Dissolved organic matter, oxygen
condition, and UV light influence the
bioavailability, uptake, and toxicity
of chemical contaminants in the
aquatic environment. In this thesis,
chemical partitioning processes,
and sublethal toxic responses were
studied in freshwater species under
different exposure conditions.
Complementary techniques were
applied to investigate the underlying
toxicity mechanisms, with particular
emphasis on transcriptional and
metabolic changes.
Publications of the University of Eastern Finland
Dissertations in Forestry and Natural Sciences
isbn 978-952-61-1014-1 (nid.)
issnl 1798-5668
issn 1798-5668
isbn 978-952-61-1015-8 (pdf)
issn 1798-5676 (pdf)
dissertations | No 95 | Stanley Ozoemena Agbo | Impact of environmental factors on chemical bioavailability and toxicity to...
Stanley Ozoemena Agbo
Impact of environmental
factors on chemical bioavailability and toxicity to aquatic
invertebrates: consequences
on transcriptional and
metabolic regulations
Stanley Ozoemena Agbo
Impact of environmental factors
on chemical bioavailability and
toxicity to aquatic invertebrates:
consequences on transcriptional
and metabolic regulations
Publications of the University of Eastern Finland
Dissertations in Forestry and Natural Sciences No 95
STANLEY OZOEMENA AGBO
Impact of environmental
factors on chemical
bioavailability and toxicity to
aquatic invertebrates:
consequences on
transcriptional and metabolic
regulations
Publications of the University of Eastern Finland
Dissertations in Forestry and Natural Sciences
No 95
Academic Dissertation
To be presented by permission of the Faculty of Science and Forestry for public
examination in the Auditorium C2 in Carelia Building at the University of Eastern
Finland, Joensuu, on 25th of January 2013, at 12 o’clock noon.
Department of Biology
Kopijyvä Oy
Joensuu, 2013
Editors: Profs. Pertti Pasanen,
Pekka Kilpeläinen, and Matti Vornanen
Distribution:
Eastern Finland University Library / Sales of publications
[email protected]
www.uef.fi/kirjasto
ISBN: 978-952-61-1014-1 (nid.)
ISSNL: 1798-5668
ISSN: 1798-5668
ISBN: 978-952-61-1015-8 (PDF)
ISSN: 1798-5676 (PDF)
Author’s address:
University of Eastern Finland
Department of Biology
P.O. Box 111
80101 JOENSUU
FINLAND
email: [email protected]
Supervisors:
Professor Jussi V.K. Kukkonen, Ph.D.
University of Jyväskylä
Department of Biological and Environmental Science
P.O. Box 35 (Survontie 9)
40014 JYVÄSKYLÄ
FINLAND
email: [email protected]
Matti T. Leppänen, Ph.D.
Finnish Environment Institute (SYKE)
Research and Innovation Laboratory
P.O. Box 35 (Survontie 9)
40014 JYVÄSKYLÄ
FINLAND
email: [email protected]
Juha Lemmetyinen, Ph.D.
University of Eastern Finland
Department of Biology
P.O. Box 111
80101 JOENSUU
FINLAND
email: [email protected]
Jarkko Akkanen, Ph.D.
University of Eastern Finland
Department of Biology
P.O. Box 111
80101 JOENSUU
FINLAND
email: [email protected]
Reviewers:
Anna-Lea Rantalainen, Ph.D.
University of Helsinki
Department of Environmental Sciences
15140 LAHTI
FINLAND
email: [email protected]
Docent Kari Lehtonen, Ph.D.
Finnish Environment Institute (SYKE)
Marine Research Centre/Marine Spatial Planning Unit
P.O. Box 140
00560 HELSINKI
FINLAND
email: [email protected]
Opponent:
Docent Olli-Pekka Penttinen, Ph.D.
University of Helsinki
Department of Environmental Sciences
15140 LAHTI
FINLAND
email: [email protected]
ABSTRACT
Aquatic organisms face multiple threats that include exposure
to chemicals whose toxicity can vary with the prevailing
environmental conditions (DOC, UV light, and hypoxia).
Traditionally, estimations of toxicity are derived from total
environmental concentrations, and these alone may grossly
overestimate the actual risk of exposure to contaminants. In part,
this is because typical effect concentration values do not provide
sufficient evidence to assess the magnitude of toxic effects, since
critical volume estimates at target sites are seldom determined.
On the other hand, environmental factors that influence
chemical bioavailability and stability can exacerbate the toxicity
of contaminants in their natural conditions. Therefore, it is
crucial to investigate the influence of these factors on
bioconcentration and kinetics, but also to elucidate the
mechanisms of species’ response to waterborne contaminants.
In this thesis, the effects of environmental variables on
bioavailability of benzo(a)pyrene (B(a)P) were studied in
Lumbriculus variegatus and Chironomus riparius. The freely
dissolved fraction in test water was determined via solid phase
micro extraction (SPME) and used for the prediction of uptake
and elimination kinetics. The body burdens were assessed by
measuring [14C] in extracts of tissue homogenates. In a related
study, aqueous preparations of pentachlorophenol (PCP) and
phenanthrene were irradiated continuously while a decline in
toxicity was monitored by the increased mobility of Daphnia
magna neonates. Alterations in transcripts; endogenous
metabolites; and DNA adducts were studied in L. variegatus
exposed to cadmium, (B(a)P) or PCP, to further unravel some of
the underlying mechanistic processes in worms subjected to
chemical stress. Additionally, metabolic changes were assessed
in worms exposed to model chemicals, hypoxia (< 30% O2), or
the combination of both stressors.
The dissolved concentration of B(a)P correlated negatively
with the DOC contents, even though the kinetic modelpredicted BCF values were different between DOC treatments.
Looking at the distribution of B(a)P in tissue extract
homogenates, it is reasonable to infer that biotransformation
was faster in midges than uptake from solution. The
disappearance of PCP was evident under continuous irradiation
while phenanthrene was rather persistent. It underlines the
application of solar radiation for the degradation of waterborne
chemicals but also offers a possibility for toxicity monitoring of
environmental samples using D. magna.
Actin related processes, proteolysis, and antioxidant defense
systems were prominently affected at the transcriptional and
metabolic levels. Hypoxia appeared to be the more dominant
stressor compared to chemical treatments within the range of
concentrations that were tested. Altogether, it appears that the
complementary methods augment each other in predicting
substance toxicity at concentrations that normally would not
cause adverse effects in a whole organism. In a successful test,
combined methodologies can distinguish specific mechanistic
processes from the general unspecific response of test organisms.
This approach revealed toxic responses that otherwise would be
masked by the exclusive use of a single assessment method.
Despite the current knowledge of transcriptional and metabolic
alterations, more research is needed to explain the link between
the different biological responses in physiological terms.
Universal Decimal Classification:
502.51, 504.5, 574.64, 577.214, 591.05, 595.142
CAB Thesaurus: toxicity; toxic substances; water; water pollution;
bioavailability; transformation; organic matter; organic compounds; aquatic
environment; aquatic invertebrates; gene expression; transcription; hypoxia;
metabolism; metabolites; DNA microarrays; photolysis
Yleinen suomalainen asiasanasto: ekotoksikologia; haitalliset aineet;
ympäristömyrkyt; myrkyllisyys; vesieläimistö; selkärangattomat;
geeniekspressio; transkriptio; hypoksia; aineenvaihdunta; DNA-sirut
Preface
This research would not have been possible without the help
and support of many individuals. I would like to express my
gratitude to Professor Jussi Kukkonen for the opportunity to
conduct this work in his research group. Since then, Jussi’s
tutelage has enabled me to progress to achieve considerable
independence in the course of my work. Being able to consult
with you has always been a magnificent source of strength. I am
very thankful to Merja Lyytikäinen who has been very helpful
since my early days and relocation period but also throughout
the duration of this work.
Even though many more people than could be listed here
contributed in one way or another towards the completion of
this work, Victor Carrasco Navarro’s valuable support,
understanding and jokes, have provided me with a studious
environment in our shared office space. I wish to thank Kimmo
Mäenpää and Jani Honkanen for their patience while
introducing me to the laboratory instruments during my early
days in the department. Everyone was welcoming and assisted
me in one way or another. Marja, Julia and Riitta Pietarinen
helped me throughout with their superb technical expertise. The
friendly laughter that we shared between laboratory work and
in the coffee room helped tremendously and helped me to relax.
This work would have been impossible without the spirit of
team-work; useful conversations and support from Sari P., Kaisa,
Greta, Inna, Kukka, Suvi, Sebastian, Elijah, Anna-Maija, Juho,
Heikki, Paula, Arto; but also from the other members of staff in
the Department of Biology. Anita Tuikka was immensely
generous in sharing her wealth of experience throughout this
work.
I am indebted to my supervisors and co-authors for their
invaluable contributions and constructive criticisms from the
development and planning of the experiments up to their
publication. Special thanks go to Matti Leppänen, Jarkko
Akkanen, Juha Lemmetyinen, Sarita Keski-Saari, and Markku
Keinänen, for guiding me meticulously and with infinite
patience throughout the duration of this work. I am also
indebted to Eberhard Küster and Anett Georgi for their
magnificent supervisory support and warm hospitality during
my research at the Helmholtz Centre for Environmental
Research (UFZ), Leipzig, Germany. Several other colleagues
with whom I have remained in contact until now, helped to
make my stay memorable. We have remained friends forever! I
extend warm gratitude to all the members of the Keybioeffects
project but also to our research collaborators; particularly Rolf
Vogt, Dag Olav Andersen, Philipp Mayer, Zhixin Wang, Hailin
Wang and David Price, for their invaluable contributions.
This work has been supported mainly by the Marie Curie
research fellowship of the EU’s 6th framework programme
(MRTN-CT-2006-035695); the Academy of Finland projects
214545 and 123587; grants from the Finnish Cultural Foundation;
and the Department of Biology at the University of Eastern
Finland. The Finnish Doctoral Programme in Environmental
Science and Technology (EnSTe) is gratefully acknowledged for
their financial support enabling me to attend conferences and
training courses both locally and abroad.
Thomas ter Laak, Joop Hermens and Kees van Gestel were
very inspirational and continuously offered their support
during my formative years in the Netherlands. Finally, I wish to
thank my family, particularly my beloved wife Racheal for her
profound support, sacrifice and understanding in the course of
this work. A very special thank you goes to my brother Eddy for
helping to shape my interest in biology. It brings endless joy to
my life.
LIST OF ABBREVIATIONS
ADaM
AFW
ASTM
B(a)P
BAF
BCF
BPDE-DNA
CE-LIF
DOC
EC50
EST
GC-MS
HPLC
IARC
ISO
KDOC
ke
ku
LSC
MEOX
MSTFA
OECD
PAH
PCA
PCP
PDMS
PVC
qPCR
ROS
rpm
SPME
TLC
TMS
US EPA
UV
Aachener daphnien medium
artificial fresh water
American society for testing and materials
benzo(a)pyrene
biota accumulation factor
bioconcentration factor
benzo(a)pyrene diolepoxide-deoxyribonucleic acid
capillary electrophoresis laser induced fluorescence
dissolved organic carbon
effective concentration that immobilizes 50% of the
test population
expressed sequence tag
gas chromatography-mass spectrometry
high performance liquid chromatography
international agency for research on cancer
international standard organization
partitioning coefficient to DOC
elimination rate constant
uptake rate constant
liquid scintillation counter
methoxyaminated derivative
N-methyl-N-trimethylsilyltrifluoroacetamide
organization for economic co-operation and
development
polycyclic aromatic hydrocarbon
principal component analysis
pentachlorophenol
polydimethyl siloxane
polyvinyl chloride
quantitative polymerase chain reaction
reactive oxygen species
rotation per minute
solid phase microextraction
thin layer chromatography
trimethylsilylated derivative
United States environmental protection agency
ultra violet light
LIST OF TABLES
Table 1
Summary of the study designs depicting the
stressors and environmental variables tested in
various species.
Table 2
Euclidean distances between data points used as
the biological scores to depict the behavior of
genes per sample treatment, particularly during
cluster
analysis.
The
smaller
the
relative
geometric distance between data points, the more
the genes are likely to be grouped together. This
implies that samples with similar responses to
each other upon chemical treatments have
smaller values, whereas high values depict
increased disparity between the different samples.
Table 3
Levels of BPDE-DNA adducts in replicates of
DNA
samples obtained from
L. variegatus
exposed to B(a)P at different points in time.
LIST OF FIGURES
Figure 1
Schematic
illustration
of
the
experimental
designs and the processing of samples.
Figure 2
Venn diagrams for differentially expressed genes
in L. variegatus exposed to B(a)P, Cd or PCP for
the duration of 2, 6, 24 and 48 h. Values indicate
the number of common/different genes with a
differential expression of more than twofold
either up- or down-regulated as depicted by the
direction of the arrow (p-value < 0.01, LIMMA
“decide Tests” function).
Figure 3
GOEAST analysis of molecular function category
of genes in L. variegatus exposed to PCP for 24 h.
Boxes represent GO terms designated with the
appropriate GO ID, term definition, and extra
details organized as “q/m\t/k(p – value)”, where q
is the count of genes associated with the listed
GO ID (directly or indirectly) in our dataset; m is
the count of genes associated with the listed GO
ID (directly or indirectly) on the chosen platform;
t is the total number of genes on the chosen
platform; k is the total number of genes in our
dataset; p-value is the significance for the
enrichment in the dataset of the listed GO ID.
Boxes
marked
in
yellow
represent
the
significantly enriched GO terms. The enrichment
significance of each GO term is positively
correlated with the level of color saturation of
each node. White colored boxes depict the non
significant GO terms, while hierarchical trees not
having any significantly enriched GO terms were
not shown. Red arrows represent relationships
between two enriched GO terms; black solid
arrows represent relationships between enriched
and unenriched terms; while black dashed
arrows represent relationship between
unenriched GO terms.
two
LIST OF ORIGINAL PUBLICATIONS
This thesis is based on data presented in the following articles,
referred to by the Roman numerals I-IV.
I
Agbo S O, Akkanen J, Leppänen M T, Kukkonen J V K.
Bioconcentration of benzo(a)pyrene in Chironomus riparius
and Lumbriculus variegatus in relation to dissolved organic
matter and biotransformation.
Aquatic Ecosystem Health & Management. In press.
II
Agbo S O, Küster E, Georgi A, Akkanen J, Leppänen M T,
Kukkonen J V K. Photostability and toxicity of
pentachlorophenol and phenanthrene.
Journal of Hazardous Materials 189: 235-240, 2011.
III
Agbo S O, Lemmetyinen J, Keinänen M, Keski-Saari S,
Akkanen J, Leppänen M T, Wang Z, Wang H, Price D A,
Kukkonen J V K. Response of Lumbriculus variegatus
transcriptome and metabolites to model chemical
contaminants. Comparative Biochemistry & Physiology, Part C –
Toxicology & Pharmacology 157: 183-191, 2013.
IV
Agbo S O, Keinänen M, Keski-Saari S, Lemmetyinen J,
Akkanen J, Leppänen M T, Mayer P, Kukkonen J V K.
Changes in Lumbriculus variegatus metabolites under hypoxic
exposure to benzo(a)pyrene, chlorpyrifos and
pentachlorophenol: consequences on biotransformation.
Manuscript submitted to Chemosphere.
All the publications are reprinted with kind permission from
the publishers. The copyright for publication I is held by
Taylor & Francis Group, and for II and III by Elsevier.
AUTHOR’S CONTRIBUTION
The publications listed in this thesis were produced in
collaboration with all the authors. The contributions of each author
are stated as follows:
Paper I: The study was planned by JA, ML, JK, and SA. SA
performed the experimental work; processed and analyzed the
samples and data; interpreted the results; and wrote the article
under the supervision of coauthors.
Paper II: The study was planned by EK and SA. SA irradiated the
samples; conducted the animal exposures and sampling; the HPLC
analysis; and wrote the article with constructive comments from
the coauthors. EK and AG helped with the LC-MS analysis of
pentachlorophenol.
Paper III: The study was jointly planned by JL, JA, ML, JK, and SA.
SA performed the animal exposure and the sampling and
processing of samples. DP supplied the EST library, while JL
constructed the microarray slides and hybridized the RNAs with
support from Riitta Pietarinen. JL scanned the slides and processed
the array data to study differential expression changes. SA
extracted tissue samples and analyzed them on GC-MS with help
from SK-S. MK identified and quantified the endogenous
metabolites and offered excellent advice on the handling of the
metabolites data. ZW and HW analyzed DNA adducts. SA wrote
the manuscript with constructive comments from coauthors.
Paper IV: The study was planned by JL, JA, ML, JK, and SA. PM
supplied the passive dosing jars. SA performed the experimental
work, while MK identified and quantified the endogenous
metabolites. SK-S, MK, and SA interpreted the results. SA wrote
the manuscript with constructive comments from coauthors.
Jarkko Akkanen, JA; Matti Leppänen, ML; Jussi Kukkonen, JK; Eberhard
Küster, EK; Anett Georgi, AG; Markku Keinänen, MK; Sarita KeskiSaari, SK-S; Juha Lemmetyinen, JL; Philipp Mayer, PM; Zhixin Wang,
ZW; Hailin Wang, HW; David Price, DP; Stanley Agbo, SA
Contents
1 Background ................................................................................. 21
1.1 Impact of Industrialization on the Environment ........................ 21
1.2 Chemical Accumulation Index ................................................... 23
1.3 Chemical Distribution and Effect Assessment ........................... 24
2 Materials and Methods ............................................................. 29
2.1 Test Organisms .......................................................................... 29
2.1.1 Daphnia magna .......................................................................... 29
2.1.2 Lumbriculus variegatus.............................................................. 30
2.1.3 Chironomus riparius .................................................................. 31
2.2 Chemicals .................................................................................... 33
2.2.1 Phenanthrene and B(a)P............................................................. 33
2.2.2 Chlorpyrifos and PCP................................................................. 34
2.2.3 Cadmium .................................................................................... 35
2.3 Bioconcentration and Uptake of B(a)P in L. variegatus ............. 36
2.4 Photodegradation and Toxicity of PCP and Phenanthrene ........ 39
2.5 Molecular Changes in Worms Exposed to Chemicals ................ 40
2.5.1 Tissue Sample Preparation for GC-MS Analysis....................... 41
2.5.2 CE-LIF Immunoassay for DNA Adducts Analysis ................... 42
2.5.3 Microarrays ................................................................................ 42
2.5.4 PCR Validation of Array-based Expression Profiles .................. 44
2.6 Metabolic Response to Chemical and Hypoxic Stress ................ 44
2.6.1 Test Water and Exposure Conditions ......................................... 44
2.6.2 Tissue Residue Analysis of [14C]-labeled Chemicals ................... 46
3 Results and Discussion ............................................................. 47
3.1 Bioconcentration and Kinetics of B(a)P in L. variegatus ........... 47
3.2 Photodegradation and Toxicity of PCP and Phenanthrene ........ 51
3.3 Molecular Response to Toxicants and Hypoxia ......................... 55
3.3.1 Transcriptional and Metabolic Response in L. variegatus ......... 55
3.3.2 Metabolic Response to Chemical Treatment and Hypoxia ......... 59
4 Concluding remarks .................................................................. 65
5 Toxicity Assessment Methods and Future Perspectives .... 67
6 References ................................................................................... 69
20
1 Background
1.1
IMPACT OF INDUSTRIALIZATION ON THE
ENVIRONMENT
Rapid industrialization and the global economic changes
contribute to existing environmental challenges. Human
activities have an overwhelming impact on the overall quality of
the environment due to greenhouse gas emissions and the
release of complex chemical pollutants into the air, water and
land (Vitousek et al., 1997). In particular, the greenhouse gases
(e.g. ozone, carbon dioxide, methane) contribute to the present
global climate change and can potentially alter the reactive
chemistry of the troposphere (Vitousek et al., 1997).
On the other hand, the diversity and complexity of synthetic
chemicals have increased substantially since the industrial
revolution. Similarly, both the volume and variety of pollutants
that are released into the environment have equally increased in
recent decades. In regulating the distribution of these pollutants,
various government agencies have established standards for
controlled emissions as well as the application of biocides in
agriculture. Even though the existing practices intend to
ameliorate the degradation of the environment, both the surface
and ground water resources have continued to be polluted via
direct discharge and surface runoffs from waste dumps and
agricultural sites (Wan, 1994; Arfsten, 1996). However, these
pollutants usually occur in nature as complex mixtures that are
prone to impacts of environmental factors, thereby contributing
to their bioavailability, toxicity and elimination from exposed
organisms. Some of the newly emerging pollutants including
21
nanoparticles are generally recalcitrant and take a longer time to
undergo transformation or complete degradation.
Ecological
health
is commonly used
to predict
the
consequences of human exposure to potentially harmful
chemical
pollutants.
The
Silent
Spring,
though
slightly
exaggerated, was a wake-up call to the dangers of pesticide
applications and consequences on non-target organisms, notably
birds (Carson, 1962). The book criticized the existing practices in
the chemical industry and cited that the uncontrolled use of
pesticides was detrimental to the environment. Interestingly, the
publication sparked a political movement in many countries.
The contents and title were inspired by the assumption that
birds would all vanish due to toxicant exposures, leaving none
to sing bird songs anymore. Later, this viewpoint became even
stronger with the discovery of traces of synthetic chemicals in
remote places on Earth, including the Arctic (Butt et al., 2010;
Hung et al., 2010). However, what was not emphasized was that
the release of pollutants into the aquatic ecosystems was just as
threatening for benthic organisms as they were for birds. This is
important because of the tendency of water bodies to transport
waterborne pollutants over long distances but also the greater
sensitivity of smaller benthic invertebrate species that are lower
in the food chain. Notably, sediments act as a repository for
numerous environmental pollutants that threaten the overall
survival of benthic communities. Given the recent advancement
in science, new ways of looking at the global threats from
potentially harmful pollutants have continued to evolve. There
is increasing interest in assessing the adverse effects under
realistic environmental scenarios.
22
1.2
CHEMICAL ACCUMULATION INDEX
Chemical pollutants are distributed between multiple phases via
processes that depend on the property of surrounding matrices
and the prevailing environmental conditions (Mackay & Fraser,
2000; Calamari, 2002). A property index is used to express the
accumulation of pollutants in fresh weight biological tissue
relative to the concentration in exposure medium (Equation 1).
BCF = Corganism /Cmatrix or water
………….. (1)
It is a ratio that can vary between different species and
expressed
as
bioconcentration
factors
(BCFs),
or
biota
accumulation factors (BAFs) to depict uptake from water or
multiple exposure routes respectively. Typically, these measures
are expressed as laboratory (BCF) or field (BAF) based
determinations. The ratio can be determined under a steadystate condition after test organisms have been sufficiently
exposed to a pollutant such that the value of the ratio does not
change significantly. The uptake of pollutants in organisms can
occur via several different routes including passive absorption;
transport across respiratory surfaces; dietary uptake; and
predation (Gobas & Morrison, 2000). The partitioning of
pollutants to DOC affects their distribution kinetics in the
aqueous phase, and ultimately the fraction that is available for
uptake in organisms. It is one of many reasons why BCF values
are routinely used to express the accumulation of pollutants in
organisms without any consideration for freely dissolved
concentrations in the exposure medium. Apart from several
extrinsic factors, the capacity of an organism to transform
organic chemicals into other metabolic intermediates can vary
across species. Therefore, variability in species’ metabolisms
may affect the amount of chemical residues bound to their
tissues, and in the end the values for BCF determinations. It
23
highlights the need to account for species’ intrinsic variability in
biotransformation during BCF estimations.
1.3
CHEMICAL DISTRIBUTION AND EFFECT ASSESSMENT
Chemicals have a tendency to contaminate the aquatic
environment because of their numerous sources and diverse
physicochemical properties (Webster et al., 1998; Fent, 2003;
2004). The global biogeochemical cycle, including the surface
weathering of rocks, contributes to factors that govern inorganic
substance distribution and essentially the recalcitrance of metals.
These factors promote the widespread occurrence of inorganic
pollutants in the aquatic ecosystems (Witters, 1998; De
Schamphelaere & Janssen, 2004). In addition, mining and
smelting activities have considerable influence on the overall
contamination of water resources. They help to explain the
distribution of inorganic chemicals that mostly occur in trace
quantities, even though a few are present in higher amounts
(Marín-Guirao et al., 2005).
The combustion of fuels and organic matter under limited
oxygen can generate several different chemicals including
polycyclic aromatic hydrocarbons (PAHs). An increasing variety
of these substances is continuously released into the aquatic
environment as a result of human activity and hence
complicates the assessment of threats from industrial chemicals.
Agrochemicals constitute a large part of anthropogenic
chemicals whose original purpose was to manage insect pests in
order to improve crop and livestock yields. However, these
chemicals can be detected in the aquatic environment where
contamination may result from industrial waste discharges or
surface runoff from agricultural sites (Neff, 1985; Wan, 1994;
Arfsten, 1996). Many of the chemicals have been found to cause
24
deleterious effects on non-target organisms, thereby increasing
the complexity of ecological assessment of exposure to multiple
pollutants.
Despite
the
current
restrictions
in
biocide
applications, water bodies still receive considerable amounts,
including the newly emerging contaminants and nanoparticles.
While some of the chemicals undergo transformation when
present
in
a
multimedia
environment,
others
exhibit
considerable persistence to chemical, but also microbial
degradation (Webster et al., 1998). In fact, it is the property of
the contaminants that influence the overall fate and distribution
in the ambient environment, enabling micro-layers of sediments
and clay particles to provide binding sites for metal complexes.
By contrast, lipophilic contaminants have a tendency to bind
strongly to organic matter, which as a major component of the
sediment matrix, may affect the availability of chemicals to
benthic organisms. Although increased contact time enables
contaminants to sequester into the micro-layer of sediments,
considerable physical disturbances or bioturbation can mobilize
and redistribute them in the overlying water. Considering that
sediment bound fractions are routinely disturbed, it can be
argued that sorption processes and the release into the overlying
(pore) water constitute the key factors that determine the
bioavailability of contaminants (Atkinson et al., 2007).
Current knowledge has shown that a toxicity assessment that
is based on total rather than the dissolved concentrations is not a
reliable prediction of bioavailability. For ionizable chemicals
that may exert toxicity based on valence state and chemical
species, there is evidence that neither total nor dissolved
concentrations would reliably predict bioavailability (Witters,
1998; De Schamphelaere & Janssen, 2004). Additionally,
prevailing environmental factors influence the partitioning and
release of contaminants, which in certain cases may lead to
increased bioavailability and toxicity. Despite the existing
25
conventional methods, passive sampling devices are good
alternatives for the prediction of bioavailability of waterborne
contaminants. The solid phase micro extraction (SPME) utilizes
thin
polymer-coated
fiber
to
predict
freely
dissolved
concentrations, which is the relevant fraction that is taken up in
organisms. The properties of polymer coatings ensure that a
particular class of chemicals is selected. It has been reported that
sorption between a sampler and surrounding matrices correlate
with contaminant accumulation in organisms (Leslie et al., 2002;
Van der Wal et al., 2004). This notwithstanding, direct
measurement of contaminant residues in exposed organisms has
proven to be a realistic estimation since it accounts for multiple
uptake routes, as well as species differential biotransformation
and elimination capabilities.
Commonly used end points such as survival, fecundity and
biomarkers (e.g. vitellogenin, DNA adducts), each have their
own drawbacks. These indicators are suitable for general
screening in order to ascertain the severity of various
contaminants. However, there are credible arguments that these
indicators may not serve as suitable early warning system in
contaminated aquatic ecosystems. This is because contaminants
may have induced a considerable level of toxicity in exposed
organisms. There are concerns that biomarkers are not adequate
for assessing a wide range of contaminants due to apparent
limitations in distinguishing between chemical specific effects.
For this reason, the OMICS technology provides a robust and
complementary platform for a genome-wide assessment of
contaminant mediated toxicity. Improved sensitivity, selectivity
and the capacity to monitor thousands of genes and gene
products simultaneously, inevitably favors the modern OMICS
technology for a wider application in aquatic ecotoxicology
(Steinberg et al., 2008; Garcia-Reyero & Perkins, 2010).
26
In this thesis, the various processes that may influence
bioavailability, stability or toxicity of model chemicals in aquatic
systems were examined with particular emphasis on three
widely studied invertebrate organisms (Table 1, Figure 1). The
main objectives were to investigate:
(i)
Bioconcentration in organisms with different capacities
for biotransformation (Paper I);
(ii) The stability of PCP and phenanthrene under simulated
solar radiation;
(iii) The possibility for toxicity monitoring using D. magna
(Paper II);
(iv) Transcriptional and metabolic changes in L. variegatus
exposed to chemicals of different modes of toxic
action (Paper III);
(v) The influence of hypoxia on metabolic composition of L.
variegatus exposed to B(a)P, chlorpyrifos, or PCP
(Paper IV).
27
PAPER I
Animal culture and exposure
WORM,
MIDGE
Sampling at 2, 8, 24, 48, 72, 96 h
0 mg/L DOC
Exposure water
20 mg/L DOC
Animal tissue
40 mg/L DOC
PDMS fiber (SPME)
Analyses
Data processing
Uptake and
elimination kinetics
Liquid
Scintillation
Counting
Whole tissue
concentrations
Tissue extracts
(aqueous, solvent
phase, and pellet)
fractions
+
[14C] BaP
PAPER II
PCP or Phenanthrene
No UV filter
Daphnia magna
Immobilization
Assay
Photodegradation
PCP or Phenanthrene
UV filter
DATA
HPLC, GC-MS
Analysis
2, 6, 24, 48 h
PAPER III
■ Microarrays
Cd, BaP or PCP
■ DNA adducts (BaP)
Control
■ GC-MS analyses of
worm tissue extract
metabolites
WORM
DATA
8, 48, 96 h
BaP, PCP or CPF
PAPER IV
HYPOXIA
■ Uptake kinetics
Control
WORM
BaP, PCP or CPF
NORMOXIA
■ GC-MS analyses of
worm tissue extract
metabolites
Control
Figure 1. Schematic illustration of the experimental designs and the processing of samples.
28
DATA
Materials and Methods
2.1
TEST ORGANISMS
The selected test organisms are widely used in ecotoxicology
with each of them occupying a different microhabitat in the
aquatic ecosystems, but they also vary in their capacity for
biotransformation. Differences in species’ feeding habits and
fecundity affect their interactions with the freely dissolved and
bound fractions in exposure systems. Even so, each species
plays a crucial role in the aquatic food chain and serves as food
sources for several higher organisms including fish.
2.1.1 Daphnia magna
D. magna is a small planktonic crustacean of the order
Cladocera, which is commonly known as a water flea. It lives in
the water column and forages on smaller invertebrates
including green algae, plankton and other organic detritus.
Daphnia reproduces by parthenogenesis with young neonates
having to undergo moulting in order to grow, increasing in size
up to 3 mm. The species has a relatively short life span of two
months when maintained at 25˚C, but can live even longer at
lower temperatures in the surrounding water (Venkataraman &
Job, 1980). For this reason, laboratory cultures are usually
maintained within a narrow range of water chemistry
characteristics either as individual Daphnia in a small volume of
growth medium, or as a population in a large aquarium. Young
neonates mature early to reproduce within the first week of life,
whereas their subsequent offspring can occur every two to four
days. Over the years, several different synthetic media have
29
been formulated and used to culture the organism. However,
various national and international guidelines have been
recommended together with some suitable media, including the
Aachener Daphnien Medium (ADaM) (Kluttgen et al., 1994), A
Elendt M7, and M4 medium. As a standard organism, Daphnia is
included in several guidelines (e.g. OECD 202 for acute
immobilization and reproduction tests; OECD 211 for 21 day
chronic toxicity testing) particularly because of the species’
sensitivity and suitability for toxicity testing.
2.1.2 Lumbriculus variegatus
The oligochaete, L. variegatus, is a suitable benthic worm that is
widely used in ecotoxicology. Its overall convenience and ease
of handling; the nature of its microhabitat and niche; the good
yields of nucleic acids; and general availability, among other
reasons make them the organism chosen in this study. Although
it is recommended to maintain multiple cultures, these
organisms can be easily kept in a large laboratory aquarium
either in sediment, or as a water-only culture. Usually, black
worms burrow through sediments and this makes them suitable
test species since sediments act as a repository for numerous
contaminants (Kukkonen & Landrum, 1994; Jantunen et al.,
2008). Worms live at the interface between sediment and water
where a considerable part of the tail reaches out of the sediment,
extending into the water column to obtain respiratory oxygen.
Therefore, it is important to account for not only exposure to
waterborne contaminants, but also the fraction that is bound to
the sediment. Even though they are routinely maintained in
laboratory conditions, these worms can tolerate a wide range of
different water and sediment characteristics. Some of the typical
sub-lethal toxicity end points are feeding rate and faecal pellet
production, growth, fecundity, and heat dissipation (Penttinen
& Kukkonen, 1998). As for many oligochaetes, L. variegatus plays
30
an important role in the actual turnover and bioturbation of
organic matter. Taking into account the wide distribution in
freshwater ecosystems, these worms constitute a significant food
source for higher organisms. Standard guidelines for testing the
accumulation and toxicity of chemicals in L. variegatus have been
published by several agencies including the OECD, ISO,
American Society for Testing and Materials (ASTM), and the
United States Environmental Protection Agency (U.S. EPA).
2.1.3 Chironomus riparius
C. riparius is a standard organism that is commonly used in
testing the toxicity of environmentally relevant chemicals. An
adult midge is terrestrial while the other life stages of the insect
live in a sediment-water system. The lifecycle and maturation
time makes it suitable for acute as well as for chronic laboratory
tests (Ristola et al., 1999). Cultivation of the organism can be
easily performed in laboratory conditions (Penttinen et al., 1996),
which makes it possible to generate age-synchronized larvae for
sediment or water-only testing. Ideally, its lifecycle comprises
the egg, larval, pupal and the adult stage. Considering that C.
riparius spends the longest developmental phase of the first until
the fourth instar stages in the sediment, the larvae are routinely
used in experiments, especially for chronic testing. Despite
burrowing through the sediment during feeding, the larvae are
usually found swimming in the overlying water. Laboratory
grown adults are usually housed in netted cages where they
deposit egg sacs in the surrounding water. Brood sacs with
several eggs attach themselves to the walls of aquarium vessels
and hang down to be immersed in water to prevent desiccation.
Emerging larvae can be maintained in an aqueous medium, or
more suitably in a sediment-water system. The adult males can
be distinguished from female insects by the length of antennae
and abdominal terminalia. Commonly used end points for
31
toxicity testing include the fertility of the egg sacs; emergence
time of the larvae; sex ratio of the emerged adults; and mouthpart deformation (Dias et al., 2008). Although there are existing
guidelines (e.g. OECD 218, 219) for ecotoxicity testing using C.
riparius, the designs can easily be modified to suit the purpose of
study and exposure scenarios.
Table 1. Summary of the study designs depicting the stressors and environmental variables
tested in various species.
Paper
I
II
III
Toxicant
B(a)P
PCP, Phenanthrene
B(a)P, Cadmium, PCP
Assessment
Stressor/
endpoint
variable
Species
Biotransformation rate,
DOC,
C. riparius,
Tissue residue
Toxicant
L. variegatus
Photodegradation rate,
UV light,
D. magna
Toxicity (EC50)
Toxicants
Gene expression,
Toxicants
L. variegatus
Tissue residue,
Hypoxia,
L. variegatus
Metabolic changes
Toxicants
DNA adducts,
Metabolic changes
IV
32
B(a)P, Chlorpyrifos, PCP
2.2
CHEMICALS
2.2.1 Phenanthrene and B(a)P
Phenanthrene and B(a)P belong to a class of chemicals known as
PAHs that are composed of fused benzene rings. The chemicals
are produced with difficulty in commercial quantities; however,
they are ubiquitous in the environment because of their
continuous release from the incomplete combustion of organic
materials
including
fossil
fuels
(Wild
&
Jones,
1995).
Phenanthrene, a three-ringed lower molecular weight PAH is
relatively less lipophilic than five-ringed B(a)P. Phenanthrene is
a major component of the overall PAHs in the environment
(Wild & Jones, 1995), where it receives a considerable amount of
solar radiation annually. On the other hand, B(a)P binds more
strongly to organic matter and the mineral contents of soil and
sediment to sequestrate into remote sites within organic
matrices. It is the characteristic binding properties that
contribute to the overall bioavailability, persistence and
accumulation of PAHs in organisms. Uptake of potentially
harmful chemicals in organisms can induce various homeostatic
changes, which include enzymatic transformations that are
required
in
detoxification
processes.
Metabolic
enzymes
facilitate the excretion of contaminants by converting a part of
their internalized residues into water-soluble forms (Livingstone,
1998). However, transformation products resulting from the
metabolism of B(a)P may become even more potent than the
parent chemical and can interfere with the integrity of the DNA
(Gao et al., 2005). Therefore, it is classified by the International
Agency for Research on Cancer (IARC) as a probable carcinogen
(IARC, 1983). It implies that metabolic activation can occur
when animals are exposed to B(a)P since it can lead to
genotoxicity and carcinogenicity (Buhler & Williams, 1988).
Although
previous studies documented the
partitioning
33
behavior and bioavailability of B(a)P and phenanthrene in an
aquatic environment (White, 2005; Jensen et al., 2012), not many
of them considered the fact that substance accumulation can
vary in lower invertebrates due to species’ differential
biotransformation capabilities. This is interesting considering
that the uptake of chemicals can vary widely across different
species and in relation to DOC contents among other sorbent
materials in nature. Despite the fact that enhanced toxicity
resulting from the metabolism of B(a)P is relatively well known,
acute toxicity in organisms exposed to phenanthrene is seldom
reported. It is important in future studies to address the
combined impact of environmental conditions, particularly solar
radiation and animal exposure to contaminants.
2.2.2 Chlorpyrifos and PCP
Chlorpyrifos and PCP were once widely used as broadspectrum biocides in the agriculture and wood industry. As an
organophosphorus insecticide, chlorpyrifos has public health
applications in the control of mosquitoes, but is also used in
agriculture for managing crop pests. Chlorpyrifos does not
persist in soil or water, and to some extent is known to have a
relatively low aqueous solubility. A moderate tendency to bind
to soil and sediment organic matter appears to increase the risk
of groundwater contamination that is seldom reported.
However, chlorpyrifos is so highly potent that as low as 0.08
μg/L and 1.3 μg/L have been reported to cause adverse effects in
daphnids and fish, respectively (Barron & Woodburn, 1995). The
most acute effects due to chlorpyrifos exposure in natural
waters result from spray drift or surface runoffs from
agricultural sites. A portion of the chemical partitions to
sediments,
and
thereby
organisms
chronically
to
exposing
fish
and
desorbing fractions.
invertebrate
Terrestrial
organisms are less affected in the ambient environment.
34
PCP is relatively lipophilic, and hence can induce polar
narcosis
or
interfere
with
intermediary
metabolism
by
uncoupling oxidative phosphorylation in aerobic organisms
(Terada, 1990; Kukkonen, 2002). In addition to the pH of
aqueous preparations, the extent of chlorination and the various
isomeric positions of chlorine atoms are crucial in determining
the toxicity of PCP (Kukkonen, 2002). Its persistence coupled
with a widespread application in the past, has caused
considerable contamination of the aquatic ecosystems. PCP
residues that are bound to soil and sediment matters can desorb
over time to contaminate surface water and ground water. For
example, as much as 25.7 mg/L was reported in ground water at
a former pesticide manufacturing plant in Taiwan and up to 27
mg/L near a rubbish-dump site in Brazil (Tse et al., 2001; do
Nascimento et al., 2004). Stackelberg et al. (2004) have reported
that several contaminants including PCP were present in
streams and raw-water supplies in a drinking water treatment
plant in the United States. Equally, low levels (0.1 – 1 μg/L) of
waterborne PCP have been reported to cause adverse effects on
exposed organisms (Kim et al., 2007). Despite the restricted use
in many countries, PCP remains an important pesticide from the
viewpoint of ecotoxicology.
2.2.3 Cadmium
Cadmium is a nonessential toxic metal that can be found in the
Earth’s crust, where it forms compounds with zinc, copper and
lead (Fassett, 1975). The natural weathering of rock surfaces,
coupled with industrial applications, contributes to the release
of cadmium into the environment. Cadmium is produced
mainly as a byproduct of mining and smelting of zinc ore but
has several industrial uses in cadmium-nickel batteries; plastic
stabilizers in PVC products; and coating and plating of metal
surfaces (Valko et al., 2005). Due to its broad applications,
35
cadmium occurs in natural water but can also bind strongly to
soil and sediment organic matter (Järup, 2003). As a highly
persistent contaminant, cadmium is relatively immune to
chemical and microbial degradation. This may explain its
overall accumulations in the ambient environment where
exposure to humans has been linked to adverse effects (Fassett,
1975). Cadmium uptake in organisms varies with environmental
pH and for this reason contributes to the toxicity challenges
associated with waste water and industrial effluent discharges.
Exposure to low cadmium concentrations has been linked to
immune system disorders and DNA damage, which can lead to
cancer in higher organisms (Järup, 2003; Jin et al., 2003).
2.3
BIOCONCENTRATION AND UPTAKE OF B(a)P IN L.
VARIEGATUS
In paper I, the uptake kinetics of waterborne [14C] B(a)P was
studied in L. variegatus and C. riparius at different levels of
dissolved organic matter. B(a)P was introduced into artificial
fresh water (AFW) in drops and with gentle stirring. The
organic matter that was recovered from a dystrophic lake by
reverse osmosis was used to prepare humic water. Quantities of
organic matter were weighed into the test water and stirred
sufficiently to yield humic waters of 20 mg/L and 40 mg/L DOC.
The temperature and pH of the test water were 20±1°C and
7.2±0.2 respectively. All exposures were conducted in 0.5 mM
AFW with constitutive inorganic salts (MgSO4 + 7H2O, CaCl2 +
2H2O, KCl, NaHCO3), aerated until oxygen saturation was
achieved.
Two species with differing capabilities for biotransformation
and excretion were tested, since these factors strongly influence
the accumulation of contaminants. Oligochaete worms were
taken from the main culture aquaria and acclimated overnight
36
in AFW to purge their guts of ingested sediment food particles.
On the other hand, age-synchronized C. riparius larvae were
generated from egg sacs that were obtained from adults reared
in main laboratory culture aquaria at 20±1°C and 16/8 h
light/dark photoperiod. The larvae were reared in Petri dishes
having a thin layer of acid-washed quartz sand as a substrate for
attachment, with constant aeration until the fourth instar. Also,
larvae were acclimated in AFW to purge their guts prior to
exposure, for the controls as well as the B(a)P exposed
individuals. Organisms were co-exposed in the same vial with a
pair of 4 cm length fiber having a coating thickness of 30 µm
PDMS for up to 96 h. Each vial contained 10 oligochaetes; four
individuals were sampled for total body residue analyses, while
three were analyzed for tissue contents of the parent B(a)P; total
metabolites that partitioned into the water phase as well as in
the extraction solvent phase. Animals were sampled at each
point in time, and were blotted dry on paper towels prior to
weighing. Samples were stored at minus 20°C until use. Similar
experimental designs and extraction protocols were used for L.
variegatus exposure as for C. riparius.
The partitioning coefficient of 4.75 (log KPDMS-WATER) was
determined from the short lengths of PDMS-coated fiber, where
k was estimated as a function of measured B(a)P concentrations
in PDMS and in corresponding test water. DOC-water
partitioning coefficients (KDOC) were calculated according to
Jonker and Koelmans (2001). For determining the radioactivity,
exposure water was sampled into scintillation vials in three
replicates per point in time. An equal volume of LSC InstaGel
cocktail was added, gently shaken, and stored at room
temperature until the analyses. Worm samples were solubilized
overnight at room temperature in 0.5 mL of Soluene tissue
dissolvent and mixed with 5 mL of UltimaGold cocktail prior to
analyses. To analyze the tissue extract distribution of B(a)P in
37
extraction liquids, samples were extracted according to Kane
Driscoll and McElroy (1997), but with slight modifications.
Tissue extractions were conducted based on the perception that
the solvent phase fractions would constitute mainly the parent
B(a)P, while the water-phase fraction represents the polar
metabolites (Leversee et al., 1982; Borchert et al., 1997; Schuler et
al., 2003). The distribution of parents and corresponding
metabolic
derivatives
was
determined
by
thin
layer
chromatography (TLC). A detailed description of the procedure
has been given elsewhere (Paper I). The activity of the tissue
pellet represents the total radioactivity at the final centrifugation
of homogenates and hence consists of a non-extractable pellet,
metabolites and parent chemical. The levels of B(a)P in exposure
water, tissue extract fractions, and in non-extractable pellet were
determined by LSC.
Toxicokinetic parameters were determined by fitting a first
order kinetic model to the bioaccumulation data according to
equation 2:
Ca (t) = ku/ke•Cw•(1-e-ke•t)
……….. (2)
Where Ca(t) refers to the solvent phase tissue extract
concentration of B(a)P in animal mass at time t; Cw is PDMS
fiber-predicted concentration in test water; ku is the uptake
clearance coefficient from the test medium; ke is the elimination
rate constant. Bioconcentration values were expressed on the
basis of the parent B(a)P (BCFPAR); total [14C] activity (BCFTOT);
and radioactivity counts in tissue and exposure water
(BCFCOUNT). A univariate analysis of variance (ANOVA) was
performed to test whether DOC played any role in the uptake of
[14C]. The data were subjected to factorial ANOVA to determine
whether DOC has any influence on species biotransformation
capacity for B(a)P.
38
2.4
PHOTODEGRADATION AND TOXICITY OF PCP AND
PHENANTHRENE
Paper II focused on the application of simulated solar radiation
to remediate waterborne organic chemicals under laboratory
conditions. The survival of D. magna neonates was used to
determine the degradation efficacy of the parent chemicals.
Stock solutions of PCP and phenanthrene were prepared in
doubly distilled water, and afterwards the exposure water was
constituted by being diluted with the appropriate volume of
ADaM. Aqueous concentrations ranged from 0.1 to 2.1 mg/L for
PCP, and 0.11 to 0.22 mg/L for phenanthrene. A 30 mL round
bottomed glass vial having the test solution allowed light
penetration from the top via a thin UV-transparent cap, with a
surface diameter of 1.1 cm. As a result, only about 65% of the
overall light intensity (27.3 Wm-2) reached the test water under
continuous mixing (400 rpm). Vials having test water were
maintained at 21±1°C in a thermostat-controlled recirculating
water bath throughout each set of tests. Irradiation lamps were
mounted at a distance of 1.36 m from the water bath, while three
replicate vials were taken at intervals from where analytical
samples were drawn. The remaining water samples were used
for the D. magna immobilization test.
The absolute intensity of UV light was measured with a
quantum photometer linked to an immersion probe, whereas
the spectral composition of light sources was determined in the
presence and absence of UV filter using a TRISTAN
spectroradiometer.
The aqueous concentrations of PCP and phenanthrene were
analyzed by high performance liquid chromatography (HPLC),
and a detailed description of the analysis is provided elsewhere
(Paper II). Rate constants (k) at different concentrations were
determined by fitting the photodegradation data to the firstorder equation 3 (Sabaté et al., 2001):
39
ln C0/Ct = kt
……………
(3)
where C0 and Ct are concentrations of the test substances at
times zero and t respectively. I calculated the half-lives from
equation 4 by substituting Ct with C0/2:
t1/2 = ln 2/k = 0.6931/k
……..
(4)
Daphnia neonates used for bioassays were exposed in pyrex
glass containers. The animals were exposed in test water preirradiated under direct illumination with and without a UV
filter. Control animals were exposed in ADaM without the test
chemicals and monitored for immobilization for up to 72 h.
Solutions of each chemical were sampled at intervals and used
for immobilization tests. Five neonates were tested in each glass
tube in four replicates of 10 mL at 21±1°C and pH of 7.5±0.2 for ≤
48 h. The animals were observed for movement at intervals as
well as monitoring the endpoint to test the efficacy of
photodegradation.
Fewer than 10% of neonates were expected to become
immobilized or show other signs of stress and abnormal
behavior in the control exposures. Animals were considered to
be immobilized when they did not to swim within 15 s after a
gentle agitation of the test vessel, even if they could still move
their antennae (OECD 2004; Persoone et al., 2009). Median
effective concentration (EC50) values were calculated for 24 h
and 48 h exposure time points using REGTOX-EV7.0.3.
2.5
MOLECULAR CHANGES IN WORMS EXPOSED TO
CHEMICALS
In paper III, the response of L. variegatus exposed to toxic
chemicals were examined by measuring changes in gene
expression, endogenous metabolites and DNA integrity. Tissue
samples processed for microarrays and metabolites analysis
40
were obtained from L. variegatus exposed to PCP (32.8 µg/L),
B(a)P (0.3 µg/L), or cadmium chloride (54 µg/L) and their
respective controls. BPDE-DNA measurements were obtained
from worms treated with B(a)P. The samples for metabolites
and DNA adducts analyses were from the same exposure
vessels as the source template RNA for microarrays.
2.5.1 Tissue Sample Preparation for GC-MS Analysis
Frozen worms were homogenized in 2 mL round-bottomed
Safe-Lock Eppendorf tubes in TissueLyser (Qiagen, Helsinki,
Finland) as described elsewhere (Paper III). The internal
standard comprised a mixture of benzoic-D5 acid (0.33 mg/mL),
salicylic acid-carboxyl-13C (0.33 mg/mL), D-glucose
C (0.167
13
mg/mL), and glycerol-D8 (0.167 mg/mL) in methanol/water (1/1,
v/v). The resulting tissue homogenates were extracted for 15
min at 4°C and at a mixing speed of 1000 rpm using a
thermomixer (Eppendorf, Hamburg-Eppendorf, Germany).
Supernatants from repeated extractions were combined and two
aliquots of 150 µL were evaporated to a dry state in a vacuum
evaporator (Thermo Fisher Scientific Inc, Vantaa, Finland). The
dry extracts were degassed under a gentle stream of nitrogen
gas before the vials were capped and stored at –70°C until
analyses. The mass spectrometry analysis was done according to
Roessner et al. (2000) with the following modifications. The
details of the methoxymation and silylation steps, including the
analysis of tissue extract fractions, were given in Paper III.
Analytes were tentatively identified by comparing their spectra
and retention indices to standard compounds; or entries from
the NIST Mass Spectral Database (Mass Spectral Search Program,
version 2.0f, Agilent technologies); and the GOLM Metabolome
Database.
The
compounds
were
quantified
using
a
MetaboliteDetector (Hiller et al., 2009). All of the peak areas
were normalized to glycerol-D8 and the fresh weight of samples.
41
2.5.2 CE-LIF Immunoassay for DNA Adducts Analysis
Total DNA was extracted from the L. variegatus exposed to B(a)P,
and from the unexposed chemical controls. The samples were
purified with wizard genomic DNA purification kits (Promega,
Madison, WI), while the analyses were performed on a lab-built
CE-LIF system. A detailed description of the DNA sample
purification steps and CE separations were given in Paper III
and according to Wang et al. (2009). The adduct frequency for
each sample was calculated by standard calibration by using
well characterized BPDE-DNA adducts standards.
2.5.3 Microarrays
Microarrays were conducted using a custom made 60-mer
oligonucleotide 8 x15K microarray manufactured with Agilent’s
SurePrint Inkjet technology. The sequence information (5710
EST clones) was based on the L. variegatus EST-library obtained
as a side product from the frog (Xenopus tropicalis) EST project
(Price et al., 2005). In the frog EST project the whole body of a
frog, which had been fed live oligochaete worms, was used as
the source of mRNA for the construction of cDNA libraries for
EST sequencing. The worm ESTs were identified by comparing
them to the sequenced frog genome. In addition, probes for 1712
Lumbricus rubellus genes that were published in Genbank were
added in order to fill the remaining slots on the arrays (Paper
III). The annotation of genes used on the array was based on the
analysis
by
Blast2GO
software
(Götz
et
al.,
2008;
http://www.blast2go.de/). Two different oligo probes were
designed for each gene using eArray (Agilent Technologies),
which gave rise to two technical replicates per array. The Array
design and all of the microarray data have been submitted to the
ArrayExpress
accession
respectively.
42
(http://www.ebi.ac.uk/arrayexpress/)
numbers
A-MEXP-2200
and
with
E-MEXP-3588,
Total RNA of the Lumbriculus variegatus was extracted from
14 individuals in Trizol (Invitrogen, San Diego) for each sample
of the four chemical treatments and their corresponding controls.
All the RNA samples from each treatment were hybridized
against the corresponding control samples having the same
treatment and timeframe. All hybridizations were conducted
with pooled samples and in technical duplicates (dye swap).
The integrity of the RNA was verified by gel electrophoresis,
while the quantity and purity were analyzed using a NanoDrop
ND-1000 UV/Vis spectrophotometer. The RNA samples (25 µg)
were labeled according to Brosché et al. (2005) by the coupling
of Cy® dyes to aminoallyl-dUTP labeled cDNA.
Other post labeling steps including purification of cDNA;
hybridization; and the washing and scanning of slides have
been explained in detail in Paper III. Spot intensities were
quantified
with
ScanAlyze
software
(Michael
Eisen,
http://rana.lbl.gov/EisenSoftware.htm), while the mean signal
and median background values were used subsequently in
calculations. Due to the high density layout of the arrays,
quantitations were performed in two grid position steps: first to
the spots in the rows of odd numbers and then to the spots in
the rows of even numbers. Later, the right positioned data from
two ScanAlyze result files were combined to a new file. Data
normalization and statistical analysis of the microarray
experiment were carried out with R (2.7.0)/Bioconductor
package LIMMA (2.12.0) (Linear Model for Microarray Data;
Smyth, 2005). GO-enrichment was calculated by the Gene
Ontology Enrichment Analysis Software Toolkit and GOEAST.
(http://omicslab.genetics.ac.cn/GOEAST/index.php,
Zheng
&
Wang, 2008).
43
2.5.4 PCR Validation of Array-based Expression Profiles
Real time qPCR was performed on selected genes in relation to
18S ribosomal RNA in order to verify the microarrays data as
explained in paper III. Template cDNA for qPCR was from the
same primary sample extract of total RNA that was used for the
microarrays. Real time qPCR were run in three technical
replicates and performed using DyNAmoTM HS SYBR® Green
qPCR kit (Finnzymes, Vantaa, Finland) on a thermal cycler
supplied with Chromo4 Continuous Fluorescence Detector (MJ
Research, Waltham, MA), in 40 cycles of 15 sec at 95°C; 10 sec at
94°C; and a further 20 sec at the primer annealing temperature
of 57°C using customized primers (Table 2 Paper III). Melting
curve analysis within a range of 72 - 95°C confirmed the
presence of a single major amplification product.
2.6
METABOLIC RESPONSE TO CHEMICAL AND HYPOXIC
STRESS
2.6.1 Test Water and Exposure Conditions
The main goal of paper IV was to investigate the changes in the
endogenous metabolites of L. variegatus exposed to chemicals
under different oxygen conditions. Hypoxic exposure water was
prepared by passing a stream of nitrogen gas to saturate AFW.
However, filtered air was allowed to saturate AFW that was
used to constitute the exposure medium in normal oxygen
(normoxic) conditions. High-capacity passive dosing jars were
loaded with B(a)P according to Mayer and Holmstrup (2008),
but with some modifications. The silicone-coated amber jars
were loaded as a methanol suspension using [14C] B(a)P as tracer.
A methanol suspension was introduced to the silicone and
shaken at 100 rpm for 96 h to attain saturation. After the
methanol was removed, exposure jars were left in a fume
cupboard to vaporize the solvent. A small amount of AFW was
44
added to the loaded jars to cleanse any solvent droplets and
remnant crystals of the toxicant. Ten milliliters of AFW was
added and shaken overnight at 100 rpm in order to reconstitute
the stable test water that was within the limit of aqueous
solubility. Meanwhile, the worms were purged overnight in
AFW to empty their gut contents. An additional volume of
oxygen-saturated AFW was added to test jars up to 50 mL and
equilibrated overnight in an aerated exposure chamber (Plas
Labs Inc., Lansing, Michigan) prior to the addition of the
animals. Exposure was similar in both oxygen conditions, except
that the test jars were filled up to 50 mL using nitrogensaturated AFW upon hypoxia, while a gentle stream of nitrogen
was allowed into the exposure chamber. Each chemical
exposure was conducted in three replicates per oxygen
condition; having 10 animals per jar for 8, 48 and 96 h. The
average exposure temperature was maintained at 19.5±2°C. A
tolerable level of 100 μg/L for PCP (Sigma Chemical Co, St Louis,
MO) or chlorpyrifos (Dr Ehrenstorfer GmbH, Augsburg) was
obtained from the preliminary screening of the animals below
acute toxicity. B(a)P was tested within the limit of aqueous
solubility (3 µg/L). The exposure water was spiked with the
[14C]-labeled chemical as a tracer and stirred overnight under
normoxic (≥ 49%) or hypoxic (< 30% O2) conditions in the
exposure chamber. Dissolved oxygen was monitored in test
water using an oxygen meter that was linked to an immersion
probe
(Wissenschaftlich-Technische
Werkstätten
GmbH,
Weilheim, Germany). In order to minimize a considerable
depletion of chlorpyrifos and PCP, the test water was renewed
every 24 h throughout the exposure period. All of the exposures
were conducted in a conditioning chamber to maintain the
required oxygen condition. The presence of ambidextrous
neoprene gloves facilitated the handling of the contents of the
45
exposure chamber, whereas the duplex inlet-outlet channels
allowed the transfer of items through the chamber.
2.6.2 Tissue Residue Analysis of [14C]-labeled Chemicals
Associating toxicity endpoints with quantitative values from
chemical
residue
analysis
alleviates
the
challenges
of
bioavailability predictions. Worms were sampled at each point
in time and blotted dry on paper towels. Samples were weighed
and stored at –70°C until analysis. In order to ensure that
chemicals were actually taken up in worms, a fraction of
sampled tissues was solubilized overnight at room temperature
in 0.5 mL Soluene-350 and mixed with 5 mL of UltimaGold
cocktail. The radioactivity of the tissue homogenate was
measured in a Wallac WinSpectral liquid scintillation counter
(Wallac Finland OY, Turku, Finland). The measurement time
was 10 min per sample, and the radioactivity value was
corrected for quench using the internal overlay technique, an
inbuilt method in Wallac LSC after subtracting the backgrounds.
46
3 Results and Discussion
This
study
utilized
different
methods
to
assess
the
bioavailability, uptake and toxicity of contaminants to aquatic
invertebrates in relation to external environmental conditions. In
addition to a molecular level assessment of transcripts and
metabolites, chemical residues in tissues were related to time
and
inherent
physiological
traits
of
the
organisms. A
combination of assessment methods from different levels of
biological
organization
was
crucial
since
each
one
complemented the other and improved our understanding of
underlying toxicity mechanisms in the studied organisms.
3.1
BIOCONCENTRATION AND KINETICS OF B(a)P IN L.
VARIEGATUS
Synthetic polymer coatings are good alternatives to predict
bioavailable concentrations in the aquatic environment (Vrana
et al., 2005). This is important because surface water contains
both the dissolved (DOC) and particulate carbon that
sequestrate contaminating chemicals and thereby influence the
proportion in the aqueous phase. Usually, the aqueous phase is
considered to be the relevant portion that is available for uptake
in organisms (Paper I). An aspect of this study utilized short
lengths
of
PDMS-coated
fiber
to
monitor
chemical
concentrations in the same exposure vial as the test organisms.
The results from this study revealed that B(a)P were bound to
the DOC in a partitioning process that reduced the freelydissolved concentrations in exposure water. The B(a)P attained
sorption equilibrium between the fiber and the surrounding
47
water in a shorter time with gentle shaking than under static
biota/fiber exposure. The sorption of B(a)P to PDMS increased
linearly in the setup with co-exposed organisms, but it does not
suggest that equilibrium was imminent. The results were in
agreement with the increased [14C] levels in the tissues of the
oligochaetes. However, these findings differed from the
corresponding accumulation pattern in worms in the presence
of DOC.
Concentrations were non-linear in a comparable exposure of
C. riparius to the test chemical. Despite the complexity of the test
systems, the capacity to detect variability in exposure
concentrations highlights the efficacy of fiber in predicting
bioavailability and accumulation of chemicals in biota. It is
noteworthy that the presence of animals in the same exposure
vial as the fiber appeared to have prolonged the sorption
equilibrium time. On the other hand, a gentle agitation in the
absence of test animals enhanced the chemical equilibrium
between the fiber and the surrounding exposure water. It is
possible that a small amount of B(a)P may have been removed
from the system by the animals or bound to organic matter,
which might contribute to the difference in KDOC values
compared to the literature (Mott, 2002).
Total [14C] in the tissue extracts of the L. variegatus exposed
together with the fiber was nonlinear. This is contrary to the
linearity of B(a)P uptake in the oligochaetes exposed to the test
chemical in the absence of DOC. The variability in patterns of
chemical uptake highlights the significance of DOC interactions
with the soluble fraction, which became even more evident
under a prolonged exposure period. Nonlinear kinetic model
parameters ku and ke were used to describe the uptake and
elimination of B(a)P from the exposure water and L. variegatus
respectively (Table 1 Paper I). It was evident from the TLC
measurements that the tissue extracts consisted of mainly the
48
parent B(a)P with less than 10% of the total extracts attributed to
the aqueous phase fraction of the tissue homogenate in both
DOC treatments. A non-extractable portion did not exceed 20%
of the total radioactivity of the tissue extracts. By contrast,
radioactivity that corresponded to the tissue extracts of the
parent were considerably low and underlines the greater
metabolic and elimination efficiency of C. riparius. The nonextractable portion ranged between 14 – 37% of total
radioactivity. The C. riparius tissue extract activity of the parent
B(a)P from the reference test water without DOC differed
significantly (p < 0.05) from each of the DOC treatments. The
activity of the parent tissue extracts varied markedly between
both DOC treatments in the C. riparius exposure. On the other
hand, the extractable parent fraction did not reveal any
significant difference in oligochaetes despite a two-fold
difference in DOC levels.
Dissolved concentrations are considered to be the dominant
exposure route, even though estimations of BCFs based on total
concentrations are common. Model estimations of BCFs as a
function of uptake and elimination rate constants in many cases
do not only consider parent chemicals, but also rely on total
concentrations. For this reason, it is essential to account for the
species’ differential capabilities for biotransformation, since
chemical fractions either as the parent or the metabolites
influence the overall BCF values. A refinement of BCF
determination based on available concentrations as a function of
parent fraction rather than total tissue residue is necessary and
possibly a realistic estimation. BCF values that are based on total
[14C] rather than tissue activity corresponding to the parent
chemical contribute to the widespread variations in publications.
As
opposed
to
the
parent
chemical
that
accumulated
substantially in L. variegatus, similar studies in C. riparius
showed the accumulation of metabolic products (Leversee et al.,
49
1982; Schuler et al., 2003). These values are in agreement with
the relative distribution of metabolites in the aqueous phase, but
also reflect an apparent difference in the species’ capacity for
biotransformation. These results do not indicate that C. riparius
successfully excreted potentially harmful products from the
body; rather, the overall accumulation was reduced which can
be deduced from the lower BCF (ku/ke) values (Table 1 Paper I).
In fact, the accumulation of metabolic products may predispose
organisms including C. riparius to narcotic effects, and a further
risk to the metabolism enhanced toxicity cannot be ruled out.
The BCF values differed unexpectedly between DOC
treatments, while the values derived from the models decreased
by a factor of nearly two as DOC concentration doubled.
Arguably, if fiber predictions are optimal and biotransformation
slower than B(a)P uptake, any variability in dissolved
concentrations due to DOC treatments would not affect the BCF
values. The differences in test conditions and in lipid contents of
the organisms may contribute to the discrepancy in BCF of 292
(96 h) from this study, compared to the publication value of 200
(8 h) (Leversee et al., 1982). The determination of BCFs based on
substance toxic fraction either as the parent chemical or
metabolite may represent bioconcentration more appropriately.
It is possible to overestimate the partitioning between water and
animals from total [14C] activity, or underestimate them
technically due to sub-optimal tissue extractions for substances
exerting toxicity due to the parent form. In view of the overall
complexity of aquatic ecosystems, a reliable assessment of risks
due to contaminant exposures requires that environmental
factors and species’ inherent traits are taken into account.
Therefore, it is important to consider species’ differential
capabilities for biotransformation in order to reliably determine
the toxicity data (EC50) and accumulation index (BCFs).
50
3.2
PHOTODEGRADATION AND TOXICITY OF PCP AND
PHENANTHRENE
The initial t0 concentrations declined by 5 – 25% after continuous
radiation of the aqueous solution of PCP for 2 h (Figure 2A
Paper II). Only 71 – 77%; 44 – 58%; and 15 – 27% of the test
concentrations were measurable after 6, 12 and 25 h,
respectively. The residual amounts of PCP in the exposure water
having t0 concentrations of 0.41, 0.59, 1.1, and 2.1 mg/L were not
measurable beyond 25, 36, 50 and 70 h of continuous radiation
respectively. The disappearance of PCP from the solution may
suggest that contaminating pentachlorophenol can be reduced
considerably from waste water by prolonged direct radiation.
Calculated values of PCP degradation rate constants (k) at
concentrations of 0.41, 0.59, 1.1, and 2.1 mg/L were 7.1 x 10-2 h-1;
7.2 x 10-2 h-1; 5.5 x 10-2 h-1; and 4.9 x 10-2 h-1 respectively, and
showed a slight decrease with increasing initial concentrations.
Values of k for aqueous solutions of phenanthrene at 0.12 and
0.22 mg/L were only 0.42 x 10-2 h-1 and 0.29 x 10-2 h-1 respectively.
The
rate
constants
varied
slightly
with
the
substance
concentrations, even though the calculated values for PCP were
considerably higher than for phenanthrene (Table 1 Paper II).
On
the
other
hand,
no
measurable
decrease
in
PCP
concentrations was observed following direct radiation in the
presence of the UV filter. This might suggest that seasonal
changes, sunshine duration and vegetation shields may affect
the stability of water borne PCP in the ambient environment.
The resulting irradiated water was tested for photoproduct
mediated toxicity to D. magna neonates. At each of the
concentrations that were tested, the mobility of D. magna
increased correspondingly with radiation time up to 91 h. The
neonate mobility was used to determine the fitness of the
organisms. Even though the radiation of the exposure water
influenced the test concentrations and correlated positively with
51
the neonate mobility, the results varied markedly between tests
in the presence and absence of the UV filter. Less than 16% of
the neonates were mobile after 48 h exposure in pre-irradiated
water at the highest initial PCP concentration. Survival
increased and reached 100% in nearly all of the organisms
exposed for 48 h in the test water, which was pre-irradiated for
up to 91 h. On the contrary, there was hardly any movement of
the neonates in the test water pre-irradiated in the presence of
the UV filter for the two highest PCP concentrations. These
results were consistent with the predicted EC50 values that
varied with the duration of Daphnia exposures in the preirradiated aqueous solutions of PCP (Table 2 Paper II). Also, the
effect concentrations increased with pre-radiation time until
EC50 determinations were no longer possible after 91 h. The
calculated effect concentration values helped to confirm the
disappearance of the parent PCP following extensive radiation,
even though the photoproducts’ mediated toxicity was
indiscernible. The predicted EC50 values after the exposure of
Daphnia for 24 h to the non-irradiated water were in agreement
with the published material (Ikuno et al., 2008). The
corresponding EC50s were lower after 48 h than the 24 h animal
exposures, even though the values increased with the increasing
pre-irradiation time (Table 2 Paper II). The EC50 values were
consistent with measured concentrations under the UV filter
and suggested that UV was responsible for the disappearance of
the parent PCP.
A direct irradiation of aqueous phenanthrene for 93 h caused
a slight reduction in concentration, even though about 69% and
75% of the initial t0 concentrations (0.12 mg/L and 0.22 mg/L)
were still measurable (Figure 5 Paper II). This showed that the
degradation of phenanthrene under extensive UV treatment was
less dramatic compared to the equivalent treatments with PCP.
The identification of the photoproducts was beyond the scope of
52
this
study.
However,
the
maximum
level
of
aqueous
phenanthrene was not sufficient to immobilize D. magna
neonates. These results were different from a previous study
that reported immobilized D. magna neonates at EC50 of 0.48
mg/L. Although the treatment level (maximum aqueous
concentration) in this present study might have been too low to
immobilize the animals, it is possible that their mobility
improved, albeit slightly, with increasing the irradiation time
given the stability of phenanthrene. The significance of these
findings is that geographical location and seasonal changes may
influence contaminant degradation under natural sunlight. The
susceptibility or persistence of chemicals to photodegradation is
a property that could be harnessed in making decisions for
remediating waterborne organic contaminants. The application
of UV irradiation is cost effective and the outcome may be
substantial given the vast amount of annual tropical solar
radiation for local remediation. Nonetheless, filter effects from
soil and sediment organic matter may slow down the
degradation of strongly bound contaminants in natural waters.
53
Table 2. Euclidean distances between data points used as the biological scores to depict the
behavior of genes per sample treatment, particularly during cluster analysis. The smaller the
relative geometric distance between data points, the more the genes are likely to be grouped
together. This implies that samples with similar responses to each other upon chemical
treatments have smaller values, whereas high values depict increased disparity between the
different samples.
BaP2
BaP6
BaP24
BaP48
Cd2
Cd6
Cd24
Cd48
PCP2
PCP6
BaP6
40.0
BaP24
48.9
40.0
BaP48
61.5
59.8
53.0
Cd2
70.0
68.4
71.2
44.4
Cd6
52.8
46.8
43.1
68.9
85.7
Cd24
53.8
50.2
25.2
54.9
76.3
41.4
Cd48
69.9
66.1
69.8
37.4
40.2
89.0
77.3
PCP2
39.4
44.2
43.0
57.7
69.9
39.7
41.7
74.5
PCP6
45.8
42.8
43.3
62.2
76.7
30.7
41.0
80.6
30.5
PCP24
63.0
56.0
55.9
39.7
41.0
77.1
63.4
35.5
65.2
70.7
PCPP48
81.3
74.4
82.9
58.6
41.2
101.0
91.4
37.1
86.4
92.0
54
PCP24
41.6
3.3
MOLECULAR RESPONSE TO TOXICANTS AND HYPOXIA
3.3.1 Transcriptional and Metabolic Response in L. variegatus
In view of the toxic actions for the different chemical treatments,
a set of genes showed a differential expression. They include
genes that are involved in mitochondrial function and oxygen
transport, cytoskeleton function, reactive oxygen species (ROS)
and oxidative stress regulation, as can be visualized by gene
ontology enrichment analysis (Figure 3). As shown here, the
patterns of gene expression changes were different between
chemical treatments, and varied markedly in the kinetics of
response (Figure 2, Figure 2 Paper III). The Euclidean distance
scores depict samples that are more likely to be classified
together, or display a clear distinction from others (Table 2).
Worms exposed to toxicants besides cadmium showed a general
tendency for elevated expression within 6 h, but subsequently
declined at 48 h (Figure 2 Paper III, Table 1 Paper III). Given the
kinetics of response, it is possible that the amounts of cadmium
reaching the target sites in the exposed animals were minimal,
and as a result have contributed to delayed transcriptional
changes at 2 h.
55
Figure 2. Venn diagrams for differentially expressed genes in L. variegatus exposed to
B(a)P, Cd or PCP for the duration of 2, 6, 24 and 48 h. Values indicate the number of
common/different genes with differential expression of more than twofold either up- or
down-regulated as depicted by the direction of the arrow (p-value < 0.01, LIMMA
“decide Tests” function).
56
One of the challenges of combining transcriptomics with
metabolites analysis is finding the link between differentially
expressed
genes;
altered
metabolites;
and
the
general
physiology of species. In this present study however, it is
apparent that genes with roles in muscular contractions and
microfilament structures were differentially expressed in
exposed organisms. Most of the differentially expressed genes
involved in actin processes including troponin; intermediate
filament proteins; thaumatin-like protein; and ferritin were
effectively down-regulated in all chemical treatments (Table 1
Paper III). In particular, the textures of worms treated with PCP
were unusually softer than at pre-exposure conditions. The
decreased expressions were strongest until after 6 h in the
exposed individuals and suggests that PCP may have affected
the outer wall integrity of the worms.
On the other hand, the elevated expression of genes that are
involved in redox processes including glutathione peroxidase;
glutathione S-transferase; and thioredoxin peroxidase is an
indication of an active antioxidant defense system. It is
consistent with the changes in the antioxidant vitamin
precursors including alpha-tocopherol and cholecalciferol. The
increased amount of BPDE-DNA adducts in the B(a)P exposed
worms appeared to correlate with the variability in uridine,
inosine and xanthine, which are key components of a nucleoside
metabolism. It suggests that DNA damage was imminent, or
peaked within 6 h (Table 3). Taking into account the
transcriptional changes in proteases (Table 1 Paper III), the
observed variability in free amino acid levels were not very
surprising. Ideally, proteases activate protein functions via the
cleavage of specific peptides or entirely deactivating a complete
peptide. A decline in free amino acid levels (e.g. glycine,
threonine, tyrosine, and phenylalanine) conformed to the
increased levels of urea, but also for some partially identified
57
amine-like compounds. The reduced expression of proteases
may be linked to changes in amine metabolisms. Looking at the
variability in free amino acid homoeostasis, it can be argued that
the normal energy pathways may have been disrupted since
these compounds are known precursors for energy metabolism
in higher organisms (Chace & Odessey, 1981). Also, phosphates
and some key precursor energy compounds were decreased in
individuals exposed to the test chemicals. Glucose 3-O-methyl
(1MEOX)
(4TMS);
inosine-5'-monophosphate-like;
and
disaccharides-like compounds showed a marked decrease in the
exposed worms. We found a strong cadmium response for the
last two metabolites but also for eicosadienoic acid methyl ester;
octadecanoic acid; octadecenoic acid; hexadecanoic acid; and
their methyl ester derivatives. It is noteworthy that most
transcriptional changes occurred within a 6 h exposure to the
chemicals. Overall, the results from the analysis of the
cadmium-exposed individuals appeared to be the most dramatic
and evident in the characteristic delay in transcriptional
response at 2 h, but also the effective response of several
metabolites.
58
Table 3. Levels of BPDE-DNA adducts in replicates of DNA samples obtained from L.
variegatus exposed to B(a)P at different points in time.
BPDE-DNA adduct/109nt
Time (h)
2
6
24
48
Control
B(a)P exposed worms
0.0±0.5
19.6±3.2
2.6±2.5
17.2±1.8
2.8±1.8
17.3±3.4
2.5±1.5
17.7±1.8
2.4±1.5
19.3±6.7
0.9±1.7
18.2±3.5
1.8±1.6
7.7±1.9
2.5±1.5
10.8±3.9
3.2±1.0
12.3±5.3
--
13.5±3.7
4.7±1.2
12.0±4.6
1.8±1.2
11.7±4.1
-- sample was lost due to a broken centrifugation tube
3.3.2 Metabolic Response to Chemical Treatment and Hypoxia
Many studies to test species’ responses to waterborne
contaminants are routinely conducted under normal oxygen
conditions despite the prevalence of hypoxia in the aquatic
environment. In this present study, the whole tissue contents of
the exposed chemicals differed in the worms under different
oxygen conditions, while the relative amount of toxicants
increased with time in all cases. However, hypoxia did not
enhance the accumulation of PCP in the worms as much as in
the other chemicals and hence underlines the significance of
oxygen in chemical uptake (Figure 1 Paper IV). The measured
activity of the different tissue extract fractions was similar in
both oxygen conditions for chlorpyrifos and PCP, but varied
59
markedly between both chemicals. On the other hand, the
radioactivity of the aqueous phase tissue extract from the B(a)P
exposed worms almost doubled under normoxia prior to a time
variable decrease (Figure 1 Paper IV). Given that the parent
forms of the test chemicals may stimulate metabolites that are
different from those by their transformation products, it was
crucial to determine their relative proportions in the aqueous
phase and hexane. These findings suggest that increased oxygen
availability can accelerate the biological transformation of B(a)P,
which may predispose worms to metabolism enhanced toxicity.
The relative amount of the endogenous metabolites analyzed
varied with the chemical treatment, as with the different oxygen
conditions (Figure S1 Paper IV, Table S1 Paper IV). A couple of
metabolites were tentatively designated by the retention index
and quantitation ion. For example, 1229.00_126 was barely
detected under normoxia, but in many cases accumulated
significantly (ANOVA, p ≤ 0.0001) in worms exposed to hypoxic
conditions. Equally, succinic acid and glycerol-3-phosphate
increased significantly (p < 0.0001) in individuals exposed under
hypoxia, while cysteine was effectively repressed (Figure S1
Paper IV, Table S1 Paper IV). It is thought that the hypoxia
inducible factor mechanisms may have become activated under
hypoxia, taking into account the decline in cellular levels of 2oxo and 2-hydroxy glutaric acid in relation to a corresponding
accumulation of succinate. The relative amount of glucose,
maltose, phosphoric acid or xanthurenic acid was considerably
higher compared to the other metabolites. Cysteine is an
essential component in the synthesis of glutathione, which in
turn scavenges free radical species to alleviate oxidative stress
conditions (Mansfield et al., 2004). Given the lack of consensus
opinion on the exact mechanisms how hypoxia influences ROS
production, it is possible that the decline of intracellular cysteine
in individuals exposed to hypoxia may have consequences on
60
glutathione synthesis, and hence increase their sensitivity to
oxidative damage.
The worms climbed the walls of the test jars within 2 minutes
of exposure in the hypoxic water. It suggests that the animals
can sense reduced oxygen levels in a natural environment and
respond promptly to escape from the existing stressor. This
initial avoidance behavior spends a considerable amount of
energy in worms and may contribute to the depletion of sugars
and other high energy derivatives. Interestingly, worms
appeared to recover from hypoxic shock by accumulating
sugars at the later period of exposure. A decline in sugar level
has been previously reported in potato tubers under hypoxic
conditions (Zhou & Solomos, 1998). Although in different
species, it is in contrast with glycerol 3-phosphate (3TMS) that
increased significantly (p < 0.0001) in worms exposed under
hypoxia. Others have observed similar increase and proposed
that glycerol 3-phosphate can serve as a realistic indicator for
monitoring hypoxia (Bachelard et al., 1993). Despite the fact that
hypoxic influences were particularly strong and the most
dominant of the stressors (Figure 2 Paper IV), inositol-like
compounds
and
their
phosphate
derivatives
showed
considerable stability. This class of compounds was hardly
influenced by hypoxia; chemical treatment; or a combination of
both stressors. The results showed the separation of the samples
according to oxygen conditions, even though the B(a)P
treatment effects were not clearly apparent (Figure 2 Paper IV).
It suggests that external factors including hypoxia can influence
species’ response under toxicant exposure by interfering with
their endogenous metabolite compositions. Moreover, it is
possible that subjecting the animals to the stressors in sequence
(chemicals followed by hypoxia or vice versa) may affect the
overall composition of metabolites in exposed individuals. The
results from this study revealed that the PCP (100 µg/L) exposed
61
worms showed the highest sensitivity among the chemicals that
were tested, with the least response in individuals exposed to
B(a)P.
62
Figure 3. GOEAST analysis of molecular function category of genes in L. variegatus
exposed to PCP for 24 h. Boxes represent GO terms designated with the appropriate
GO ID, term definition, and extra details organized as “q/m\t/k(p – value)”, where q
is the count of genes associated with the listed GO ID (directly or indirectly) in our
dataset; m is the count of genes associated with the listed GO ID (directly or
indirectly) on the chosen platform; t is the total number of genes on the chosen
platform; k is the total number of genes in our dataset; p-value is the significance for
the enrichment in the dataset of the listed GO ID. Boxes marked in yellow represent
the significantly enriched GO terms. The enrichment significance of each GO term is
positively correlated with the level of color saturation of each node. White colored boxes
depict the non significant GO terms, while hierarchical trees not having any
significantly enriched GO terms were not shown. Red arrows represent relationships
between two enriched GO terms; black solid arrows represent relationships between
enriched and unenriched terms; while black dashed arrows represent relationship
between two unenriched GO terms.
63
64
4 Concluding remarks
The main purpose of this study was to investigate the
bioavailability, stability and toxicity of waterborne contaminants
with considerations for various inherent traits of the species.
The results revealed that the uptake kinetics between the species
were partly influenced by the DOC contents, which in turn
correlated negatively with the dissolved fractions of B(a)P.
Furthermore, they showed that degradation of contaminants
under continuous solar radiation can be assessed by species’
fitness and mobility. It implies that the presence of vegetation
and forest canopies over natural waters can affect the
degradation of chemical contaminants. The application of a
direct UV light to remediate contaminated sites depends on the
overall
property
consequences
transcriptomics
of
on
and
contaminants
natural
and
remediation.
metabolites
analysis
altogether
A
has
combined
revealed
some
mechanistic processes that are crucial in the homeostasis of L.
variegatus exposed to waterborne contaminants. Compared to
chemical treatment, hypoxia was prominently the dominant
stressor even though the effects of their interactions were
evident only in few analyzed metabolites. Taken together, the
observed molecular changes can act as early signs in the event of
pollution. Also, they augment the current understanding of
underlying toxicity mechanisms for the model contaminants.
65
66
5 Toxicity Assessment
Methods and Future
Perspectives
Traditional methods to assess chemical risks were based on total
concentrations and accumulation indices, including the EC50,
LC50, and BCF values. To improve the reproduction of test
results, controlled conditions are constituted in the laboratory
despite undermining the value of environmental realism.
However, regulators still utilize toxicity indices to set threshold
limits that hardly provide any warnings against imminent
toxicity. In practice, toxicity indices rely on the evidence of
bioaccumulation; measurable endpoints; and in the worst cases
the death of the organisms, to establish effects mediated by
chemicals. This approach is not preventive and despite the
numerous drawbacks, is still used to assess chemical risks.
Therefore, it is imperative to develop new methods that can
alert the relevant authorities to imminent exposure to chemical
contaminants.
Biomarkers including the OMICS technology are good
alternatives to detect minute levels of potentially harmful
pollutants prior to exerting toxicity to an exposed population.
However, several assumptions are associated with microarrays
including the fact that it does not explain cellular events beyond
the expression of genes. The technique assumes that mRNA
retains stability throughout transcription but also that mRNA is
transcribed into cDNA in equal proportions. The fact that
cellular events leading to the expression of genes including
67
transcription, translation and post-translation modifications are
inefficient, is a good basis to explore subcellular metabolic
changes in addressing the risk of chemicals. Considering the
above limitations, it is only natural to explore other assessment
methods that are not only regulated beyond the expression of
genes, but also have direct involvement in numerous cellular
processes. Accordingly, endogenous metabolites are the key
drivers of response in several different species above other
biological molecules, including transcripts and proteins.
In this thesis, different toxicity assessment methods were
explored in a manner that revealed the toxicity mechanisms for
the studied chemicals. In spite of these efforts, multiple methods
applied here have limited relevance if they are not explained in
physiological
terms
in
whole
organisms.
Even
more
importantly, adequate considerations for fluctuations in the
environmental conditions will further improve the relevance of
future research in this field. Hence, external variables such as
temperature, dissolved oxygen levels, pH, and sunlight are
crucially relevant in freshwater ecosystems. Future studies that
will consider these factors are required to validate existing
toxicity data but also to explain the influence of prevailing
environmental conditions on the fate of chemicals. Finally, it
will offer more reliable predictions of global climate change
impacts
on
toxic
chemical
distribution
and
consequences on individual species and population.
68
the
likely
6 References
Arfsten D P, Schaeffer D J & Mulveny D C (1996) The effects of
near ultraviolet radiation on the toxic effects of polycyclic
aromatic hydrocarbons in animals and plants: a review.
Ecotoxicology and Environmental Safety 33: 1-24.
Atkinson C A, Jolley D F & Simpson S L (2007) Effect of overlying
water pH, dissolved oxygen, salinity and sediment disturbances
on metal release and sequestration from metal contaminated
marine sediments. Chemosphere 69: 1428-1437.
Bachelard H, Bader-Goffer R, Ben-Yoseph O, Morris P & Thatcher
N (1993) Studies on metabolic regulation using NMR
spectroscopy. Developmental Neuroscience 15: 207-215.
Barron M G & Woodburn K B (1995) Ecotoxicology of chlorpyrifos.
Reviews of Environmental Contamination and Toxicology 144: 1-93.
Borchert J, Karbe L & Westendorf J (1997) Uptake and metabolism
of benzo(a)pyrene absorbed to sediment by the freshwater
invertebrate species Chironomus riparius and Sphaerium corneum.
Bulletin of Environmental Contamination and Toxicology 58: 158-165.
Brosché M, Vinocur B, Alatalo E, Lamminmaki A, Teichmann T,
Ottow E, Djilianov D, Afif D, Bogeat-Triboulot M -B, Altman A,
Polle A, Dreyer E, Rudd S, Paulin L, Auvinen P, Kangasjärvi J
(2005) Gene expression and metabolie profiling of Populus
euphratica growing in the Negev desert. Genome Biology 6: R101.
Buhler D R & Williams D E (1988) The role of biotransformation in
the toxicity of chemicals. Aquatic Toxicology 11: 19-28.
Butt C M, Berger U, Bossi R & Tomy G T (2010) Levels and trends
of
poly-and
perfluorinated
compounds
in
the
arctic
environment. Science of the Total Environment 408: 2936-2965.
69
Calamari D (2002) Assessment of persistent and bioaccumulating
chemicals in the aquatic environment. Toxicology 181: 183-186.
Carson R (1962) Silent Spring, Houghton Mifflin, USA.
Chace K V & Odessey R (1981) The utilization by rabbit aorta of
carbohydrates, fatty acids, ketone bodies, and amino acids as
substrates for energy production. Circulation Research 48: 850-858.
De Schamphelaere K A C & Janssen C R (2004) Effects of dissolved
organic matter concentrations and source, pH and water
hardness on chronic toxicity of copper to Daphnia magna.
Environmental Toxicology and Chemistry 23: 1115-1122.
Dias V, Vasseur C & Bonzom J -M (2008) Exposure of Chironomus
riparius larvae to uranium: Effects on survival, development
time, growth, and mouthpart deformities. Chemosphere 71: 574581.
do Nascimento N R, Nicola S M C, Rezende M O O, Oliveira T A &
Oberg
G
(2004)
Pollution
by
hexachlorobenzene
and
pentachlorophenol in the coastal plain of Sao Paulo state, Brazil.
Geoderma 121: 221-232.
Fassett D W (1975) Cadmium: biological effects and occurrence in
the environment. Annual Review of Pharmacology 15: 425-435.
Fent
K
(2003)
Ecotoxicological
problems
associated
with
contaminated sites. Toxicology Letters 140: 353-365.
Fent K (2004) Ecotoxicological effects at contaminated sites.
Toxicology 205: 223-240.
Gao D, Luo Y, Guevara D, Wang Y, Rui M, Goldwyn B, Lu Y,
Smith E C A, Lebwohl M & Wei H (2005) Benzo(a)pyrene and its
metabolites combined with ultraviolet A synergistically induce
8-hydroxy-2-deoxyguanosine via reactive oxygen species. Free
Radical Biology and Medicine 39: 1177-1183.
Garcia-Reyero N & Perkins E J (2010) Systems biology: leading the
revolution in ecotoxicology. Environmental Toxicology and
Chemistry 30: 265-273.
70
Gobas F A P C & Morrison H A (2000) Bioconcentration and
biomagnifications in the aquatic environment. In Boethling R S
& Mackay D (Eds) Handbook of Property and Estimation Methods
for Chemicals. CRC Press, Boca Raton, FL, pp. 189-231.
GOLM Metabolome Database:
http://csbdb.mpimp-golm.mpg.de/csbdb/gmd/gmd.html
Götz S, García-Gómez J -M, Terol J, Williams T D, Nueda M J,
Robles M, Talón M, Dopazo J & Conesa A (2008) Highthroughput functional annotation and data mining with the
Blast2GO suite. Nucleic Acids Research 36: 3420-3435.
Hiller K, Hangebrauk J, Jäger C, Spura J, Schreiber K & Schomburg
D (2009) MetaboliteDetector: Comprehensive Analysis Tool for
Targeted and Nontargeted GC/MS Based Metabolome Analysis.
Analytical Chemistry 81: 3429-3439.
Hung H, Kallenborn R, Breivik K, Su Y, Brorström-Lundén E,
Olafsdottir K, Thorlacius J M, Leppänen S, Bossi R, Skov H,
Manø S, Patton G W, Stern G, Sverko E & Fellin P (2010)
Atmospheric monitoring of organic pollutants in the Arctic
under the Arctic Monitoring and Assessment Programme
(AMAP): 1993–2006. Science of the Total Environment 408: 28542873.
IARC (1983) IARC Monographs on the Evaluation of Carcinogenic
Risks to Humans: Polynuclear Aromatic Compounds, Part 1:
Chemical, Environmental and Experimental Data Vol. 32. pp.
33-224, Lyon, France.
Ikuno E, Matsumoto T, Okubo T, Itoi S & Sugita H (2008)
Differences in the sensitivity to chemical compounds between
female and male neonates of Daphnia magna. Environmental
Toxicology 23: 570-575.
Jantunen A P K, Tuikka A, Akkanen J & Kukkonen J V K (2008)
Bioaccumulation of Atrazine and Chlorpyrifos to Lumbriculus
variegatus from lake sediments. Ecotoxicology and Environmental
Safety 71: 860-868.
71
Järup L (2003) Hazards of heavy metal contamination. British
Medical Bulletin 68: 167-182.
Jensen L K, Honkanen J O, Jæger I & Carroll J (2012)
Bioaccumulation of phenanthrene and benzo(a)pyrene in
Calanus finmarchicus. Ecotoxicology and Environmental Safety 78:
225-231.
Jin Y H, Clark A B, Slebos R J C, Al-Refai H, Taylor J A, Kunkel T A,
Resnick M A & Gordenin D A (2003) Cadmium is a mutagen
that acts by inhibiting mismatch repair. Nature Genetics 34: 326329.
Jonker M T O & Koelmans A A (2001) Polyoxymethylene solid
phase extraction as a partitioning method for hydrophobic
organic chemicals in sediment and soot. Environmental Science
and Technology 35: 3742-3748.
Kane Driscoll S B & McElroy A E (1997) Elimination of sedimentassociated benzo(a)pyrene and its metabolites by polychaete
worms exposed to 3-methylcholanthrene. Aquatic Toxicology 39:
77-91.
Kim J, Choi K, Cho I, Son H & Zoh K (2007) Application of a
microbial toxicity assay for monitoring treatment effectiveness
of pentachlorophenol in water using UV photolysis and TiO2
photocatalysis. Journal of Hazardous Materials 148: 281-286.
Kluttgen B, Dulmer U, Engels M & Ratte H T (1994) ADaM, an
artificial freshwater for the culture of zooplankton. Water
Research 28: 743-746.
Kukkonen J & Landrum P F (1994) Toxicokinetics and toxicity of
sediment-associated
pyrene
to
Lumbriculus
variegatus
(oligochaeta). Environmental Toxicology and Chemistry 13: 14571468.
Kukkonen J V K (2002) Lethal body residue of chlorophenols and
mixtures of chlorophenols in benthic organisms. Archives of
Environmental Contamination and Toxicology 43: 214-220.
72
Leslie H A, Ter Laak T L, Busser F J M, Kraak M H S & Hermens J L
M (2002) Bioconcentration of organic chemicals: Is a solid-phase
microextraction fiber a good surrogate for biota? Environmental
Science and Technology 36: 5399-5404.
Leversee G J, Giesy J P, Landrum P F, Gerould S, Bowling J W,
Fannin T E, Haddock J D, Bartell S M (1982) Kinetics and
biotransformation of benzo(a)pyrene in Chironomus riparius.
Archives of Environmental Contamination and Toxicology 11: 25-31.
Livingstone D R (1998) The fate of organic xenobiotics in aquatic
ecosystems:
quantitative
and
qualitative
differences
biotransformation by invertebrates and fish.
in
Comparative
Biochemistry and Physiology, Part A 120: 43-49.
Mackay D & Fraser A (2000) Bioaccumulation of persistent organic
chemicals: mechanisms and models. Environmental Pollution 110:
373-391.
Mansfield K D, Simon M C & Keith B (2004) Hypoxic reduction in
cellular glutathione levels requires mitochondrial reactive
oxygen species. Journal of Applied Physiology 97: 1358-1366.
Marín-Guirao L, Cesar A, Marín A, Lloret J & Vita R (2005)
Establishing the ecological quality status of soft-bottom miningimpacted coastal water bodies in the scope of the Water
Framework Directive. Marine Pollution Bulletin 50: 374-387.
Mayer P & Holmstrup M (2008) Passive dosing of soil invertebrates
with polycyclic aromatic hydrocarbons: limited chemical
activity explains toxicity cutoff. Environmental Science and
Technology 42: 7516-7521.
Mott H V (2002) Association of hydrophobic organic contaminants
with soluble organic matter: evaluation of the database of Kdoc
values. Advances in Environmental Research 6: 577-593.
Neff J M (1985) Polycyclic aromatic hydrocarbons. In Rand G M &
Petrocelli S R (Eds) Fundamentals of Aquatic Toxicology.
Hemisphere Publ. Corp., New York, pp. 416-454.
73
OECD Guidelines for the testing of chemicals - Guideline 202:
Daphnia magna reproduction test, adopted in April 2004.
Penttinen O, Kukkonen J & Pellinen J (1996) Preliminary study to
compare body residues and sublethal energetic responses in
benthic
invertebrates
exposed
to
sediment-bound
2,4,5-
trichlorophenol. Environmental Toxicology and Chemistry 15: 160166.
Penttinen O & Kukkonen J (1998) Chemical stress and metabolic
rate
in
aquatic
invertebrates:
threshold,
dose-response
relationships, and mode of toxic action. Environmental Toxicology
and Chemistry 17: 883-890.
Persoone G, Baudo R, Cotman M, Blaise C, Thompson K Cl,
Moreira-Santos M, Vollat B, Törökne A & Han T (2009) Review
on the acute Daphnia magna toxicity test – Evaluation of the
sensitivity and the precision of assays performed with
organisms from laboratory cultures or hatched from dormant
eggs. Knowledge and Management of Aquatic Ecosystems 393: 1-29.
Price D A, Tulsian J O, Saunders M E & Fend S V (2005) An
incidental EST library for a polyploid oligochaete. Florida
Genetics 2005 November 30 – December 1, 2005 J. Wayne Reitz
Union, University of Florida, Gainesville, FL. Poster #54.
http://www.ufgi.ufl.edu/Symposium/FG2005Program_rev.pdf
Ristola T, Pellinen J, Ruokolainen M, Kostamo A & Kukkonen J V K
(1999) Effects of sediment type, feeding level, and larval density
on growth and development of a midge (CHIRONOMUS
RIPARIUS). Environmental Toxicology and Chemistry 18: 756-764.
Roessner U, Wagner C, Kopka J, Trethewey R N & Willmitzer L
(2000) Simultaneous analysis of metabolites in potato tuber by
gas chromatography-mass spectrometry. Plant Journal 23: 131142.
Sabaté J, Bayona J M, Solanas A M (2001) Photolysis of PAHs in
aqueous phase by UV irradiation. Chemosphere 44: 119-124.
74
Schuler L J, Wheeler M, Bailer A J & Lydy M J (2003) Toxicokinetics
of sediment-sorbed benzo[a]pyrene and hexachlorobiphenyl
using the freshwater invertebrates Hyalella azteca, Chironomus
tentans, and Lumbriculus variegatus. Environmental Toxicology and
Chemistry 22: 439-449.
Smyth G K (2005) Limma: linear models for microarray data. In
Gentleman R, Carey V, Dudoit S, Irizarry R & Huber W (Eds)
Bioinformatics and Computational Biology Solutions using R and
Bioconductor. Springer, New York, pp. 397-420.
Stackelberg P E, Furlong E T, Meyer M T, Zaugg S D, Henderson A
K & Reissman D B (2004) Persistence of pharmaceutical
compounds and other organic wastewater contaminants in a
conventional drinking-water-treatment plant. Science of the Total
Environment 329: 99-113.
Steinberg C E W, Stürzenbaum S R & Menzel R (2008) Genes and
environment – striking the fine balance between sophisticated
biomonitoring and true functional environmental genomics.
Science of the Total Environment 400: 142-161.
Terada H (1990) Uncoupler of oxidative phosphorylation.
Environmental Health Perspectives 87: 213-218.
Tse K C, Lo S –L & Wang J W H (2001) Pilot study of in situ
thermal treatment for the remediation of pentachlorophenolcontaminated aquifers. Environmental Science and Technology 35:
4910-4915.
Valko M, Morris H & Cronin M T D (2005) Metals, Toxicity and
Oxidative Stress. Current Medicinal Chemistry 12: 1161-1208.
Van der Wal L, Jager T, Fleuren R H L J, Barendregt A, Sinnige T L,
van Gestel C A M & Hermens J L M (2004) Solid-phase
microextraction to predict bioavailability and accumulation of
organic micropollutants in terrestrial organisms after exposure
to a field-contaminated soil. Environmental Science and Technology
38: 4842-4848.
75
Venkataraman K & Job S V (1980) Effect of temperature on the
development, growth and egg production in Daphnia carinata
king (Cladocera-Daphnidae). Hydrobiologia 68: 217-224.
Vitousek P M, Mooney H A, Lubchenco J & Melillo J M (1997)
Human Domination of Earth’s Ecosystems. Science 227: 494-499.
Vrana B, Mills G A, Allan I J, Dominiak E, Svensson K, Knutsson J,
Morrison G & Greenwood R (2005) Passive sampling techniques
for monitoring pollutants in water. Trends in Analytical Chemistry
24: 845-868.
Wan M T (1994) Utility right-of-way contaminants: Polycyclic
aromatic hydrocarbons. Journal of Environmental Quality 23: 12971304.
Wang Z, Lu M, Wang X, Yin R, Song Y, Le X C & Wang H (2009).
Quantum dots enhanced ultrasensitive detection of DNA
adducts. Analytical Chemistry 81: 10285-10289.
Webster E, Mackay D & Wania F (1998) Evaluating environmental
persistence. Environmental Toxicological Chemistry 17: 2148-2158.
White K E (2005) Ph.D Dissertation. Bioavailability of Polycyclic
Aromatic Hydrocarbons in the Aquatic Environment. North
Carolina State University, Raleigh, USA.
Wild S R & Jones K C (1995) Polycyclic Aromatic Hydrocarbons in
the United Kingdom environment: a preliminary source
inventory and budget. Environmental Pollution 88: 91-108.
Witters H E (1998) Chemical speciation dynamics and toxicity
assessment in aquatic systems. Ecotoxicology and Environmental
Safety 41: 90-95.
Zheng Q & Wang X J (2008) GOEAST: a web–based software
toolkit for Gene Ontology enrichment analysis. Nucleic Acids
Research May 16.
Zhou D & Solomos T (1998) Effect of hypoxia on sugar
accumulation, respiration, activities of amylase and starch
phosphorylase, and induction of alternative oxidase and acid
76
invertase during storage of potato tubers (Solanum tuberosum cv.
Russet Burbank) at 1⁰C. Physiologia Plantarum 104: 255-265.
77
Dissolved organic matter, oxygen
condition, and UV light influence the
bioavailability, uptake, and toxicity
of chemical contaminants in the
aquatic environment. In this thesis,
chemical partitioning processes,
and sublethal toxic responses were
studied in freshwater species under
different exposure conditions.
Complementary techniques were
applied to investigate the underlying
toxicity mechanisms, with particular
emphasis on transcriptional and
metabolic changes.
Publications of the University of Eastern Finland
Dissertations in Forestry and Natural Sciences
isbn 978-952-61-1014-1 (nid.)
issnl 1798-5668
issn 1798-5668
isbn 978-952-61-1015-8 (pdf)
issn 1798-5676 (pdf)
dissertations | No 95 | Stanley Ozoemena Agbo | Impact of environmental factors on chemical bioavailability and toxicity to...
Stanley Ozoemena Agbo
Impact of environmental
factors on chemical bioavailability and toxicity to aquatic
invertebrates: consequences
on transcriptional and
metabolic regulations
Stanley Ozoemena Agbo
Impact of environmental factors
on chemical bioavailability and
toxicity to aquatic invertebrates:
consequences on transcriptional
and metabolic regulations
Publications of the University of Eastern Finland
Dissertations in Forestry and Natural Sciences No 95