensis, which is native to the Dnieper and Bug
Limans of the northern Black Sea, was found
limited to the lower Danube River (Annex 1).
The majority of recorded non-native species are
of Ponto-Casian origin (14 species). Beside
these, aliens from New Zealand (the mudsnail,
P. antipodarum) and Eastern Asia (the Chinese
pond mussel, Sinanodonta woodiana, Asian
clams, C. fluminea and Corbicula fluminalis, and
the tubificid worm, B. sowerbyi) were recorded.
Assessments of the alien contamination and
ecological status of the lower and middle Danube
River and its tributaries (five AUs) suggest a
high level of biocontamination and low status of
ecological quality along the main course of the
river, as well as in its tributaries, the Sava River
and the Tisa River. With respect to IBCI, all
AUs showed severe biocontamination and
consequently indicated bad ecological status
(Annex 3).
In the northwestward-located parts of the
Southern invasion corridor, the Main-Danube
Canal (SC5), Main River (SC6) and Rhine River
(SC7), 3, 9 and 10 AS were recorded (Annex 1).
In common to the Danube River, the Main and
Rhine rivers indicated severe biological contami-
nation and bad ecological status. Whereas in the
Main-Danube canal, since 1992 connecting these
rivers, the ecological status with respect to AS
was estimated as poor, consequently one rank
better than in the rivers (Annex 3). Such an
estimate was derived from data collected in 1998
(since then no surveys were done). Within the
last decade, at least 5 new AS from the Black
Sea basin (including amphipods Chelicorophium
robustum and Chaetogammarus trichiatus) have
spread via the canal to the Main River, and
further on to the Rhine River (Bernauer and
Jansen 2006). The dispersal of AS in the
opposite direction also is ongoing, and at least
one macroinvertebrate species, the Chinese
mitten crab, Eriocheir sinensis, has used the
canal to penetrate into the Danube River
(Rabitsch and Schiemer 2003). Currently the
biocontamination of the Main-Danube Canal
probably is substantially higher, and the ecologi-
cal status of this water body has changed from
poor to bad.
Assessment of biocontamination was applied
also to Sukhoy Liman (SMC1), the AU which
includes marine and freshwater environments. In
marine part of Sukhoy Liman, which includes the
nearby Commercial Sea Port of Illichivsk, that is
known to harbor numerous non-indigenous
species (Koshelev and Son 2007), low densities
of AS were detected. In contrast, only the New
Zealand mudsnail, P. antipodarum, was present
in small rivers and streams emptying into the
liman, however, this species was very abundant.
As a result, ACI estimates, and consequently
SBCI estimates, for freshwater locations were
higher than those for marine locations (Annex 3).
The survey of main European waterways
clearly suggests that benthic macroinvertebrate
communities in all studied AUs are biologically
contaminated, with integrated biocontamination
indices ranging from 2 (moderate biocontami-
nation or moderate ecological status) to 4 (severe
biocontamination or bad ecological status).
Highly biocontaminated water bodies include the
littoral zone of Lake Ladoga, the lower section
of the Nemunas River, the Oder River, the Rhine
River and its delta, the Main River and the
Danube River and its sampled tributaries, with
mean relative abundance of AS exceeding 50%
of the macroinvertebrate community. Only the
Pripyat River, the Bug River, their joining canal,
and the Neva Bay (during 1999 when sampling
was performed) were found to be moderately
biocontaminated systems (Figure 1, Annex 3).
The relationship between abundance contami-
nation and richness contamination at different
ranks was analysed over those AUs for which
estimates on all taxonomic resolutions were
available, i.e. the Rhine River Delta, Nemunas
River, Pripyat River, Pripyat-Bug Canal, and the
Danube River and its tributaries. A significant
correlation was observed between abundance
contamination and richness contamination for
each taxonomic rank (Figure 3). As specific con-
ditions within different waterways may influence
these relationships, an ANCOVA with waterway
as a fixed factor and ACI as a covariate was
applied. Partial correlations between ACI and
RCI were significant for specific (r=0.52,
P<0.001) familial (r=0.53, P<0.001) and ordinal
(r=0.34, P<0.017) ranks with the weakest
correlation for ordinal level. The later may be
interpreted also as the indication by ordinal
richness contamination of different aspect of
biocontamination in comparison to that measured
by abundance contamination.
Of interest is how indices of abundance and
richness contamination varied within different
biocontamination classes. For this analysis, data
were used from those AUs for which information
on ordinal and familial RCI and spatial variation
of site-specific estimates were available (see

K. Arbač i auskas e t a l . , Assessment o f b iocontaminat ion

219
Annex 3). Variation of metrics showed that the
separation between moderate, high and severe
biocontamination resulted primarily from abun-
dance contamination estimates, whereas, diffe-
rentiation between good and moderate ecological
status with respect to AS depended upon
estimates of richness contamination (Figure 4).
This is in accordance with above described tem-
poral and spatial trends of biocontamination (see
Figure 2) which definitely indicate that richness
contamination precedes abundance contami-
nation. Consequently, holistic assessment of
biocontamination should consider and integrate
both measures of contamination by AS.
Variation of richness contamination at familial
rank within SBCI classes suggests that this index
actually may be used as a proxy SBCI for
separation between low, moderate and high bio-
contamination (Figure 4). The threshold values
between adjacent quality classes may be set as
means between 75 and 25 percentiles of variation
within higher and lower biocontamination
classes, respectively. From current data, familial
RCI values for good, moderate and poor
ecological status range 0-0.07, 0.08-0.15 and

Figure 3. Relationships between abundance contamination
index (ACI) and richness contamination index at specific
(RCIs), familial (RCIf) and ordinal (RCIo) ranks. Lines
indicate linear fit. Correlations between ACI and RCIs, RCIf
and RCIo are 0.60, 0.65 and 0.59 (n=52, P<0.001),
correspondingly.

Figure 4. Variation of abundance contamination index (ACI) and richness contamination index at ordinal (TCIo) and familial
(TCIf) ranks within different biocontamination classes: low (1), moderate (2), high (3) and severe (4).

>0.15, respectively. Such a procedure is in
accordance with recommendations for the
establishment of thresholds for ecological quality
classes (see European Communities 2003).
When assessing biocontamination another
important aspect is that severe abundance
contamination of recipient communities may be
caused by a single invasive AS. Although the

K. Arbač i auskas e t a l . , Assessment o f b iocontaminat ion
220
correlation between the number of non-
indigenous species in a community and
abundance contamination was significant in the
current study, the number of AS explained just
9% of total ACI variation (Figure 5). In
particular, high abundance contaminations
caused by just one or two AS were observed in
Ladoga Lake, the Vistula River, the Oder River
and the Sukhoj Liman (Annex 3).

Figure 5. Relationship between number of alien species in a
community and abundance contamination index. One
estimates per study sites involved, i.e. for the Rhine River
Delta only the last estimate was used to address pseudo-
replication.
The second purpose of this study was to
investigate the relationship between biocontami-
nation and ecological quality status assessed by
conventional methods. Estimates of ecological
status may depend upon biogeographical reasons,
sampling methods and other local factors,
therefore, BMWP estimates were subjected to
nested ANOVA with SBCI ranks nested within
waterway classes. Since the number of BMWP
score estimates for high, good and moderate
quality classes (with respect to biocontami-
nation) were few, the BMWP estimates for those
classes were merged into one group and only
those waterways wherein estimates for at least 3
SBCI classes were available were involved in
this analysis. As country effect when comparing
BMWP scores for study sites located in Belarus
and Poland was not significant (nested ANOVA,
country effect: F1,26=0.7, P=0.41), those BMWP
estimates were further merged into one group.
Results of the analysis showed that the
ecological quality status estimated by BMWP
method was significantly influenced by bioconta-
mination (Table 2, Figure 6). Sites with higher
biocontamination had lower ecological quality.
Furthermore, the highly significant effect of
waterway primarily reflects the different
sampling effort (sampling method) used in
Table 2. ANOVA assessing the impact of site-specific bio-
contamination index (SBCI) class nested in waterway effect
on ecological quality status measured as BMWP scores.
Effect MS df F P
Intercept 192397 1 250.0 <0.001
Waterway 24223 2 31.5 <0.001
SBCI 2482 6 3.2 0.009
Error 770 50


Figure 6. BMWP score weighted means in 3 groups of
differently biocontaminated study sites (SBCI: 0-2, 3 and 4)
located within Ladoga Lake and Neva Bay, the Pripyat, Bug
and Oder rivers and Pripyat-Bug canal, and the Nemunas
River. For results of nested ANOVA see Table 2. Vertical
bars denote 0.95 confidence intervals.
different countries. It seems that the larger
sampling effort applied in the Nemunas River
resulted in larger numbers of recorded taxa and,

K. Arbač i auskas e t a l . , Assessment o f b iocontaminat ion

221
consequently, higher BMWP estimates in
comparison to other AUs.
The above analysis only revealed that study
sites with higher biocontamination receive lower
ecological quality estimates by BMWP method,
however, it does not answer the question whether
biocontamination directly affects ecological
status estimates by conventional methods.
Consequently, ANCOVAs using waterway (i.e.
country) as a fixed factor and indices of
biocontamination (ACI and ordinal and familial
RCI) as covariates were conducted. To enable
the traditional ANCOVA design, i.e. to remove
the interaction of continuous and categorical
predictors, which was detected when using
familial RCI as covariate (homogeneity-of-slope
model, interaction effect: F4,64=5.26, P<0.002),
prior to analysis, BMWP scores were log-
transformed. This analysis clearly shows that
biocontamination significantly affects estimates
of ecological status by BMWP method (Table 3,
Figure 7). As in the previous analysis, the
waterway effect also was significant. Estimates
of ACI and RCI at familial and ordinal ranks
were negatively correlated with BMWP scores
(partial correlations were −0.36, −0.62 and
−0.60, correspondingly), with a stronger negative
correlation between indices of richness
contamination and BMWP. The later suggests
that the negative influence of richness conta-
mination on BMWP estimates is stronger than
that for abundance contamination.

Table 3. ANCOVAs assessing the impact of abundance
contamination index (ACI) and richness contamination
indices at ordinal (TCIo) and familial (TCIf) ranks, as
covariates, on log-transformed BMWP scores measured in
different waterways.
Effect MS df F P
ACI 0.3770 1 10.0 0.002
Waterway 0.9247 4 24.4 <0.001
Error 0.0378 68
TCIo 1.0561 1 37.9 <0.001
Waterway 0.9777 4 35.1 <0.001
Error 0.0279 68
TCIf 1.1354 1 42.5 <0.001
Waterway 0.9128 4 34.2 <0.001
Error 0.0237 68




Figure 7. Relationships between abundance contamination index (ACI) and richness contamination indices at ordinal (RCIo) and
familial (RCIf) ranks, and ecological status estimated by BMWP scores. Data for the Nemunas River (1), and the Pripyat, Bug
and Oder river including Pripyat-Bug Canal (2) are highlighted. Note logarithmic scale. Solid and dashed lines indicate linear fit
for 1 and 2, respectively.

K. Arbač i auskas e t a l . , Assessment o f b iocontaminat ion
222

Discussion
We attempted in this study to assess the bio-
contamination of benthic macroinvertebrate
communities in various inland waterways of
Europe. To achieve this it was first necessary to
develop a method for assessing biocontami-
nation. The method was designed to be simple
and applicable to data collected during routine
monitoring of water ecological quality. We
hypothesized that an integrated index of bio-
contamination should include two measures
characterizing different aspects of community
structural organization, a measure of community
dominance by AS which can be measured by
relative abundance of alien individuals in the
community (abundance contamination), and a
measure of alien contribution to community
disparity which can be assessed by the propor-
tion of alien taxonomic orders in the community
(ordinal richness contamination or disparity
contamination). The utility of richness contami-
nation at familial and specific ranks for biocon-
tamination evaluation also was investigated.
Data on spatial and temporal variation of AS
presence in benthic communities from different
European inland waterways suggest that richness
contamination precedes abundance contami-
nation. During initial phases of invasions,
ordinal RCI appears to be more sensitive to
changes in comparison to estimates of richness
contamination at familial or specific ranks (see
Figure 2), and that is in accordance with the
precautionary principle in environment manage-
ment. Estimates of specific richness contami-
nation (which also can be interpreted as diversity
contamination) appeared to be of low value for
assessment of biocontamination, at least with the
method described here. Moreover, identification
of all benthic macroinvertebrates to species level
requires a substantial sample processing effort.
This, however, does not imply that identification
of AS to species is not required. Meanwhile,
richness contamination at familial rank was
capable to separate between classes of low,
moderate and high biocontamination (see Figure
4), and after a more detailed analysis it may be
found to be applicable in the assessment of
biocontamination when only qualitative data are
available or for the fast screening of previously
collected information on benthic macro-
invertebrates.
Our biocontamination index derived from
estimates of abundance and richness composition
allows a biocontamination evaluation of specific
study sites as well as entire ecosystems or
assessment units. It can be recommended for
implementation in routine water ecological qua-
lity monitoring when multiple-habitat sampling
is applied, and data are sufficient to reflect the
taxonomic richness of benthic macroinverte-
brates. While we applied this method to benthic
macroinvertebrates, which are among the main
indicators of ecological status of flowing waters,
it may be extended to include other aquatic
organisms such as fish or macrophytes in order
to receive a more comprehensive assessment of
biocontamination in aquatic ecosystems.
Our assessment of biocontamination clearly
shows that the main inland waterways of Europe,
at least with respect to benthic macroinverte-
brates, are highly contaminated with AS. Since
data in a few assessment units were collected
some time ago, the current status of biocontami-
nation therein may be even higher. Highly
contaminated assessment units were identified in
all three European invasion corridors (Figure 1).
Severe biocontamination was observed in the
littoral area of Lake Ladoga, the lower section of
the Nemunas River, the Oder River, the Rhine
River, the Main River, and the Danube River and
its tributaries. In all these waterways, abundance
contamination was over 50%, and bad ecological
status with respect to biocontamination definitely
can be stated. Only the Pripyat River, the Bug
River and the Pripyat-Bug Canal can be
considered to be just moderately biologically
contaminated among the studied waterways.
This study also indicates that high abundance
contamination of recipient communities may be
caused by just a single ecologically aggressive
AS (Figure 5). In Lake Ladoga, for example, the
Baikalian amphipod, G. fasciatus, was capable
alone of causing very high biological contami-
nation of the littoral area. Other highly eco-
logically aggressive species include Ponto-
Caspian amphipods such as P. robustoides,
which frequently was responsible for severe
biocontamination in the Nemunas River, and also
D. villosus and D. haemobaphes. High abun-
dance contamination caused by the Ponto-
Caspian snail, L. naticoides, was also observed,
e.g. in the Pripyat and Danube rivers.
Information from the Pripyat River suggests that
river ports may facilitate the spread of AS across

K. Arbač i auskas e t a l . , Assessment o f b iocontaminat ion

223
inland waterways and promote richness
contamination. Meanwhile, artificial water
bodies, such as canals, are likely to be more
susceptible to biocontamination relative to
natural waterbodies.
The ecological quality of study sites assessed
by the BMWP method significantly varied
between studied waterways due not only to
biogeographical factors, but also because of
variation in sampling procedures applied in
different countries. However, study results also
clearly suggest that biocontamination and
ecological status of water bodies assessed by
conventional methods are in fact related. Study
sites with higher biocontamination have lower
estimates of ecological quality (Figure 6). This
observation may suggest that sites of lower
ecological status with respect to water quality
and hydromorphology are more susceptible to
biological invasions. The increase in industrial
and agricultural pollution along large European
rivers was hypothesized to be a trigger of mass
AS invasions (e.g. Jazdzewski and Konopacka
2002). However, since indicators of biocontami-
nation were negatively correlated with ecological
quality by BMWP method (Figure 7), this may
also indicate that biocontamination directly
affects ecological quality. Moreover, the
negative effect of richness contamination on
BMWP estimates seems to be stronger than that
originating from abundance contamination.
The negative correlation between biocontami-
nation and ecological status does not imply that
alien invaders are affecting water or hydro-
morphological quality of ecosystems (although
this theory may not be totally excluded) for
which metrics of ecological status, based on
native fauna, have been developed. Instead, it
implies that AS may impact native communities
and may cause a decreased estimate of ecological
status by suppressing local species and distorting
the true quality status. Negative impacts of aliens
on native species have been well documented.
For example, a decline in the macroinvertebrate
fauna following the arrival of D. villosus was
reported from the waters of the Netherlands
(Dick and Platvoet 2000; Van der Velde et al.
2000; Dick et al. 2002; Van Riel et al. 2006) and
France (Devin et al. 2001). The extermination of
a native amphipod by invaders was observed in a
large number of water bodies (Jazdzewski et al.
2004; Arbačiauskas 2005; Grabowski et al.
2006). In experiments, Krisp and Maier (2005)
have showed that D. villosus and C. ischnus
effectively consume the larvae of Ephemeroptera
that form the main indicator group for estimation
of ecological water quality. Negative impacts by
P. robustoides on abundance and diversity of
lake littoral communities have also been
documented (Arbačiauskas and Gumuliauskaitė
2007; Gumuliauskaitė and Arbačiauskas 2008).
Parts of European inland waterways that are
highly biologically contaminated are probably
irreversibly changed with respect to benthic
fauna composition. Communities formerly con-
sisting of native species are now alien-dominated
communities. During the period of species
composition change, they more properly may be
defined as assemblages. In some water bodies,
however, alien-dominated communities have
shown very stable composition of dominant
species for over a decade. When addressing the
dominance of non-native species, such newly
established communities may be defined as
xenocommunities (in analogy to xenodiversity,
sensu Leppäkoski and Olenin 2000). The
improvement of ecological status, i.e. bad status
with respect to biocontamination, in such water
bodies with alien-dominated communities is
unlikely, or too expensive. Meanwhile, the
concept of water ecological status has been
developed to consider the water physico-
chemical parameters and the hydromorphological
quality of river basins (European Community
2000). The water body status conferred by these
characteristics may be substantially higher, with
implementable management options in contrast
to the status defined by biocontamination.
Therefore, the water body status, with respect to
biocontamination, is probably more appropria-
tely defined as biological quality status, in order
to exclude the dual interpretation of ecological
quality status.
Although biocontamination in parts of main
European inland waterways have irreversibly
changed native communities, the problem of
biological pressures from AS must be considered
in water management strategies in order to
prevent, as much as possible, the further spread
of unwanted AS. These tasks are potentially
manageable as dispersal of non-native aquatic
species is by definition tightly associated with
human activities. Other urgent tasks for the
implementation of EU Water Framework
Directive are to: 1) specify methods of water and
hydromorphological quality assessment with
respect to AS presence; and 2) develop holistic
estimates of ecological quality status that
incorporate biocontamination of aquatic eco-
systems.

K. Arbač i auskas e t a l . , Assessment o f b iocontaminat ion
224
Only those impacts of AS invasions which
cause “fitness for survival” decrease are to be
considered the negative ecological impacts, i.e.
biological pollution (Elliott 2003). If we agree
on that, then translating the presence of AS into
an integrated ecological quality assessment of
aquatic ecosystem is no simple challenge. A
presence of AS may not only suppress but also
enhance ecosystem functions and services.
Hence, a rethink of the concept of ecological
quality status of inland waters should not be
excluded.
Conclusions
The biocontamination of study sites and aquatic
ecosystems or selected assessment units can be
assessed by site-specific biocontamination index
and integrated biocontamination index, respecti-
vely, which classifies water bodies into 5 quality
classes. These indices are derived from two
metrics, abundance contamination index and
richness contamination index at ordinal rank.
Metrics of contamination in abundance and
richness reflect the extent of alien contamination
in the structural organisation of communities.
This method of biocontamination assessment can
be applied to data collected during routine water
quality monitoring.
Severe biocontamination was observed in
most parts of European inland waterways and,
consequently, their status was classified as bad.
The quality status with respect to biocontami-
nation should be defined as biological quality
status. Prevention of biocontamination should be
considered in water management policy.
Estimates of ecological quality status
determined by conventional methods appear to
depend upon biocontamination. A specification
of methods for water and hydromorphological
quality assessment considering the presence of
alien species, and an elaboration of holistic
estimates of ecological quality status of aquatic
ecosystems that incorporate biocontamination
are warranted.
Acknowledgements
We are especially grateful to Ronald Griffiths for
constructive comments, valuable advice and
linguistic improvements. The manuscript also
profited from comments by Wolfgang Jansen and
Frances Lucy. This study has been supported by
the European Commission 6th Framework
Programme Integrated Project ALARM (contract
GOCE-CT-2003-506675) (Settele et al. 2005).
Study of the Nemunas River has been granted by
Lithuanian Environment Protection Agency.
References
AQEM consortium (2002) Manual for the application of the
AQEM system. A comprehensive method to assess
European streams using benthic macroinvertebrates,
developed for the purpose of the Water Framework
Directive. Version 1.0, February 2002. 202 рр.
Arbačiauskas K (2002) Ponto-Caspian amphipods and mysids
in the inland waters of Lithuania: history of introduction,
current distribution and relations with native
malacostracans. In: Leppäkoski E, Gollasch S, Olenin S
(eds) Invasive Aquatic Species of Europe - Distribution,
Impacts and Management. Kluwer Academic Publishers,
Dordrecht, The Netherlands, pp. 104-115
Arbačiauskas K (2005) The distribution and local dispersal of
Ponto-Caspian Peracarida in Lithuanian fresh waters
with notes on Pontogammarus robustoides population
establishment, abundance, and impact. Oceanological
and Hydrobiological Studies 34 (Suppl. 1): 93-113
Arbačiauskas K, Gumuliauskaitė S (2007) Invasion of the
Baltic Sea basin by the Ponto-Caspian amphipod
Pontogammarus robustoides and its ecological impact.
In: Gherardi F (ed.), Biological Invaders in Inland
Waters: profiles, distribution and threats Invading
Nature - Springer Series in Invasion Ecology, Vol. 2: pp.
463-477
Armitage PD, Moss D, Wright JF, Furse MT (1993) The
performance of a new biological water quality score
system based on macroinvertebrates over a wide range of
unpolluted running water sites. Water Research 17: 333-
347
Bernauer D, Jansen W (2006) Recent invasions of exotic
macroinvertebrates and loss of native species in the
upper Rhine River, Germany. Aquatic Invasions 1: 55-71
Bij de Vaate A, Jazdzewski K, Ketelaars HAM, Gollasch S,
Van der Velde G (2002) Geographical patterns in range
extension of Ponto-Caspian macroinvertebrate species in
Europe. Canadian Journal of Fisheries and Aquatic
Sciences 59: 1159–1174
Cardoso AC, Free G (2008) Incorporating invasive alien
species into ecological assessment in the context of the
Water Framework Directive. A paper for discussion at
the ECOSTAT workshop on alien species and the EC
Water Framework Directive, 2-3 April 2008, Bordeaux,
France. 7 pp
Devin S, Beisel JN, Bachmann V, Moreteau JC (2001)
Dikerogammarus villosus (Amphipoda: Gamma-
ridae):another invasive species newly established in the
Moselle river and French hydrosystems. Annales de
Limnologie 37: 21–27
Dick JTA, Platvoet D (2000) Invading predatory crustacean
Dikerogammarus villosus eliminates both native and
exotic species. Proceedings of the Royal Society,
London, Series B, 267: 977–983
Dick JTA, Platvoet D, Kelly DW (2002) Predatory impact of
the freshwater invader Dikerogammarus villosus
(Crustacea: Amphipoda). Canadian Journal of Fisheries
and Aquatic Sciences 59: 1078-1084

K. Arbač i auskas e t a l . , Assessment o f b iocontaminat ion

225
Elliott M (2003) Biological pollutants and biological
pollution––an increasing cause for concern. Marine
Pollution Bulletin 46: 275–280
European Community (2000) Directive 2000/60/EC of the
European Parliament and of the Council of 23 October
2000 establishing a framework for Community action in
the field of water policy. Official Journal of the
European Communities L327:1-72
European Communities (2003) Common Implementation
Strategy for the Water Framework Directive
(2000/60/EC) Guidance document no 10: Rivers and
lakes – typology, reference conditions and classification
systems. Produced by working group 2.3 – REFCOND.
Luxembourg: Office for Official Publications of the
European Communities, 87 pp
Foote M (1997) The evolution of morphological diversity.
Annual Review of Ecology and Systematics 28: 192-152
Galil BS, Nehring S, Panov VE (2007) Waterways as invasion
highways. Impact of climate change and globalization
In: Nentwig W (ed) Biological Invasions. Ecological
Studies Nr. 193. Berlin, Germany: Springer, pp.59–74
Gherardi F (ed) (2007) Biological invaders in inland waters:
profiles, distribution, and threats. Springer, Dordrecht,
The Netherlands, 733 pp.
Grabowski M, Konopacka A, Jazdzewski K, Janowska E
(2006) Invasions of alien gammarid species and retreat
of natives in the Vistula Lagoon (Baltic Sea, Poland).
Helgoland Marine Research 60: 90-97
Gumuliauskaitė S, Arbačiauskas K (2008) The impact of the
invasive Ponto-Caspian amphipod Pontogammarus
robustoides on littoral communities in Lithuanian lakes.
Hydrobiologia 599: 127-134
Jazdzewski K, Konopacka A (2002) Invasive Ponto-Caspian
species in waters of the Vistula and Oder basins and of
the Southern Baltic Sea. In: Leppäkoski E, Gollasch S
and Olenin S (eds) Invasive Aquatic Species of Europe -
Distribution, Impacts and Management. Kluwer
Academic Publishers, Dordrecht, The Netherlands, pp
384-398
Jazdzewski K, Konopacka A, Grabowski M (2004) Recent
drastic changes in the gammarid fauna (Crustacea,
Amphipoda) of the Vistula River deltaic system in
Poland caused by alien invaders. Diversity and
Distributions 10: 81-87
Koshelev AV, Son MO (2007) Some aspects of biodiversity
formation in isolated estuarine reservoir within Odessa
agglomeration. In: Natural and invasive processes of
formation of biodiversity of water and terrestrial
ecosystem’s biodiversity, SSC RAS Publishing, Rostov-
on-Don, p 172 [in Russian]
Krisp H, Maier G (2005) Consumption of macroinvertebrates
by invasive and native gammarids: a comparison.
Journal of Limnology 64(1): 55-59
Leeds-Harrison PB, Quinton JN, Walker MJ, Harrison KS,
Tyrrel SF, Morris J, Mills HT (1996). Buffer Zones in
headwater catchments. Report on MAFF/English Nature
Buffer Zone Project CSA 2285. Cranfield University,
Silsoe, UK, 22 pp.
Leppäkoski E, Gollasch S, Olenin S (eds) (2002) Invasive
Aquatic Species of Europe. Distribution, Impacts and
Management. Kluwer Academic Publishers, Dordrecht.
583 pp.
Leppäkoski E, Olenin S (2000) Xenodiversity of the European
brackish water seas: the North American contribution.
In: Pederson J (ed.) Marine bioinvasions. Proceedings of
the First National Conference, Massachusetts Institute of
Technology, January 24-27. pp 107-119
Molnar JL, Gamboa RL, Revenga C, Spalding MD (2008)
Assessing the global threat of invasive species to marine
biodiversity. Frontiers in Ecology and the Environment
6: doi:10.1890/070064
Olenin S, Minchin D, Daunys D (2007) Assessment of
biopollution in aquatic ecosystems. Marine Pollution
Bulletin 55: 379–394
Panov VE (1996) Establishment of the Baikalian endemic
amphipod Gmelinoides fasciatus Stebb. in Lake Ladoga.
Hydrobiologia 322: 187-192
Panov V, Gollasch S, Leppäkoski E, Olenin S (2002)
International Cooperation in Aquatic Invasive Species
Research, Information Exchange and Management in
Europe. Aquatic Invaders 13 (4): 1-4
Panov VE, Alexandrov B, Arbačiauskas K, Binimelis R, Copp
GH, Grabowski M, Lucy F, Leuven RSEW, Nehring S,
Paunović M, Semenchenko V, Son MO (2008) Assessing
the risks of aquatic species invasions via European
inland waterways: from concepts to environmental
indicators. IEAM (in press)
Paunovic M, Miljanovic B, Simic V, Cakic P, Djikanovic V,
Jakovcev-Todorovic D, Stojanovic, B, Veljkovic A
(2005) Distribution of non-indigenous tubificid worm
Branchiura sowerbyi (Beddard, 1892) in Serbia.
Biotechnology and Biotechnological Equipment 3. 91-97
Rabitsch W, Schiemer F (2003) Chinesische Wollhandkrabbe
(Eriocheir sinensis) in der österreichischen Donau
festgestellt. Österreichs Fischerei 56: 61-65
Semenchenko VP, Moroz MD (2005). Comparative analysis
of biotic indices in the monitoring system of running
water in a Biospheric Reserve. Water Resources 32(2):
200–203
Settele J, Hammen V, Hulme PE, Karlson U, Klotz S, Kotarac
M, Kunin WE, Marion G, O'Connor M, Petanidou T,
Peterseon K, Potts S, Pritchard H, Pyšek P, Rounsevell
M, Spangenberg J, Steffan-Dewenter I, Sykes MT, Vighi
M, Zobel M, Kühn I (2005) ALARM: Assesing LArge
scale environmental Risks for biodiversity with tested
Methods. GAIA - Ecological Perspectives for Science
and Society 14: 69-72
Simon KS, Townsend CR (2003) Impacts of freshwater
invaders at different levels of ecological organization,
with emphasis on salmonids and ecosystem
consequences. Freshwater Biology 48(6): 982-995
Van der Velde G., Rajagopal S, Kelleher B, Musko IB, Bij de
Vaate A (2000) Ecological impact of crustacean
invaders: general consideration and examples from the
Rhine River. In: The biodiversity crisis and crustacean.
Proceedings of the Fourth International Crustacean
Congress. Crustacean, Issues 12. A.A. Balkema,
Rotterdam, the Netherlands: 3-33
Van Riel MC, Van der Velde G, Rajagopal S, Marguillier S,
Dehairs F, Bij de Vaate A (2006). Tropic relationships in
the Rhine food web during invasion and after
establishment of the Ponto-Caspian invader
Dikerogammarus villosus. Hydrobiologia 565: 39-58