biomarkers in experimental ecotoxicology
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Dissertation thesis
Biomarkers in experimental ecotoxicology
Veronika Pašková
2012
Masaryk University, Faculty of Science
Research Centre for Toxic Compounds in the Environment
Brno, Czech Republic
Supervisor: Mgr. Klára Hilscherová, Ph.D.
Consultant: Doc. RNDr. Luděk Bláha, Ph.D.
BIBLIOGRAPHIC IDENTIFICATION
Author Veronika Pašková
Title of dissertation Biomarkers in experimental ecotoxicology
Title of dissertation (in Czech) Využití biochemických markerů v
experimentální ekotoxikologii
Ph.D. study program Chemistry
Specialization Environmental chemistry
Supervisor Mgr. Klára Hilscherová, Ph.D.
Year of defense 2012
Keywords biomarkers, detoxification, oxidative stress,
polycyclic aromatic hydrocarbons, azaarenes,
cyanobacterial biomass, pesticides, plants, fish,
birds, amphibian, embryotoxicity
Keywords (in Czech) biomarkery, detoxifikace, oxidativní stres,
polycyklické aromatické uhlovodíky, azaareny,
sinicová biomasa, pesticidy, rostliny, ryby, ptáci,
obojživelníci, embryotoxicita
© VERONIKA PAŠKOVÁ, MASARYK UNIVERSITY, 2012
AKNOWLEDGEMENTS
I would like to express my thanks to my supervisor, Dr. Klára Hilscherová, for
precious guidance and valuable support during my postgraduate study.
I would also like to thank Assoc. Prof. Luděk Bláha for offering me the scientific
background and motivation during my postgraduate study.
Next I would like to thank my friends and colleagues from the Research Centre for
Toxic Compounds in the Environment, from both the Ecotoxicology and
Environmental chemistry division, for helpful and unselfish advice and support
and friendly working-space.
Special thanks belong to my family and friends for invaluable support during my
studies.
ABSTRACT
The environment is continuously loaded with chemical compounds released by
urban communities and industry. Not only anthropogenic substances may pose a
risk to organisms. The intensification of agricultural and industrial activities is
associated with the increase of eutrophication in surface freshwater bodies
supporting expansion of phytoplanktonic blooms, which can produce secondary
metabolites with adverse effects on organisms. Biochemical markers can be
examined to assess the exposure or the effects of toxicants. They can provide
information about the health status of organisms and can be thus used as early
warning signals of general or particular stress. This dissertation thesis focuses on
the biomarkers of exposure and effects of various anthropogenic and natural
compounds. The phase I (cytochrome P-450 monooxygenases) and II
biotransformation enzymes (glutathione-S-transferases) and antioxidants
(superoxide dismutase, catalase, glutathione peroxidase, glutathione reductase and
glutathione) were studied together with oxidative stress parameters (lipid
peroxides).
This thesis is divided into four thematic parts covering assessment of biomarkers
in experiments with plants exposed to polycyclic aromatic hydrocarbons and
azaarenes, fish species exposed to cyanobacterial biomass, birds exposed next to
cyanobacterial biomass to heavy metals and vaccine and frogs exposed to
pesticides. Presented results are based on four scientific publications, one review-
paper and one manuscript.
The first part of the thesis documented responses of three plant species to exposure
to polycyclic aromatic compounds (PAHs) and their N-heterocyclic derivatives
(NPAHs) during 4-day germination. Standard phytotoxic parameters together with
biochemical responses were determined after exposure to three parental PAHs and
seven NPAHs. NPAHs were significantly more phytotoxic than parent PAHs,
however all chemicals modulated activity of plant detoxification and antioxidative
enzymes.
The second part of the thesis characterized the responses of selected biomarkers
after four and nine-week exposure of two fish species to the natural cyanobacterial
water-bloom. Modulations of biomarkers, especially activities of glutathione
reductase and glutathione-S-transferase and level of glutathione, have confirmed
an important role of oxidative stress in the toxicity of complex cyanobacterial
bloom. Changes of biomarkers preceded any signs of toxicity and may thus serve
as sensitive markers of stress caused by cyanobacterial exposure.
The third part of this thesis showed effects of single cyanobacterial exposure on
standard bird model species Japanese quail in 10-day and 30-day study and also of
30-day multistressor exposure to cyanobacteria, heavy metals and vaccination. The
study brought unique data from the first controlled experiments with the exposure
to cyanobacterial biomass in birds. Birds reacted to cyanobacterial exposure as to
xenobiotics, which was documented by the activation of general detoxification
mechanisms.
The fourth part of the thesis reviewed the involvement of oxidative stress in the
process of teratogenic action of pesticides in relation to their adverse effects on the
non-target organisms - amphibians, fish and aquatic invertebrates. Further, toxic
effects of paraquat and diquat on the early phases of amphibian development were
described using African clawed frog in the standard FETAX scheme supplemented
with the assessment of sublethal biochemical markers. The baseline developmental
profile of antioxidative and detoxification compounds and the effects of pesticides
on these parameters were evaluated in 24 hour-intervals. The protective effect of
external addition of antioxidant ascorbic acid supported the theory of oxidative
stress involvement in bipyridyl pesticides teratogenicity.
The present thesis demonstrates the involvement of oxidative stress in toxicity of
several important types of environmental stressors. Biomarkers reflect toxic
mechanisms and major processes protecting tissues from oxidative stress. In our
studies, namely glutathione reductase and glutathione-S-transferase responded to
low concentrations of stressors preceding any signs of toxicity and can be
successfully used as sensitive markers of effects of various environmental
stressors.
ABSTRAKT
Do životního prostředí jsou vlivem lidské činnosti a průmyslu neustále vnášeny
chemické látky. Avšak rizika pro lidské zdraví představují nejen antropogenní
chemikálie, ale i sloučeniny vylučované do prostředí vodními květy sinic, jejichž
růst je umocňován eutrofizací vod v důsledku intenzivní hospodářské činnosti a
průmyslových aktivit. Sinice produkují sekundární metabolity, u kterých bylo
prokázáno negativní působení na organismy. K hodnocení účinků a expozice
chemickým látkám lze použít biochemické markery, které indikují zdravotní stav
organismu a mohou být použity jako časné varovné signály narušení organismu.
Dizertační práce se zabývá studiem biomarkerů expozice a účinků různých
antropogenních i přírodních látek na organismy. Konkrétně je zaměřena na
enzymy první a druhé fáze biotransformace (cytochrom P-450 monooxygenázu;
glutation-S-transferázu), antioxidativní sloučeniny (superoxid dismutázu, katalázu,
glutation peroxidázu, glutation reduktázu a glutation) a parametr oxidativního
stresu (lipidní peroxidaci).
Dizertační práce je rozčleněna do čtyř tematických celků, které se zabývají
hodnocením biomarkerů oxidativního stresu a detoxifikace v experimentech s
rostlinami exponovanými polycyklickými aromatickými sloučeninami, rybami
exponovanými biomasou sinic, ptáky exponovanými biomasou sinic a také
těžkými kovy a patogeny a žábami exponovanými pesticidy. Předložené výsledky
byly publikovány ve čtyřech vědeckých a jednom rešeršním článku a dále jsou
součástí jednoho manuskriptu.
První část dizertační práce zkoumala vliv polycyklických aromatických sloučenin
(PAHs) a jejich N-heterocyklických derivátů (NPAHs) na tři druhy vyšších rostlin
ve čtyřdenním testu klíčivosti. Kromě standardních fytotoxických parametrů byly
sledovány biochemické markery jako odpověď na expozici třem parentálním
PAHs a sedmi NPAHs sloučeninám. Výraznější fytotoxické účinky byly
prokázány u NPAHs, avšak antioxidativní a detoxifikační parametry byly
významně ovlivněny všemi testovanými chemickými sloučeninami.
Druhá část dizertační práce se zabývala stanovením vybraných biomarkerů ve
čtyř- a devítitýdenním experimentu se dvěma druhy ryb vystavenými účinkům
přírodní biomasy sinic. Modulace biomarkerů, a to zejména aktivit glutation
reduktázy a glutation-S-transferázy a hladiny glutationu, potvrdila důležitou úlohu
oxidativního stresu v toxicitě komplexní biomasy sinic. Modulace biomarkerů v
rybích tkáních předcházely jakékoli známky toxicity, a mohou proto sloužit jako
senzitivní časné signály stresu způsobeného sinicovou expozicí.
Třetí část dizertační práce zkoumala účinky deseti- a třicetidenního působení
biomasy sinic na křepelku japonskou a dále účinky třicetidenní kombinované
expozice biomasou sinic, těžkými kovy a vakcinace. Tyto první experimenty s
kontrolovanými dávkami sinicové biomasy přinesly unikátní výsledky o toxicitě
sinic u ptáků. Ptáci reagovali na sinice podobně jako na xenobiotika, což bylo
dokumentováno aktivací detoxifikačních mechanismů.
Čtvrtá část dizertační práce shrnula úlohu oxidativního stresu v procesu
teratogeneze pesticidů u necílových organismů – obojživelníků, ryb a vodních
bezobratlovců. Podkapitolu tvoří výzkum toxických účinků paraquatu a diquatu na
raná vývojová stádia obojživelníků s použitím drápatky vodní ve standardním testu
FETAX doplněném o hodnocení biochemických markerů. Výstupem experimentu
je kromě toxických účinků pesticidů také vývojový profil antioxidativních a
detoxifikačních parametrů ve 24-hodinových intervalech. Pozitivní účinky
přídavku antioxidantu kyseliny askorbové podpořily teorii o úloze oxidativního
stresu v teratogenitě bipyridylových pesticidů.
Tato dizertační práce dokumentuje úlohu oxidativního stresu v toxicitě několika
důležitých environmentálních stresorů. Biomarkery odrážejí mechanismus
toxického působení a také důležitých buněčných procesů, které ochraňují tkáně
před oxidativním stresem. Ve všech studiích byly nejcitliv ější enzymy glutation
reduktáza a glutation-S-transferáza, které reagovaly na nízké koncentrace stresorů
a předcházely jakékoli známky toxicity, a mohou být proto využívány jako citlivé
markery působení různých environmentálních stresorů.
LIST OF ABBREVIATIONS
ANOVA Analysis of Variance
CAT Catalase
CDNB 1-chloro-2,4-dinitrobenzene
DMNS Dimethylsulfoxide
DNA Deoxyribonucleic Acid
DTNB 5,5′-dithiobis-2-nitrobenzoic Acid
EDTA Ethylenediaminetetraacetic Acid
EROD Ethoxyresorufin-O-deethylase
FETAX Frog Embryo Teratogenesis Assay Xenopus
G-6-P Glucose-6-phosphate
GPx Glutathione Peroxidase
GSSG Glutathione Disulphide
GSH Glutathione
GST Glutathione-S-Transferase
GR Glutathione Reductase
hCG Human Chorionic Gonadotropin
IU International Unit
MDA Malondialdehyde
NADP+ Nicotinamide Adenine Dinucleotide Phosphate
NADPH Nicotinamide Adenine Dinucleotide Phosphate (reduced form)
NBT Nitrobluetetrazolium
NPAH N-heterocyclic Aromatic Hydrocarbons
P450 Cytochrome P450
PAH Polycyclic Aromatic Compounds
PBS Phosphate Buffered Saline
PUFAs Polyunsaturated Fatty Acids
SOD Superoxide Dismutase
RNA Ribonucleic Acid
ROS Reactive Oxygen Species
TBA Thiobarbituric Acid
TCA Trichloroacetic Acid
TI Teratogenic Index
UV Ultra Violet
VTG Vitellogenin
LIST OF ORIGINAL ARTICLES AND AUTHOR’S CONTRIBUTION TO
THE ARTICLES
This thesis is based on original publications listed below:
Paper I. Pašková, V., Hilscherová, K., Feldmannová, M. and Bláha, L. (2006).
Toxic effects and oxidative stress in higher plants exposed to polycyclic
aromatic hydrocarbons and their N-heterocyclic derivatives.
Environmental Toxicology and Chemistry, Vol. 25, No. 12, pp. 3238-
3245.
Veronika Pašková performed the experiments with plant germination, measured
the parameters of phytotoxicity and biomarkers in these experiments, evaluated
and interpreted the data, prepared and finalized the manuscript.
Paper II. Adamovský, O., Kopp, R., Hilscherová, K., Babica, P., Palíková, M.,
Pašková, V., Navrátil, S., Maršálek, B.and Bláha, L. (2007).
Microcystin kinetics (bioaccumulation and elimination) and
biochemical responses in common carp (Cyprinus carpio) and silver
carp (Hypophthalmichthys molitrix) exposed to toxic cyanobacterial
blooms. Environmental Toxicology and Chemistry, Vol. 26, No. 12,
pp. 2687-2693.
Veronika Pašková performed the measurement of biochemical markers of
detoxification and oxidative stress in fish tissues, evaluated and interpreted the
data and participated in the manuscript preparation.
Paper III. Pašková, V., Adamovský, O., Pikula, J., Skočovská, B., Banďouchová,
H., Horáková, J., Babica, P., Maršálek, B. and Hilscherová, K. (2008).
Detoxification and oxidative stress responses along with microcystins
accumulation in Japanese quail exposed to cyanobacterial biomass.
Science of the Total Environment, Vol. 398, Is. 1-3, pp. 34-47.
Veronika Pašková participated in the experiments with quails, performed the
measurement of biochemical markers of detoxification and oxidative stress in
quail tissues, evaluated and interpreted the data and prepared and finalized the
manuscript.
Paper IV. Pašková, V., Paskerová, H., Pikula, J., Banďouchová, H., Sedláčková, J.
and Hilscherová, K. (2011). Combined exposure of Japanese quails to
cyanotoxins, Newcastle virus and lead: Oxidative stress responses.
Ecotoxicology and Environmental Safety 74 (7): 2082-2090.
Veronika Pašková participated in the experiments with quails, performed the
measurement of biochemical markers of detoxification and oxidative stress in
quail tissues, evaluated and interpreted the data and prepared and finalized the
manuscript.
Paper V. Pašková, V., Hilscherová, K. and Bláha, L. (2011). Teratogenicity and
embryotoxicity in aquatic organisms after pesticide exposure and the
role of oxidative stress. Reviews of Environmental Contamination and Toxicology, Vol. 211, pp. 25-61.
Veronika Pašková collected literature, analyzed the current state of knowledge and
prepared and finalized the review manuscript.
Paper VI. Pašková, V., Moosová, Z. and Hilscherová, K. (2012). Embryotoxicity
and induction of oxidative stress after exposure to bipyridyl herbicides
paraquat and diquat on the model non-targer aquatic organism African
clawed frog (Xenopus laevis). Manuscript in preparation.
Veronika Pašková performed the FETAX experiments with frog embryos,
measured biochemical markers of detoxification and oxidative stress in frog
tissues, evaluated and interpreted the data and prepared the manuscript.
TABLE OF CONTENT
CHAPTER 1 ...................................................................................................................................19
1.1 PREFACE.............................................................................................................................20 1.2 SCOPE AND OBJECTIVES OF THE THESIS..............................................................................22
CHAPTER 2 ...................................................................................................................................25
2.1 INTRODUCTION TO BIOMARKERS........................................................................................26
2.2 BIOMARKERS OF BIOTRANSFORMATION..............................................................................27 2.3 OXIDATIVE STRESS.............................................................................................................30 2.4 BIOMARKERS OF EXPOSURE................................................................................................36
2.5 BIOMARKERS OF EFFECT.....................................................................................................37
CHAPTER 3 – PAPER No.1 .........................................................................................................43
DETOXIFICATION AND OXIDATIVE STRESS RESPONSES IN PLANTS AFTER EXPOSURE TO POLYCYCLIC AROMATIC COMPOUNDS AND THEIR N-HETEROCYCLIC DERIVATES
3.1 HYPOTHESES OF THE STUDY...............................................................................................44
3.2 RESULTS AND DISCUSSION..................................................................................................46
CHAPTER 4 – PAPER No.2 .........................................................................................................51
FISH EXPOSURE TO CYANOBACTERIAL BIOMASS - DETOXIFICATION AND ANTIOXIDATIVE RESPONSES
4.1 HYPOTHESES OF THE STUDY...............................................................................................52
4.2 RESULTS AND DISCUSSION..................................................................................................54
CHAPTER 5 – PAPER No.3 .........................................................................................................57
DETOXIFICATION AND OXIDATIVE STRESS RESPONSES IN JAPANESE QUAIL EXPOSED TO CYANOBACTERIAL BIOMASS
5.1 HYPOTHESES OF THE STUDY...............................................................................................58
5.2 RESULTS AND DISCUSSION..................................................................................................60
CHAPTER 6 – PAPER No.4 .........................................................................................................65
DETOXIFICATION AND OXIDATIVE STRESS RESPONSES IN JAPANESE QUAIL EXPOSED TO MULTIPLE STRESSORS
6.1 HYPOTHESES OF THE STUDY...............................................................................................66
6.2 RESULTS AND DISCUSSION..................................................................................................67
CHAPTER 7 – PAPER No.5 .........................................................................................................73
DETOXIFICATION AND OXIDATIVE STRESS RESPONSES OF EARLY STAGES OF AQUATIC ORGANISMS EXPOSED TO PESTICIDES
7.1 INTRODUCTION.....................................................................................................................1 7.2 PESTICIDES TERATOGENICITY IN INVERTEBRATES..............................................................75 7.3 PESTICIDES TERATOGENICITY IN FISH.................................................................................76
7.4 PESTICIDES TERATOGENICITY IN AMPHIBIAN......................................................................77 7.5 ROLE OF OXYGEN AND ANTIOXIDANTIVE COMPOUNDS IN EMBRYOGENESIS.......................79 7.6 CONCLUSIONS.....................................................................................................................83
CHAPTER 8 – PAPER No.6......................................................................................................... 85
EMBRYOTOXICITY AND INDUCTION OF OXIDATIVE STRESS AFTER EXPOSURE OF MODEL NON-TARGET AQUATIC ORGANISM AFRICAN CLAWED FROG (XENOPUS LAEVIS) TO BIPYRIDYL HERBICIDES PARAQUAT AND DIQUAT
8.1 INTRODUCTION .................................................................................................................. 86 8.2 MATERIALS AND METHODS................................................................................................ 90
8.3 RESULTS............................................................................................................................ 94 8.4 DISCUSSION..................................................................................................................... 103
CHAPTER 9 – GENERAL DISCUSSION................................................................................ 109
9.1 BIOMARKERS AFTER EXPOSURE TO PAHS IN PLANTS....................................................... 110 9.2 BIOMARKERS OF EXPOSURE TO CYANOBACTERIAL BIOMASS IN FISH AND BIRDS AND
MULTIPLE–STRESSOR EXPOSURE IN BIRDS................................................................................. 111
9.3 BIOMARKERS OF EXPOSURE TO PESTICIDES IN AQUATIC INVERTEBRATES, FISH AND
AMPHIBIANS .............................................................................................................................. 114 9.4 CONCLUSIONS.................................................................................................................. 117
REFERENCES…………………………………………………………………………………..119 ANNEXES………………………………………………………………………………………..129
PAPER I PAPER II. PAPER III. PAPER IV. PAPER V. Curriculum vitae
1 CHAPTER 1
Preface
CHAPTER 1
20
1.1 Preface
The environment is continuously intentionally or unintentionally loaded with
foreign organic chemicals (xenobiotics) and metals released by urban communities
and industry. Since the 20th century, thousands of organic pollutants have been
produced and, in part, released into the environment (Helm et al. 2011). Many of
these chemicals that are released into the environment are extremely stable and
persistent and pose a hazard to the wild living organisms that are exposed directly
in their habitats. Persistent organic pollutants have toxic properties, resist
degradation, bioaccumulate in terrestrial and aquatic ecosystems and are
transported through air, water and migratory species and deposited far from the
place of their production (Choi and Wania 2011).
Moreover, not only anthropogenic substances may pose a risk to organisms in the
environment. The intensification of agricultural and industrial activities associated
with the growth of human population is associated with the increase of
eutrophication in surface freshwater bodies. These phenomena of high inputs of
nutrients into the waters together with particular temperature, environmental and
light conditions support the expansion of phytoplanktonic blooms that are
becoming more frequent worldwide. Cyanobacteria (blue-green algae) are known
to produce secondary metabolites with adverse effects on mammals, birds and fish
and have been recognized as human and animal health hazards (Codd 1996). The
production of natural cyanobacterial toxins (cyanotoxins) and other secondary
metabolites is thus being influenced by the anthropogenic activities accelerating
their production.
Organisms including human can be exposed to the environmental concentrations
of xenobiotics and also natural toxins and other compounds with possible
toxicological adverse effects. Next to the body burdens of pollutants in tissues the
biochemical markers may be examined to assess the exposure to or the effects of
toxicants (Van der Oost et al. 2003). Biochemical markers can provide information
about the health status of organisms and can be thus used as early warning signals
of general or particular stress (Korte et al. 2000).
PREFACE
21
Complex modern approaches of toxicogenomics or proteomics are used in some
ecotoxicology studies nowadays and are bringing new beneficial information about
sensitive markers. Toxicogenomic approach based on gene expression evaluation
is suggested to be more sensitive than conventional markers in detecting toxicity
signals in experimental studies and in looking for the new sensitive markers
(Ellinger-Ziegelbauer et al. 2008). On the other hand, the specificity of thousands
of genomic biomarkers may be sometimes confusing and not significant enough
(Zhang et al. 2012).
In this dissertation thesis, the conventional biomarkers of exposure and/or effects
of various anthropogenic and natural compounds on aquatic and terrestrial
organisms were studied. The phase I (cytochrome P-450 monooxygenases) and
II biotransformation enzymes (glutathione-S-transferases) and antioxidants
(superoxide dismutase, catalase, glutathione peroxidase, glutathione reductase and
glutathione) were studied. Together with the oxidative stress parameters (lipid
peroxides) they have been chosen as promising easily measurable parameters for
this thesis. Moreover, they can be used as the early-warning signals reflecting the
adverse biological responses to the environmental stressors.
In addition to the measurement of contaminants accumulating in tissues,
biomarkers can offer more complete and biologically more relevant information on
the potential impact of toxic pollutants on the health of organisms (Stegeman et al.
1992). Moreover, biomarkers should be sensitive and quick measurements that can
indicate the exposure to xenobiotics in biomonitoring experiments just by an
increase in enzymatic activity during the biotransformation phase I and II (Roy et
al. 1995).
On the other hand, more complex results of ecotoxicological experiments can be
obtained when combining the biomarker approach with the analytical
measurements. The combination of biomarkers measurement with supplementary
approaches was used to enhance the data interpretation in studies within this
dissertation thesis. The role of biotransformation, oxidative stress, antioxidative
parameters and detoxification was studied in model organisms exposed to selected
environmental stressors using various experimental approaches.
CHAPTER 1
22
1.2 Scope and objectives of the thesis
This dissertation work was focused on biomarkers of biotransformation,
detoxification and oxidative stress in various organisms after exposure to chosen
environmental stressors.
The goal of the dissertation work was to assess the sensitivity of the biochemical
markers across the chosen ecotoxicological experiments in various model
organisms under the influence of different stressors. The research also aimed to
determine which of the parameters responded most strongly and frequently and
could be thus used as the most appropriate early-warning signals of sublethal
toxicity.
Important aim was to optimize the methods for the assessment of biomarkers
(glutathione and total protein level, glutathione-S-transferase, glutathione
peroxidase, glutathione reductase, ethoxyresorufin-O-deethylase, superoxide
dismutase and catalase activities, lipid peroxides level) in various types of samples
(plant roots and hypocotyls, fish hepatopancreas, bird liver, heart and brain tissue
and amphibian embryos and larvae).
Various approaches have been applied to assess the biotransformation,
antioxidative responses and oxidative stress processes in several representatives of
autotrophic and heterotrophic organisms:
– in plants (study No. 1 with higher terrestrial plants Sinapis alba, Triticum
aestivum and Phaseolus vulgaris)
– in fish (study No. 2 with common carp (Cyprio carpio) and silver carp
(Hypophthalmichthys molitrix))
– in birds (studies No. 3 and 4 with Japanese quail (Coturnix coturnix
japonica))
– in embryos and larvae of aquatic cold-blooded organisms (study No.5 –
review on amphibians, fish and aquatic invertebrates), especially in
amphibians (study No.6 with African clawed frog (Xenopus laevis))
PREFACE
23
Another criterion - the point of maturity of model organisms - can be used when
summarizing the approaches used in this work:
– early stages (study No.1 examining the plant germination and
polyaromatics phytotoxicity; study No.5 reviewing biomarkers in
embryo-larval development of aquatic organisms exposed to pesticides;
study No. 6 assessing pesticides embryotoxicity and teratogenity in the
frog early development)
– juveniles (studies No. 3 and 4 with acute and subchronic exposures of
four month old birds to cyanobacterial biomass and multiple stressors)
– mature organisms (study No. 2 focused on cyanobacterial exposure to
two years old fish)
Biochemical markers of exposure/effects and other ecotoxicological parameters
were studied using model environmental stressors:
– polycyclic aromatic hydrocarbons and azaarenes (study No. 1 with
three homocyclic parental compounds and their seven N-heterocyclic
derivates)
– cyanobacterial biomass (studies No. 2, 3 and 4 with natural
cyanobacterial biomasses with controlled cyanotoxins concentrations);
Pb and Newcastle virus vaccination (study No. 4)
– pesticides (studies No. 5 and 6 examining the role of pesticides in
embryotoxicity and teratogenity)
Moreover, multistressor exposure was adopted in one of the experiments to
simulate the ecological situation (study No. 4 – experiment with single and
combined exposures to three stressors – cyanobacterial biomass, lead and
vaccination).
CHAPTER 1
24
2 CHAPTER 2
Biomarkers and oxidative stress
CHAPTER 2
26
2.1 Introduction to biomarkers
Biomarkers can be defined as measurable changes of cellular or biochemical
compounds, structures or functions caused by xenobiotics, namely after exposure
to environmental contaminants. Biological response can occur on molecular,
cellular, tissue or organ level and can be measurable in biological systems such as
tissues, cells and biological fluids (Kimmenade and Januzzi 2012). These changes
are related to the exposure or to the effects of toxicants and have been successfully
applied to monitor the presence and the effects of contaminants in various
toxicological and ecotoxicological studies (Adonis et al. 2003; Almroth et al.
2005; Risom et al. 2005). Monitoring the parameters of an initial change caused by
the interaction of organism and xenobiotic compound can characterize the level of
exposure or toxic effect (Smith and Warner 1992). Biomarkers can provide
information on the health of organisms, and can be used as early warning signals
for general or particular stress (Korte et al. 2000). High sensitivity of biomarkers
enables their applications for detection of the early changes in pathogenesis or
physiological adaptation mechanisms. Moreover, molecular and biochemical
markers of biological response to chemical compounds can be used as diagnostic
or prognostic tools for assessing the effects of pollutants in the environment
(Saint-Denis et al. 1999).
From the toxicokinetics point of view, the fate of each chemical after entering the
organism is influenced by four main processes: adsorbtion, distribution,
metabolisation and excretion. The interaction of chemical with the organism can
be evaluated at various levels. Generally, various biomarkers may be used for
evaluation of the interaction of chemicals with the organisms in ecotoxicological
studies: biotransformation enzymes (phase I and II), antioxidative compounds,
oxidative stress parameters, biotransformation products, stress proteins,
metallothioneins, multixenobiotic resistance proteins, hematological and
histopathological parameters, immunological, reproductive and endocrine
parameters, genotoxic parameters, neuromuscular parameters, physiological and
morphological parameters and many others, as summarized by Van der Oost et al.
(2003).
BIOMARKERS AND OXIDATIVE STRESS
27
2.2 Biomarkers of biotransformation
Many xenobiotics are lipophilic, readily absorbed and can accumulate to reach
toxic levels in organism (Coleman et al. 1997). Xenobiotics are taken up into the
organism by several ways (Halliwell and Gutterdige 2007). Because they cannot
be used for nutrition or as a source of energy and because of their possible toxic
effects, the organisms had to develop sufficient mechanisms to avoid these toxic
effects. Biotransformation reactions belong to the general cellular mechanisms
protecting against possible toxic effects of xenobiotics, which are linked with
reactive oxygen species (ROS) and sequent damage to macromolecules and
tissues. Cells are equipped with endogenous enzymatic and non-enzymatic
defenses against oxidative damage, and also many small molecule antioxidants
from the dietary intake (Scandalios 1997). After xenobiotics enter the organism,
the major pathways of metabolisation and elimination are activated to defend
against the toxic effects of xenobiotics. Some of them are directly metabolized and
eliminated from the organism (Okuno et al. 2001). Metabolism of xenobiotics
proceeds generally in three phases including the elimination phase. The first phase
includes mixed-function-oxidase reactions and other reactions catalyzed by haem
proteins cytochromes P450, as for example oxidation, reduction, hydratation or
dehalogenation reactions. Mostly less toxic products are formed by reactions
catalyzed by P450. On the other hand, free radical intermediates can be also
formed in the phase I by the hydroxylation reactions (Halliwell and Gutterdige
2007). The intermediates are further metabolized in the second phase of
biotransformation. In the phase II the highly toxic electrophiles bind with the
endogenous nucleophilic centers resulting in formation of more soluble products
or inactivation of toxic intermediates. Glucuronidation, acetylation, methylation,
sulfation, aminoacids or glutathione conjugation reactions (simplified by the Fig.1)
belong to the most important reactions of this conjugation phases of
biotransformation.
CHAPTER 2
28
Figure 1.: Biotransformation phases I and II
Many xenobiotics are metabolized by conjugation with tripeptide glutathione
(GSH), catalysed by glutathione-S-transferase (GST) enzymes (Zimniak 2008).
Conjugation with GSH plays a key role in detoxification of many xenobiotics.
GSTs isoenzymes can be cytosolic or membrane-bound (mitochondria,
endoplasmic reticulum) proteins in all eukaryotes. Compounds metabolized by
GST enzymes include wide range of organic xenobiotics, drugs and toxins.
Moreover, GSTs may be important protectors against lipid peroxidation, they can
metabolize some of its toxic end-products. Except of their catalytic function they
serve also as intracellular carrier proteins for important bio-molecules as for
example haem, hormones or steroids (Halliwell and Gutterdige 2007). In case of
non-sufficient elimination or the excessive xenobiotics exposure level, reactive
intermediates (O2.-, OH. , H2O2) can arise and bind to cellular macromolecules.
Similarly, xenobiotic free radical reactive intermediates can react directly or
indirectly with molecular oxygen and initiate the formation of ROS. Reactive
oxygen species may cause oxidative stress - oxidatively damage cellular
macromolecules such as lipids, proteins, RNA and DNA, if not detoxified by
FAT-SOLUBLE
TOXINS PHASE I (Cytochrome P450 enzymes)
Oxidation Reduction Hydravion Dehalogenation Hydrolysis
INTERMEDIARY METABOLISM
PHASE II (Conjugation pathways)
Glucuronidation Glutatione conjugation Acetylation Methylation Sulfation Aminoacid conjugation
WASTE
ELIMINATION (via gall bladder and kidneys)
Hydratation Glutathione conjugation
ELIMINATION (via gall bladder and kidneys)
BIOMARKERS AND OXIDATIVE STRESS
29
antioxidants or antioxidative enzymes, as described above and graphically in
Figure 2 (Wells et al. 1997).
Figure 2.: The role of selected enzymes and non-enzymatic compounds in
biotransformation and detoxification of xenobiotics (including pathways of ROS
formation; (Wells et al. 1997)).
H2O2
xenobiotics P450
P450 GSH GSSG
GST
quinone
O2
O2.
semiquinone
covalent binding
- DNA
- protein
HO.
SOD
oxidative damage
- lipids
- DNA
- protein
H2O
GSH
GPx
G-6-P
NADP+ NADPH
GR
GSSG
G-6-P dehydrogenase
free radical intermediate
CAT
H20 O2
NA
DP
H P
450
redu
ctas
e
CHAPTER 2
30
2.3 Oxidative stress
Oxidative stress occurs as a result of an imbalance between the pro-oxidants and
the ability of the antioxidants to scavenge the excess ROS production in case of
impaired antioxidant defence mechanisms (Wells et al. 1997). Oxidative stress can
be characterized by an oxidative ‘burst’, or a rapid and transient production of high
quantities of ROS, such as superoxide radical, hydrogen peroxide, hydroxyl
radical, singlet oxygen, and hydroxyperoxyradicals (Halliwell and Gutterdige
2007).
The production of ROS is a natural phenomenon, triggered by various external
factors (Rijstenbil et al. 1994), and generally reduced in organisms under normal
conditions of growth (Ferrat et al. 2003). The oxidative burst can be induced
directly by various pollutants or indirectly by their metabolisation (Droege 2002).
Organic compounds and transition metals were shown to be pro-oxidants and to
accelerate the formation of oxy-radicals (Halliwell and Gutterdige 2007), and their
excess increased lipid peroxidation (see further details below) via loss of
membrane integrity (Rijstenbil et al. 1994).
On the other hand, at moderate concentrations, nitric oxide (NO) and ROS play an
important role as regulatory mediators in signaling processes. Higher organisms
have evolved the use of NO and ROS also as signaling molecules for other
physiological functions including regulation of vascular tone, monitoring of
oxygen tension in the control of ventilation and erythropoietin production, and
signal transduction from membrane receptors in various physiological processes
(Droege 2002).
The term “Reactive oxygen species” includes both oxygen radicals and certain
non-radicals, which are oxidizing agents and/or are easily converted into radicals
(HOCl, HOBr, O3, ONOO-, 1O2, H2O2). Generally, all oxygen radicals are ROS,
but not all ROS are oxygen radicals. The three major forms of ROS, the most
important reactive molecules derived from oxygen, include superoxide (O2.-),
hydrogen peroxide (H2O2) and hydroxyl radical (OH.).
BIOMARKERS AND OXIDATIVE STRESS
31
Hydroxyl radical
Hydroxyl radical can be generated from H2O2 by redox reactions involving metal
ions (Fenton reaction; see below), mostly by quinones and semiquinones.
Quinones can be derived from aromatic compounds by conversion of even number
of –CH= groups into –C(=O)– groups with any rearrangement of double bond.
SQ.- + H2O2 OH. + OH-
.
+ Q
Moreover, chemicals as for example some chlorinated compounds can generate
hydroxyl radical from H2O2 by direct, metal ion-independent reactions (Halliwell
and Gutterdige 2007). Hydroxyl radical can be also formed by UV-induced
homolytic fusion of the O-O bond in H2O2
H-O-O-H 2OH.
Ionising radiation is other source of OH.. Exposure to high-energy radiation can
result in hydroxyl radical production by homolytic fusion of water, also in living
cells, where the radicals often cause damage to cellular DNA, proteins and lipids.
Hydroxyl radical is strongly reactive with biomolecules and is able to cause more
damage to biological systems than any other ROS. Hydrogen peroxide can be
enzymatically metabolised to dioxygen and water by a number of different enzyme
systems or converted to hydrogen peroxide, which is extremely reactive, via a
chemical reaction catalysed by transition metals (Betterigde 2000).
Superoxide anion
The superoxide anion formed from molecular oxygen by the addition of an
electron does not readily cross membranes, although it can pass through the anion
exchange proteins present in some cells, for example erythrocytes and lung
(Halliwell and Gutterdige 2007). Superoxide can be in some cases produced in
vivo by the enzymes xanthine and hypoxanthin oxidase when the tissue is damaged
or by activated fagocyting cells. Two molecules of superoxide rapidly dismutate,
either spontaneously or via superoxide dismutases to dioxygen and hydrogen
peroxide (Khatisashvili et al. 1997).
CHAPTER 2
32
O2.- + O2
.- + 2H+ H2O2 + O2
Superoxide is far less reactive than OH. and does not react with most biological
molecules in aqueous solution. But it reacts with some other radicals and non-
radicals.
SO2 + O2.- O2 + SO2
.-
Hydrogen peroxide
Hydrogen peroxide is not a free radical but is nonetheless highly important
because of its ability to penetrate biological membranes. It plays a radical forming
role as an intermediate in the production of more reactive ROS molecules, such as
hydroxyl radical, via oxidation of transition metals. A variety of chemicals, mostly
aromatic compounds, can be enzymatically reduced to H2O2 or/and other free
radicals by the redox cycling reaction. Hydrogen peroxide can be removed by at
least three antioxidant enzyme systems, namely catalases, glutathione peroxidases
and peroxiredoxins (Nordberg and Arnér 2001).
2H2O2 2H2O + O2
Role of xenobiotics in oxidative stress
Reactive species have been suggested to be involved in the actions of many
xenobiotics as for example some pesticides, therapeutics or environmental
pollutants. There are several mechanisms of involvement of xenobiotics in the
formation of oxidative stress (Halliwell and Gutterdige 2007):
a) the xenobiotic already exists in a form of reactive species (oxides,
peroxides etc.)
b) the xenobiotic is metabolized to a reactive species
c) the xenobiotic undergoes redox cycling, i.e. it is reduced by a cellular
system and the reduction product is then reoxidized by O2, producing O2.-
BIOMARKERS AND OXIDATIVE STRESS
33
and regenerating the original compound; the cycle then repeats
d) the xenobiotic interferes with antioxidant defenses (many compounds are
metabolized by conjugation with GSH; a large dose may deplete GSH and
lead to secondary oxidative damage by failure to adequately remove
endogenous reactive species).
e) the xenobiotic stimulates endogenous generation of reactive species
(affecting mitochondrial electron transport, activating phagocytes)
f) the xenobiotic or its metabolite binds to biomolecules to create new
antigen, provoking an immune response involving reactive species
g) combination of the above listed mechanisms (cigarette smoke etc.).
Free radicals and ROS can readily react with most biomolecules, starting a chain
reaction of free radical formation. In order to stop this chain reaction, a newly
formed radical must either react with another free radical, eliminating the unpaired
electrons, or react with a free radical scavenger - a chain-breaking or primary
antioxidant (Nordberg and Arnér 2001). Oxidative stress implicates the damage of
plasma membranes because of high content of polyunsaturated fatty acids
(PUFAs). Cells have developed protection against oxidative stress but in case of
energy lack this protection potential is insufficient leading to oxidative damage to
biomolecules (Zhang et al. 2004).
Oxidative DNA damage
Oxidative stress attacks not only the fluidity of the membrane but also the integrity
of DNA in cellular nucleus. Two factors protect DNA from oxidative insult:
characteristic tight packaging of the DNA and antioxidants (Twigg et al. 1998).
Studies with gametes exposure to artificially produced ROS resulted in a
significant increase of DNA damage in the form of modification of all bases,
production of base-free sites, deletions, frameshifts, DNA cross-links, and
chromosomal rearrangements (Duru et al. 2000). Single and double DNA strand
CHAPTER 2
34
breaks were formed in association with oxidative stress (Aitken and Krausz 2001).
Enzymatic antioxidants catalase (Jeulin et al. 1989) or glutathione peroxidase
(Alvarez and Storey 1989), may shield DNA from damage by acting as ROS
scavengers (Toyokuni and Sagripanti 1992). The serious and irreversible DNA-
damage initiates the process of apoptosis (Vaux and Korsmeyer 1999). Oxidative
stress and deficiencies in natural processes such as chromatin package have been
identified as the main factors involved in the etiology of DNA damage (Dietrich et
al. 2005). Further, ROS have been shown to be mutagenic (Marnett 2000) as
suggested by chemical modification of DNA caused by ROS. Generally, damage
to DNA determined as for example chromosomal aberrations, DNA breaks or
micronuclei appearance could be understood as biomarker of effect of oxidative
stress.
Peroxidation of lipid membranes
Oxidative stress can also damage lipid membranes in the process of lipid
peroxidation (Cakmak and Horst 1991; Korte et al. 2000). The hydroxyl radical is
a powerful initiator of lipid peroxidation. Most membrane polyunsaturated fatty
acids have unconjugated double bonds separated by methylene groups, which
makes the methylene carbon–hydrogen bonds weaker, and therefore hydrogen is
more susceptible to abstraction. Once this abstraction has occurred, the radical is
stabilized and a conjugated diene is formed. Conjugated dienes react with O2 to
form a lipid peroxyl radical (ROO•), which abstracts hydrogen atoms from other
lipid molecules resulting in lipid hydroperoxides.
Lipid hydroperoxides are stable until they come into contact with transition metals,
such as iron or copper. These metals catalyze the generation of alkoxyl and
peroxyl radicals from lipid hydroperoxides, which then continue the chain reaction
within the membrane and propagate the damage throughout the cell. Propagation
of lipid peroxidation depends on the antioxidant strategies. Chain-breaking
antioxidants inhibit this process by scavenging peroxyl (RO•) and alkoxyl (ROO•)
radicals. The prevention of excessive ROS formation belongs to other antioxidant
defence mechanisms. For example the binding of metal ions can prevent them
BIOMARKERS AND OXIDATIVE STRESS
35
from initiating a chain reaction. Also the reduction of some enzymes producing
ROS (NADPH oxidase) can decrease tissue damage (Agarwal et al. 2003;
Yamamoto et al. 2001).
TBARS (thiobarbituric acid reactive substances) assay can be used for
quantification the naturally-occurring end-product of lipid peroxidation,
malondialdehyde (MDA) spectrophotometrically (Uchiyama and Mihara 1978;
Livingstone et al. 1990).
Peroxidation of proteins
ROS are also known to convert amino groups of protein to carbonyl moieties
(Parihar and Pandit 2003). ROS have been shown to react with several amino acid
residues in vitro, leading to modified and less active enzymes or even to denatured,
non-functional proteins (Butterfield et al. 1998). Oxidative modification of protein
leads to increased recognition and degradation by proteases and loss of enzymatic
activity (Rivett and Levine 1990). Among the most susceptible amino acids are
sulfur- (or selenium)-containing residues. General antioxidant systems such as
thioredoxines, glutathione reductase or glutathione protect proteins from such
modifications (Nordberg and Arnér 2001).
As summarized above, oxidative reactions play very important role in the fate of
chemicals and their toxic effects to the organisms. Responses of endogenous
compounds connected to these oxidative reactions can thus be used as biomarkers
indicating exposure to chemicals or their toxic effects. Based on their functions,
biomarkers can be divided into three categories – biomarkers of exposure, effect or
sensitivity (NRC 1987; WHO 1993).
CHAPTER 2
36
2.4 Biomarkers of exposure
Biomarkers of exposure cover the detection and measurement of an exogenous
substance or its metabolite or the product of an interaction between a xenobiotic
agent and some target molecule or cell that is measured in a compartment within
an organism (Van der Oost et al. 2003). These biomarkers strongly indicate
exposure of organism to a toxicant. They characterize the amount of toxicant that
has entered the organism. Besides the measurement of concentration of chemical
compounds including their metabolites in body fluids, body tissues or cells,
biomarkers of exposure can be assessed using measurement of the products of
interaction between the xenobiotic compound and endogenous substance.
Biomarkers of exposure to some chemical compound include metallothioneins,
heat shock proteins, acetylcholinesterase inhibition, biotransformation enzymes
and many others (Berglund et al. 2007).
Metallothioneins form a family of cysteine-rich, low molecular weight proteins
localized in the membrane of the Golgi apparatus. They have the capacity to bind
both physiological (such as zinc, copper, selenium) and xenobiotic (such as
cadmium, mercury, silver, arsenic) heavy metals through the thiol group of its
cysteine residues. Metallothioneins may provide protection against metal toxicity,
are involved in regulation of physiological metals (Zn and Cu) and provide
protection against oxidative stress (House 2009).
Heat shock proteins are a class of functionally related proteins involved in folding
and unfolding of other proteins. Their expression is increased when cells are
exposed to elevated temperatures or other stress. They are upregulated in reaction
to the exposure to different kinds of environmental stress conditions, exposure to
xenobiotics or contaminants (ethanol, arsenic or trace metals among many others),
or water deprivation and others. As a consequence, the heat shock proteins are also
referred to as stress proteins and their upregulation is sometimes described more
generally as a part of the stress response (Santoro 2000).
Acetylcholinesterase is a serine protease enzyme that hydrolyzes the
neurotransmitter acetylcholine. It is found mainly at neuromuscular junctions and
cholinergic brain synapses, where its activity serves to terminate synaptic
BIOMARKERS AND OXIDATIVE STRESS
37
transmision. Activity of this enzyme can be employed as a biomarker of exposure
to organophosphates or some other pesticides and chemical compounds because of
their ability to inhibit the catalytic activity of this very important enzyme (Taylor
and Radic 2004).
The most important families of enzymes involved in the first and second phase of
biotransformation can be used as biomarkers of exposure to chemical compounds.
Changes in catalytic functions of P450 can be monitored using for example the
ethoxyresorufin-O-deethylase (EROD) fluorimetric bioassay, where the induction
of cytochrome P4501A catalytic activity is measured (Prough et al. 1978). A
multitude of chemicals induce EROD activity in a variety of organisms and this
biomarker has proven its value in a number of experiments and field investigations
of industrial effluents, contaminated sediments or chemical spills (Whyte et al.
2000).
Glutathione-S-transferase (GST) enzymes play a crucial role in the metabolisation
of xenobiotics by conjugation with glutathione (more details about this reaction
see in the next Chapter). Compounds metabolized by GST enzymes include wide
range of organic xenobiotics, drugs and toxins (Halliwell and Gutterdige 2007).
The catalytic activity of GST as the biomarker of exposure to xenobiotics can be
easily measured spectrophotometrically using 1-chloro-2,4-dinitrobenzene as an
substrate (Habig et al. 1974).
2.5 Biomarkers of effect
Biomarkers of effect are morphological, physiological or biochemical changes that
can occur as a result of exposure to xenobiotics in body fluids or tissues. These
biomarkers are dose responding and dependent on homeostasis and the bio-
effective or critical dose, which can be accepted by the tissue or cell. In extreme
situation they can be recognized as associated with an established or possible
health impairment or disease (Van der Oost et al. 2003). Biomarkers of effect
include measurable biochemical, physiological or other alterations within tissues
or body fluids of an organism that can be recognized as associated with an
CHAPTER 2
38
established or possible health impairment or disease.
Biomarkers of effect include wide range of biomarkers; biomarkers of endocrine
disruption, biomarkers of genotoxicity, oxidative stress, histopathological markers
or for example biomarkers of energetic and metabolic balance expressed by the
content of macromolecules.
A number of xenobiotics with widespread distribution in the environment are
reported to have endocrine activity that might affect reproduction. Synthesis of
yolk proteins precursor vitellogenin (VTG) is controlled by estradiol and can be
affected also by pollutants with affinity to estrogenic receptor. Plasma VTG level
can be thus used as biomarker. Zona radiata protein (vitelline envelope protein)
synthesis in males, decreased sperm count or decreased gonadosomatic index can
be used as other biomarkers of endocrine disruption (Van der Oost et al. 2003).
Biomarkers of genotoxicity as results of pollutant-induced changes in genetic
material belong to important group of biomarkers of effect. DNA adducts can be
formed by CYP1A bioactivation after exposure mostly to PAHs and DNA adducts
are thus considered as biomarkers of PAH exposure. Damage to DNA caused by
genotoxic compounds may result in DNA strand breaks that can be used as other
genotoxic biomarker (Van der Oost et al. 2003).
Biomarkers of effect relevant to oxidative stress and detoxification of xenobiotics
include various measurable parameters of oxidative stress, DNA damage (see the
chapter 2.3) or antioxidants.
Apart from the direct measurement of oxygen radicals production also the
activities of antioxidative enzymes and non-enzymatic antioxidants can be used as
biomarkers of oxidative stress. Increases in their enzymatic activities or levels
occur as a result of exposure to xenobiotics (Roy et al. 1995). Similarly, products
of oxidative damage to macromolecules (more details in Chapter 2.3) as for
example lipid peroxides, DNA adducts and breaks or products of protein
peroxidation can be observed as biomarkers of oxidative stress.
Non-enzymatic antioxidants present the first group of compounds that can be
studied as biomarkers of oxidative stress. Many important compounds including
hydrophilic and lipophilic radical scavengers or for example thioredoxins and
BIOMARKERS AND OXIDATIVE STRESS
39
peroxiredoxins belong to this group of antioxidants. Hydrophilic radical
scavengers such as ascorbate, urate and glutathione play an important role due to
their thiol-disulfide exchange reactions. Tocopherols, flavonoids, carotenoids and
ubiquinol are lipophilic radical scavengers (Scandalios 1997). Polypeptides
thioredoxins undergo redox reactions with multiple proteins and play a key role as
reducing agents for enzymes repairing oxidative damage in proteins.
Peroxiredoxins are a family of peroxidases that reduce H2O2 and organic peroxides
(Halliwell and Gutterdige 2007).
Glutathione (GSH), thiol-containing tripeptide (glutamic acid-cysteine-glycine) is
present in mM intracellular concentrations in almost all animals and plants and can
be considered as the most important non-enzymatic antioxidative compound. GSH
is synthesized in the cell cytoplasm, the liver being the most active organ.
Glutathione can scavenge the reactive species (OH., , HOCl, RO., , RO2. etc.),
carbon-centered radicals or 1O2.
Because high levels of thiol-tripeptides are present, GSSG/GSH (oxidized to
reduced glutathione) couple is a major contributor to the redox state of the cell.
Moreover, GSH is involved in many other metabolic processes, including
ascorbate metabolism, maintaining cell gap-junctional communication and
generally preventing protein-SH groups from oxidizing and cross-linking
(Halliwell and Gutterdige 2007). GSH can regulate gene expression depending on
the environmental stress or pathogenic attack (Dron et al. 1988). GSH acts as a
disulphidic reductant detoxifying many xenobiotics by the conjugation –
spontaneously or in cooperation with GST.
RX + GSH → RSG + HX
This reaction is mostly mediated by selenium-containing enzyme glutathione
peroxidase and more polar and less toxic product is formed (Pflugmacher et al.
1998). It is, however, associated with a dose- and exposure time-dependent
depletion in the glutathione pool. The glutathione pathway, the conversion of
oxidized glutathione to the reduced one is catalysed by the NADPH-dependent
CHAPTER 2
40
flavoenzyme glutathione reductase. Glutathione reductase and other enzymes
involved in regeneration of oxidized antioxidants such as dehydroascorbate
reductase or glucose-6-phosphate dehydrogenase can be also used as biomarkers
of effect (Scandalios 1997).
Glutathione reductase is an enzyme catalyzing the conversion of glutathione
disulphide GSSG back to GSH.
GSSG + NADPH + H+ → 2GSH + NADP+
The main source of NADPH for this reaction is provided by several mechanisms, a
major one being the pentose phosphate pathway (Halliwell and Gutterdige 2007).
NADPH oxidation of GSSG can be used as one method for spectrophotometric
measurement of glutathione reductase activity (Carlberg and Mannervik 1975).
Concentration of GSH can be determined using many approaches. The
spectrophotometric method using 5,5′-dithiobis-2-nitrobenzoic acid (DTNB) as a
substrate (Ellmann 1959) was applied in experiments within this dissertation.
Glutathione peroxidase enzymes (GPx) are widely distributed in animal tissues
and are mostly specific for GSH as hydrogen donors. They remove H2O2 by
coupling its reduction to H2O with oxidation of reduced glutathione.
H2O2 + 2GSH → GSSG + 2H2O
GPx can also act on peroxides other than hydrogen peroxide.
LOOH + 2GSH → GSSG + H2O + LOH
They catalyze GSH-dependent reduction of fatty acid hydroperoxides and various
synthetic hydroperoxides such as τ – butylhydroperoxide that can be used as
substrate in spectrophotometric GPx assay (Flohé and Gunzler 1984).
Hydrogen peroxides that are generated by dismutation of O2.- in oxidase reactions
in the first step of biotransformation can be removed by activity of glutathione
peroxidases and also catalases.
BIOMARKERS AND OXIDATIVE STRESS
41
Catalases (CAT) are enzymes that catalyze the direct decomposition of H2O2 to
ground state O2.
2H2O2 → 2H2O + O2
In animals catalase is present in all types of organs, but especially concentrated in
liver (Halliwell and Gutterdige 2007). The initial rate of hydrogen peroxide
removal catalyzed by catalase is proportional to the hydrogen peroxide
concentration, which can be used in the catalase assay (Aebi 1984).
Superoxide dismutase (SOD) catalyzes dismutation reaction where one H2O2 is
reduced to H2O and the other oxidized to O2. There are at least four types of SOD –
CuZnSODs, MnSOD, FeSOD and hybrid SOD. CuZnSODs are present in almost
all eukaryotic cells. In animal cells, most CuZnSOD is located in cytosol.
MnSODs are widespread in all organisms and in most animal tissues they are
located in mitochondria. FeSODs can be found in plants, bacteria and algae. Metal
ions are on active sites of enzyme and help to undergo the dismutation reaction or
stabilize the enzyme (Halliwell and Gutterdige 2007).
Activity of SOD can be measured for example by a spectrophotometric assay
using nitrobluetetrazolium (NBT) reduction by O2.- to a deep-blue-coloured
formazan (Ewing and Janero 1995).
Selected biomarkers of exposure (cytochrome P450 and GST activity) and effect
(enzymatic activities of GPx, GR, CAT and SOD together with level of GSH and
lipid peroxides) were chosen for the purposes of this dissertation thesis.
Biotransformation and detoxification reactions belong to the general cellular
mechanisms protecting against possible toxic effects of xenobiotics, which are
linked with reactive oxygen species imbalances and sequent damage to
macromolecules and tissues.
CHAPTER 2
42
3 CHAPTER 3
PAPER No.1
DETOXIFICATION AND OXIDATIVE STRESS
RESPONSES IN PLANTS AFTER EXPOSURE TO
POLYCYCLIC AROMATIC COMPOUNDS AND THEIR
N-HETEROCYCLIC DERIVATES
Published as:
Toxic effects and oxidative stress in higher plants exposed to polycyclic aromatic
hydrocarbons and their N-heterocyclic derivatives (2006)
Veronika Pašková, Klára Hilscherová, Marie Feldmannová and Luděk Bláha
Environmental Toxicology and Chemistry, Vol. 25, No. 12, pp. 3238-3245
CHAPTER 3
44
3.1 Hypotheses of the study
Plants have versatile detoxification systems to counter the phytotoxicity of the
wide variety of xenobiotics present in the environment. This is very often utilized
in the bioremediation of xenobiotic-contaminated areas (Gerhardt et al. 2009;
Abhilash et al. 2009). Polycyclic aromatic compounds are a class of ubiquitous
environmental organic compounds chosen as a model group of important
contaminants in this study. Some of them are of great concern namely due to their
mutagenic and carcinogenic properties (IARC 1983). Moreover, environmental
and toxicological importance of their N-derivatives – azaarenes (NPAHs) has been
also recognized (Santodonato and Howard 1981). They have received attention
because of their higher polarity and water solubility when compared to PAHs and
their carcinogenic potential (Sims and Loughlin 1989). It was also suggested that
polyaromatic compounds cause significant phytotoxicity (Henner et al. 1999).
Toxicity of PAHs was observed in multiple plant species; the documented effects
include inhibition of germination and growth as well as of physiological processes
such as photosynthesis or mineral uptake (Kummerova et al. 2001; Marwood et al.
2001; Sverdrup et al. 2003; Alkio et al. 2005). Moreover, it has been shown that
organic pollutants including PAHs accumulate in vegetation (Simonich and Hites
1994). A correlation of plant PAHs accumulation and soil PAHs concentration as
well as the occurrence of elevated PAH levels in vegetables grown in PAH-
contaminated soil was found (Samsoe-Petersen et al. 2002). Vegetation plays an
important role in the global cycling of polyaromatic compounds (Collins et al.
2006), but heretofore the various processes of accumulation, migration, and
transformation of PAHs within plants have not been well understood. On the other
hand, it is known that the chemical modification of xenobiotics by covalent
linkage to the endogenous tripeptide, glutathione, belongs to the important plant
detoxification mechanisms (Coleman et al. 1997). However, there is limited
information available on the activity of detoxification enzymes and antioxidative
molecules playing role in terrestrial plants and transgenic plants after PAH and
NPAH exposure (Roy et al. 1994; Sverdrup et al. 2003; Muratova et al. 2009;
Abhilash et al. 2009; Dixit et al. 2011).
PAPER No.1 - DETOXIFICATION IN PLANTS AFTER PAHs EXPOSURE
45
N
N
phenanthrene benzo[h]quinoline phenanthridine
N
N
N
N
N
N
1,10-phenanthroline 1,7-phenanthroline
4,7-phenanthroline
N
anthracene acridine
N
fluorene carbazole
Figure 3.: Chemical structure of tested compounds.
This study was focused on investigation of detoxification and oxidative stress
responses after exposure of three different plant species to polycyclic aromatic
compounds and their N-heterocyclic derivatives (see chemical structure of tested
chemical in Fig.3) during the 4-day germination. Several biochemical responses to
acute PAHs (phenanthrene, anthracene and fluorene) and NPAHs (phenanthridine,
1,10-phenanthroline, 4,7-phenanthroline, 1,7-phenanthroline, benzo[h]quinoline,
acridine and carbazole) exposure were determined. Measurements included
enzymatic activities of glutathione-S-transferase, glutathione peroxidase and
glutathione reductase and the levels of glutathione and lipid peroxidation. Slightly
modified OECD Guideline 208 (OECD/OCDE 2006) has been applied and
standard test parameters such as plant germination and hypocotyl and root
CHAPTER 3
46
elongation have been also evaluated. They were further compared with the
measured biomarkers of oxidative stress and detoxification. Selected test species
represent different plant classes and also groups with different carbon metabolism
including both dicotyledonous Phaseolus vulgaris and Sinapis alba and
monocotyledonous plant Triticum aestivum.
3.2 Results and discussion
This study brought new information on the phytotoxicity and biochemical effects
of important organic contaminants PAHs and relatively poorly characterized group
of their N-heterocyclic derivatives. The concentrations used in this study (0.02 –
200 µM ≈ 3.3 µg L-1 – 36 mg L-1) were within or close to the environmentally
relevant range, when compared to the concentrations reported by the U.S. EPA
public health assessment program study (ATSDR-PHA 1984) and the effects
observed at lower doses could be of general concern.
The acute phytotoxic parameters (germinability, weight and length of roots and
hypocotyle) used for testing responded differently after exposure to parental PAHs
and their heterocyclic derivatives and generally, NPAHs were significantly more
phytotoxic than parent PAHs. Interestingly, 1,7-phenanthroline was the most toxic
among all tested compounds, it affected most parameters already at the
concentration 0.02 µM, while the changes of all measured parameters were
observed at 2 µM.
Correspondingly to our results, several studies reported phytotoxicity of PAHs or
NPAHs to various plant species (Sverdrup et al. 2003; Alkio et al. 2005; Gissel-
Nielsen and Nielsen 1996; Van Vlaardingen et al. 1996). Various phytotoxic
effects included inhibition of growth and root development and induction of leaf
lesions in Arabidopsis thaliana exposed to phenanthrene (Alkio et al. 2005).
Similarly, negative effects of acridine on Brassica campestris, Lolium multiflorum
and Hordeum vulgare seedlings germination and growth were reported (Gissel-
Nielsen and Nielsen 1996). Acridine was also the most toxic of NPAHs tested in
study with alga Scenedesmus acuminatus (Van Vlaardingen et al. 1996). On the
other hand, study with Sinapis alba, Trifolium pratense and Lolium perenne
PAPER No.1 - DETOXIFICATION IN PLANTS AFTER PAHs EXPOSURE
47
(Sverdrup et al. 2003) reported only minor differences between the toxicity of
homocyclic and heterocyclic PAHs (fluoranthene, pyrene, phenanthrene, fluorene,
carbazole, dibenzothiophene, acridine). The differences in results of various
experimental studies and our study might be explained either by experimental
variability but more likely by different sensitivities of plant species as
demonstrated above.
Ger
min
abili
ty(%
) b
enzo
[h]q
uino
line
20
40
60
80
100
120
acr
idin
e
20
40
60
80
100
120
phe
nant
hrid
ine
20
40
60
80
100
120
∗∗∗∗ ∗∗∗∗∗∗∗∗∗∗∗∗
∗∗∗∗
∗∗∗∗∗∗∗∗∗∗∗∗
∗∗∗∗
∗∗∗∗∗∗∗∗
Figure 4.: Germination of Triticum aestivum after 96 h exposure to selected N-heterocyclic polyaromatic hydrocarbons (benzo[h]quinoline, acridine, phenanthridine). Box includes 50% values, middle point is median and whiskers show extremes. Asterisks indicate the statistically significant difference from control [∗∗∗∗ = p<0.05; ∗∗∗∗∗∗∗∗ = p<0.01]
0 0.02 0.2 2 20 200 Concentration [µM]
CHAPTER 3
48
In contrast to apparently higher acute phytotoxicity of NPAHs, the effects of both
PAHs and NPAHs on biochemical parameters were comparable. All tested
chemicals modulated activity of plant detoxification and antioxidative enzymes
(Table 1). Several PAHs and NPAHs induced lipid peroxidation and also increased
activities of GST, GR and GPx and modulated concentrations of GSH, and the
effects were often observed even at low 0.02 µM concentrations. The most
pronounced modulations were in general observed after exposure to
phenanthridine, benzo[h]quinoline (Figure 4), and 1,7-phenanthroline. S. alba and
T. aestivum were more sensitive plant species than P. vulgaris. The most sensitive
biomarker among those analyzed was the activity of glutathione reductase,
NADPH-dependent flavoenzyme maintaining GSH/GSSG homeostasis and
playing thus very important role in the xenobiotics-detoxification pathway in
plants.
A limited number of studies also documented sensitive biochemical responses in
plants exposed to various PAHs. For example, modulations of GSH and increased
activities of GR, GST, SOD and ascorbate peroxidase were reported in aquatic
plant Fontinalis antipyretica exposed to prototypical PAHs benzo[a]pyrene and
benzo[a]anthracene (Roy et al. 1994). The correlations between elevated
antioxidative enzyme activities in this species and accumulated PAHs were also
observed in the field (Roy et al. 1996). Study of Alkio et al. (2005) with
Arabidopsis exposed to relatively high concentrations of phenanthrene (≥ 50 µM)
showed induced H2O2 production and modulated GR and ascorbate peroxidase.
Similarly, GR activity was increased dramatically in Lemna gibba exposed to the
mixture of copper and oxo-PAH dihydroxyanthraquinone (Babu et al. 2005).
To conclude this study, biomarkers do not only reflect toxic mechanisms and
major processes protecting plant tissues from oxidative stress but they can also be
successfully used as early warnings of in vivo phytotoxic effects. Biochemical
changes were in general more sensitive and occurred mostly at concentrations
about an order of magnitude lower than those causing signs of toxicity.
PAPER No.1 - DETOXIFICATION IN PLANTS AFTER PAHs EXPOSURE
49
Table 1.: Summary of the effects of N-heterocyclic PAHs and their unsubstituted analogues on
biochemical parameters in plants (- no effect, + statistically significant effect at > 2 µM, ++ effect
at 0.2-2 µM, +++ effect at 0.02 µM); p < 0.05). Biomarkers: thiobarbituric acid reactive substance
(TBARS), glutathione (GSH), glutathione-S-transferase (GST), glutathione peroxidase (GPx) and
glutathione reductase (GR).
toxicant species Biochemical parameter TBARS GSH GST GPx GR
T. aestivum - - - - +++ S. alba - - +++ ++ +++
1,10-phenanthroline
P. vulgaris - - + + +++ T. aestivum - - - ++ ++ S. alba - - - - +++
4,7-phenanthroline
P. vulgaris - - + ++ +++ T. aestivum ++ - ++ ++ ++ S. alba - ++ ++ ++ +++
1,7-phenanthroline
P. vulgaris ++ ++ + + +++ T. aestivum - ++ ++ ++ ++ S. alba - ++ ++ ++ ++
benzo[h]quinoline
P. vulgaris ++ - +++ +++ ++ T. aestivum - - - - +++ S. alba - ++ + ++ ++
phenanthrene
P. vulgaris ++ ++ ++ ++ +++ T. aestivum +++ + ++ +++ +++ S. alba +++ +++ ++ +++ +++
phenanthridine
P. vulgaris - - + + + T. aestivum + - + ++ ++ S. alba - ++ ++ - -
acridine
P. vulgaris - + ++ - - T. aestivum - ++ - - ++ S. alba - + - - -
anthracene
P. vulgaris - +++ ++ + - T. aestivum +++ - +++ +++ ++ S. alba +++ - - - +++
fluorene
P. vulgaris +++ ++ +++ - - T. aestivum ++ + - - - S. alba ++ - +++ ++ +
carbazol
P. vulgaris +++ ++ - ++ -
CHAPTER 3
4 CHAPTER 4
PAPER No.2
FISH EXPOSURE TO CYANOBACTERIAL BIOMASS - DETOXIFICATION AND ANTIOXIDATIVE
RESPONSES
Published as:
Microcystin kinetics (bioaccumulation and elimination) and biochemical responses
in common carp (Cyprinus carpio) and silver carp (Hypophthalmichthys molitrix)
exposed to toxic cyanobacterial blooms (2007)
Ondřej Adamovský, Radovan Kopp, Klára Hilscherová, Pavel Babica, Miroslava
Palíková, Veronika Pašková, Stanislav Navrátil, Blahoslav Maršálek and Luděk
Bláha
Environmental Toxicology and Chemistry, Vol. 26, No. 12, pp. 2687-2693
CHAPTER 4
52
4.1 Hypotheses of the study
Environmental conditions such as higher temperature and pH values, low
turbulence, and high nutrient inputs, above all phosphorus and nitrogen, enhance
the development of planktonic cyanobacteria in lakes and reservoirs, leading to
worldwide formation of cyanobacterial water-blooms (de Figueiredo et al. 2004).
However, the problem of eutrophication in freshwater bodies is caused mainly by
the intensification of agricultural and industrial activities associated with the
growth of human population. Cyanobacteria are known to produce secondary
metabolites, which have been recognized as human and animal health hazards,
since they have been shown to cause adverse effects in various organisms
including fish (Malbrouck et al. 2003). The toxins are synthesized during the
growth phase of cyanobacteria and the greatest amounts of best-known
cyanotoxins microcystins (MCs) are released after cell lysis or from actively
expanding cyanobacterial populations into the water (Pearson et al. 2004).
Cyanotoxins are very diverse in their chemical structure and toxicity, usually being
classified as dermatotoxins (lipopolysaccharides, lyngbyatoxin-a, and
aplysiatoxins), neurotoxins (anatoxin-a, homoanatoxin-a, anatoxin-a(s), and
saxitoxins), and hepatotoxins (nodularin, cylindrospermopsin and microcystins)
(Codd et al. 2005), according to their toxic effects on animals. The most frequently
occurring cyanobacterial toxins are microcystins, monocyclic heptapeptides
composed of D-alanine at position 1, two changeable L-amino acids at positions
2 and 4, γ-linked D-glutamic acid at position 6, D-methylaspartic acid (D-MeAsp)
at position 3, (2S, 3S, 8S, 9S)-3-amino-9-methoxy-2, 6, 8-trimethyl-10-
phenyldeca-4, 6-dienoic acid (ADDA) at position 5 and N-methyl dehydroalanine
(MDha) at position 7. There are over 70 MCs variants differing mainly in
demethylation of D-MeAsp and/or MDha and in two mentioned L-amino acids at
position 2 and 4, respectively (Dawson 1998). The most extensively studied and
the most common MCs are MC-LR (2-Leu, 4-Arg), MC-RR (2-Arg, 4-Arg) and
MC-YR (2-Tyr, 4-Arg). The substitution of hydrophobic L-Leu with another
hydrophobic L-amino acid (e.g., tryptophan, alanine or phenylalanine) does not
change its toxicity, but replacement with a hydrophilic amino acid (e.g., arginine)
degrades toxicity. The least toxic MCs like MC-RR or MC-M(O)R contain polar
PAPER No.2 - DETOXIFICATION AFTER CYANOBACTERIAL EXPOSURE OF
FISH
53
substitutions in both variable amino acid positions (Zurawell et al. 2005).
Microcystins production by cyanobacteria is a serious public health issue
(Carmichael 1997), because of its ability to cause acute poisonings and to promote
cancer in humans by chronic exposure to low concentrations in drinking water.
Another discussed route of human exposure might come from the nutrition,
especially in countries with high fish and seafood consumption and occurrence of
water-blooms in fish-reservoirs (Nyakairu et al. 2010). This was the reason for the
selection of two widespread edible fish species – common and silver carp - for this
experiment aiming partly at investigation of accumulation and elimination of
microcystins in various fish tissues and at detoxification and oxidative stress
responses after exposure to MC-containing cyanobacterial bloom. There are many
studies reporting the toxic effects of cyanotoxins on various fish species, mainly
focusing on histopathological, immunological and behavioural changes (Cazenave
et al. 2006). There are also studies investigating the accumulation and
biotransformation of cyanobacterial toxins in fish, however the route of exposure
in these studies is mostly differing from the natural cyanotoxins intake
(intraperitoneal (Prieto et al. 2006), single acute dosing (Cazenave et al. 2006)) or
embryolarval exposure (Wiegand and Pflugmacher 2001; Best et al. 2002).
Previous studies have demonstrated the conversion of microcystin in animal liver
to more polar compound in correlation with a depletion of glutathione pool of the
cell (Kondo et al. 1996). The existence of MC-LR glutathione conjugate, the first
step in the detoxification of microcystins, formed enzymatically via soluble
glutathione-S-transferase (GST) was showed in various aquatic organisms
(Pflugmacher and Wiegand 2001) indicating the involvement of glutathione-
related compounds in detoxification of microcystins.
The task of this study was to evaluate the role of several detoxification and
antioxidative compounds (glutathione, glutathione-S-transferase, glutathione
peroxidase and glutathione reductase) in biotransformation of cyanobacterial
metabolites after four and nine-week exposure of two fish species to natural
cyanobacterial water-bloom in outdoor fish-pond. Aim of this work was also to
compare biochemical responses of chosen fish models (benthophagous common
carp and phytophagous silver carp).
CHAPTER 4
54
4.2 Results and discussion
Set of glutathione-related biomarkers has been investigated in hepatopancreas of
common and silver carp after four and nine weeks of cyanobacterial exposure. The
exposure has simulated the natural situation in the environment with water
concentration of total MCs reaching 13.8 to 22.7 µg/L without external feeding.
This study showed correspondence in response of tripeptide glutathione and
enzymatic activity of glutathione-S-transferase, which demonstrates their
cooperation in microcystins conjugation (Wiegand and Pflugmacher 2001;
Pflugmacher et al. 1998). Activity of glutathione reductase was the most sensitive
biomarker; it was significantly elevated in most experimental variants, especially
in common carp. On the other hand, changes in glutathione peroxidase activity
were less sensitive in this experiment. Elevated glutathione concentrations and
activities of the flavoenzyme glutathione reductase, which plays crucial role in the
GSH/GSSG homeostasis (Van der Oost et al. 2003), further reveal increased
demands for reduced GSH because of enhanced detoxification and/or oxidative
stress induced by toxic cyanobacteria (Li et al. 2003; Jos et al. 2005).
Benthophagous common carp and phytophagous silver carp respond differently to
the cyanobacterial exposure. There would be longer exposure needed to investigate
this phenomenon in more detail but it seems that the phytophagous fish is better
adapted for the active ingestion of cyanobacterial cells and its detoxification is
quick and less energy demanding.
This study, however, demonstrates that biochemical adaptations can be only
temporary and that prolonged exposures may result in signs of general toxicity
(when comparing the four- and nine-week exposures in case of silver carp as
shown in Fig. 5).
PAPER No.2 - DETOXIFICATION AFTER CYANOBACTERIAL EXPOSURE OF
FISH
55
G
ST
G
Px
GR
G
SH
C. EXP. C. EXP. C. EXP. C. EXP. 4 weeks 9 weeks 4 weeks 9 weeks common carp silver carp
.
.
30 .
.
20.
a
aa
10aaa
0.
240
200a
160
aa
120
a
80
3
2a
1
a
0
a
12a
8
4a
0
**
*
**
Figure 5.: Responses of detoxification and antioxidative compounds in fish hepatopancreas after
four and nine weeks of exposure to cyanobacterial biomass (box includes the 25th to 75th
percentiles, with the middle point representing the median and the whiskers showing the extremes.
An asterisk indicates a statistically significant difference from control (p < 0.05, Student’s t test).
Biomarkers: glutathione (GSH), glutathione-S-transferase (GST), glutathione peroxidase (GPx)
and glutathione reductase (GR).
CHAPTER 4
56
Apparent time-, species-, and MC variant–dependent variability exists in
biochemical responses of organisms to MCs (Prieto et al. 2006). Inductions of
GST are among the most often reported responses (Wiegand et al. 1999; Pietsch et
al. 2001) (present study), but other authors have reported rapid, 24-h inhibitions of
GST in Corydoras paleatus exposed to purified MC-RR (Cazenave et al. 2006).
Modulations of biomarkers in the present study confirm an important role for
oxidative stress in the toxicity of complex cyanobacterial bloom, and it also
demonstrates that biochemical parameters (especially GR, GST, and GSH) elicit
sensitive reponses. Further research would be needed to characterize both natural
variability and temporal changes in responses to toxicants.
Modulations of biomarkers in the present study confirm an important role of
oxidative stress in the toxicity of complex cyanobacterial bloom, and it also
demonstrates that biochemical parameters (especially GR, GST, and GSH) may
serve as sensitive early markers of oxidative stress in fish.
5 CHAPTER 5
PAPER No.3
DETOXIFICATION AND OXIDATIVE STRESS
RESPONSES IN JAPANESE QUAIL EXPOSED TO
CYANOBACTERIAL BIOMASS
Published as:
Detoxification and oxidative stress responses along with microcystins
accumulation in Japanese quail exposed to cyanobacterial biomass (2008)
Veronika Pašková, Ondřej Adamovský, Jiří Pikula, Blanka Skočovská, Hana
Banďouchová, Jana Horáková, Pavel Babica, Blahoslav Maršálek and Klára
Hilscherová
Science of the Total Environment, Vol. 398, Is. 1-3, pp. 34-47
CHAPTER 5
58
5.1 Hypotheses of the study
The role of cyanobacterial toxins in poisonings of wild life and especially their
possible connection with the mass mortalities of wild birds have been investigated
over recent years. Tens of thousands Lesser Flamingos in Kenya and Tanzania
dyed most probably after exposure to hot spring cyanobacterial hepato- and
neurotoxins from drinking water, which were observed in stomach contents and
faecal pellets of flamingos (Krienitz et al. 2003; Lugomela et al. 2006). Greater
Flamingo chick deaths, also attributed to microcystins, occurred at wetlands
lagoon in Spain after the sudden development of bloom with predominant content
of Microcystis aeruginosa and Anabaena flos-aquae (Alonso-Andicoberry et al.
2002). Another report of unnatural bird death (Matsunaga et al. 1999) is being
connected with occurrence of toxic freshwater cyanobacterial bloom of
Microcystis aeruginosa in eutrophicated lake in Japan, following the untreated
sewage. The increased concentration of phosphorus in lake water joint with the
enhanced growth of Microcystis and Aphanizomenon was also observed in Canada
(Murphy et al. 2000), where anoxic conditions caused the development of water
bloom with indirect linkage between sediment and Clostridium botulinum resulting
in avian botulism. Cyanobacterial blooms of Anabaena lemmermannii were
implicated in bird kills at lakes in Denmark, where the content of neurotoxin with
anticholinesterase activity was shown (Onodera et al. 1997).
Next to the above mentioned and other few reported examples of toxic effects of
complex cyanobacterial exposure in birds, cyanotoxins microcystins are known to
affect microalgae, zooplankton, aquatic and terrestrial plants, terrestrial insects,
fish and mammals (de Figueiredo et al. 2004). The primary mechanism of
microcystins toxicity is probably the inhibition of the eukaryote serine/threonine
protein phosphatases 1 and 2A, which leads to the hyperphosphorylation of the
major cytoskeletal intermediate filament proteins keratin 8 and 18, resulting in the
destruction of cytoskeleton directly causing the cytotoxic effects, bleeding and
disappearance of hepatocytes in liver, cytolysis or apoptosis of hepatocytes, but
also of glomeruli and renal proximal tubule cells (Gehringer 2004). Moreover,
PAPER No.3 – DETOXIFICATION IN BIRDS AFTER CYANOBACTERIAL
EXPOSURE
59
cloudy swelling of hepatocytes, vacuolar dystrophy, steatosis and hyperplasia of
lymphatic centres were documented in laboratory experiments with Japanese
quails performed by Skočovská et al. (2006). These histopathological observations
were supported by the changes on the subcellular level where the shrunken nuclei
of hepatocytes containing ring-like nucleoli, cristolysis within mitochondria and
vacuoles with pseudomyelin structures were shown. MCs-induced DNA
fragmentation and degradation along with the deregulation of cell division, leading
to the tumor-promoting activity has also been observed (Carmichael 1997).
It has been documented that target organs of microcystins are particularly liver and
brain (Fischer et al. 2005), which requires the uptake of microcystins across the
sinusoidal plasma membrane of hepatocytes and its transport crosswise the blood-
brain barrier. This process is followed through the substrates of organic anion
transporting polypeptide, superfamily of membrane transporters, which are
expressed in brain as well as in liver (Kullak-Ublick et al. 1995). Various organic
anion transport proteins are present also in gastrointestinal tract or kidney
(Hagenbuch and Meier 2003). Interestingly, there is a growing body of evidence
for toxic effects of microcystins in mammalian reproductive system and testes
seem to be another target organ for these biotoxins (Ding et al. 2006; Li et al.
2008). Laboratory studies with bird males correspondingly documented vacuolar
degeneration of the testicular germinative epithelium (Skocovska et al. 2007),
moderate to marked atrophy of the seminiferous tubular epithelium and only
sparse developmental stages of spermatozoa and Sertolli cells (Damkova et al.
2011). On the other hand, the lower weight of eggs produced by exposed parental
hens was not reflected in their biological quality and surprisingly, reproductive
parameters in cyanobacterial-biomass-exposed birds were better than in the control
group (Damkova et al. 2009). The ability of microcystins (or structurally related
nodularins) to bioaccumulate was reported not only in liver, but also in intestines,
kidneys, brain, heart, gonads and muscles of fish and mammals (Kankaanpaa et al.
2005; Cazenave et al. 2006; Adamovsky et al. 2007; Kagalou et al. 2008). The
above presented toxicity of microcystins to various bird tissues together with their
CHAPTER 5
60
bioaccumulation potential support the choice of tested organs (liver, heart, brain,
gonads) in studies, which were included into this dissertation thesis.
Exposure to cyanobacterial biomass and/or purified microcystins has been shown
to cause oxidative stress in various organisms (Ding et al. 2000; Pietsch et al.
2001; Li et al. 2003; Wiegand and Pflugmacher 2005), but there is only little data
on oxidative responses in adult warm-blooded vertebrates after cyanobacterial
exposure (Gehringer et al. 2004; Moreno et al. 2005). To my best knowledge, there
was no accessible information about the role of oxidative stress responses in
cyanobacteria-biomass-exposed birds except of studies, which have been
performed within this dissertation thesis. However, modification of blood
biochemical parameters such as increased lactate dehydrogenase activity and drop
in glucose level were reported in birds exposed to natural cyanobacterial biomass
(Skocovska et al. 2007; Damkova et al. 2009). Other study with Japanese quail
chicks exposed to cyanobacterial biomass resulted in hypoproteinaemia, increased
concentrations of triglycerides, uric acid and the total antioxidant capacity and a
drop in high-density lipoprotein cholesterol in blood (Peckova et al. 2009).
The aim of this study was to assess the effect of cyanobacterial exposure on
standard bird model species Japanese quail (Coturnix coturnix japonica). The
study focused on activation (P450-dependent 7-ethoxyresorufin-O-deethylase
activity) and conjugation (glutathione-S-transferase, glutathione) phase of
detoxication metabolism, lipid peroxidation and further antioxidant activities
(glutathione peroxidase, glutathione reductase). Part of this experiment was
concerned also with the accumulation of microcystins in bird muscle and liver.
5.2 Results and discussion
Acute 10-day and sub-chronic 30-day studies with controlled doses of natural
cyanobacterial bloom on 4-month old Japanese quails were performed according
to OECD Guideline for the testing of chemicals 205 – Avian Dietary Toxicity Test
(OECD 1984) and the responses of detoxification and antioxidative parameters
PAPER No.3 – DETOXIFICATION IN BIRDS AFTER CYANOBACTERIAL
EXPOSURE
61
together with lipid peroxidation (summarized in Table 2) as a measure of damage
to macromolecules were evaluated in liver, heart and brain of each individual.
Four treatment groups of quails (E1-E4) were fed three times a day to reach 10 mL
of a natural cyanobacterial biomass equivalent in amount of 0.123 to 123 mg dry
biomass using crop probe. The birds also received standard bird food and drinking
water ad libitum during the study. The biomass originated from the Brno reservoir
and was collected using plankton net at the end of cyanobacterial vegetation
season in year 2004 to reach the maximal cyanotoxin concentration in this
biomass.
Table 2.: Summary of the effects of cyanobacterial biomass on the bird antioxidative and
detoxification system (statistically significant increase ↑ and decrease ↓; P < 0.05). Biomarkers:
glutathione (GSH), glutathione-S-transferase (GST), glutathione peroxidase (GPx), glutathione
reductase (GR), thiobarbituric acid reactive substance (TBARS) and 7-ethoxyresorufin-O-
deethylase (EROD).
GSH GST GPX GR TBARS EROD Acute test liver ↑ - ↑ ↑ ↑ -
heart ↑ - - ↑ ↑ ↑
brain ↓ - - - ↑ ↑
Sub-chronic test liver ↑ ↑ ↓ - - -
heart - ↑ - - ↑ -
brain ↑ - ↑ - ↑ ↑
There was no mortality neither in the acute nor the sub-chronic study. After the
end of exposure the birds were euthanized and tissues chosen for further
biochemical analyses were stored at –80º C. This study brought first information
about the effect of controlled doses of natural cyanobacterial biomass on birds
related to detoxification and oxidative stress responses. Activities of cytochrome
P-450 dependent 7-ethoxyresorufin-O-deethylase (EROD) were increased namely
in the acute 10-day test in heart and brain tissue. In the sub-chronic experiment,
CHAPTER 5
62
there was significant increase of this enzyme activity only in brain tissue. This
sensitive biomarker belongs to the large family of P-450 biotransformation
enzymes, which play a crucial role in the first step of xenobiotics detoxification
and can reflect exposure to various contaminants in birds (Walker and Ronis
1989).
The activity of glutathione-S-transferase, enzyme catalysing the conjugation of
some microcystins with glutathione (Pflugmacher and Wiegand 2001) showed
more distinct changes in the sub-chronic exposure in comparison with acute test.
Activities of this enzyme increased in all studied organs, but namely in liver, only
in the sub-chronic exposure, while liver glutathione level was increased in both
acute and sub-chronic exposure. Increases in glutathione level were detected in all
tissues with the exception of brain of birds from acute test, where significant
decline was ascertained. Moreover, in correspondence with the enhanced activity
of glutathione-S-transferase, there was an increased level of glutathione in all
tested organs in the sub-chronic test confirming the importance of these two
biomolecules in protection against the harmful effects of MCs-containing
cyanobacterial biomass.
In respect of glutathione peroxidase activity, different response was detected in
acute and sub-chronic test. There was an increased activity observed in liver from
acute test and brain from sub-chronic test, while there was a significant decrease of
GPx activity detected in liver tissue from sub-chronic test. Similar results were
obtained also in case of glutathione reductase activity with declines of in sub-
chronic test and significant increases in acute test. Different responses of GR and
GPx activities in the acute and sub-chronic test may indicate the potential
adaptation of these glutathione-related biomarkers to the cyanobacterial exposure
with increasing time of exposure.
This study confirmed that lipid peroxidation, measured as TBARS (thiobarbituric
acid reactive species), can be induced by exposure to MCs-containing
cyanobacteria, as reported by Halliwell and Gutterdige (2007). Increases of
PAPER No.3 – DETOXIFICATION IN BIRDS AFTER CYANOBACTERIAL
EXPOSURE
63
TBARS levels were observed in all tested organs mostly at the lowest tested
concentrations.
To conclude, the complexity and interdependence of chosen biomarkers was
showed (documented by correlation analysis presented in the full text of the
article). The liver was confirmed as an important organ for detoxification of
xenobiotics including natural compounds in birds, which has been similarly
reported in Riviere et al. (1985). Moreover, significant responses of glutathione-
related biomarkers and also oxidative stress were presented also in heart and brain,
indicating them as other targets of the toxicity of MCs-containing cyanobacteria.
Birds coming to contact with the eutrophicated ecosystem react to the
cyanobacterial metabolites as to the xenobiotics. Their general mechanism of
detoxification is activated in case of contact with cyanotoxins. As shown in this
study, this antioxidative and detoxification mechanism can get adapted to
cyanobacterial exposure with increasing time of exposure, as it has been
characterized by the increases of glutathione level and activity of glutathione-S-
transferase.
Moreover, the results of microcystins accumulation (presented in the article) and
mainly the comparison of accumulation in the acute and sub-chronic test support
this conclusion. Briefly, significantly higher accumulation in quails exposed for
10 days to the cyanobacterial biomass compared to the sub-chronic exposure may
support the hypotheses of the important role of glutathione-detoxification pathway
of cyanobacterial metabolites. It seems that the longer-term exposure to MCs-
containing cyanobacterial biomass activated the detoxification and elimination
pathway, which has been documented by six times lower accumulation at longer
exposure time.
The exposure of model birds to natural cyanobacterial biomass caused significant
changes in levels and activities of antioxidative and detoxification compounds and
accumulation of cyanotoxins mainly in liver and little accumulation in the
CHAPTER 5
64
muscles. Cyanobacteria are thus capable to induce oxidative stress responses in
birds linked with activation or inhibition of detoxification compounds. The
generation of oxidative stress combined with insufficiency of defense mechanisms
could in sensitive species at prolonged exposure potentially result in effects on the
health status, especially if other stressors are involved at the same time, which is
often the case in the environment.
6 CHAPTER 6
PAPER No.4
DETOXIFICATION AND OXIDATIVE STRESS RESPONSES IN JAPANESE QUAIL EXPOSED TO
MULTIPLE STRESSORS
Published as:
Combined exposure of Japanese quails to cyanotoxins, Newcastle virus and lead:
Oxidative stress responses (2011)
Veronika Pašková, Hana Paskerová, Jiří Pikula, Hana Banďouchová, Jana
Sedláčková and Klára Hilscherová
Ecotoxicology and Environmental Safety, Vol. 74, Is. 7, pp. 2082-2090
CHAPTER 6
66
6.1 Hypotheses of the study
As mentioned in the previous study within this dissertation thesis, there are many
mass mortalities of wild-living organisms, especially birds, worldwide being
reported in connection to the cyanobacterial toxicity (Krienitz et al. 2003;
Lugomela et al. 2006; Alonso-Andicoberry et al. 2002; Matsunaga et al. 1999;
Murphy et al. 2000; Onodera et al. 1997; Park et al. 2001; Wirsing et al. 1998).
But in the environment the birds have to face multiple stressors and can be subject
to mortality because of the effects of natural toxins, pathogens, industrial and
agricultural chemicals such as pesticides and also other various anthropogenic
contaminants (Norris and Evans 2000; Sagerup et al. 2009; Rattner 2009).
Heavy metals constitute group of important widespread contaminants capable due
to their high concentration and bioavailability cause toxic effects to wild-living
organisms (Sanchez-Chardi et al. 2007). Some cases of mortalities of aquatic and
other bird species due to the toxic effects of heavy metals have been reported
(Degernes et al. 2006; O'Connell et al. 2008). Contamination by heavy metals
occurs often in regions with former industrial activities or, interestingly, with
frequent hunting (Guillemain et al. 2007; Blus et al. 1995). It has been reported
that wild birds are exposed to heavy metals by oral ingestion of spent lead shot or
bullet fragments (Fisher et al. 2006), which was the impulse to use lead shots
ingestion as a model example for heavy metal exposure in this multi-stressor
study.
Originally, the wild-living organisms had to fight against various infectious
diseases and up today, the most common causes of mass mortalities of organisms
are bacterial, fungal or viral infections (Waller and Underhill 2007). Newcastle
virus is one of the known viruses affecting aquatic birds (Liu et al. 2008) with an
effective vaccination (Shebannavar et al. 2010), which may be used for immune
response induction such as in this study. The vaccinated organisms become more
susceptible to other stressors when investing energy to avoid pathological side-
effects caused by an elevated immune response (Costantini and Moller 2009).
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67
Hypotheses of this study was that combined exposure to cyanobacterial biomass,
heavy metal and vaccination may enhance the toxic effects in birds and more
relevantly simulate the situation in the environment than in the previous single
exposure to cyanobacterial biomass. Among many negative effects of
cyanobacteria and heavy metals in the environment, there is at least one shared
mechanism of action—their ability to increase the generation of reactive oxygen
species (ROS) (Stohs and Bagchi 1995; Li et al. 2003). Formation of ROS and
oxidative stress are also associated with the development of many pathological
states and damage, including immuno-pathology (Costantini and Moller 2009),
which may result from high dose of ROS released during the immune response.
Aims of this study were to evaluate the oxidative stress responses together with
detoxification and antioxidative compounds activity in quail liver and heart after
single and combined exposures to the chosen stressors. The following set of
biomarkers was measured in this study – the level of glutathione and activities of
glutathione-S-transferase, glutathione peroxidase, glutathione reductase, catalase
and superoxide dismutase together with the parameter of damage – lipid
peroxidation.
6.2 Results and discussion
Sub-chronic 30-day multi-stressor exposure of 4-month old Japanese quail males
to controlled doses of natural cyanobacterial bloom, lead and vaccination strain
was performed according to OECD Guideline for the testing of chemicals 205 –
Avian Dietary Toxicity Test (OECD 1984).The responses of detoxification and
antioxidative parameters together with lipid peroxidation were evaluated in liver
and heart of each individual. The quails were randomly divided into 8 groups –
exposed to cyanobacterial biomass, Newcastle vaccination, Pb and their
combinations as illustrated in Table 3. Briefly, quails from B groups were fed
with natural cyanobacterial biomass (dominated by Microcystis sp.) twice a day
using the crop probe to reach the daily microcystins concentration of 46 µg. The
biomass came from the same reservoir as in our previously mentioned study and
the dosage was equal (in MC content) to the highest concentration applied in that
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68
study. Quails from the V groups were vaccinated with the Newcastle disease
vaccination at the beginning of the exposure to induce antigenic stress and immune
response. Quails from Pb groups were given six lead shots into the crop to induce
lead toxicosis. Commercially available feeds and drinking water were supplied ad
libitum.
Table 3.: Labeling and characterization of experimental groups.
Abbrev. Exposure Dosing
C control 10 mL of control water / day
B cyanobacterial biomass 10 mL of cyanobacterial biomass / day
Pb lead 6 lead shots at the beginning of experiment + 10 mL of control water / day
V vaccination vaccination at the beginning of experiment + 10 mL of control water / day
BPb cyanobacterial biomass + lead 6 lead shots at the beginning of experiment + 10 mL of cyanobacterial biomass / day
BV cyanobacterial biomass + vaccination
vaccination at the beginning of experiment + 10 mL of cyanobacterial biomass / day
PbV lead + vaccination 6 lead shots at the beginning of experiment + vaccination at the beginning of experiment + 10 mL of control water / day
BPbV cyanobacterial biomass + lead + vaccination
6 lead shots + vaccination at the beginning of experiment +10 mL of cyanobacterial biomass / day
In accordance with the previous single-cyanobacterial exposures to quails, there
was no mortality in B group and also in Control, V, BV and BPbV groups. One
bird died in the single lead-exposure (Pb) and the combined exposures to
cyanobacterial biomass and lead (BPb) and lead and Newcastle vaccination (PbV)
resulted in the death of two out of five birds.
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The responses of the non-enzymatic antioxidant glutathione differed in liver and
heart (Fig. 6). The glutathione level was threefold increased in combined
exposures BPb and BPbV in comparison with the other groups in liver. But there
was significant elevation compared to control in heart in most single and combined
exposed groups except V and BV groups.
Figure 6.: Level of glutathione (nmol/mg protein) in liver (A) and heart (B). Box includes 50%
values, middle point is median and whiskers show non-outlier range. Letters indicate the
statistically significant difference from control (C) or other treatment groups (lead Pb,
cyanobacterial biomass B, vaccination V) [LSD test].
The pattern of hepatic glutathione-S-transferase activity was similar to the hepatic
glutathione level being significantly higher in BPbV group than in all the other
groups. The higher detoxification is obvious in PbV and BPbV groups in heart,
where the significantly higher GST activity was documented when compared with
both control and lead-exposed birds.
According to the liver glutathione level, the activities of glutathione peroxidase
were elevated in BPb and BPbV groups. Also in heart, exposure to these combined
groups and also BV group resulted in significant increases of GPx activity in heart.
GS
H n
mol
/ m
g pr
otei
n
CPb
EV
Pb+VE+V
E+PbE+Pb+V
16
20
24
28
32
36
40
44
GS
H n
mol
/ m
g pr
otei
n
CPb
EV
Pb+VE+V
E+PbE+Pb+V
4
8
12
16
20
24
28
GS
H n
mo
l / m
g
pro
tein
C B PbV BPb C B PbV BPb Pb V BV BPbV Pb V BV BPbV
A.
C,Pb,B,V
PbV,BV
C,Pb,B,V PbV,BV
B. C
C
C C C
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70
Interestingly, increased glutathione reductase activities were documented in liver
after exposures to all groups with exception of BPbV group, which resulted in
lower GR activity than single V and B groups. Other results were shown in heart,
where glutathione reductase activity in B, BPbV and PbV groups was elevated
against control and single V and Pb groups (Fig. 7).
Figure 7.: Activity of glutathione reductase (nmol NADPH/min/mg protein) in liver (A) and heart
(B). Box plot parameters as in Fig. 6. [LSD test]
In addition, activities of superoxide dismutase and catalase, two other
antioxidative enzymes, were measured in liver of experimental birds. The three-
stressor exposure resulted in fivefold increased SOD activity, when compared to
the single biomass exposure. No other significant modulations were observed in
other groups regarding SOD activity and no differences were found in the catalase
activity among exposure groups.
Biomarker of damage to lipid macromolecules was also evaluated and only in case
of liver of BPbV-exposed quails there was significant induction of lipid
peroxidation.
* C,E
CPb
EV
Pb+VE+V
E+PbE+Pb+V
5
6
7
8
9
10 B. C,Pb,V C,Pb,V
C,Pb,V
CPb
EV
Pb+VE+V
E+PbE+Pb+V
A. C C
B,V
C B PbV BPb C B PbV BPb Pb V BV BPbV Pb V BV BPbV
GR
nm
ol N
AD
PH
/ min
/ m
g p
rote
in
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Table 4.: Summary of the effects of exposure to cyanobacterial biomass (B), lead (Pb), Newcastle disease vaccine (V) and their combinations on bird antioxidative, detoxification and oxidative stress parameters. Statistically significant increase ▲ and decrease ▼ (p < 0.05) of parameter in specific group compared to control; ■ no statistically significant effect, - not measured. Biomarkers: glutathione (GSH), glutathione-S-transferase (GST), glutathione peroxidase (GPx), glutathione reductase (GR), lipid peroxidation (LP), superoxidedismutase (SOD) catalase (CAT).
GSH GST GPx GR LP stim. LP SOD CAT
liver
▲
BPbV
▲
BPb
▲
BPbV
▲
BPb
BPbV
▲
B
V
▼
BPbV
▲
PbV
BPbV
▲
BPbV
▲
BPbV
■
heart
▲
Pb
B
PbV
BPb
BPbV
▲
PbV
BV
BPbV
▲
BV
BPb
BPbV
▲
B
PbV
BPbV
▲
PbV
BV
BPbV
■
-
-
To summarize the results, general stimulation of antioxidative system with the
greatest modulations of sublethal parameters in the individuals from the groups
with combined exposures was (see Table 4). The greater modulation of biomarkers
in combined exposures was also confirmed by the principal component analysis,
which clearly separated the co-exposure groups from the other groups. Oxidative
stress has been documented in various bird species after cyanobacterial exposure
(our previous study), lead exposure (Douglas-Stroebel et al. 2004) and also
bacterial infections (Georgieva et al. 2006). This study confirmed this unspecific
biochemical process also in combined exposures to all the mentioned stressors.
When comparing all results, positive correlations among different biomarkers were
found only in liver (more detail presented in the full text of article) confirming
thus the major role of liver in detoxification of xenobiotics in birds (Riviere et al.
1985).
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72
The accumulation of microcystins within this experiment, which has been
published separately (Pikula et al. 2010), may support the diverse responses of
oxidative stress biomarkers. Briefly, there was higher accumulation of
microcystins documented in groups of combined lead-exposures than in non-lead
groups, which may indicate somewhat greater uptake of the cyanobacterial
metabolites in birds weakened by the lead exposure (Pikula et al. 2010). Finally,
these higher levels of microcystins along with toxic effects of lead (and effects of
immunological challenge) could contribute to the greatest modulations of almost
all examined biomarkers in these groups.
The role of glutathione and glutathione-S-transferase in detoxification of
cyanobacterial metabolites has been confirmed in correspondence with our
previous study with cyanobacterial biomass. Increases of GSH and GST after lead
exposure have been published (Douglas-Stroebel et al. 2004) and our study
similarly documented modulation of glutathione-related biomarkers after Pb
exposure. The correspondence among the activity of GST, GPx and the increased
level of GSH confirms the cooperation of the enzymes and previously reported
crucial role of glutathione in detoxification after exposure to lead (Berglund et al.
2007) and cyanobacterial biomass (Pašková et al. 2008) and the significance of
these biomolecules in the protection from harmful effects. The most significant
changes after multiple stressors exposure confirm our hypothesis that effects of
cyanobacterial biomass, lead and immunological challenge may combine to
enhance their influence. This study brought unique information on the effects of
combined exposure on important sublethal parameters in birds. General activation
of the antioxidant enzymatic system in exposed quails documents the greater need
of antioxidative protection in the studied organs and their ability to produce
molecules protecting cells against adverse oxidation processes.
73
7 CHAPTER 7
PAPER No.5
DETOXIFICATION AND OXIDATIVE STRESS RESPONSES IN EARLY STAGES OF AQUATIC
ORGANISMS EXPOSED TO PESTICIDES
Published as:
Teratogenicity and embryotoxicity in aquatic organisms after pesticide exposure
and the role of oxidative stress (2011)
Veronika Pašková, Klára Hilscherová and Luděk Bláha
Reviews of Environmental Contamination and Toxicology, Vol. 211, pp. 25-61
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7.1 Introduction
Generally, various developmental abnormalities have been documented in natural
populations of aquatic vertebrates. The intensification of agriculture with
application of huge amounts of fertilizers and crop protective agents next to the
loss of breeding habitats and other factors may contribute to the decline in aquatic
populations worldwide. Pesticides used in agricultural habitats are thus one of
possible factors contributing to occurrence of malformations in wild populations of
frogs and other aquatic vertebrates. Namely amphibians may potentially be target
of various environmental stressors and toxic exposure due to their biphasic life
cycles and skin permeability. There is some evidence about the oxidative
mechanism of action of certain pesticides (not only the pesticides designed to
produce ROS such as bipyridyls) being together with insufficient antioxidative and
detoxification potential one of the suggested mechanisms of pesticide
teratogenicity in non-target aquatic organisms.
In the review, which is a part of this dissertation thesis, the involvement of
oxidative stress in the process of teratogenic action of some pesticides is discussed
in relation to the adverse effects of pesticides on the non-target organisms -
amphibians, fish and aquatic invertebrates. Moreover, to my best knowledge, no
consistent overview of pesticide embryotoxicity in aquatic invertebrates and
vertebrates is available. Hence, in the review, the existing knowledge on this topic
is summarized and an overview of available information on the general teratogenic
and embryotoxic effects of pesticides in aquatic biota such as fish, amphibia and
invertebrates is presented. The toxic effects that are related to oxidative stress and
lines of evidence to support the view that it is a possible toxicity mechanism are
emphasized in this review.
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Any external factor, being harmless below threshold level, can become
embryotoxic at high concentration and have adverse effect on embryo or lead to
embryonic death. Teratogens can cause congenital malformations or abnormalities
in postnatal development, when acting during gravidity, without disturbing
maternal organism. Teratogens can alter cell division, proliferation, differentiation
or apoptosis (Gilbert 2006). Congenital malformation is a permanent structural or
functional abnormality or a biochemical change exceeding the boundaries of
normal species variability. Malformations, growth and functional retardation
represent the final manifestations of abnormal development (Meteyer 2000).
Genotype and the developmental phases in the moment of exposure can influence
the embryonic susceptibility to the teratogenic compound. Organogenesis
constitutes the most sensitive phase of development, when the bases of all organs
originate and the cellular impairment can lead to extreme structural changes (Wells
et al. 2005). Ultraviolet radiation, infections, parasitic trematodes, some
pharmaceuticals, aromatic compounds and other environmental contaminants
including pesticides have been reported as the conventional teratogenic factors
(Bilski et al. 2003; Blaustein and Johnson 2003; Ankley, et al. 2004; Hayes et al.
2006) in aquatic environment.
7.2 Pesticides teratogenicity in invertebrates
Pesticides can alter development and reproduction functions of various aquatic
organisms (see details in the fulltext of the review-article) including invertebrate
populations. The most toxic and teratogenic effects on embryos and larvae of
invertebrate organisms together with decreased hatching success and delayed
hatching time were observed for gastropods, bivalve molluscs, echinoids and
decapod crustaceans (Harper et al. 2008; Key et al. 2007; Lee and Oshima 1998;
Sawasdee and Köhler 2009; Bhide et al. 2006; Buznikov et al. 2007). These
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studies were performed with organophosphate and organochlorine insecticides and
the effects of other pesticide classes (e.g., synthetic pyrethroids, chlorocetanilides
and terpenoids; triazines, carbamates, azoles and phenylpyrazoles) were also
studied. The toxic effects included decreased rate of fertilization, increased
polyspermy and alterations of mitotic divisions, increased embryonic and larval
mortality, abnormal cleavage and disruptions of gastrulation and other basic
morphogenetic processes, altered early embryonic development together with
larval deformities. Also other species as for example Cladoceran can be used for
embryolarval testing but embryolethality can be masked by changes in other
parameters, such as adult immobilization or number of offspring (Abe et al.
2001). On the other hand the Ascidian (Phallusia mammillata) embryos and larvae
have been used to evaluate effects of pesticides imazalil and triadimefon on sperm
viability, fertilization and embryogenesis. Alterations of the anterior structures of
the trunk, incorrectly differentiated papillary nerves and anterior central nervous
system have been observed after pesticides exposure (Pennati et al. 2006).
7.3 Pesticides teratogenicity in fish
Embryonic abnormalities in fish can result from direct exposure to toxic
contaminants in water column (Heintz et al. 1999) or from the bioaccumulation of
toxic compounds (organochlorines; PCBs and pesticides; heavy metals) in
reproductive tissues (Westernhagen von 1988). Early stages of fish embryonic
development were assessed as ideal pollution bioindicators. Embryolarval
abnormalities are thus considered indicators of general water quality and
embryonic malformation rates can be useful when expressing the toxic impact of
specific pollutants with teratogenic action (Klumpp et al. 2002). Species from
families Cyprinidae, Adrianichthyidae and Salmonidae are most common in
studies on embryonic and teratogenic effects in fish (Osterauer and Köhler 2008;
Villalobos et al. 2000; Sylvie et al. 1996). Wide variety of pesticides was used in
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these studies as for example organophosphates, triazines, synthetic pyrethroids,
carbamates, organochlorines or organosulphur pesticides (see the detailed
summarization in the fulltext of the review-article). Gastrulation and early
segmentation phases were shown to be sensitive to pesticides exposure causing
larval deformities such as twisted notochord and notochord distortions (Haendel et
al. 2004). Other adverse effects caused by pesticides exposure included embryonic
malformations as wavy notochord, disorganized somites, and shortened yolk sac
extension (Teraoka et al. 2006), reduced larval survival together with growth and
eye diameters, opaque skin, exophthalmia (Cook et al. 2005), poor yolk resorption,
cephalic and spinal deformities together with bradycardia, pericardial edema
(Villalobos et al. 2000) and myoskeletal abnormalities (Viant et al. 2006). There
were also documented adverse effects of pesticides on the reproduction and
hatching success (Köprücü and AydIn 2004) as well as on behavioural changes
and uncoordinated muscle contractions along the body axis in response to touch
(Stehr et al. 2006).
7.4 Pesticides teratogenicity in amphibian
Amphibians can be negatively affected by many factors (chemical, physical,
habitat etc.) or by the combination of these stressors in the environment (Boone et
al. 2007; Bridges et al. 2004). Moreover, many organic compounds and metals can
accumulate in their tissues. For example, the average concentrations of
polychlorinated biphenyls and chlorinated pesticides were in tenths and units of
ng/g DW value in wild-living frogs Rana spp. (Russell et al. 1997). Amphibians
do not tolerate intensive agricultural activity and destruction of habitats, which has
resulted in the reduction of amphibian populations over recent decades (Ouellet et
al. 1997). Moreover, amphibians are particularly vulnerable to environmental
contaminants including pesticides because of their semipermeable skins and
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78
biphasic life cycles (Mahaney 1994). Pesticides are among a number of proposed
causes of amphibian malformations and decline worldwide (Muths et al. 2006).
The direct linkage between agricultural use of pesticides and amphibian
populations decline has not been clearly driven, but there is some evidence about
the exposure to pesticides during the embryogenesis linked with modifications of
the normal developmental processes (Anguiano et al. 2001). Moreover, most
amphibian species breed during spring, when huge quantities of pesticides are
sprayed onto the land becoming accessible for amphibian embryos and potentially
leading to malformations in their early developmental phases (Greulich and
Pflugmacher 2003). Morphological abnormalities and injuries, for example
clinodactyly (congenital curly toes), ectrodactyly (missing digit), brachydactyly
(short toe), polydactyly (supernumerary toes), polymely (redundant or
supernumerary digits), hemimelea (short tibia or fibula), ectromelea (incomplete
limb with missing lower part), abnormal webbing of toes, tail projections and
unilateral anophthalmia (congenital absence of one eye) or microphthalmia occur
physiologically in wild amphibian populations, but only at low frequencies ranging
from 0 to 2% (Ouellet 2000). In contaminated ponds, the occurrence of malformed
individuals can reach 60% of the newly metamorphosed frog population (Meteyer
2000). The deleterious effects of agricultural pesticides and fertilizers are one of
the hypothetical causes of deformities and mortality of amphibians and other
aquatic wildlife. There are many studies reporting the occurrence of malformed
aquatic organisms in the environment as for example the study of morphological
abnormalities in natural populations of Limnonectus limnocharis, L. keralensis, L.
brevipalmata and Spherotheca rufescens inhabiting Indian ecoagrosystem
(Gurushankara et al. 2007). Limb and eye abnormalities together with tumours on
the femur and bulged abdomen were found in linkage with organochlorine,
carbamate and organophosphorus pesticides used in agriculture.
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Embryotoxicity and teratogenicity of various pesticides have been documented to
occur in amphibians in laboratory studies and in field observation studies (see
details in the fulltext of the review-article). Most of the studies used Xenopus
laevis from the Pipidae family as the model organism, but species from other
families that included Ranidae, Bufonidae, Microhylidae and other amphibians as
for example salamanders from the family Ambystomatidae, were also used.
Similarly to the effects occurring in fish and also invertebrates, amphibians are
known to be highly sensitive to several pesticides affecting their development.
Myoskeletal defects, abnormal tail formation and limb differentiation are among
the most frequently reported effects caused by pesticide exposure (Bacchetta et al.
2008). Further alterations include incomplete neurulation, edemas, epidermal
defects or gut malformations (Robles-Mendoza et al. 2009), as well as
dysmorphogenesis, embryonic and larval lethality, delayed hatching, growth
retardations or altered metamorphosis (Vismara et al. 2000).
7.5 Role of oxygen and antioxidantive compounds in embryogenesis
Mechanisms acting in the process of teratogenesis are not sufficiently known. The
possible mechanisms differ for various chemicals, including biochemical,
physiological, structural or gene-expression alteration. The DNA damage, for
example gene mutation, chromosomal aberration, mitotic interference, modulation
of nucleic acid metabolism and cell energetic supply, inhibition of enzymes and
membrane alteration as well as disruption of retinoic acid signaling or oxidative
stress are considered important mechanisms by which xenobiotics may induce
developmental effects (Beckman and Brent 1984; Wells et al. 2005). Thus, various
factors with various mechanisms of action can affect the normal development of
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80
organisms in the environment; probably their combination in connection with
environmental conditions may induce malformations (Blaustein and Johnson
2003). Malformations represent primary errors in development, errors in chemical
communication or translation of genetic information. The type of error or insult as
well as the timing of the error (developmental stage at which the error occurred)
influences the occurrence and the type of malformation. The morphology of the
malformation does not define the cause. Therefore investigations that aim to
determine the cause of malformation need to look at agents (chemical, physical,
biological) that are present in the animals or their habitat at early developmental
stages (Meteyer 2000).
Oxygen plays a key role in metabolism, and is critical to the early developmental
stages of organisms. Several oxygen derivatives, known as ROS, are known to
have signaling functions and may affect several physiological and pathological
processes in an organism (Covarrubias et al. 2008). At the level of embryogenesis,
sensitive regulation of ROS has been linked to control of oocyte cleavage (Allen
and Balin 1989), as well as oocyte maturation, ovarian steroidogenesis, ovulation,
implantation, formation of blastocysts (Guerin et al. 2001).
ROS play also an important role in cell signaling mechanisms that control gene
expression and minor changes in the redox-status of cells can alter gene expression
at the decisive stages of embryonic development potentially resulting in
teratogenic events (Hilscherova et al. 2003). Only small amounts of ROS are
necessary to maintain normal cell functions. The levels of ROS must be
continuously controlled to prevent them from becoming highly toxic to biological
macromolecules (e.g., proteins, DNA and membrane lipids) (Agarwal et al. 2003).
The resulting teratogenic effect of xenobiotics thus depends also on detoxification,
macromolecule repair and other protective mechanisms (Wells et al. 2005).
General antioxidant defenses were recently shown to play an important role in
PAPER No.5 - DETOXIFICATION IN AQUATIC ORGANISMS AFTER
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81
protecting both early aquatic larval stages (Maria et al. 2009; Tilton et al. 2008),
later developmental phases, as well as the metamorphosis process (Dandapat et al.
2003).
The imbalance between oxidative intermediates (pro-oxidants) and the ability of
the antioxidants to scavenge the excess ROS production in case of impaired
antioxidant defence mechanisms may result in oxidative stress (Wells et al. 2005)
and be thus the main factor affecting the normal embryogenesis. Most research
that has been conducted on this topic has included studies with model compounds
such as hydrogen peroxide, and has employed laboratory rodents or human
embryos. In these studies, pro-oxidants induced severe oxidative stress damage to
oocytes, mitochondrial alterations, ATP depletion, DNA damage and lipid
peroxidation, apoptosis or delays in whole embryo development (Aitken and
Krausz 2001; Duru et al. 2000). The importance of oxidative stress in causing
embryotoxicity or teratogenicity was also indirectly confirmed in mammalian and
human studies, in which external additions of antioxidants prevented damage to
embryos (Feugang et al. 2004). Studies with model pro-oxidants have also
demonstrated detrimental effects in fish embryos and larvae (Westernhagen von
1988), as well as in the larvae of the giant prawn Macrobrachium rosenbergii
(Dandapat et al. 2003). Moreover, the addition of antioxidants protected fish
embryonal development against the effects of oxidative stress (Tilton et al. 2008).
There are only few laboratory studies available documenting the role of oxidative
stress in pesticide-induced teratogenicity in aquatic organisms. The exposure of
embryos of fish Danio rerio to atrazine lead to retardation of organogenesis
(especially eyes, somites, otolithes and melanophores), dysfunctions of the
circulatory system, edemas and a delay in embryonic development; interestingly,
these effects occurred in parallel with alterations of GST activities (Wiegand et al.
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82
2000; Wiegand and Pflugmacher 2001).
Mortality in embryos and developmental abnormalities along with oxidative stress
markers were observed in two studies with embryos of the toad Bufo arenarum.
Anguiano et al. (2001) discovered that the organochlorine insecticide lindane
caused abnormal segmentation of furrows, along with irregular blastomeres,
profuse scaling, dropsy, organ displacements and bent tail. Interestingly, only
moderate alterations of embryonic morphology and ahemorrhagia were observed
after exposure to another organochlorine insecticide - dieldrin (Anguiano et al.
2001). In the same study Anguiano et al. (2001) also showed that the
organophosphate insecticides malathion and parathion were highly embryotoxic
and caused a pathological curvature of the antero-posterior axis, tail folding
edema, frequent dropsy and also induced circle-swimming movements. Ferrari et
al. (2009) studied the effects of carbaryl and azinphos methyl on the embryos of
Bufo arenarum, and demonstrated progressive dropsy, notochord malformations,
gill atrophy, paralysis and delayed development. The above described effects were
also correlated with modulations of glutathione levels and elevated activities of
antioxidants GST, SOD, CAT and GR (Anguiano et al. 2001; Ferrari et al. 2009).
In studies with invertebrates the oxidative stress and disruptions of development
were shown after exposure to heptachlor in grass shrimp (Snyder and Mulder
2001). Similarly, larval toxicity and modulation of antioxidant and detoxification
parameters were shown also after exposures to complex media contaminated with
pesticides in oysters (Damiens et al. 2004).
On the other hand, direct toxic effects of pesticides on developing embryos were
not found in other studies, but signs of oxidative stress and variable modulation of
the antioxidative system were observed (as for example (Küster and Altenburger
2007)).
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The studies with bipyridyl herbicides paraquat and diquat are of special interest
because the major mechanism by which they produce their toxic action in target
organisms (both animals or plants) is through lipid peroxidation. Disturbances of
normal early developmental processes after exposure to paraquat were clearly
documented in Xenopus laevis embryos (Vismara et al. 2001; Vismara et al. 2000).
These toxic effects were prevented after the addition of the water-soluble
antioxidant ascorbic acid to the test medium (Vismara et al. 2001; Vismara et al.
2006).
7.6 Conclusions
To conclude this review, many pesticides have been documented to induce
embryotoxicity and teratogenicity in non-target aquatic biota such as fish,
amphibians and invertebrates. This review of the existing available literature
showed that a broad range of pesticides, representing several different chemical
classes, induce variable toxic effects in aquatic species. The observed effects
include diverse morphological malformations as well as physiological and
behavioural effects. When developmental malformations occur, the myoskeletal
system is among the most sensitive targets. Myoskeletal effects that have been
documented to result from pesticide exposures include common notochord and
vertebrate column degeneration and related abnormalities. Pesticides were also
shown to interfere with the development of organ systems including eyes or heart
and are also known to often cause lethal or sublethal edema in exposed organisms.
The physiological, behavioral and population endpoints affected by pesticides
include low or delayed hatching, growth suppression, as well as embryonal or
larval mortality. The risks associated with pesticide exposure increases particularly
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84
during spring. This is the period of time in which major pesticide applications take
place, and this period unfortunately also coincides with many sensitive
reproductive events such as spawning, egg-laying and early development of many
aquatic organisms.
Only few experimental studies with pesticides have directly linked developmental
toxicity with key oxidative stress endpoints, such as lipid peroxidation, oxidative
DNA damage or modulation of antioxidant mechanisms. On the other hand, it has
been documented in many reports that pesticide-related oxidative damage occurs
in exposed adult fish, amphibians and invertebrates. Moreover, the contribution of
oxidative stress to the toxicity of pesticides has been emphasized in several review
papers concerned with this topic (Valavanidis et al. 2006; Monserrat et al. 2007;
Debenest et al. 2010).
In conclusion, the available experimental data, augmented by several indirect lines
of evidence, support the concept that oxidative stress is a highly important
mechanism in pesticide-induced reproductive or developmental toxicity. Other
stressors may also act by oxidative mechanisms. This notwithstanding, there is
much yet to learn about the details of this phenomenon and further research is
needed to more fully elucidate the effects that pesticides have, and the
environmental risks they pose in the early development of aquatic organisms.
8 CHAPTER 8
PAPER No.6
EMBRYOTOXICITY AND INDUCTION OF OXIDATIVE STRESS AFTER EXPOSURE OF MODEL
NON-TARGET AQUATIC ORGANISM AFRICAN CLAWED FROG ( XENOPUS LAEVIS) TO BIPYRIDYL
HERBICIDES PARAQUAT AND DIQUAT
Veronika Pašková, Zdena Moosová and Klára Hilscherová manuscript
CHAPTER 8
86
8.1 Introduction
The potential impact of pesticides and other contaminants on aquatic organisms
has been widely discussed in connection with worldwide decline of frog
populations. As reviewed above, amphibians are particularly vulnerable to
environmental contaminants because of their skin semipermeability and biphasic
life cycles (Mahaney 1994). Moreover, many factors and their combinations
present in the environment (Boone et al. 2007) can negatively affect them. The
effects of pesticides are widely discussed (Anguiano et al. 2001), however, the
direct linkage between agricultural use of pesticides and amphibian populations
decline has not been clearly driven, even though there is evidence about the
exposure to pesticides during the embryogenesis and modifications of the normal
developmental processes. The fact that most amphibian species breed during
spring, when many pesticides are sprayed onto the land becoming accessible to
amphibian embryos and potentially leading to malformations in their early
developmental phases (Greulich and Pflugmacher 2003) belongs to important
factors playing role when evaluating the potential negative effects of pesticides.
In embryos, various metabolic pathways and enzymes can produce endogenous
ROS, but also some environmental pollutants can contribute to generation of these
reactive intermediates. Bipyridyl pesticides form an important class of aquatic
pollutants that can induce oxidative stress in cells. Moreover, they are toxic at low
environmental concentrations and may thus influence the non-target organisms as
for example amphibians (Winston and Di Giulio 1991; George et al. 2000).
Generally, toxicological and herbicidal properties of bipyridyls depend on the
ability of their parent cation to undergo a single electron addition in the presence
of NADPH-reductases to form a free bipyridyl radical, which reacts with
molecular oxygen to reform the cation and concomitantly produce a superoxide
STUDY No.6 – EMBRYOTOXICITY AND OXIDATIVE STRESS IN FROGS
EXPOSED TO PESTICIDES
87
anion (Vale and Meredith 1981). The reoxidized bipyridyl compound is capable of
accepting another electron and continuing the electron transfer reactions and
formation of reactive species.
Paraquat, diquat, chlormequat, difenzoquat and morfamquat are considered the
most important members of the bipyridyl herbicides. On these grounds two non-
selective contact herbicides – paraquat and diquat – were used for this experiment.
Paraquat (PQ; 1,1'dimethyl, 4,4'bipyridyl) is used worldwide in approximately
130 countries for plantation crops (banana, cocoa-palm, coffee, oil-palm etc.) and
for citrus fruits, apples, plums, vines and tea. On certain crops (potato, pineapple,
sugar-cane, sunflower), it is used as a dessicant; it is also used as a cotton
defoliant. Diquat (DQ; 1,1'ethylene, 2,2'bipyridyl) is used to control both broad-
leaved weeds among crops and submerged and floating weeds in water bodies, for
potato haulm destruction, and for seed crop desiccation (rice, sunflower etc.).
Generally, PQ is marketed as an aqueous solution of the dichloride salt
Gramoxone®, DQ as an aqueous solution of the dibromide salt Reglone®, both in
concentration of 200 ± 10 g/litre. Rates for various PQ applications usually range
1.1 to 2.2 L/acre and for DQ 0.5 to 1.5 L/acre (acre – 0.4 ha – 4000 m2) (Syngenta
2012; Syngenta 2012), respectively, with working dilutions 1 - 5 g PQ or DQ/L in
water (WHO 1991; WHO 1984).
PQ and DQ are rapidly absorbed by green plant tissue and act as contact herbicides
and dessicants with limited systemic properties (Peterson et al. 1997); they are
chemically reduced in plants by replacing NADP as an electron acceptor in
photosynthesis and when oxidized, highly phytotoxic H2O2 is produced. The
continued production of H2O2 is dependent upon the maintenance of
photosynthetic electron transport (Harris and Dodge 1972). In mammals, they
cause principally lung and kidney damage sometimes coupled with extensive
CHAPTER 8
88
degeneration and fibrosis of skeletal muscles (Tabata et al. 1999) and generally the
major cause of death in bipyridyl poisonings is respiratory failure due to an
oxidative insult to the alveolar epithelium with subsequent obliterating fibrosis
(Suntres 2002).
The oxidative damage caused by bipyridyls has been studied mostly in vitro and
the addition of antioxidants into media helping to scavenge the excessive
production of ROS has been tested (Zychlinski et al. 1987; Yant et al. 2003;
Lawlor and O´Brien 1995). The decrease of NADPH oxidation rate in presence of
ascorbic acid in media with paraquat was observed in rat lung microsomal
fractions (Zychlinski et al. 1987). The essential role of other antioxidant -
glutathione peroxidase 4 (GPX4) - in protection against oxidative damage induced
by stressors including pesticide paraquat has been ascertained in a study with mice
embryonic stem cells (Yant et al. 2003). GPX4 is the only major antioxidant
enzyme known to directly reduce phospholipid hydroperoxides within membranes
and lipoproteins, acting in conjunction with α-tocopherol to inhibit lipid
peroxidation. Reduction of oxidative stress caused by paraquat by addition of
antioxidants astaxanthin and β-carotene was accordingly observed in primary
cultures of chicken embryo fibroblasts (Lawlor and O´Brien 1995). Studies with
co-exposure to antioxidants in coherence with protection against oxidative damage
were carried out also in vivo. The enhancement of the antioxidant status of
hatching chicks through the supplementation of vitamin E to the maternal diet was
found in connection with the protection of tissues of the progeny from oxidative
injury (Lin et al. 2005). The ability of two antioxidants, hydrophilic analogue of
vitamin E and β-mercaptoethanol, to prevent to some extent blastocyst
degeneration induced by prooxidants was observed in a study with bovine embryos
culture from the morula stage (Feugang et al. 2004).
STUDY No.6 – EMBRYOTOXICITY AND OXIDATIVE STRESS IN FROGS
EXPOSED TO PESTICIDES
89
The protective role of ascorbic acid was confirmed in frogs as well (Vismara et al.
2001). The embryotoxic effects of paraquat seemed to be most likely linked to
oxidative damage of embryos since the addition of antioxidant ascorbic acid was
followed by the drastically reduced embryotoxicity. Also in human the addition of
ascorbic acid to a sperm culture significantly protected its development (Fraga et
al. 1991). In mammalian embryo culture, addition of another antioxidant SOD to
the medium increased embryonic SOD activity and SOD or enzymatic antioxidant
CAT also blocked oxidative damage as embryolesions and embryolethality caused
by the anticonvulsant drug phenytoin and tobacco carcinogen benzo[a]pyrene
(Wells et al. 1997). Accordingly, in this study we investigated the embryotoxic and
teratogenic effects of two chosen bipyridyls with/without the presence of ascorbic
acid in FETAX media in experiments with Xenopus laevis embryos.
Embryos are protected against oxidative stress by ROS scavengers. Low activities
of scavenging enzymes resulting from pathological metabolism imbalances may
reduce protection potential of embryo, potentially resulting in teratogenic event.
The scope of this experiment was to study the effect of paraquat and diquat on the
early phases of amphibian development using African clawed frog (Xenopus
laevis) in the standard testing FETAX scheme supplemented with assessment of
sublethal toxic effects on modulation of antioxidative and detoxification
compounds. The study design with detoxification parameter glutathione-S-
transferase, antioxidative parameters as glutathione, glutathione reductase,
glutathione peroxidase, catalase and superoxidedismutase (only in assessing the
developmental profile during development) together with the lipid peroxidation as
the damage indicator was chosen to assess the role of oxidative stress in the
embryotoxicity and teratogenity of tested bipyridyls. For this purpose, the
developmental profile of these antioxidative and detoxification compounds were
CHAPTER 8
90
evaluated in a 24 hour-interval. These basal activities of biomarkers were
determined to illustrate their physiological developmental changes occurring
within the normal development of frog embryos. Further, the changes of these
biomarkers (in 24-hour intervals) were studied after exposure to paraquat and
diquat. Sequentially, the effect of external addition of antioxidant ascorbic acid
was evaluated to test the hypothesis of oxidative stress involvement in bipyridyl
pesticides teratogenicity as suggested in similar investigations (Vismara et al.
2001; Feugang et al. 2004; George et al. 2000; Lawlor and O´Brien 1995; Wang et
al. 2002).
8.2 Materials and methods
Chemicals
Paraquat dichloride (CAS 1910-42-5), diquat dibromide (CAS 85-00-7),
biochemicals and enzymes were purchased from Sigma-Aldrich (Prague, Czech
Republic). Other chemicals used for preparation of media as well as solvent
(dimethylsulfoxide [DMSO]) were of the highest quality available.
Bioassay
Frog embryos were obtained from adult pairs of X. laevis injected with human
chorionic gonadotropin (HCG; N.V. Organon, Oss, Holland) in the dorsal lymph
sac, whereas the animals were separately pre-injected with 50-100 IU one week
prior to mating. To induce mating, the male and female received 150 and 300 IU,
respectively. Amplexus normally ensued within 2 to 6 h and the deposition of eggs
occurred from 9 to 12 h after injection. Embryotoxicity tests were conducted using
the standard guide for the Frog Embryo Teratogenesis Assay Xenopus (ASTM
1998). Mid-blastula (stage 8) to early gastrula (stage 11) embryos (Nieuwkoop and
Faber 1994) were selected for testing. Groups of 25 embryos were randomly
STUDY No.6 – EMBRYOTOXICITY AND OXIDATIVE STRESS IN FROGS
EXPOSED TO PESTICIDES
91
placed in covered 60-mm plastic Petri dishes into 10mL of standard FETAX
solution (625 mg NaCl, 96 mg NaHCO3, 75 mg MgSO4, 60mg CaSO4·2H2O,
30mg KCl and 15 mg CaCl2 per liter of distilled water; pH 7.6–7.9) and exposed to
chemicals. The bipyridyl compounds were dosed in solvent DMSO (final
concentration 0.5 % v/v) and their exposure concentrations were 0.03 to 0.5 mg/L
for PQ and up to 5 mg/L in case of DQ. Each concentration was tested in three
parallels. Control groups were exposed to the standard FETAX medium only. For
the assessment of the developmental profile of the antioxidative and detoxification
compounds three independent groups of eggs from three parental pairs were used.
The solvent control was also performed in final DMSO concentration of 0.5 % v/v.
The addition of ascorbic acid at final concentration of 200 mg/L into the pesticide-
treated FETAX-medium was done as the parallel experiment. The FETAX assay
was performed at 23±1 ◦C for 96 h, every 24 hours the exposure solutions were
changed, dead embryos were recorded and removed and the embryos for
biochemical analyses were kept at −80 ◦C. At the end of the assay (96 h),
surviving embryos were fixed in 3% (v/v) formaldehyde, the length of embryos
was determined by a ruler and the embryos were assessed for morphological
abnormalities under the dissecting microscope. The deep-frozen embryos were
then homogenized on ice in phosphate buffer saline (PBS, pH 7.2). The
postmitochondrial supernatant was collected after centrifugation (30min at
30 000 g at 4°C for CAT and SOD and 15 min at 10 000 g at 4°C for the other
parameters) and stored frozen at -80°C prior to analyses.
Biochemical methods
Glutathione-S-transferase (GST) activity was measured spectrophotometrically at
340 nm using 1 mM 1-chloro-2, 4-dinitrobenzene (CDNB) as a substrate and
2 mM GSH in PBS (Habig et al. 1974). Specific activity was expressed as nmoles
CHAPTER 8
92
of evolved product per minute per milligram protein. The concentration of
glutathione was determined by spectrophotometric method using 5,5´-dithiobis-2-
nitrobenzoic acid (DTNB) as a substrate (Ellmann 1959). Tissues were treated
with trichloracetic acid (TCA, 25% w/v) and centrifuged (6 000 g for 10 min at
4°C). Then 50 µl of supernatant was mixed with 230 µl of TRIS-HCl buffer (0.8 M
TRIS, 0.02 M EDTA, pH 8.9) and 20 µl of 0.01 M DTNB and incubated for 5 min
at room temperature. Absorbance was measured at 420/680 nm and the
concentrations (nmol GSH/mg protein) were calculated according to the standard
calibration with reduced GSH. Activity of glutathione peroxidase (GPx) was
determined (Flohé and Gunzler 1984) from the rate of NADPH oxidation recorded
as the decrease in absorbance at 340 nm. GPx activity was assayed in microplates
with final concentrations of 3 mM GSH, 1 U glutathione reductase (GR) (1 unit
[U] will reduce 1.0 µmole of oxidized glutathione per min at pH 7.6 at 25ºC),
0.15 mM NADPH in 0.02 M potassium phosphate/0.2 mM EDTA buffer (pH 7).
Substrate used for the assay was 1.2 mM butylhydroperoxide. Also the activity of
GR was determined by spectrophotometric measurement of NADPH oxidation
(Carlberg and Mannervik 1975) in microplates. The final mixture contained
0.05 M potassium phosphate/1 mM EDTA buffer (pH 7.0), 1 mM oxidized
glutathione (GSSG), 0.1 mM NADPH and the supernatant (0.25 % v/v). Specific
activities of both GPx and GR were expressed as nmoles NADPH oxidized per
minute per milligram protein. Activity of superoxide dismutase (SOD) was
determined spectrophotometrically at 560 nm according to the method using
nitroblue tetrazolium (NBT) as a substrate (Ewing and Janero 1995). The reaction
mixture contained 60 µM NBT, 100 µM NADH and 35 µM phenazine
methosulfonate in 50 mM potassium phosphate/1 mM EDTA buffer. Specific
activity was expressed as nmol NBT oxidized per minute per milligram protein.
Activity of catalase (CAT) was evaluated spectrophotometrically at 240 nm in
cuvettes as a rate of break down of 0.09% hydrogen peroxide in 50 mM
STUDY No.6 – EMBRYOTOXICITY AND OXIDATIVE STRESS IN FROGS
EXPOSED TO PESTICIDES
93
TRIS/0.1 mM EDTA buffer and the specific activity was thus expressed as µmol
hydrogen peroxide oxidized per minute pre milligram protein (Aebi 1984). The
level of lipid peroxidation as a potential damage of lipid membranes in frog
embryonic tissues was assessed as total thiobarbituric acid (TBA) reactive species
(TBARS) (Uchiyama and Mihara 1978); (Livingstone et al. 1990). The
homogenates were mixed with trichloracetic acid (TCA, 6% w/v) and butylated
hydroxytoluene (0.6% w/v) and centrifuged (1 500 g for 20 min). Supernatant was
further mixed with 0.06 N HCl and 40 mM TBA prepared in 10 mM TRIS (pH
7.4). The mixture was boiled at heating plates for 45 min and then cooled to room
temperature. Absorbance of the sample was measured at 550/590 nm and the
concentration of TBARS (nmol TBARS per milligram protein) was calculated
according to standard calibration curve generated with malondialdehyde prepared
by acidic hydrolysis of 1,1,3,3-tetraethoxypropane. The protein concentrations
were determined by the method using Folin-Ciocalteu phenol reagent that forms
red-colored complex measurable at 680 nm in reaction with proteins (Lowry et al.
1951). Bovine serum albumin was used as standard for protein calibration. The
microplate spectrophotometer PowerWave (BioTek, Winoosa, USA) was used to
measure the absorbance in microplates and spectrophotometer VARIAN CARY
50 Bio (Varian, USA) was used for the measurement of absorbances of solutions
in cuvettes.
Statistical evaluation
Statistical analyses were performed with Statistica for Windows® 7.0 (StatSoft,
Tulsa, OK, USA). The homogeneity of variances prior to ANOVA was assessed
by the Levene’s test. Differences among the total embryo lengths and biochemical
parameters were evaluated by ANOVA and LSD post-hoc test. Differences in
frequencies of mortalities and malformations were compared by χ2 test. P values
CHAPTER 8
94
less than 0.05 were considered statistically significant. Bipyridyl concentrations
causing 50% lethality (LC50) and concentrations eliciting malformations in 50% of
surviving embryos (EC50) were analysed by the programme GraphPad Prism by
non-linear regression (GraphPad Software, San Diego, CA, USA) .The teratogenic
index (TI) was calculated as a ratio of LC50 and EC50 for tested compounds.
8.3 Results
Mortality
Significant mortality of Xenopus embryos after paraquat (PQ) exposure was
shown. This toxic effect was ten times higher when compared to diquat (DQ)
embryolethality, respectively (Fig.7A). More than 50% mortality was observed at
0.12 mg/L PQ; 100% mortality at 0.25 mg/L after 96 hours and at 0.5 mg/L
already after 72 hours of exposure. In case of diquat, the same percentual
mortalities occurred at ten times higher concentrations.
STUDY No.6 – EMBRYOTOXICITY AND OXIDATIVE STRESS IN FROGS
EXPOSED TO PESTICIDES
95
Figure 7.: Mortality of Xenopus laevis embryos during first 96 hours of development after
paraquat and diquat exposure (A) with the addition of ascorbic acid (AA) at concentration of
200 mg/L (B). Asterisks indicate statistically significant increase in mortality in comparison with
the control [χ2; * = P < 0.05].
The external addition of 200 mg/L ascorbic acid (AA) to the FETAX media
strongly reduced the rate of bipyridyls embryolethality as shown in Fig.7B.
Significant mortality (88% mortality after 96 hours of exposure) of the mixture of
PQ with 200 mg/L AA was documented only at the highest PQ concentration
tested. Only 32% mortality after 96 hours of exposure was detected in case of DQ
mixture with 200 mg/L AA at the highest DQ concentration. The addition of
0
25
50
75
100
% m
ort
ality
I II III IV
days of development
paraquat + ascorbic acid 200 mg/L
control
AA 200 mg/L
0.12 mg/L + AA
0.25 mg/L + AA
0.5 mg/L + AA
*
B.
0
25
50
75
100
% m
ort
ality
I II III IV
days of development
paraquat
control
0.12 mg/L
0.25 mg/L
0.5 mg/L
*
*
*
* *
*
0
25
50
75
100
% m
ort
ality
I II III IV
days of development
diquat
control
0.5 mg/L
1.25 mg/L
2.5 mg/L
5 mg/L
0
25
50
75
100
% m
ort
ality
I II III IV
days of development
diquat + ascorbic acid 200 mg/L
control
AA 200 mg/L
0.5 mg/L + AA
1.25 mg/L + AA
2.5 mg/L + AA
5 mg/L + AA
A.
*
*
*
*
* *
*
*
CHAPTER 8
96
200 mg/L AA caused an increase of LC50 values from 0.123 to 0.389 mg/L and
0.726 to 4.99 mg/L (LC30 value) for PQ and DQ, respectively.
In conclusion, the embryotoxic effects of paraquat seemed to be linked to
oxidative damage of embryos and the addition of antioxidant ascorbic acid lead to
significantly reduced embryotoxicity.
Growth reduction
Significant reduction of length of the surviving embryos was documented after
both paraquat and diquat exposures. Statistically significant differences between
the total body lengths of control and treated embryos were found already at PQ
concentration of 0.125 mg/L; DQ concentration of 0.5 mg/L, and higher (Fig.8).
Lesser growth inhibition was observed in the cotreatment of both bipyridyls with
200 mg/L ascorbic acid. Moreover, the total body lengths of the embryos from the
groups with AA in some cases almost restored to the lengths of the control
embryos.
Figure 8.: Growth inhibition of Xenopus laevis embryos after paraquat and diquat 96 hour-
exposure and the effect of ascorbic acid added into the FETAX media on this growth inhibition.
Asterisks indicate the statistically significant differences from the body length of the control
embryos [LSD test; * = P < 0.05].
diquat
0
2000
4000
6000
8000
10000
control 200mg/LAA
0.5mg/LDQ
0.5mg/LDQ +200mg/LAA
1.25mg/LDQ
1.25mg/LDQ +200mg/LAA
2.5mg/LDQ
2.5mg/LDQ +200mg/LAA
5 mg/L 5 mg/L+ 200mg/L AA
tota
l len
gth
(um
)
.
* * * * *
*
paraquat
0
2000
4000
6000
8000
10000
control 200 AA 0.06mg/L PQ
0.125mg/L PQ
0.125mg/L PQ
+ 200mg/L AA
0.250mg/L PQ
0.250mg/L PQ
+ 200mg/L AA
0.500mg/L
PQ+ 200mg/L AA
tota
l len
gth
(um
)
.
* *
* *
STUDY No.6 – EMBRYOTOXICITY AND OXIDATIVE STRESS IN FROGS
EXPOSED TO PESTICIDES
97
Malformations
The exposure to pesticides paraquat and diquat induced embryonal malformations.
About 30% induction of malformations was observed at PQ concentration of
0.06 mg/L and DQ concentration of 0.25 mg/L, respectively. Interestingly, 100%
malformed individuals occurred at PQ concentration of 0.25 mg/L and DQ
concentration of 2.5 mg/L (Fig. 9A). Axial abnormalities, oedemas, microphtalmia
and asymmetric eye formation were the most frequent malformations. Moreover,
in case of abnormal gut coiling and more complex axial deformities, there was
obvious dose-response trend. Significantly lower occurrence of malformations was
ascertained after addition of ascorbic acid to the FETAX media and interestingly,
at some testing concentrations the ascorbic acid addition resulted in lower
occurrence of malformation than in control or in no malformations at all (Fig.9B).
The teratogenic index (TI) was calculated according to the FETAX methodology
as a ratio of LC50 and EC50 for tested compounds; TI > 1.5 indicates embryotoxic
and teratogenic potential and TI > 3 indicates strong teratogenic potential. Low TI
value of 1.4 was calculated for paraquat but higher TI (2.2) was ascertained in case
of diquat.
CHAPTER 8
98
Figure 9.: Frequency of embryolarval malformations of Xenopus laevis after 96 hours of paraquat
and diquat exposure (A.) and the effect of ascorbic acid added into the FETAX media on
malformations occurrence (B.). Asterisks indicate statistically significant differences from the
appropriate testing concentration after addition of ascorbic acid [LSD test; * = P < 0.05].
Early developmental profile of antioxidative and detoxification compounds
and lipid peroxides
The 4-day profile of chosen glutathione-related biomarkers and other antioxidants
together with the level of lipid peroxides have been determined during the first
96 hours of Xenopus laevis embryolarval development (Fig. 10).
0
25
50
75
100
control 200 mg/L
AA
0.06 mg/L 0.125 mg/L
PQ
0.125 mg/L
PQ+200
mg/L AA
0.25 mg/L 0.250
mg/l+200
mg/L AA
0
25
50
75
100
control 200 mg/L
AA
0.5 mg/L
DQ
0.5 mg/L
DQ +
200 mg/L
AA
1.25
mg/L DQ
1.25
mg/L DQ
+ 200
mg/L AA
2.5 mg/L
DQ
2.5 mg/L
DQ +
200 mg/L
AA
5 mg/L
DQ
5 mg/L
DQ +
200 mg/L
AA
*
* * *
* *
% m
alfo
rma
tion
s
paraquat diquat
0
25
50
75
100
control 0.06 mg/L 0.12 mg/L 0.25 mg/L
n= n=n=
0
25
50
75
100
control 0.25 mg/L 0.5 mg/L 1.25 mg/L 2.5 mg/L
n=57 n=15 n=46 n=45 n=17
% m
alfo
rma
tion
s
paraquat diquat A.
B.
control 0.06 mg/L 0.12 mg/L 0.25 mg/L control 0.25 mg/L 0.5 mg/L 1.25 mg/L 2.5 mg/L
control 200 mg/L 0.06 0.125 0.125 0.25 0.25 control 200 0.5 0.5 1.25 1.25 2.5 2.5 5 5 AA mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L +200 +200 +200 +200 +200 +200 mg/L AA mg/L AA mg/L AA mg/L AA mg/LAA mg/LAA
STUDY No.6 – EMBRYOTOXICITY AND OXIDATIVE STRESS IN FROGS
EXPOSED TO PESTICIDES
99
GS
H n
mol
/mg
prot
ein
control
24 hod 48 hod 72 hod 96 hod
4
8
12
16
20
24*
*
* * *
* *
24 hours 48 hours 72 hours 96 hours G
ST
nm
ol/m
in/m
g pr
otei
n
control
24 hod 48 hod 72 hod 96 hod0
10
20
30
40
50
60
*
* * *
* *
24 hours 48 hours 72 hours 96 hours
Figure 10A.: Profile of glutathione and glutathione-S-transferase during the early phases of
Xenopus laevis development. Asterisks indicate the statistically significant differences from the 24-
hour-activity/level [LSD test; * = p<0.05; *** = p<0.001].
The level of glutathione (Fig. 10A) significantly increased during the 96hours of
development from 6 to 15 nmoles/mg protein from the first to the fourth day of
development, interestingly the most distinct increase was obvious in the first two
days of development. Similarly, almost threefold significant increase of enzymatic
activity was documented in case of GST. Its activity increased linearly from 15 to
40 nmoles/min/mg protein from the first to the fourth day of development.
CHAPTER 8
100
G
R n
mol
NA
DP
H/m
in/m
g pr
otei
n
control
24 hod 48 hod 72 hod 96 hod0
1
2
3
4
5
6
7
*
* * *
* *
24 hours 48 hours 72 hours 96 hours G
Px
nmol
NA
DP
H/m
in/m
g pr
otei
n
control
24 hod48 hod
72 hod96 hod
10
20
30
40
24 hours 48 hours 72 hours 96 hours G
Px
nm
ol N
AD
PH
/min
/mg
pro
tein
Figure 10B.: Profile of glutathione reductase and glutathione peroxidase during the early phases
of Xenopus laevis development. Asterisks indicate the statistically significant differences from the
24-hour-activity/level [LSD test; * = p<0.05; *** = p<0.001].
Also the activity of GR significantly increased from 2 to 4 nmoles
NADPH/min/mg protein during the 96 hours of development (Fig. 10B). On the
other hand, GPx, CAT and SOD activities did not change significantly during the
first 96 hours of development. And finally, the level of lipid peroxides was high on
the first day and decreased during the period of the experiment (Fig. 10C), which
might be related to the greater activities of the antoxidative enzymes.
STUDY No.6 – EMBRYOTOXICITY AND OXIDATIVE STRESS IN FROGS
EXPOSED TO PESTICIDES
101
CA
T u
mol
H2O
2/m
in/m
g pr
otei
n
control
24 hod48 hod
72 hod96 hod
0
1
2
3
4
CA
T µ
mol
H2O
2/m
in/m
g p
rote
in
24 hours 48 hours 72 hours 96 hours
TB
AR
S n
mol
/mg
prot
ein
control
24 hod48 hod
72 hod96 hod
0,0
0,4
0,8
1,2
1,6
*
TB
AR
S n
mo
l/mg
pro
tein
1.6
1.2
0.8
0.4
0
24 hours 48 hours 72 hours 96 hours
*
SO
D n
mol
NB
T/m
in/m
g pr
otei
n
control
24 hod48 hod
72 hod96 hod
2
4
6
8
10
12
24 hours 48 hours 72 hours 96 hours
SO
D n
mo
l NB
T/m
in/m
g p
rote
in
Figure 10C.: Profile of catalase and superoxidedismutase together with content of lipid
peroxides during the early phases of Xenopus laevis development. Asterisks indicate the
statistically significant differences from the 24-hour-activity/level [LSD test; * = p<0.05].
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Biomarker responses to pesticides exposure
Glutathione, glutathione-S-transferase, glutathione reductase and glutathione
peroxidase together with lipid peroxidation were chosen as the biomarkers for the
evaluation of pesticides effects and the same set of biomarkers was used also for
the experiments with external addition of ascorbic acid.
Only enzymatic antioxidants GST and GR responded to pesticides exposures with
significant tendency within the first four days of frog development (see the
summarizing Table 4). Activity of GST was sensitive to DQ exposure and was
significantly enhanced from the day two of the exposure. On the other hand, there
were no significant modulations of GST activity observed in experiment with PQ.
Addition of AA caused significant decrease of GST activity when compared to the
single pesticides exposure.
GR activity responded to both DQ and PQ exposures. DQ exposure caused
significant enhancement of GR activity from the day three of the exposure in a
wide range of its concentrations. In case of PQ exposure, GR activity responded
from the day two of exposure; there were significant modulations after PQ
exposures and their mixtures with AA, the addition of AA mostly reversed the
modulation when compared with the single PQ effect.
The level of lipid peroxidation as the parameter of damage to lipids was not
affected by the pesticides exposure in exception of toxic concentration of PQ
causing its decrease.
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Table 4: Biomarker significant responses (↑↓) after single diquat (DQ) and paraquat (PQ)
exposures and the effects of addition of ascorbic acid compared to the single pesticide exposure.
DQ 24 hours 48 hours 72 hours 96 hours
GSH - ↑DQ 1.25mg/L ↓AA+DQ 1.25mg/L
- -
GST - ↑DQ 0.12-2.5mg/L ↓AA+DQ 1.25mg/L
↑DQ 0.12-2.5 mg/L ↓AA+DQ 0.5mg/L
↑DQ 0.06-0.5mg/L
GPx ↑DQ 0.03-0.06 mg/L ↑AA+DQ 0.5mg/L
- - -
GR ↑AA+DQ 2.5 mg/L - ↑DQ 0.25-1.25 mg/L ↑AA+DQ 5 mg/L
↑DQ 0.06 mg/L
TBARS - - - -
PQ 24 hours 48 hours 72 hours 96 hours
GSH - - - ↓AA+PQ 0.12mg/L
GST - - - -
GPx - - - ↓AA+PQ 0.12mg/L
GR - ↑PQ 0.06 mg/L ↑AA+PQ 0.06mg/L
↑PQ 0.03-0.06 mg/L ↓PQ 0.12 mg/L ↓AA+PQ 0.03-0.06 mg/L ↑AA+PQ 0.12 mg/L
↑PQ 0.06 mg/L ↓AA+PQ 0.06 mg/L
TBARS - ↓PQ 0.12 mg/L ↓AA+PQ 0.06 mg/L
- -
8.4 Discussion
This study confirmed the acute embryotoxic and teratogenic effects of paraquat
and diquat using FETAX test on the early developmental stages of Xenopus laevis.
Based on the calculated teratogenic index, the paraquat can be considered
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embryotoxic rather than teratogenic compound (TI=1.4), on the other hand diquat
can be characterized as the compound with moderate teratogenic properties
(TI=2.2). Concentration-related growth inhibition was documented in larvae
exposed to both model pesticides with the LOEC at concentration 0.125 mg/L PQ
and 0.5 mg/L DQ, respectively. Significant concentration-dependend induction of
malformations was observed. Interestingly, with increasing pesticide concentration
more severe malformations, as for example the complex malformation of the axial
part of the body connected with gut abnormalities, were documented. Accordingly,
inductions of malformations such as ventral tail flexure, abnormal somites,
necrotized myocytes together with high embryolethality and growth retardations
were documented in a similar study with paraquat exposure to X.laevis (Vismara et
al. 2000).
Our experiment was further focused on characterization of the developmental
profile of chosen detoxification and antioxidative compounds related to oxidative
stress. In case of glutathione, glutathone-S-transferase and glutathione reductase, a
strong activation was detected within the first four days of frog development.
Similarly, glutathione content increased during the metamorphic progression of the
giant prawn larvae, Macrobrachium rosenbergii, and developing grass shrimp,
Palaemonetes pugio (Dandapat et al. 2003; Winston et al. 2004). The accrual of
the glutahione level was also documented during the development of toad Bufo
arenarum embryos (Anguiano et al. 2001). Our data correspond to the studies
presenting a gradual increase of antioxidant enzyme activities during
embryogenesis, accompanied by a sudden rise of these enzymes in freshly hatched
larvae of aquatic invertebrates - prawn Macrobrachium malcolmsonii and grass
shrimp Palaemonetes pugio (Arun and Subramanian 1998; Winston et al. 2004).
The antioxidants glutathione peroxidase, catalase and superoxide dismutase did
not show any significant change during the early phases of development in our
STUDY No.6 – EMBRYOTOXICITY AND OXIDATIVE STRESS IN FROGS
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105
experiment. This result differes from studies performed on freshwater fish (Aceto
et al. 1994) where an important role of catalase in developing embryos was shown
(Mourente et al. 1999). On the other hand, in the same study, Mourente et al.
(1999) showed that the titers of two other detoxification enzymes (GST and SOD)
reached their highest levels in eggs, compared to later developmental stages. The
statistically significant decrease of MDA content during the early development,
which was further reported in that study, corresponds also to our results, where the
same decline was observed, indicating sufficient natural embryonic potential to
scavenge and detoxify free radicals in their tissues.
The results from the parallel experiment with external addition of ascorbic acid
confirm significantly lower embryotoxicity, teratogenity and induction of growth
retardations caused by bipyridyl pesticides in Xenopus laevis embryos when
ascorbic acid is supplied to the experimental media. This corresponds to a similar
study with paraquat on X.leavis (Vismara et al. 2001). Our study showed also the
embryotoxic and teratogenic effects of diquat and has involved more parameters
into the study design to provide more detailed research in the oxidative mechanism
of pesticides-teratogenesis in frogs. Ascorbic acid acts as an important water-
soluble antioxidant that reduces sulphydryls, scavenges free radicals and can
protect against endogenous oxidative DNA damage. The protective role of
ascorbic acid was confirmed also in studies with mammals where the ascorbic acid
significantly protected the development of sperm culture (Fraga et al. 1991). In
mammalian embryo culture, addition of another antioxidant SOD to the medium
increased embryonic SOD activity and SOD or enzymatic antioxidant CAT also
blocked oxidative damage as embryolesions and embryolethality caused by the
anticonvulsant drug phenytoin and tobacco carcinogen benzo[a]pyrene (Wells et
al. 1997).
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Further, in our study, there were different biomarker-responses after pesticides
exposure observed, when compared to the developmental profile of these
biomarkers. Various concentrations of pesticides caused significant enhancements
of levels or activities of biomarkers at various days of FETAX test. These effects
were in some cases even increased by the addition of ascorbic acid to the
experimental media or reduced assuming that the mixture with ascorbic acid
exhibited the opposite effect to the single-pesticide. These modulations were
observed depending on the phases of development and the concentration of single
pesticide. Generally, enhancement of antioxidant activity was observed after lower
PQ and/or DQ exposure concentrations. Addition of ascorbic acid to experiments
with pesticides enhanced survival of embryos at higher pesticides concentrations
and enabled to assess the biomarkers also at these higher concentrations. However,
mostly decreasing tendency of antioxidant activities was observed at greater toxic
concentrations when compared with lower pesticides concentrations. This was for
example documented by responses of GSH and GPx to PQ exposure. Significant
decreases of GSH level and GPx activity were observed after four days of
exposure to toxic concentration (0.12 mg/L) of PQ causing about 50% lethality
(see Fig. 7A).
Glutathione reductase and glutathione-S-transferase were the most sensitive
biomarkers responding to very low concentrations of the pesticides, ascorbic acid
and their mixture. In experiment with diquat, significantly different responses of
biomarkers in comparison with their basal levels were observed already on the first
day but in experiment with paraquat the responses were not detectable till the
second day of exposure. On the other hand, most biomarkers did not differ from
control levels on the first day of development; they responded after enhanced need
for detoxification in later phases of development.
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107
Interestingly, even though there were lethal effects and malformations
documented, there was not significant level of lipid peroxides detected.
Surprisingly, there was significant decrease of lipid peroxidation after exposure to
low concentrations of PQ.
Most of the biomarkers responded already at lower (sub-lethal) tested
concentrations especially in case of PQ exposures. In these cases, modifications of
biomarkers preceded any signs of toxicity or malformations.
Ascorbic acid significantly prevented embryotoxicity, growth retardations,
malformations and modulated detoxification and antioxidative parameters in
embryos. The study has confirmed the involvement of oxidative stress in
teratogenity and embryotoxicity of bipyridyl pesticides and indicated the role of
selected antioxidative and detoxification compounds in this process. Further work
is, however, needed to better characterize the mechanisms involved in the normal
development.
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9 CHAPTER 9
GENERAL DISCUSSION
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110
Biomarker-measurements have been applied as common parameters in studies
involving various experimental organisms including plants, fish, birds and frogs
exposed to several important groups of environmental stressors. Various stressors
can lead to the production of ROS (Rijstenbil et al. 1994). The role of ROS in
toxic action of many anthropogenic compounds together with possible
mechanisms of involvement of these compounds in the formation of oxidation
stress have been documented (Halliwell and Gutterdige 2007). Correspondingly,
Studies within this dissertation were focused on potential oxidative and
glutathione-involving detoxification of polyaromatic compounds, pesticides,
complex cyanobacterial biomass, heavy metals and vaccination. The role of
antioxidative and detoxification compounds in the reaction of organisms to various
stressorswas studied; the measurement of biomarkers has been employed also after
multistressor exposure. Both mature and juvenile organisms have been studied to
assess the role of chosen biomarkers at different stages of development. Sensitivity
of selected conventional biomarkers of biotransformation, detoxification and
oxidative stress was studied. The research also aimed to determine which of the
parameters responded most strongly and frequently preceding signs of toxicity in
model ecotoxicological organisms.
9.1 Biomarkers after exposure to PAHs in plants
This study brought new information on the phytotoxicity and biochemical effects
of important organic contaminants – polycyclic aromatic compounds - and
relatively poorly characterized group of their N-heterocyclic derivatives within or
close to the environmentally relevant concentration range. The parameters
reflecting acute phytotoxicity (germinability, weight and length of roots and
hypocotyle) used for testing responded differently to parental PAHs and their
heterocyclic compounds. Generally, NPAHs were significantly more phytotoxic
than parent PAHs, which corresponds to the results on higher plants and algae
GENERAL DISCUSSION
111
(Gissel-Nielsen et al. 1996; Van Vlaardingen et al. 1996). On the contrary, study
with Sinapis alba, Trifolium pratense and Lolium perenne (Sverdrup et al. 2003)
reported only minor differences between the toxicity of homocyclic and
heterocyclic PAHs. On the other hand, the effects of both PAHs and NPAHs on
biochemical parameters were comparable and all tested chemicals modulated
activity of plant detoxification and antioxidative enzymes. The most sensitive
biomarker from those analyzed was the activity of glutathione reductase.
Glutathione reductase was similarly enhanced in studies with Arabidopsis and
Lemna gibba (Alkio et al. 2005; Babu et al. 2005). Biochemical changes were in
general more sensitive and occurred already at concentrations about an order of
magnitude lower than those causing signs of phytotoxicity.
9.2 Biomarkers of exposure to cyanobacterial biomass in fish and birds and
multiple–stressor exposure in birds
Cyanobacteria are known to produce secondary metabolites, which have been
recognized as human and animal health hazards, since they have been shown to
cause adverse effects in various organisms including fish or birds (Malbrouck et al.
2003; Krienitz et al. 2003; Lugomela et al. 2006). A set of studies focused on
cyanobacterial exposure has been performed on model organisms aiming to
describe the glutathione-mediated detoxification and oxidative mechanisms of
cyanobacterial toxicity in wild-living fish and birds coming potentially to contact
with this stressor.
Firstly the effect of cyanobacterial biomass on modulation of glutathione-related
biomarkers has been investigated in common and silver carp in four and nine
week-study simulating the natural situation. Benthophagous common carp and
phytophagous silver carp responded differently to the cyanobacterial exposure and
different adaptation of the detoxification system to cyanobacteria has been
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suggested. Species-dependent variability in fish biochemical responses to
microcystins exposure was correspondingly reported in Prieto et al. (2006). The
correspondence between tripeptide glutathione and the catalysing enzyme
glutathione-S-transferase has been confirmed in fish model similarly to the
published connection between these compounds in conjugation of cyanotoxins
microcystins in plants, invertebrates, fish embryos or mature fish (Wiegand and
Pflugmacher 2001; Pflugmacher et al. 1998). Activity of glutathione reductase was
the most sensitive biomarker in this study. Correspondingly, crucial role of this
enzyme in maintaining the GSSG/GSH homeostasis in experiments with toxic
cyanobacteria in fish was documented in studies of Li et al. (2003) and Jos et al.
(2005). Modulations of biomarkers, especially GR, GST and GSH, have confirmed
an important role of oxidative stress in the toxicity of complex cyanobacterial
bloom in this study. Changes of biomarkers preceded any signs of toxicity.
Subsequently the glutathione-mediated detoxification and oxidative mechanisms
of toxicity of the natural cyanobacterial biomass have been studied on Japanese
quails in a 10-day and 30-day experiment. The study brought unique sublethal
ecotoxicological data from the first controlled experiments with the exposure to
cyanobacterial biomass in birds. Birds reacted to the cyanobacterial exposure as to
xenobiotics, which was documented by the activation of the general detoxification
mechanisms. It was moreover shown that antioxidative and detoxification
mechanisms are interdependent and can get adapted to cyanobacterial exposure
with increasing time of exposure. It was suggested that the prolonged exposure to
cyanotoxins activated the microcystins detoxification and elimination, which was
documented by six times lower accumulation at the longer exposure time.
Glutathione and glutathione-S-transferase were the most sensitive biomarkers. This
is in correspondence with their important role in conjugation and elimination of
microcystins in birds as discussed by Pflugmacher et al. (1998) and Wiegand et al.
GENERAL DISCUSSION
113
(2001). Our study also documented the cyanobacterial biomass-caused induction
of lipid peroxidation in bird liver, heart and brain indicating them as some of the
targets of the cyanobacterial toxicity corresponding to cyanotoxins
bioaccumulation reported not only in liver, but also in intestines, kidneys, brain,
heart, gonads and muscles of fish and mammals (Kankaanpaa et al. 2005;
Cazenave et al. 2006; Adamovsky et al. 2007; Kagalou et al. 2008).
Finally, 30-day exposure to cyanobacterial biomass, lead and vaccination strain
was further performed to simulate the environmental conditions and to test the
hypothesis of modulation of toxic effects by combined exposure. General
stimulation of antioxidative system with the greatest modulations of sublethal
parameters in the individuals from the groups with combined exposures was
shown. The results of microcystins accumulation (Pikula et al. 2010) support the
diverse responses of oxidative stress biomarkers. Briefly, there was higher
accumulation of microcystins documented in groups of combined exposures with
lead than in combinations without lead, which may indicate somewhat greater
uptake of cyanobacterial metabolites in birds weakened by lead exposure. These
higher levels of microcystins along with toxic effects of lead (and effects of
immunological challenge) could have contributed to the greatest modulations of
almost all examined biomarkers in these groups. The role of glutathione and
glutathione-S-transferase in detoxification of cyanobacterial metabolites has been
confirmed correspondingly to previously discussed experiment with quails and
also the study with fish exposed to cyanobacterial biomass. The correspondence of
GSH and GST with GPx documents the cooperation of the enzymes and the
crucial role of GSH in detoxification after exposure to lead and cyanobacterial
biomass (Berglund et al. 2007; Pašková et al. 2008). Moreover, significance of
these biomolecules in the protection from harmful effects in birds has been
documented similarly to study of Douglas-Stroebel et al. (2004). The most
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significant changes after multiple stressors exposure confirm our hypothesis that
effects of cyanobacterial biomass, lead and immunological challenge may combine
to enhance their influence.
Modulations of biomarkers have confirmed an important role of oxidative stress in
the toxicity of complex cyanobacterial bloom in studies on fish and birds, and it
has also demonstrated that biochemical parameters (especially GR, GST, and
GSH) may serve as sensitive early markers of adverse effects in these species
preceding any signs of toxicity. Cyanobacteria are capable to induce oxidative
stress responses in fish and birds linked with activation or inhibition of
detoxification compounds. The generation of oxidative stress combined with
insufficiency of defense mechanisms could in sensitive species at prolonged
exposure potentially result in effects on the health status, especially if other
stressors are involved at the same time. This is often the case in the environment
and it has been modelled by the multistressor exposure to birds.
9.3 Biomarkers of exposure to pesticides in aquatic invertebrates, fish and
amphibians
A review of teratogenicity and embryotoxicity in aquatic organisms after pesticide
exposure together with experiments with African clawed frog embryos exposed to
pesticides have been performed to study the role of oxidative stress in the process
of development.
Our review of the existing available literature showed that a broad range of
pesticides, representing several different chemical classes, induce variable toxic
effects in aquatic species. Many pesticides have been documented to induce
embryotoxicity and teratogenicity in non-target aquatic biota such as fish,
amphibians and invertebrates. However, only few experimental studies with
GENERAL DISCUSSION
115
pesticides have directly linked developmental toxicity with key oxidative-stress
endpoints, such as lipid peroxidation, oxidative DNA damage or modulation of
antioxidant mechanisms in aquatic organisms (Wiegand et al. 2000; Wiegand and
Pflugmacher 2001; Anguiano et al. 2001; Snyder and Mulder 2001; Ferrari et al.
2009). On the other hand, it has been documented in many reports that pesticide-
related oxidative damage occurs in exposed adult fish, amphibians and
invertebrates. Moreover, the contribution of oxidative stress to the toxicity of
pesticides has been emphasized in several recent review papers dealing with this
topic (Valavanidis et al. 2006; Monserrat et al. 2007; Debenest et al. 2010). It has
been concluded in our review that oxidative stress is a highly important
mechanism in pesticide-induced reproductive or developmental toxicity and
further research is needed to more fully elucidate the effects that pesticides have,
and the environmental risks they pose in the early development of aquatic
organisms.
The experiments with the African clawed frog (Xenopus laevis) embryos
confirmed the acute embryotoxic and teratogenic effects of paraquat and diquat
corresponding to study of Vismara et al. (2001) and brought new information
about biomarker-responses to pesticides in the early developmental phases of
frogs. The developmental profile of chosen biomarkers has been evaluated during
the first 96 hours of frog development. The strong activation of glutathione,
glutathione-S-transferase and glutathione reductase in the first 96 hours was found
out in the X.laevis larvae and the most important role of these three biomarkers
within the range of biomarkers tested has been suggested. The role of these
biomarkers has been similarly documented in several studies with invertebrates,
fish and amphibians (Dandapat et al. 2003; Winston et al. 2004; Mourente et al.
1999; Anguiano et al. 2001). Significant modulations of biomarkers were also
documented after paraquat and diquat exposure as indicators of mechanism of
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toxic action of these bipyridyl pesticides. These modulations depended on the
phases of development and the concentration of each pesticide. They were mostly
activated in later phases of development where the exposure to pesticides resulted
in enhanced need for detoxification. Generally, biomarkers were elevated at lower
concentrations preceding signs of embryotoxicity or teratogenity; higher
concentrations were mostly toxic making the measurement of biomarkers due to
low number of survivals impossible. Glutathione reductase was the most sensitive
biomarker responding to very low concentrations of the pesticides. Elevated GR
together with GST and CAT was correspondingly documented in study with Bufo
arenarum exposed to pesticides (Anguiano et al. 2001; Ferrari et al. 2009).
The addition of ascorbic acid was also investigated to confirm the oxidative
mechanism of teratogenesis of bipyridyls. The results of significantly lower
embryotoxicity, teratogenity and fewer growth retardations after bipyridyls
exposure, when the ascorbic acid was supplied to the experimental media, were in
accordance to the article published by Vismara et al. (2001). Lower embryonic
mortality at toxic concentrations of pesticides due to addition of ascorbic acid
enabled measurement of biomarkers in these embryos. In those cases, biomarkers
were often decreased probably as the result of pesticides toxicity and deficiency in
energy. Moreover, the experiment with involvement of detoxification and
oxidative stress markers brought new approach in assessment of oxidative
mechanism of action. Especially in case of paraquat, most of the biomarkers
responded already at lower (sub-lethal) tested concentrations preceding any signs
of embryotoxicity or teratogenity. Glutathione reductase and glutathione-S-
transferase were the most sensitive biomarkers responding to very low
concentrations of the pesticides, ascorbic acid and their mixture.
To conclude the experiment, ascorbic acid significantly prevented embryotoxicity,
growth retardations, malformations and enhanced or modulated detoxification
GENERAL DISCUSSION
117
parameters in embryos confirming thus the oxidative stress involvement of
teratogenity and embryotoxicity of bipyridyl pesticides
9.4 Conclusions
To conclude this dissertation work, biomarkers reflect toxic mechanisms and
major processes protecting tissues from oxidative stress. They sensitively respond
to low concentrations of stressors preceding any signs of toxicity and can be thus
successfully used as sensitive markers of adverse effects of various environmental
stressors. In particular, glutathione reductase was shown to be the most sensitive
biomarker of sublethal toxicity in plant exposure to polyaromatics and fish
exposure to cyanobacterial biomass. Glutathione reductase was also the most
frequently responding biomarker in pesticides-exposed frog embryos and played
an important role in development of aquatic organisms next to glutathione and
glutathione-S-transferase. In case of cyanobacterial biomass-exposed Japanese
quails, glutathione and glutathione-S-transferase responded most strongly to the
exposure. As presented in this work, biomarkers can be also valuable beneficial-
parameters in ecotoxicological studies complementing the results. They provide
evidence about the general activation of the antioxidative system in exposed
organisms. This documents the greater need of antioxidative protection in the
studied organs and their ability to produce molecules protecting cells against
adverse oxidation processes. However, their responses strongly depend on many
factors including experimental design, sensitivity of the model species, stages of
development and properties of the tested compound. New approaches as
toxicogenomics represent promising tool when identifying sensitive markers in
ecotoxicology. These resources-demanding multibiomarker systems could be
suggested as more sensitive and informative than selected conventional biomarkers
used in this thesis. On the other hand, because of the confusing specificity of
genomic biomarkers (Zhang et al. 2012) and the very high cost of toxigenomic
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analysis for each sample the information on conventional biomarkers is still
valuable, cost-effective and brings important knowledge.
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ANNEXES
Paper I.
Pašková, V., Hilscherová, K., Feldmannová, M. and Bláha, L. (2006).
Toxic effects and oxidative stress in higher plants exposed to polycyclic aromatic hydrocarbons and their N-heterocyclic derivatives.
Environmental Toxicology and Chemistry, Vol. 25, No. 12, pp. 3238-3245.
3238
Environmental Toxicology and Chemistry, Vol. 25, No. 12, pp. 3238–3245, 2006� 2006 SETAC
Printed in the USA0730-7268/06 $12.00 � .00
TOXIC EFFECTS AND OXIDATIVE STRESS IN HIGHER PLANTS EXPOSED TOPOLYCYCLIC AROMATIC HYDROCARBONS AND THEIR
N-HETEROCYCLIC DERIVATIVES
VERONIKA PASKOVA,† KLARA HILSCHEROVA,*†‡ MARIE FELDMANNOVA,†‡ and LUDEK BLAHA†‡†RECETOX—Research Centre for Environmental Chemistry and Ecotoxicology, Masaryk University, Kamenice 3, CZ 625 00 Brno,
Czech Republic‡Institute of Botany, Academy of Sciences of the Czech Republic, Kvetna 8, CZ 603 65 Brno, Czech Republic
(Received 3 April 2006; Accepted 22 June 2006)
Abstract—N-heterocyclic derivatives of polycyclic aromatic hydrocarbons (NPAHs) are widespread concomitantly with their parentanalogues and have been detected in air, water, sediments, and soil. Although they were shown to be highly toxic to some organisms,our understanding of their occurrence, environmental fate, biological metabolism, and effects is limited. This study evaluated toxiceffects of three homocyclic aromatic hydrocarbons (PAHs—phenanthrene, anthracene, fluorene) and their seven N-heterocyclicderivates on higher terrestrial plants Sinapis alba, Triticum aestivum, and Phaseolus vulgaris. Germinability, morphological end-points, parameters of detoxification, and antioxidant components of plant metabolism as well as lipid peroxidation were studied inacute phytotoxicity tests. Phytotoxicity of NPAHs was generally more pronounced than the effects of parent PAHs, and it significantlydiffered with respect to the structure of individual NPAHs. Sinapis alba and T. aestivum were more sensitive plant species thanP. vulgaris. Chemicals with the strongest inhibition effect on germination and growth of plants were phenanthridine, acridine,benzo[h]quinoline, and 1,10- and 1,7-phenanthroline. All tested chemicals significantly induced activities of detoxification andantioxidant enzymes (glutathione reductase, glutathione peroxidase, and glutathione-S-transferase) at nanomolar to low micromolarconcentrations. Levels of reduced glutathione were induced by all tested chemicals except 1,10- and 4,7-phenanthroline. Furthermore,fluorene, carbazole, acridine, phenanthrene, phenanthridine, benzo[h]quinoline, and 1,7-phenanthroline significantly increased lipidperoxidation. The results of our study newly demonstrate significant toxicity of NPAHs to plants and demonstrate suitability ofmultiple biomarker assessment to characterize mechanisms of oxidative stress and to serve as an early warning of phytotoxicity invivo.
Keywords—Phytotoxicity N-heterocyclic polyaromatic hydrocarbons Lipid peroxidation Detoxification enzymesOxidative stress
INTRODUCTION
Most research on polycyclic aromatic hydrocarbons(PAHs), an important group of ubiquitous environmental pol-lutants, has been focused on homocyclic compounds. However,two-thirds of the known aromatic compounds are heterocyclicwith oxygen, sulfur, and/or nitrogen in-ring substitutions ofone or more carbon atoms. Environmental and toxicologicalimportance of nitrogen heterocyclic derivatives of PAHs(NPAHs) has been recognized [1]. Large differences in chem-ical characteristics and biological reactivity are likely to existamong PAHs and their NPAHs. The substitution of a carbonatom by a nitrogen atom makes the substances more polar andincreases their water solubility [2]. The sources of NPAHs aresimilar to those of PAHs, including coal production, incom-plete combustion of organic matter, fuel exhaust, petroleum-derived products, and some industrial processes [3]. Finlayson-Pitts and Pitts [4] reported formation of NPAHs during theincomplete combustion of nitrogen-containing organic sub-stances in the presence of NOx. The presence of NPAHs hasbeen documented in air, groundwater, and both marine andfreshwater environments [1]. Concentrations of NPAHs foundin the environment are reported to be one to two orders ofmagnitude lower than PAHs concentrations [5], but their bi-ological effects can be of similar magnitude. While the toxicity
* To whom correspondence may be addressed(hilscherova@recetox.muni.cz).
of PAHs has been extensively investigated (particularly theeffects in animals and humans [6]), relatively little is knownabout the toxicity of PAH derivatives such as NPAHs. SomeNPAHs are known mutagens and/or carcinogens [7]. The fewexisting ecotoxicological studies focused primarily on the ef-fects in prototypical bioassay organisms such as algae, inver-tebrates, and fish [8,9].
However, it has been shown previously that organic pol-lutants including PAHs accumulate in vegetation [10], and theycan cause significant phytotoxicity [11]. Toxicity of PAHs wasobserved in multiple plant species; the documented effectsincluded inhibition of germination, growth, and photosynthesis[12]. Increasing phytotoxicity was shown after photomodifi-cation of parent compounds [13]. Plants have also been suc-cessfully used for phytoremediations of the sites contaminatedwith both metals and organic pollutants [14,15]. However,there is only limited knowledge about the toxicity mechanismsof PAHs and their derivatives such as NPAHs both relative tocrop species or plants with potential use in remediation.
The role of oxidative stress in the phytotoxicity of severalinorganic [16] and organic chemicals [17] has been docu-mented. However, only lately has the research interest focusedin detail on the mechanisms of PAHs toxicity in plants. Arecent study with Arabidopsis thaliana exposed to phenan-threne has shown a correlation between morphological signsof phytotoxicity and induction of oxidative stress [18]. Anincrease in activities of antioxidant enzymes as well as in
PAH and NPAH phytotoxicity and oxidative stress responses Environ. Toxicol. Chem. 25, 2006 3239
Fig. 1. Chemical structures of tested compounds.
glutathione levels was observed in Fontinalis antipyretica ex-posed to benzo[a]pyrene and benzo[a]anthracene [19] or to amixture of PAHs in the field [20]. Further, synergistic effectsof environmentally relevant concentrations of metal and PAH(copper and dihydroxyanthraquinone) on induction of oxida-tive stress have recently been reported in the studies withaquatic plant Lemna gibba [21].
The structure of the antioxidant system protecting plantcells against reactive oxygen species is generally well under-stood. Catalase, ascorbase, glutathione, monodehydroascor-bate reductase, ascorbate peroxidase, dehydroascorbate reduc-tase, glutathione reductase, glutathione peroxidase, and glu-tathione transferase are the most important biomolecules play-ing a role in a plant antioxidative system and maintainingcellular homeostasis [22]. The physiological levels of gluta-thione and antioxidant enzymes activities in plant cells areoften induced in response to various stress conditions includingchemically induced oxidative stress, and they can be used assuitable early biomarkers of toxicity. However, pronouncedintoxication can lead to irreversible protein degradation anddamage of the defense system that might result in the growthinhibition or necrosis [22].
To derive internationally acceptable toxicity results, the Or-ganization for Economic Cooperation and Development(OECD) guidelines (a collection of the most relevant inter-nationally agreed testing methods for the characterization ofpotential hazards of chemical substances and preparations/mix-tures supervised by OECD) are recommended for testing ofchemicals. Although Guideline 208 (terrestrial plants, growthtest) has been used for relatively long time since 1984 [23],there is surprisingly limited information on the phytotoxicityof other substances than pesticides and metals. In the presentstudy, we analyzed toxicity of PAHs and their derivatives toplants using the slightly modified OECD Guideline 208 andinvestigated several biochemical responses related to oxidativestress. The test species were selected to represent differentplant classes and also groups with different carbon metabolismincluding both dicotyledonous Phaseolus vulgaris and Sinapisalba (C3 metabolism) and monocotyledonous plants Triticumaestivum (C4 metabolism). We first evaluated standard mor-phological toxicity parameters such as germination and hy-pocotyl and root elongation. We further compared the acutephytotoxicity results with several biochemical responses in-cluding concentrations of glutathione (GSH), activities of glu-tathione-S-transferase (GST), glutathione peroxidase (GPx),and glutathione reductase (GR) and the level of lipid perox-idation. The effects of parental homocyclic PAHs (phenan-threne, anthracene, and fluorene) were compared with their N-heterocyclic derivates (phenanthridine, 1,10-phenanthroline,4,7-phenanthroline, 1,7-phenanthroline, benzo[h]quinoline,acridine, carbazole) to investigate structure–toxicity relation-ships.
MATERIALS AND METHODS
Chemicals
Phenanthrene (CAS 85-01-8), phenanthridine (CAS 229-87-8), 1,10-phenanthroline (CAS 66-71-7), 1,7-phenanthroline(CAS 230-46-6), 4,7-phenanthroline (CAS 230-07-9), ben-zo[h]quinoline (CAS 91-22-5), anthracene (CAS 120-12-7),acridine (CAS 260-94-6), fluorene (CAS 86-73-7), and car-bazole (CAS 86-74-8) as well as biochemicals and enzymeswere purchased from Sigma-Aldrich (Prague, Czech Repub-lic). The chemical structures of tested PAHs and NPAHs are
given in Figure 1. Other chemicals used for preparation ofmedia as well as solvent (dimethylsulfoxide [DMSO]) wereof the highest quality available.
Bioassay
The germination and root elongation test was performedaccording to the OECD 208 guideline and standard norm STN(Slovak technical norm) 83 8303 with some minor modifica-tions. The test was conducted on glass Petri dishes with fil-tration paper saturated with 5 ml standard media (294 mg/LCaCl2·2H2O, 123 mg/L MgSO4·7H2O, 65 mg/L NaHCO3, 5.8mg/L KCl; final pH 7.8 � 0.2). The compounds were dosedin solvent DMSO (final concentration 0.5% v/v). Then five toseven seeds rinsed with normal saline were randomly placedon each Petri dish (seven seeds for mustard [S. alba] and wheat[T. aestivum] and five for bean [P. vulgaris]). Six Petri disheswere used for each tested concentration and controls. The seedswere incubated for 96 h in dark at 23 to 25�C. At the end ofexposure, the number of germinated seeds and lengths andweights of the roots and hypocotyls were recorded. The seed-lings were further homogenized on ice in phosphate-bufferedsaline (PBS; pH 7.2), 1 g fresh weight in 5 ml of PBS, andthe supernatant was collected after centrifugation (15 min at3,000 g at 4�C) and stored frozen at �80�C until biochemicalanalyses.
Biomarker methods
Glutathione-S-transferase activity was measured spectro-photometrically at 340 nm using 0.32 mM 1-chloro-2,4-dini-trobenzene and 4 mM GSH in PBS [24]. Specific activity wasexpressed as nmoles of evolved product per minute per mil-ligram protein. The concentration of reduced glutathione wasdetermined by a spectrophotometric method using 5,5�-di-thiobis-2-nitrobenzoic acid as a substrate [25]. Plant sampleswere treated with trichloracetic acid (2.27% v/v) and centri-fuged (6,000 g for 10 min at 4�C). Supernatant was mixedwith Tris (tris[hydroxymethyl]aminomethane)-HCl buffer (0.5M Tris, 0.0125 M ethylenediaminetetraacetic acid [EDTA], pH
3240 Environ. Toxicol. Chem. 25, 2006 V. Paskova et al.
Table 1. Summary of the effects of N-heterocyclic polyaromatic hydrocarbons and their unsubstituted analogues on morphological parametersin plants (— no effect; � statistically significant difference from control at �2 �M, �� at 0.2–2 �M, ��� at 0.02 �M; p 0.05)
PlantRoot
lengthHypocotyl
lengthRoot
weightHypocotyl
weightTotallength
Totalweight Germinability
Phenanthrene Triticum aestivumSinapis albaPhaseolus vulgaris
———
———
———
———
———
———
———
1,10-Phenanthroline T. aestivumS. albaP. vulgaris
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4,7-Phenanthroline T. aestivumS. albaP. vulgaris
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———
1,7-Phenanthroline T. aestivumS. albaP. vulgaris
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Benzo[h]quinoline T. aestivumS. albaP. vulgaris
—���
—
——
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Phenanthridine T. aestivumS. albaP. vulgaris
—�—
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Anthracene T. aestivum — — — — — — —S. albaP. vulgaris
—��
——
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—��
——
—��
——
Acridine T. aestivumS. albaP. vulgaris
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—
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—
��——
Fluorene T. aestivumS. albaP. vulgaris
———
———
———
———
———
———
———
Carbazole T. aestivumS. albaP. vulgaris
—��
—��
———
———
—�—
———
—���
8.9) and 0.6 �M 5,5�-dithiobis-2-nitrobenzoic acid and incu-bated for 5 min at room temperature. Absorbance was mea-sured at 420/680 nm, and the concentrations (nmol GSH/mgprotein) were calculated according to the standard calibrationof reduced GSH. Activity of GPx was determined from therate of nicotinamide adenine dinucleotide phosphate (NADPH)oxidation recorded as the decline in absorbance at 340 nm[26]. The reaction mixtures contained 3 mM GSH, 1 U GR (1U will reduce 1.0 �mole of oxidized glutathione per min atpH 7.6 at 25�C), and 0.15 mM NADPH in 0.1 M potassiumphosphate/1 mM EDTA buffer (pH 7). Substrate used for theassay was 1.2 mM butylhydroperoxide. Also, the activity ofGR was determined by spectrophotometric measurement ofNADPH oxidation [27]. Assays for GR activity were per-formed in microplates, and the reaction mixtures contained0.05 M potassium phosphate/1 mM EDTA buffer (pH 7.0), 1mM glutathione oxidized disodium salt (GSSG), 0.1 mMNADPH, and the seedling extract (0.25% v/v). Specific activ-ities of both GPx and GR were expressed as nmoles NADPHoxidized per minute per milligram protein. The level of lipidperoxidation in plant tissue was assessed as total thiobarbituricacid–reactive species (TBARS) [28]. The seedling extractswere mixed with trichloracetic acid (10% w/v) and butylatedhydroxytoluene (1% w/v) and centrifuged (1,500 g for 20 min).Supernatant was further mixed with 0.06 N HCl and 40 mMthiobarbituric acid prepared in 10 mM Tris (pH 7.4). The mix-ture was boiled in water bath for 45 min and then cooled toroom temperature. Absorbance of the sample was measured at550 nm, and the concentration of TBARS (nmol TBARS/mgprotein) was calculated according to the standard calibrationcurve generated with malondialdehyde prepared by acidic hy-
drolysis of 1,1,3,3-tetraethoxypropane. The protein concentra-tions were determined by method using Folin-Ciocalteu phenolreagent that forms with proteins red-colored complex mea-surable at 680 nm [29]. Bovine serum albumin was used as astandard for protein calibration.
The Tecan GENios microplate reader (TECAN, Mannedorf,Switzerland) was used for measurement of absorbance in allassays.
Statistical evaluation
Statistical analyses were performed with Statistica for Win-dows� 7.0 (StatSoft, Tulsa, OK, USA). Data normality andhomogeneity of variances were evaluated by Kolmogorov–Smirnov test and Levene’s test, respectively. One-way analysisof variance and the nonparametric Kruskal–Wallis test wereused for statistical comparisons. Only the results of the Krus-kal–Wallis test are presented in the Results section, as thevariances among some of the treatments were not homoge-neous. Values of p 0.05 were considered statistically sig-nificant for all tests.
RESULTS
The results of the acute phytotoxicity testing of PAHs andtheir N-heterocyclic derivatives are summarized in Table 1.Exposure of the plants to most NPAHs resulted in significantlylower germinability (Fig. 2 shows effects of phenanthridine,acridine, and benzo[h]quinoline as an example). Also, growth(determined as weight or length of roots and/or hypocotyle)was significantly affected by NPAHs at various concentrations(Table 1). The 1,7-phenanthroline was the most toxic from alltested compounds; it affected most of the parameters already
PAH and NPAH phytotoxicity and oxidative stress responses Environ. Toxicol. Chem. 25, 2006 3241
Fig. 2. Germination of Triticum aestivum after 96 h of exposure toselected N-heterocyclic polyaromatic hydrocarbons (benzo[h]quinoline,acridine, phenanthridine). Box includes 50% values, middle point ismedian, and whiskers show extremes. Asterisks indicate the statisti-cally significant difference from control: [* p 0.05; ** p 0.01].
Fig. 3. Effect of 1,7-phenanthroline on total length of three differentplant species after 96 h of exposure. Box plot parameters as in Figure2. [* p 0.05; ** p 0.01; *** p 0.001].
at the concentration 0.02 �M, and the changes of all measuredparameters were observed at 2 �M (Fig. 3). Also, 2 �M 1,7-phenanthroline caused 20% decrease of mustard germinability,whereas 20- and 200-�M concentrations in S. alba and 200�M in P. vulgaris were lethal. Because of the low water sol-ubility of parental PAHs [2], only concentrations of 0.02, 0.2,and 2 �M were tested for phenanthrene and fluorene and 0.02and 0.2 �M for anthracene, and only minor effects were ob-served for these compounds. In general, NPAHs were morephytotoxic than parent PAHs. For example, phenanthrene didnot induce any effect up to its highest tested concentration of2 �M in any studied species (Table 1). On the other hand, allthe phenanthrene N-heterocyclic derivatives (1,10-phenan-throline, 4,7-phenanthroline, 1,7-phenanthroline, ben-zo[h]quinoline, and phenanthridine) induced effects at 0.2 �Mor lower concentrations. Most of the NPAHs induced morepronounced effects in S. alba and T. aestivum in comparisonwith P. vulgaris (Table 1 and example of 1,7-phenanthrolinetoxicity in Fig. 3), and these observations might indicate theirhigher sensitivities to the effects of NPAHs. Interestingly, par-ent PAH anthracene affected some growth parameters in P.vulgaris, while it was nontoxic to other two plants (Table 1).
In contrast to apparently higher acute phytotoxicity ofNPAHs, the effects of both PAHs and NPAHs on biochemicalparameters were comparable. All tested chemicals modulatedactivity of plant detoxification and antioxidative enzymes toa different extent (Table 2). The most pronounced modulationswere in general observed after exposure to phenanthridine,benzo[h]quinoline, and 1,7-phenanthroline. Concentrations ofglutathione were increased after 96 h of exposure to all testedchemicals except 1,10- and 4,7- phenanthroline. In absolutevalues, the greatest increase of GSH concentration was ob-served in the biomass of the bean cotyledon after exposure to2 �M fluorene (10-fold induction 35–350 nmol/mg protein).Also anthracene induced GSH levels in P. vulgaris at con-centration as low as 0.02 �M (Table 2). In spite of their struc-
tural similarity, there were substantial differences in the effectsof 2N-analogues of phenanthrene (Fig. 4). The activity of GSTwas significantly increased after exposure to all tested com-pounds in most studied plant tissues (Table 2). Variability inthe concentration–response curves for GST inductions after96 h is demonstrated at the selected examples in Figure 5.While for some of the compounds there was a peak around0.2 to 2 �M followed by a decline in GST activities at higherconcentrations (examples of benzo[h]quinoline and acridine),other chemicals (such as phenanthridine) caused continuousconcentration-dependent induction of GST activity within alltested concentrations. One of the most effective compoundswas fluorene, which induced the GST activity in T. aestivumfrom 20 to 120 nmol/min/mg protein at concentration as lowas 0.02 �M (Table 2). Activities of GPx were induced by mostof the tested PAHs, and the most pronounced effects weredetected in bean cotyledon (Table 2). The exception was an-thracene with no effects up to the highest tested concentration0.2 �M. The most pronounced effects were observed afterexposures to benzo[h]quinoline and phenanthridine. In gen-eral, the most sensitive biomarker from those analyzed wasthe activity of glutathione reductase (Table 2). Triticum aes-tivum was the most sensitive species with GR inductions afterexposures to all compounds except carbazole. Phenanthreneand all its N-heterocyclic derivatives caused GR inductions inall studied plant species at generally very low concentrations.For example, exposure to 0.02 �M phenanthrene caused a 19-fold (from 5–90 nmol NADPH/min/mg protein) increase inGR activity in S. alba. The highest inductions of lipid per-oxidation (TBARS content) were detected in the bean tissues.Carbazole, fluorene, and phenanthridine induced the most pro-nounced effects at low concentrations (Table 2). For example,0.02 �M carbazole caused a 2.5-fold increase in lipid per-oxidation (from 18–45 nmol TBARS/mg protein). Anthracene,acridine, and interestingly also some of the 2-N-analogs ofphenanthrene (1,10-phenanthroline and 4,7-phenanthroline)had no significant effects on TBARS content.
3242 Environ. Toxicol. Chem. 25, 2006 V. Paskova et al.
Table 2. Summary of the effects of N-heterocyclic polyaromatic hydrocarbons and their unsubstituted analogues on biochemical parameters inplants. — no effect; � statistically significant difference from control at �2 �M, �� at 0.2–2 �M, ��� at 0.02 �M; p 0.05a
Plant TBARS GSH GST GPx GR
Phenanthrene Triticum aestivumSinapis albaPhaseolus vulgaris
——
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——
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���1,10-Phenanthroline T. aestivum
S. albaP. vulgaris
———
———
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4,7-Phenanthroline T. aestivumS. albaP. vulgaris
———
———
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1,7-Phenanthroline T. aestivumS. albaP. vulgaris
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Benzo[h]quinoline T. aestivumS. albaP. vulgaris
——
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Phenanthridine T. aestivumS. albaP. vulgaris
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—
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—
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�Anthracene T. aestivum
S. alba——
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——
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P. vulgaris — ��� �� — —Acridine T. aestivum
S. albaP. vulgaris
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Fluorene T. aestivumS. albaP. vulgaris
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——
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—Carbazole T. aestivum
S. albaP. vulgaris
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—
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a TBARS total thiobarbituric acid reactive species; GSH glutathione; GST glutathione-S-transferase; GPx glutathione peroxidase; GR glutathione reductase.
Fig. 4. Level of glutathione (GSH) in biomass of cotyledon of Phas-eolus vulgaris after 96 h of exposure to selected 2-N-heterocyclicpolyaromatic hydrocarbons (4,7-, 1,10-, and 1,7-phenanthroline). Boxplot parameters as in Figure 2. [* p 0.05; ** p 0.01; *** p 0.001].
DISCUSSION
This study brings new information on the plant toxicity andbiochemical effects of important organic contaminants, PAHs,and the relatively poorly characterized group of their N-het-erocyclic derivatives. Some of the prototypical PAH repre-sentatives have been previously studied [4], and significanthealth risks for both humans and the ecosystem have beendocumented [6]. However, there is only limited informationon ecotoxicity of PAH derivatives (such as NPAHs), partic-ularly on their effects in plants. For example, in concordancewith the scientific literature [9], one of the most comprehensiveecotoxicological databases, the U.S. Environmental ProtectionAgency (U.S. EPA) AQUIRE (Aquatic Toxicity InformationRetrieval) database (http://www.epa.gov/ecotox), registers andreports the data on NPAH toxicity for only a few species offish, zooplankton, and algae.
Although NPAHs might be present in the environment inconsiderable concentrations ([30]; http://www.atsdr.cdc.gov/HAC/PHA/joslyn/jms�p1.html), the monitoring data are rathersparse with respect to the lack of suitable analytical methods.For example, Osborne et al. [3] reported total NPAHs con-centrations in suspended matter and sediments ranging be-tween low �g/kg and low mg/kg with the three-ring NPAHsbeing the dominant compounds. A U.S. EPA public healthassessment program study ([30]; www.atsdr.cdc.gov/HAC/PHA/joslyn/jms�p1.html) indicated concentrations of totalnoncarcinogenic PAHs up to 850 mg/kg in soil and up to 46�g/L in surface waters. These concentrations after remediationdecreased to 17 �g/L in surface waters to the range of 1.1 to2.1 �g/L in the case of selected individual compounds (an-
thracene and phenanthrene). These data show that concentra-tions used in our study (0.02–200 �M � 3.3 �g/L–36 mg/L)were within or close to the environmentally relevant range,and the effects observed at lower doses could be of general
PAH and NPAH phytotoxicity and oxidative stress responses Environ. Toxicol. Chem. 25, 2006 3243
Fig. 5. Activity of glutathione-S-transferase in biomass of Sinapisalba after 96 h of exposure to selected N-heterocyclic polyaromatichydrocarbons (phenanthridine, acridine, benzo[h]quinoline). Box plotparameters as in Figure 2. [* p 0.05; ** p 0.01; *** p 0.001]. GST glutathione-S-transferase.
concern. However, our plant germination assays use a liquidexposure media that partially limits direct intercomparisonswith the soil concentrations.
Several studies reported phytotoxicity of PAHs or NPAHsto various plant species [31]. In a recent study, Alkio et al.[18] reported various phytotoxic effects of phenanthrene to A.thaliana including inhibition of growth and root developmentand induction of leaf lesions. The phenanthrene concentrationsused in their study were relatively high (50–500 �M), ex-ceeding water solubility of this compound. But phytotoxicitycan also be induced by direct contact, and concentrations ofphenanthrene in soil could be very high. A complex creosotemixture containing both parent and substituted PAHs inhibitedgrowth and diminished photosynthetic parameters in aquaticmacrophytes L. gibba and Myriophylum spicatum [12]. In-creased inhibition of photosynthesis during photomodificationof anthracene was reported in the aquatic higher plant L. gibba[13]. The exposure in our study was in the dark, which procuresthe formation of photomodified compounds.
In our study, only heterocyclic NPAHs and not homocyclicPAHs inhibited germination of the tested plants and also af-fected growth parameters such as weight and length of theseedlings. The only exemption was the weak toxic effect ofanthracene (0.2 �M) on some growth parameters of P. vulgaris(Table 1). These observations seem to indicate generally great-er toxicity of NPAHs, and they correlate with greater solubilityand bioavailability of NPAHs in comparison with homologousPAHs [2]. Correspondingly, a study with Brassica campestris,Lolium multiflorum, and Hordeum vulgare showed effects ofacridine on seedlings germination and growth at concentrationsranging from 1 to 100 mg/kg [32]. Acridine was also the mosttoxic of the NPAHs tested in a study with the alga Scenedesmusacuminatus [8]. Authors observed 50% growth inhibitions at0.3 and 5.2 mg/L of acridine and phenanthridine, respectively[8]. These concentrations correspond well with our observa-tions (e.g., 70–90% inhibition of germination by acridine at2 �M, i.e., 0.36 mg/L). However, another study with S. alba,Trifolium pratense, and Lolium perenne [31] reported only
minor differences between the toxicity of homocyclic and het-erocyclic PAHs (fluoranthene, pyrene, phenanthrene, fluorene,carbazole, dibenzothiophene, acridine), and the authors alsoobserved no significant toxicity of acridine. The differencesin the results of various experimental studies might be ex-plained by experimental variability but more likely by differentsensitivities of plant species, as demonstrated in both the studyof Sverdrup et al. [31] and our report.
Our results correspond with those of Mitchell et al. [33],who documented higher sensitivity of Avena sativa and Cuc-umis sativus (Poaceae family, like T. aestivum in our report).On the other hand, representatives of the Fabaceae family(Glycine max in the Mitchell et al. [33] study and P. vulgarisin this report) were among the less sensitive species. Our in-vestigation as well as the study of Mitchell et al. [33] observedgenerally lower toxicity of anthracene to plants (only two outof six tested species had EC50 [median effective concentra-tion] under 1,000 mg/kg).
Oxidative stress is an important toxicity mechanism in bothanimals and plants. It may be induced by various toxic chem-icals including PAHs and their derivatives [18,19]. However,most of the available studies on oxidative stress in plants fo-cused almost exclusively on metals [34] or herbicides [35],and the effects of other important toxicants are less docu-mented. To the best of our knowledge, the role of oxidativestress in NPAHs phytotoxicity has not been studied so far, andthe data on parent PAHs are limited.
Several mechanisms of PAH-induced oxidative stress andoverproduction of reactive oxygen species (ROS) have beenrecognized. They include photoreactions, redox cycling ofPAH derivatives [36], and side release of ROS during oxidativePAH metabolism [37]. In our study, we observed significantinductions of oxidative stress in plants exposed to PAHs andNPAHs. The most pronounced lipid peroxidation (determinedas TBARS) was induced by low concentrations of fluoreneand its N-heterocyclic derivative carbazole and also by phen-anthrene and some of its derivatives (Table 2).
In spite of considerable research, it is still poorly understoodwhich structural features of chemicals determine which bio-chemistry process will play a major role in the xenobiotictoxicity. Biochemical responses such as changes in cellularantioxidant and/or detoxification status might be a suitable toolto trace the toxicity mechanisms. The biomarkers assessed inour study reflect major processes protecting plant tissues fromoxidative stress. Glutathione is a ubiquitous thiol that plays acentral role in scavenging ROS. Glutathione is also involvedin enzymatic removal of hydrogen peroxide (H2O2) catalyzedby GPx [37], and it serves as a key endogenous substrate forGST-mediated detoxification of organic xenobiotics. Concen-trations of reduced GSH are then regenerated by another im-portant enzyme, GR, which catalyzes reduction of GSSG. Ac-tivities of all mentioned enzymes as well as GSH concentra-tions are inducible in the presence of organic xenobiotics and/or ROS, and their profiles might reflect the involved toxicitymechanisms.
The compounds in our study induced lipid peroxidation butalso variable modulations of biomarker profiles. Several PAHsand NPAHs increased activities of GST, GR, and GPx andmodulated concentrations of GSH, and the effects were oftenobserved even at low 0.02 �M concentrations. In spite ofdiverse and species-specific responses (Table 2), interestingpatterns could be derived from the phytotoxicity data. Forexample, all phenanthrene derivatives induced GR in most
3244 Environ. Toxicol. Chem. 25, 2006 V. Paskova et al.
plants, thus indicating an increased need for reduced GSH.Chemical-induced production of H202 could be one of the caus-es since GPx was also induced in the presence of all phen-anthrene derivatives (Table 2). However, TBARS (reflectingactual toxic effects of ROS to phospholipid membranes) aswell as GSH levels were elevated after exposure to phenan-threne and three of its derivatives (1,7-phenanthroline, ben-zo[h]quinoline, and phenanthridine), and no effects on theseparameters were observed for 1,10- and 4,7-phenanthroline.Therefore, in spite of close structural similarities of testedphenanthrene derivatives, substantially different mechanismsof oxidative stress seem to play a role. Possibly, plant metab-olism could lead to a formation of structure-specific and highlytoxic redox cycling ortho-quinones. This hypothesis might besupported by documented redox cycling of a specific phen-anthrene metabolite, 9,10-phenanthrenedione [38], but furtherstudies will be required to study such differences in detail.Moreover, transformation of PAHs by mixed-function oxidasesand dehydrogenases into redox-active PAH quinones is an im-portant pathway of their metabolism in mammalian cells [39].Also, model studies with naphthalene derivatives have shownthat only 2,3-dimethoxy-1,4-naphthoquinone is a pure redoxcycler but that structurally close 2-methyl-1,4-naphthoquinone(menadione) induces ROS (via redox cycling) but acts also asan arylating reactive xenobiotic [40].
A limited number of studies also documented sensitive bio-chemical responses in plants exposed to various PAHs. Forexample, modulations of GSH and increased activities of GR,GST, superoxide dismutase, and ascorbate peroxidase werereported in aquatic plant F. antipyretica exposed for 168 h to0.5 �M of prototypical PAHs benzo[a]pyrene and ben-zo[a]anthracene [19]. The correlations between elevated an-tioxidative enzyme activities in this species and accumulatedPAHs were also observed in the field [20]. In the recent studyby Alkio et al. with Arabidopsis [18], relatively high concen-trations of phenanthrene (�50 �M) induced H2O2 productionand modulated GR and ascorbate peroxidase. Similarly, GRactivity was increased dramatically in L. gibba exposed to themixture of copper and oxo-PAH dihydroxyanthraquinone [21].
Biomarkers do not only reflect toxic mechanisms; they canalso be successfully used as early warnings of in vivo effects[41]. This was also revealed in our study. The compounds thatcaused most pronounced in vivo effects on plant germinationand growth (such as 1,7-phenanthroline, benzo[h]quinoline,and phenanthridine; Table 1) significantly induced TBARS andother biomarkers (Table 2). Biochemical changes were in gen-eral more sensitive and occurred already at concentrationsabout an order of magnitude lower than those causing signsof toxicity. Our results thus demonstrate the suitability andinterpretation of multiple biomarker assessments in plant tox-icity biotests.
CONCLUSION
Although environmental concentrations of N-heterocyclicPAHs are generally lower than those of the unsubstituted PAHs(1–10%), higher polarity and water solubility of NPAHs maylead to increased bioavailability and toxicity. Our results newlyreveal significant effects of tested compounds on germinationand growth of T. aestivum, S. alba, and P. vulgaris. Thephytotoxicities of NPAHs were in general more pronouncedthan those of homocyclic PAHs, and the effects were relatedto the oxidative stress as determined by multiple biomarkers.In spite of substantial species- and compound-specific vari-
ability, sensitive biochemical responses in plants reflected var-iable mechanisms of PAH-induced oxidative stress and pro-vided early warnings of the toxic effects observed at higherconcentrations. Further research should provide detailed eco-toxicological characterization of these relatively underesti-mated environmental contaminants with particular respect totheir bioavailability, persistence, and bioaccumulation.
Acknowledgement—Research of the new types of organic pollutants issupported by the Grant Agency of the Czech Republic (grant 525/03/0367). Support from the Czech Ministry of Education to RECETOX,Masaryk University (project INCHEMBIOL, ‘‘Interactions among thechemicals, environment and biological systems and their consequenceson the global, regional and local scales’’; MSM0021622412), is alsohighly acknowledged.
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41. Triebskorn R, Adam S, Casper H, Honnen W, Pawert M, SchrammM, Schwaiger J, Kohler HR. 2002. Biomarkers as diagnostic toolsfor evaluating effects of unknown past water quality conditionson stream organisms. Ecotoxicology 11:451–465.
Paper II.
Adamovský, O., Kopp, R., Hilscherová, K., Babica, P., Palíková, M., Pašková, V., Navrátil, S., Maršálek, B.and Bláha, L. (2007).
Microcystin kinetics (bioaccumulation and elimination) and biochemical responses in
common carp (Cyprinus carpio) and silver carp (Hypophthalmichthys molitrix) exposed to toxic cyanobacterial blooms.
Environmental Toxicology and Chemistry, Vol. 26, No. 12, pp. 2687-2693.
2687
Environmental Toxicology and Chemistry, Vol. 26, No. 12, pp. 2687–2693, 2007� 2007 SETAC
Printed in the USA0730-7268/07 $12.00 � .00
MICROCYSTIN KINETICS (BIOACCUMULATION AND ELIMINATION) ANDBIOCHEMICAL RESPONSES IN COMMON CARP (CYPRINUS CARPIO) AND SILVER
CARP (HYPOPHTHALMICHTHYS MOLITRIX) EXPOSED TO TOXICCYANOBACTERIAL BLOOMS
ONDREJ ADAMOVSKY,† RADOVAN KOPP,‡ KLARA HILSCHEROVA,† PAVEL BABICA,† MIROSLAVA PALIKOVA,§VERONIKA PASKOVA,† STANISLAV NAVRATIL,§ BLAHOSLAV MARSALEK,† and LUDEK BLAHA*†
†Centre for Cyanobacteria and Their Toxins (Institute of Botany, Czech Academy of Sciences and RECETOX, Masaryk University),Kamenice 126/3, 625 00 Brno, Czech Republic
‡Department of Fishery and Hydrobiology, Mendel University of Agriculture and Forestry, Zemedelska 1, 613 00 Brno, Czech Republic§University of Veterinary and Pharmaceutical Sciences, Palackeho 1-3, 612 42 Brno, Czech Republic
(Received 19 March 2007; Accepted 12 July 2007)
Abstract—Two species of common edible fish, common carp (Cyprinus carpio) and silver carp (Hypophthalmichthys molitrix),were exposed to a Microcystis spp.–dominated natural cyanobacterial water bloom for two months (concentrations of cyanobacterialtoxin microcystin, 182–539 �g/g biomass dry wt). Toxins accumulated up to 1.4 to 29 ng/g fresh weight and 3.3 to 19 ng/g in themuscle of silver carp and common carp, respectively, as determined by enzyme-linked immunosorbent immunoassay. Concentrationsan order of magnitude higher were detected in hepatopancreas (up to 226 ng/g in silver carp), with a peak after the initial fourweeks. Calculated bioconcentration factors ranged from 0.6 to 1.7 for muscle and from 7.3 to 13.3 for hepatopancreas. Microcystinswere completely eliminated within one to two weeks from both muscle and hepatopancreas after the transfer of fish with accumulatedtoxins to clean water. Mean estimated elimination half-lives ranged from 0.7 d in silver carp muscle to 8.4 d in common carp liver.The present study also showed significant modulations of several biochemical markers in hepatopancreas of fish exposed tocyanobacteria. Levels of glutathione and catalytic activities of glutathione S-transferase and glutathione reductase were induced inboth species, indicating oxidative stress and enhanced detoxification processes. Calculation of hazard indexes using conservativeU.S. Environmental Protection Agency methodology indicated rather low risks of microcystins accumulated in edible fish, butseveral uncertainties should be explored.
Keywords—Microcystins Bioaccumulation Toxicokinetics Biomarkers
INTRODUCTION
Hepatotoxic microcystins (MCs) are a group of peptidetoxins produced by several species of freshwater cyanobac-teria, such as Microcystis sp., Planktothrix sp., and so on [1].Microcystins occurring as several structural variants are syn-thesized nonribosomally during the growth phase and mayrepresent as much as 1% of the dry biomass. Although a por-tion of produced MCs is present extracellularly, the majorityof MCs remain inside cyanobacteria, and toxins are releasedonly after cell death [1]. Microcystins are potent inhibitors ofserine/threonine protein phosphatase 1 and 2A [1], and theytend to accumulate in liver. Hepatotoxicity, liver tumor pro-motion, as well as other types of toxicity from MCs have beenintensively studied and documented [2,3]. The World HealthOrganization suggested a limit for the tolerable daily intake(TDI) of 0.04 �g/kg body weight/d and corresponding pro-visional guideline of 1 �g/L for drinking waters for the mostoften studied MC variant, MC-LR [1,4].
Although the human toxicity has been studied in detail, therole of MCs in the aquatic environment remains questionable[5,6]. Some reports have described levels of MCs in fish, theirmetabolism, and also their toxicity [7–9], but detailed toxi-cokinetics and critical evaluation of human health risks fromaccumulated toxins remain to be resolved.
* To whom correspondence may be addressed(blaha@recetox.muni.cz).
Published on the Web 7/24/2007.
An important mechanism of MC toxicity documented invarious laboratory animals [10], including fish [11], is oxi-dative stress—that is, cell damage caused by the overproduc-tion of reactive oxygen species. Oxidative stress causes de-pletion of intracellular glutathione (GSH), lipid peroxidation,and oxidative damage to other biomolecules [12]. Several bio-markers of early toxic effects in fish after exposure to variousstressors, including MCs, have been suggested (e.g., modu-lations of glutathione S-transferase [GST], glutathione reduc-tase [GR], and glutathione peroxidase [GPx] [11–13]).
Major aims of the present study were to investigate kineticsof accumulation and elimination of MCs in the tissues of twocyprinid freshwater species, common carp (Cyprinus carpio)and silver carp (Hypophthalmichthys molitrix). Both fish spe-cies are among the most widespread fish in Europe and Asia,and they often are cultured as important edible fish. In addition,the present study examined profiles of biochemical markers inhepatopancreas after cyanobacterial exposure and evaluatedthe health risks of MCs accumulated in fish tissues.
MATERIALS AND METHODS
Experimental design
Experiments simulated the natural situation in the environ-ment. Fish (C. carpio and H. molitrix; average age, two years)were obtained from Pohorelice Fisheries (Pohorelice, CzechRepublic). Uptake and accumulation of MCs was studied inthe outdoor pond during two-month (nine-week) exposures of
2688 Environ. Toxicol. Chem. 26, 2007 O. Adamovsky et al.
Table 1. Kinetics of microcystin (MC) concentrations in the muscle and liver (ng MC/g tissue fresh wt) of common carp (Cyprinus carpio) andsilver carp (Hypophthalmichthys molitrix)a
WeekWaterMCsb
BiomassMCsc
Silver carp
Fish weight (g) Muscle MCs Liver MCs
Common carp
Fish weight (g) Muscle MCs Liver MCs
Accumulation0 22.7 539 202 � 46 0d 0d 125 � 28 0d 0d
(10) (3) (3) (10) (4) (4)4 13.8 425 319 � 78 10.6 � 9.9 93.2 � 50.7 127 � 42 9.8 � 6.4 132 � 59
(10) (10) (10) (10) (7) (7)9 14.2 182 324 � 78 5.2 � 3.4 124 � 56 128 � 37 7.3 � 4.6 68.7 � 42
(10) (7) (7) (10) (7) (7)
BCFe (mean/maximum) 0.62/1.7 7.3/13.3 0.57/1.1 7.8/12.8
Elimination0 — 421 � 92 0.9 � 0.3 21.0 � 14.8 46 � 9 1.2 � 0.3 17.2 � 7.0
— (10) (5) (5) (10) (5) (5)1 — 380 � 102 0d 9.3 � 3.7 47 � 16 0.2 � 0.1 13.7 � 2.7
— (10) (5) (5) (10) (5) (5)2 — 435 � 86 0d 0.9 � 0.8 40 � 10 0d 2.3 � 0.4
— (10) (5) (5) (10) (5) (5)
a Values represent the mean � standard error, with the number of investigated fish given in parentheses.b Water concentrations of total MCs (sum of MC-LR, -RR, and -YR; �g/L).c Biomass MCs concentrations (�g/g dry wt).d Less than the limit of detection (liver, 0.31 ng/g fresh wt; muscle, 0.13 ng/g).e Bioconcentration factors (ratio between the mean/maximum tissue concentration and the average water concentration 17 �g/L).
fish to a complex cyanobacterial bloom dominated by Micro-cystis aeruginosa (45%), Microcystis ichthyoblabe (45%), andAnabaena flos-aquae (5%). Kinetics of MC elimination (afterthe transfer to clean water) was studied in fish that naturallyaccumulated MCs in the pond with Microcystis spp. Fish werenot externally fed during experiments, and no mortalities wererecorded. Fish (n 3–10 individuals/treatment) were collect-ed, weighed, and measured on weeks 4 and 9 (accumulation)and on weeks 1, 2, 4, 6, and 8 (during elimination) (Table 1).The tissue samples were immediately frozen and stored at�80�C for analyses of MCs and biomarkers. Parameters ofwater in the exposure/elimination experiments were as follows(given for the accumulation and elimination experiments, re-spectively; mean � standard error): temperature, 18.9 � 3.8and 19.6 � 1.3�C; dissolved oxygen, 18.2 � 2.0 and 11.1 �3.2 mg/L; and pH, 9.4 � 0.4 and 9.1 � 0.2.
Toxin analyses by high-performance liquidchromatography
Concentrations of MCs in the cyanobacterial biomass andwater (Table 1) were measured by high-performance liquidchromatography (HPLC) as described by Lawton et al. [14]with methods previously used in our laboratory [15]. Briefly,extracts of lyophilized biomass (50% v/v methanol) or watersamples (MCs concentrated by solid-phase extraction usingSepPack C18 cartridges [Waters, Millford, MA, USA]) wereanalyzed with a HPLC Agilent 1100 Series (Agilent Tech-nologies, Waldbronn, Germany) on a Supelcosil ABZ� Plus(length, 150 mm; inner diameter, 4.6 mm; film thickness, 5�m; Supelco, Bellefonte, PA, USA) at 30�C. The binary gra-dient of mobile phase (flow rate, 1 ml/min) consisted of H2Oplus 0.1% trifluoroacetic acid and acetonitrile plus 0.1% tri-fluoroacetic acid (linear increase during 0–30 min from 20–59% of acetonitrile). Chromatograms at 238 nm were recordedwith an Agilent 1100 Series photodiode-array detector, andMCs were identified by the retention time and characteristicabsorption spectra (200–300 nm). Quantification was based on
external calibrations of three MC variants (MC-LR, -RR, and-YR).
Tissue extractions
Tissue extractions were performed according to the methoddescribed by Magalhaes et al. [16]. The frozen sample (0.4 gfresh wt) was homogenized with methanol (3 ml), sonicatedin an ultrasonic bath for 30 min, and centrifuged at 4,000 gfor 10 min. Supernatant was collected and the pellet re-ex-tracted three times using the same procedure. Obtained meth-anol fractions were pooled and repeatedly extracted (threetimes) with 1 ml of hexane to remove lipids (hexane layersdiscarded). Methanol extract was evaporated at 50�C, and theresidue was dissolved in 1 ml of water and analyzed for MCsusing enzyme-linked immunosorbent immunoassay (ELISA).Recovery of the method (�25%; data not shown) was notconsidered during calculations to remain consistent with valuespreviously reported in the literature [16–21].
ELISA for MCs
Concentrations of MCs in the fish tissues were analyzed bydirect competitive ELISA according to the method describedby Zeck et al. [22] using a modification described previouslyin detail [15]. Briefly, high-protein-binding, 96-well micro-plates (Nunc, Wiesbaden, Germany) were incubated overnightwith the anti-mouse immunoglobulin (ICN MP Biomedicals,Solon, OH, USA). After a wash, plates were incubated for 1h with mouse monoclonal IgG MC10E7 developed againstMC-LR (5,000-fold dilution; ALEXIS, Lausen, Switzerland).The reaction was based on the competition of MCs in thesample with the conjugate of MC-LR–horseradish peroxidase[22]. The activity of horseradish peroxidase was determinedusing the 3,3�,5,5�-tetramethylbenzidine (absorbance, 420 nm;reference, 660 nm) with a microplate reader (GENios SpectraFluor Plus; Tecan Group, Mannedorf, Switzerland). Each sam-ple was analyzed in three replicates and the results comparedwith the 0.125 to 2 �g/L calibration curve of MC-LR con-
Microcystin toxicokinetics and biomarkers in fish Environ. Toxicol. Chem. 26, 2007 2689
structed for each individual ELISA plate. Samples from bothexposed and control fish were analyzed, and no significantnonspecific interferences of the tissue extracts with ELISAwere observed. The antibody used in the present study(MC10E7) has been shown to have 100 and 96% cross-re-activity with MC-LR and MC-RR, respectively [22]. Becausethese two MC variants were dominant in the present study,detected concentrations were considered to be a sum of MCs.We cannot exclude that the ELISA also detected MC fragmentsin fish tissues, such as glutathione-MC conjugates. This wasnot studied in detail, however, our approach was comparablewith those in previous studies [16–21].
Biomarker analyses
Hepatopancreas samples (1 g) were homogenized on icewith 1 ml of phosphate buffer saline (pH 7.2), and supernatantwas collected after centrifugation (5 min, 2,500 g, 4�C) andstored at �80�C before analyses. Protein concentrations weredetermined according to the method of Lowry et al. [23] usingbovine serum albumin as a standard.
Concentration of glutathione was determined according tothe method described by Ellmann [24] using 5,5�-dithiobis-2-nitrobenzoic acid as a substrate. Before analyses, the sampleswere treated with trichloroacetic acid (25% w/v) and centri-fuged (6,000 g, 10 min). Supernatant was mixed with 0.6 �M5,5�-dithiobis-2-nitrobenzoic acid in Tris-HCl/ethylenediami-netetra-acetic acid (EDTA) buffer (0.5 M tris[hydroxymethyl]-aminomethane–hydrochloric acid, 0.5 M Tris, and 12.5 mMEDTA; pH 8.9) and incubated for 5 min at room temperature.Absorbance was measured at 420/680 nm, and the concentra-tions (nmol GSH/mg protein) were calculated from the cali-bration of standard reduced GSH.
Glutathione S-transferase activity was measured spectro-photometrically using 1 mM 1-chloro-2,4-dinitrobenzene and2 mM GSH as substrates according to the method describedby Habig et al. [25]. Specific activity was expressed as nano-moles of formed product per minute per milligram of protein.
Activity of GPx was determined from the rate of nicotin-amide adenine dinucleotide phosphate (NADPH) oxidation,recorded as the decrease in absorbance at 340 nm [26]. Thereaction mixtures contained 3 mM GSH, 1.2 mM butylhydro-peroxide, 1 U of GR (1 U of GR reduces 1.0 mmol of oxidizedglutathione per minute at pH 7.6 at 25�C), and 0.15 mMNADPH in 0.1 M potassium phosphate/1 mM EDTA buffer(pH 7.0). Also, the activity of GR in fish was determined byspectrophotometric measurement of NADPH oxidation in mi-croplates [27]. The reaction mixtures contained 0.05 M po-tassium phosphate/1 mM EDTA buffer (pH 7.0), 1 mM glu-tathione-oxidized disodium salt, 0.1 mM NADPH, and thetissue extract (0.25% v/v). Specific activities of both GPx andGR were expressed as nanomoles of NADPH oxidized perminute per milligram protein.
Statistical calculations
Significant differences were determined using Student’s ttest or analysis of variance followed by Dunnett’s post-hoctests. Data normality was checked with the Kolmogorov-Smir-nov test, and homogeneity of variances was assessed with theLevene’s test. The p values less than 0.05 were considered tobe statistically significant for all tests. Calculations were per-formed using the Statistica for Windows� 7.0 software package(StatSoft, Tulsa, OK, USA). Elimination kinetic curves andMC half-lives were calculated using the one-phase exponential
decay equation incorporated in the GraphPad Prism 4 software(GraphPad Software, San Diego, CA, USA).
RESULTS AND DISCUSSION
The present study describes toxicokinetics (accumulationand elimination) of MCs in the tissues of common carp andsilver carp. Although several authors reported MC concentra-tions in zooplankton, shellfish, or fish [28–32], the kinetics ofMC accumulation and elimination in fish have not been in-vestigated in detail.
A summary of our results is given in Table 1 and in Figures1 and 2. Microcystins accumulated in the muscle of commoncarp and silver carp up to 9.8 and 10.6 ng/g fresh weight,respectively. Concentrations approximately an order of mag-nitude higher were determined in the hepatopancreas, whichis the target organ for MCs [33,34]. The muscle to liver con-centration ratio in the present study (1:10) corresponded tothat in the previous study with Atlantic salmon [33], but ahigher ratio (1:20) was found in common carp compared withthat in the study by Li et al. [18].
Average MC concentrations in both studied species gen-erally were comparable (Table 1), but slightly higher levelswere found in the liver of common carp in comparison to thosein the liver of silver carp (compare, e.g., week 4 of the ac-cumulation experiment) (Table 1). This may be related to pos-sible resistance of phytophagous silver carp to MCs in com-parison with the benthophagous common carp (as also sug-gested by Snyder et al. [19]). Calculated bioconcentration fac-tors (BCFs; average and maximum tissue concentrationsdivided by the average water concentration of 17 �g/L) rangedfrom 0.6 to 1.7 in the muscle and from 7.3 to 13.3 in the liverof both species. To our knowledge, the BCFs for MCs in fishwere not previously reported, but our results generally cor-respond to previously reported values for aquatic macrophytes(MC BCF �0.1–5.9 [35]). Higher BCFs (range, 12–22) werereported for structurally related peptide cyanotoxin nodularinin various zooplankton species [36].
Kinetics of MC accumulation in hepatopancreas seems tobe species-specific. In common carp, a peak in MC concen-trations occurred after four weeks, followed by an apparentdecrease after nine weeks (a trend that is comparable to thechanges in muscle of both species) (Table 1). On the otherhand, continuous accumulation of MCs was recorded in he-patopancreas of silver carp during the entire exposure period(up to 124 ng/g fresh wt) (Table 1). Differences may be ex-plained, for example, by phytoplanktivorous feeding of silvercarp, which actively ingests cyanobacterial cells, whereas onlypassive MC intake can be expected in omnivorous and ben-thophagous common carp [19].
The elimination experiment demonstrated that MC is rap-idly removed from the tissues after the transfer of fish to cleanwater (Table 1). In both species, calculated elimination half-lives were shorter for muscle (0.7–2.8 d) than for liver (3.5–8.4) (Fig. 1). To our knowledge, information regarding MCdepuration from the fish is rare [20,37,38]; however, studiesof MC elimination from some invertebrates also suggest fastelimination of MCs. For example, a half-life of 8 d was re-ported for freshwater snail [39], and half-lives from 3.0 to 4.8d were observed in bivalves [40]. In contrast to the rapidelimination observed in our manipulated experiments (Fig. 1),slower MC removal from silver carp and Nile tilapia has beenreported in natural lakes (elevated MCs during the period 15–40 d after the end of the accumulation period [20,37]).
2690 Environ. Toxicol. Chem. 26, 2007 O. Adamovsky et al.
Fig. 1. Microcystin elimination from the tissues of common carp (Cyprinus carpio) (A and B) and silver carp (Hypophthalmichthys molitrix)(C and D). Presented are individual tissue concentrations, elimination curves (solid lines) with 95% confidence intervals (dashed lines), and half-lives in days (mean values with 95% confidence intervals in parentheses).
Taken together, bioaccumulation of MCs in the fish is adynamic process depending on both uptake and metaboliza-tion/elimination [6]. Interspecies variability in MC metabolismand elimination, however, as well as environmental factors(e.g., temperature [40]) that may affect MC toxicokinetics willrequire further research.
We also investigated a set of glutathione-related biomarkersin the hepatopancreas of both fish species (Fig. 2). Activity ofGR (significantly elevated in a majority of experimental var-iants, especially in common carp) was the most sensitive bio-marker of cyanobacterial exposure (Fig. 2). On the other hand,changes in GPx activity were less sensitive in our experiments.
Inductions of GST seem to correspond to detoxification ofMCs by GST-mediated conjugation with GSH [9,41,42]. El-evated GSH concentrations and activities of the GR (the en-zyme regenerating GSH from its oxidized form [13]) furtherreveal increased demands for reduced GSH because of en-hanced detoxification and/or oxidative stress induced by toxiccyanobacteria [11,12,43]. Our present study, however, dem-onstrates that biochemical adaptations are only temporary andthat prolonged exposures may result in signs of general tox-icity—that is, suppression of GSH levels and inhibition of GRactivity (compare the four- and nine-week exposures for silvercarp, as shown in Fig. 2).
Apparent time-, species-, and MC variant–dependent var-iability exists in biochemical responses of organisms to MCs[44]. Inductions of GST are among the most often reportedresponses [9,42] (present study), but other authors also havereported rapid, 24-h inhibitions of GST in Corydoras paleatusexposed to purified MC-RR [45]. Modulations of biomarkersin the present study confirm an important role for oxidative
stress in the toxicity of complex cyanobacterial bloom, and italso demonstrates that biochemical parameters (especially GR,GST, and GSH) may serve as sensitive early markers of adverseeffects in fish. Direct interpretation of biomarker responsesremains complicated, however, and further research will beneeded to characterize both natural variability and temporalchanges in responses to toxicants.
It has been suggested that accumulated MCs in edible fishmay represent a risk to human health, and it has been dem-onstrated that MCs are stable and not degraded by heat duringcooking [46]. We have calculated the hazard index (HI), a ratiobetween the estimated daily intake (EDI) and chronic TDI,based on our results using an U.S. Environmental ProtectionAgency methodology [47]. To derive the EDI, we have con-sidered a one-year exposure, 48 fish meals per year (100%contaminated), ingestion rate of 132 g per serving of meat,human body weight of 70 kg, and maximum concentration ofMCs in fish fillet observed in the present study (29.3 ng/gfresh wt in silver carp). Using this worst-case scenario andconsidering a chronic TDI (0.04 �g/kg/d for MC-LR [1]), acalculated HI of 0.19 indicates a nonsignificant risk from MCsaccumulated in fish meat (HI 1 [47]). Interestingly, relativelyhigh HIs, ranging from 2.35 to 3.66, which correspond torealistically edible critical amounts of fish food (82–545 g/serving) were reported previously by Magalhaes et al. [16].Those authors, however, compared the single-day intake ofMCs with the chronic (i.e., year-round derived) TDI value,which could overestimate the total risk. Another factor thatmay affect total risk is relatively low recovery of MCs fromanimal tissues (reported values range from 3% [48] to 25%[present study]), which usually is not considered during cal-
Microcystin toxicokinetics and biomarkers in fish Environ. Toxicol. Chem. 26, 2007 2691
Fig. 2. Modulations of biochemical parameters in fish hepatopancreas after four and nine weeks of exposure to cyanobacterial biomass. Levelof glutathione (GSH; nmol/mg protein), activity of glutathione S-transferase (GST; nmol/min/mg protein), and activities of glutathione peroxidase(GPx; nmol nicotinamide adenine dinucleotide phosphate [NADPH]/min/mg protein) and glutathione reductase (GR; nmol NADPH/min/mgprotein). Box includes the 25th to 75th percentiles, with the middle point representing the median and the whiskers showing the extremes. Anasterisk indicates a statistically significant difference from control (p 0.05, Student’s t test).
2692 Environ. Toxicol. Chem. 26, 2007 O. Adamovsky et al.
culations but may lead to possible underestimation of EDI.Taken together, MCs accumulated in edible fish tissues even-tually may pose a risk to certain groups of people (e.g., fish-ermen consuming large amounts of contaminated fish), butuncertainties remain in both analytical approaches and riskassessment calculations.
CONCLUSIONS
Our results demonstrate kinetics of MC accumulation andelimination in two common Eurasian freshwater fish species,common carp and silver carp. We found that in most cases,maximum MC concentrations accumulated within the first fourweeks of exposure, and prolonged periods (nine weeks) re-sulted in a less significant increase. Our results suggest rapidelimination of MCs from the fish tissues (half-life in days),but further research should focus on interspecies differencesin metabolization and natural factors affecting MC toxicoki-netics. The role of oxidative stress and changes of detoxifi-cation capacity in response of the fish on cyanobacterial ex-posure was confirmed by modulations of several biochemicalparameters (e.g., GR, GSH, and GST in both species). Cal-culation of hazard indexes using conservative U.S. Environ-mental Protection Agency methodology indicates a rather lowrisk of accumulated MCs in edible fish, but several uncertain-ties should be explored.
Acknowledgement—This work was supported by the Ministry of Ed-ucation of the Czech Republic (projects MSM 6215712402 andIM6798593901) and by the National Agency for Agricultural Re-search (NAZV/06/3233).
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37. Soares RM, Magalhaes VF, Azevedo SMFO. 2004. Accumulationand depuration of microcystins (cyanobacteria hepatotoxins) inTilapia rendalli (Cichlidae) under laboratory conditions. AquatToxicol 70:1–10.
38. Cazenave J, Wunderlin DA, Bistoni MDL, Ame MV, Krause E,Pflugmacher S, Wiegand C. 2005. Uptake, tissue distribution, andaccumulation of microcystin-RR in Corydoras paleatus, Jenynsiamultidentata, and Odontesthes bonariensis—A field and labo-ratory study. Aquat Toxicol 75:178–190.
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41. Pflugmacher S, Wiegand C, Beattie KA, Codd GA, SteinbergCEW. 1998. Uptake of the cyanobacterial hepatotoxin microcys-tin-LR by aquatic macrophytes. Journal of Applied Botany-An-gewandte Botanik 72:228–232.
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Paper III.
Pašková, V., Adamovský, O., Pikula, J., Sko ovská, B., Ban ouchová, H., Horáková, J., Babica, P., Maršálek, B. and Hilscherová, K. (2008).
Detoxification and oxidative stress responses along with microcystins accumulation in
Japanese quail exposed to cyanobacterial biomass.
Science of the Total Environment, Vol. 398, Is. 1-3, pp. 34-47.
Detoxification and oxidative stress responses along withmicrocystins accumulation in Japanese quail exposed tocyanobacterial biomass
Veronika Paškováa,b, Ondřej Adamovskýa,b, Jiří Pikulac, Blanka Skočovskác,Hana Band'ouchovác, Jana Horákovác, Pavel Babicaa,b,Blahoslav Maršáleka,b, Klára Hilscherováa,b,⁎aCentre for Cyanobacteria and Their Toxins (Institute of Botany, The Academy of Sciences of the Czech Republic & RECETOX, MasarykUniversity), Kamenice 126/3, CZ62500, Brno, Czech RepublicbRECETOX, Research Centre for Environmental Chemistry and Ecotoxicology, Masaryk University, Kamenice 3, CZ 625 00 Brno, Czech RepubliccUniversity of Veterinary and Pharmaceutical Sciences Brno, Faculty of VeterinaryHygiene and Ecology, Palackeho 1/3, 612 42Brno, Czech Republic
A R T I C L E I N F O A B S T R A C T
Article history:Received 19 December 2007Received in revised form4 March 2008Accepted 4 March 2008
The cyanobacterial exposure has been implicated in mass mortalities of wild birds, butinformation on the actual effects of cyanobacteria on birds in controlled studies is missing.Effects on detoxification and antioxidant parameters as well as bioaccumulation ofmicrocystins (MCs) were studied in birds after sub-lethal exposure to natural cyanobacterialbiomass. Four treatment groups of model species Japanese quail (Coturnix coturnix japonica)were exposed to controlled doses of cyanobacterial bloom during acute (10 days) and sub-chronic (30 days) experiment. The daily doses of cyanobacterial biomass corresponded to 0.2–224.6 ngMCs/g bodyweight. Significant accumulationofMCswas observed in the liver for bothtest durationsand slightaccumulationalso in themusclesof thehighest treatment group fromacute test. The greatest accumulationwas observed in the liver of the highest treatment groupin the acute test reaching average concentrationof 43.7ngMCs/g freshweight. Theparametersof detoxification metabolism and oxidative stress were studied in the liver, heart and brain.The cyanobacterial exposure caused an increase of activity of cytochrome P-450-dependent 7-ethoxyresorufinO-deethylase representing the activation phase of detoxificationmetabolism.Also the conjugation phase of detoxification, namely the activity of glutathione-S-transferase,was altered. Cyanobacterial exposure alsomodulated oxidative stress responses including thelevel of glutathione and activities of glutathione-related enzymes and caused increase in lipidperoxidation. The overall pattern of detoxification parameters and oxidative stress responsesclearly separated the control and the lowest exposure group from all the higher exposedgroups. This is the first controlled study documenting the induction of oxidative stress alongwith MCs accumulation in birds exposed to natural cyanobacterial biomass. The data alsosuggest that increased activities of detoxification enzymes could lead to greaterbiotransformation and elimination of the MCs at the longer exposure time.
© 2008 Elsevier B.V. All rights reserved.
Keywords:Avian dietary toxicity testCoturnix coturnix japonicaCyanobacteriaMicrocystinDetoxificationOxidative stress
S C I E N C E O F T H E T O T A L E N V I R O N M E N T 3 9 8 ( 2 0 0 8 ) 3 4 – 4 7
0048-9697/$ – see front matter © 2008 Elsevier B.V. All rights reserved.doi:10.1016/j.scitotenv.2008.03.001
⁎ Corresponding author. CCT & RECETOX, Kamenice 126/3, CZ62500, Brno, Czech Republic. Tel.: +420 54949 3256; fax: +420 54949 2840.E-mail address: hilscherova@recetox.muni.cz (K. Hilscherová).
ava i l ab l e a t www.sc i enced i rec t . com
www.e l sev i e r. com/ loca te / sc i to tenv
1. Introduction
Cyanobacteria are known to produce secondary metabolites,which have been recognized as human and animal health ha-zards, since they have been shown to cause adverse effects inmammals, birds, fish, invertebrates as well as plants (Codd,1996; Figueiredo et al., 2004; Wiegand and Pflugmacher, 2005;Malbrouck and Kestemont, 2006; Babica et al., 2006; Skocovskaet al., 2007).
The most frequently occurring cyanobacterial toxins aremonocyclic heptapeptides called microcystins (MCs) (Carmi-chael, 1997). MCs or MC-producing cyanobacterial strains havebeen associated with poisonings of wildlife and especially withthe mass mortalities of wild birds over recent years. MCs andcyanobacterial hepato-andneurotoxins contributedprobably tomass deaths of Lesser Flamingos in Kenya (Krienitz et al., 2003;Ballot et al., 2004, 2005;Ndetei andMuhandiki, 2005) or Tanzania(Lugomela et al., 2006). Greater Flamingo chick deaths, attrib-uted to MCs, occurred at wetlands lagoon in Spain after thesudden development of a bloom with prevailing Microcystisaeruginosa and Anabaena floss-aquae (Alonso-Andicoberry et al.,2002). MCs have been also detected in cyanobacterial blooms inBelgian (Wirsing et al., 1998), Japanese (Matsunaga et al., 1999) orCanadian (Murphyetal., 2000, 2003; Parketal., 2001) lakeswhereconspicuous deaths of wild birds occurred.
MCs primarily act as hepatotoxins (Wiegand and Pflugma-cher, 2005), because they are predominantly absorbed viailleum and transported via iliac vein and portal vein into liver(Dahlem et al., 1989; Bury et al., 1998) and also lungs and heart(Ito et al., 2000; Liu et al., 2002). The hepatocytes highly expressorganic anion transport proteins, which are responsible foractive cellular uptake ofMCs fromblood (Runnegar et al., 1995).However, various organic anion transport proteins are presentalso in other organs than liver, e.g., in gastrointestinal tract,kidney or brain (Hagenbuch andMeier, 2003). Correspondingly,accumulation of MCs (or structurally related nodularins) hasbeen reported not only in the liver, but also in intestines,kidneys, brain, heart, gonads andmuscles of fish (Kankaanpaaet al., 2005; Cazenave et al., 2005; Adamovsky et al., 2007;Kagalou et al., in press) ormammals, and there is an increasingevidence about neurological or renal toxicity of MCs invertebrates (Dietrich and Hoeger, 2005). It has been suggestedthat organ-specific distribution and toxic effects of MCs aregoverned by the presence/absence, type and expression levelof organic anion transport proteins (Dietrich andHoeger, 2005).
Exposure to cyanobacterial biomass and/or purified MCshas been shown to cause oxidative stress in various organisms(Ding et al., 2000; Pietsch et al., 2001; Li et al., 2003; Wiegandand Pflugmacher, 2005). Formation of reactive oxygenspecies (ROS) and oxidative stress is associated with thedevelopment of many pathological states. Oxidative stressmay occur either due to the decrease of cellular antioxidantlevel or to the overproduction of ROS (Ding and Ong, 2003).Exposure to MC-LR has been linked with increase of ROSproduction in mammals and fish (Ding et al., 2000; Li et al.,2003). Liver as the general detoxifying organ is considered themain region of ROS generation in mammals and birds (Prietoet al., 2006). Endogenous antioxidant defenses of enzymaticand non-enzymatic nature are critical for the control of ROS-
mediated oxidative damage of biomolecules, including pro-teins, RNA, DNA and membrane polyunsaturated lipids(Halliwell and Gutterdige, 1999). The main defense mechan-isms against ROS and their toxic by-products includeenzymes, above all glutathione-S-transferases (GST), glu-tathione reductase (GR), glutathione peroxidase (GPX), cata-lase (CAT) and superoxide dismutases (SOD), and also non-enzymatic compounds such as glutathione (GSH). Moreover,GSTs are enzymes catalyzing a conjugation of MCs with GSHand therefore responsible for detoxification of MCs (Fu andXie, 2005). Significant modulations of the antioxidative anddetoxification system (GST) or increased production of lipidperoxides upon the exposure to pure MCs, MC-containingcyanobacteria or cyanobacterial extracts have been demon-strated by numerous studies with plants (Babica et al., 2006;Pflugmacher et al., 2006), invertebrates (Pietsch et al., 2001;Chen et al., 2005; Rosa et al., 2005) or fish (Malbrouck andKestemont, 2006; Fu and Xie, 2005). However, little data isavailable for adult warm-blooded vertebrates. Only fewstudies have been carried out with mammalian cell lines(Ding and Ong 2003; Bouaicha et al., 2004) or withmammals invivo (Gupta et al., 2003; Gehringer et al., 2004; Moreno et al.,2005; Maidana et al., 2006), and there is no information on thepotential oxidative stress or detoxification in cyanobacteria-exposed birds.
Our previous report indicated histopathological hepaticchanges, modification of the biochemical parameters in bloodand bioaccumulation of MCs in the liver of Japanese quails(Coturnix coturnix japonica) exposed for 10 or 30 days tocontrolled doses of natural cyanobacterial bloom with majorcontent of MC-LR and MC-RR (Skocovska et al., 2007). In thispart of the study, we investigated the effect of cyanobacterialexposure on activation (P450-dependent 7-ethoxyresorufin-O-deethylase activity) and conjugation (GST, GSH) phase ofdetoxification metabolism, antioxidant activities and lipidperoxidation as ameasure of oxidative damage in the exposedbirds. We also studied MC levels in the liver as the primarytarget organ and in the muscles as the tissue that can be usedfor human consumption. This study brings more informationabout the effects of cyanobacteria on birds in connection withdetoxification and oxidative stress responses.
2. Materials and methods
2.1. Bioassay
The sub-lethal effects of cyanobacterial biomasswere studied inJapanese quails after exposure performed according to the Or-ganization for EconomicCo-operationandDevelopment (OECD)Guideline for the testing of chemicals 205 — Avian DietaryToxicity Test (OECD, 1984) with some minor modificationsdescribed in detail in our previous paper (Skocovska et al., 2007).
2.2. Cyanobacterial biomass
Cyanobacterial biomass with domination of Microcystis sp.was collected with plankton net (25 μm) from Brno reservoir(Czech Republic) in autumn 2004. Biomass concentration was
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determined by cell counting under a fluorescent microscopeafter disintegration of colonies using an ultrasonic probe(Bandelin Sonopuls UW2070, Bandelin Electronics, Berlin,Germany) for 10 min (70% cycle, 70% power). Dry weight ofthe biomass was determined by drying at 50°C. MCs wereextracted from 5 mL of fresh cyanobacterial biomass afteraddition of equivalent volume of methanol using ultrasonica-tion (Bandelin Sonopuls UW2070, twice 30 s, 80% cycle, 100%power). The extract was centrifuged (10 min, 2800 g) andconcentration of MCs in supernatant was measured by HPLCAgilent 1100 Series coupled with PDA detector (Agilent Tech-nologies, Waldbronn, Germany) on Supelcosil ABZ+ Pluscolumn, 150×4.6 mm, 5 μm (Supelco, Bellefonte, PA, USA) at30°C. Binary gradient of mobile phase consisted of H2O+0.1%TFA (A) and acetonitrile+0.1% TFA (B); linear increase from20% B at 0 min to 59% B at 30 min, flow rate was 1 mL min−1.UV spectra were recorded from 200 to 300 nm and chromato-grams were evaluated at 238 nm. MCs were identified bycomparison of UV spectra and retention times with standardsof MC-LR, -LF, -LW, -RR and -YR (Alexis Biochemicals,Laeufingen, Switzerland). Concentrations of MCs in naturalcyanobacterial biomass were 141.8 μg/g DW of MC-RR,141.7 μg/g DW of MC-LR, 33.7 μg/g DW of MC-YR and 56.1 μg/g DW of unidentified compound with MC-like UV spectrum.The total concentration of MCs in studied biomass was373.3 μg/g DW. The homogenization of biomass for dosingwas performed by repeated freezing and thawing (two times)and by ultrasonication (Bandelin Sonopuls UW2070, 10 min,70% cycle, 70% power). Four biomass concentrations wereprepared by dilution of biomasswith drinkingwater, aliquotedinto plastic cups and stored frozen. Drinking water for thecontrol group of quails was handled the same way.
2.3. Exposure
Experiment was conducted with 4 months old individuals ofCoturnix coturnix japonica (Japanese quail, gallinaceous birdspecies). Japanese quail belongs to the common experimentalbird species. Quails were held in standard lab cages and werefed with commercial bird food and drinking water ad libitum.The exposure design has been described in detail in thepreviously published part of our study (Skocovska et al., 2007).Briefly, the birds (mean weight 205 g) were divided into fiveexperimental groups (control group C, exposure groups E1–E4)fed various daily doses (Table 1) of the cyanobacterial biomass.The daily doses of 10 mL contained from 3×106 (group E1) to3×109 (group E4) cyanobacterial cells, which is equivalent to0.123 mg to 123 mg dry biomass, respectively. The same dailydoses have been administered and the same experimentaldesign has been carried out during the acute (10 days) and sub-chronic (30 days) exposure. After the experiment, the animalswere sacrificed by decapitation. Selected organs (liver, brain,heart and major pectoral muscles) were dissected and storedat −80°C for analyses of MC concentration and measurementof biochemical parameters.
2.4. Determination of microcystin concentration in tissues
MC concentrations were determined in the liver tissue andmajor pectoral muscles. Liver and muscles (400 mg of fresh
weight) were extracted three times with 3 mL methanol usingultrasonication bath (30 min) and centrifuged (10 min, 2800 g).Distilled water (2 mL) was added to combined supernatantsand extracts were portioned three times with 1 mL of hexane.The hexane layers were discarded and methanolic fractionwas evaporated to dryness at 50°C. The residues wereredissolved in 1 mL of distilled water on ultrasonic bath(15 min) and MC concentration was analysed by direct com-petitive ELISA (modified from Zeck et al., 2001). High protein-binding 96-well microplates (Nunc, Wiesbaden, Germany)were pre-incubated overnight with 2000-fold diluted anti-mouse anti-Fc-IgG (MP Biomedicals, Ohio, USA). Free IgG wasthen removed by washing with phosphate buffer saline (PBS,pH 7.3), and the plates were coated for 1 h with 5000-folddiluted monoclonal IgG (MC10E7, Alexis Biomedicals, SanDiego, USA) developed against MC-LR. The plate was thenwashed five times with 0.05% (v/v) Tween-20 in PBS, andnonspecific interactions were blocked by adding 20 μL of theblock solution to each well (1% v/v EDTA, 1% v/v bovine serumalbumin in 1 M TRIS–HCl, pH 7.4). The filtered samples,standards and controls were immediately added to the wells(200 μL per well) and the plate was incubated for 40 min atroom temperature. Finally, 50 μL of MC-LR conjugated withHRP prepared and purified according to Zeck et al. (2001) wasadded to each well. The reaction was then incubated at roomtemperature for another 15 min, the plates were washed fivetimes with 0.05% (v/v) Tween-20 in PBS, and 175 μL of the HRPsubstrate 3,3′,5,5′-tetramethylbenzidine was added. Develop-ment of the coloured product was stopped after 10 min byadding 50 μL of 5% (v/v) sulfuric acid. The absorbance (420 nmwith reference 660 nm) was determined with a microplatereader (GENios, Tecan Group, Switzerland). Each sample wasanalysed in three replicates and compared with 0.125–2 μg/Lcalibration curve of MC-LR constructed for each individualplate.
2.5. Biochemical methods
The biochemical analysis was focused on three importantorgans that could be directly affected by MCs and othercyanobacterial toxins and that are known to be susceptible tooxidative stress, i.e. liver, heart and brain. The tissues werehomogenized on ice in phosphate buffer saline (PBS, pH 7.2)
Table 1 – Characterization of biomass dilutions preparedfor each exposure group and recalculated daily doses ofcyanobacterial biomass related to the average weight ofexperimental birds
Biomass Daily dose/body weight
Group Amountof cells/L
mgDW/L
∑ μgMCs/L
Amount ofcells/g
μgDW/g
∑ ngMCs/g
C – – – – – –E1 3×108 12.3 4.5 14.6×103 0.6 0.2E2 3×109 123.3 46.1 14.6×104 6.0 2.24E3 3×1010 1233.6 460.5 14.6×105 60.1 22.46E4 3×1011 12334.8 4605.4 14.6×106 601.7 224.6
Experimental birds consumed 10 mL of biomass dilution per dayduring acute (10 days) and sub-chronic (30 days) study.
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using mechanical homogenizer, 100 mg of tissue in 1 mL ofPBS; postmitochondrial supernatant was collected after cen-trifugation (15 min at 10000 g at 4°C) and stored frozen at−80 °C until biochemical analyses.
All biochemicals and enzymes were purchased fromSigma-Aldrich (Prague, CR), other chemicals used for prepara-tion of buffers were of the highest commercial grade available.
Glutathione-S-transferase (GST) activity was measuredspectrophotometrically at 340 nm using 1 mM 1-chloro-2,4-dinitrobenzene (CDNB) and 2 mM GSH in PBS (Habig et al.,1974). Specific activity was expressed as nmol of evolved pro-duct per minute per milligram protein. The concentration ofreduced glutathione was determined by spectrophotometricmethod using 5,5′-dithiobis-2-nitrobenzoic acid (DTNB) as asubstrate (Ellmann, 1959). Tissues were treated with trichlor-oacetic acid (TCA, 2.5% v/v) and centrifuged (6000 g for 10 minat 4°C). Supernatant was mixed with TRIS–HCl buffer (0.6 MTRIS, 0.015 M EDTA, pH 8.9) and 0.8 mM DTNB and incubatedfor 5 min at room temperature. Absorbance was measured at420/680 nm and the concentrations (nmol GSH/mg protein)were calculated according to the standard calibration withreduced GSH. Activity of glutathione peroxidase (GPX) wasdetermined from the rate of NADPH oxidation recorded as thedecline in absorbance at 340 nm (Flohé and Gunzler, 1984). Thereaction mixtures contained 3 mM GSH, 1 U glutathionereductase (GR) (1 unit [U] will reduce 1.0 μmol of oxidizedglutathione permin at pH 7.6 at 25°C), 0.15mMNADPH in 0.1Mpotassiumphosphate/1mMEDTAbuffer (pH7). Substrate usedfor the assaywas 1.2mMbutylhydroperoxide. Also the activityof GR was determined by spectrophotometricmeasurement ofNADPH oxidation (Carlberg and Mannervik, 1975). Assays forGR activity were performed in microplates, and the reactionmixtures contained 0.05 M potassium phosphate/1 mM EDTAbuffer (pH 7.0), 1 mM oxidized glutathione (GSSG), 0.1 mMNADPH and the supernatant (0.25% v/v). Specific activities ofboth GPX and GR were expressed as nmol NADPH oxidized perminute permilligramprotein. The level of lipid peroxidation inavian tissues was assessed as total thiobarbituric acid (TBA)reactive species (TBARS) (Uchiama and Mihara, 1978; Living-stone et al., 1990). The extractsweremixedwith trichloroaceticacid (TCA, 6% w/v) and butylated hydroxytoluene (0.6% w/v)and centrifuged (1500 g for 20 min). Supernatant was furthermixedwith 0.06 NHCl and 40mMTBA prepared in 10mMTRIS(pH 7.4). The mixture was boiled in water bath for 45 min andthen cooled to room temperature. Absorbance of the samplewas measured at 550/590 nm and the concentration of TBARS(nmol TBARS per milligram protein) was calculated accordingto the standard calibration curve generated with malondial-dehyde prepared by acidic hydrolysis of 1,1,3,3-tetraethoxy-propane. The protein concentrations were determined by themethod using Folin–Ciocalteu phenol reagent that forms withproteins red-coloured complex measurable at 680 nm (Lowryet al., 1951). Bovine serum albumin was used as a standard forprotein calibration. The activity of cytochrome P-450-depen-dent 7-ethoxyresorufin O-deethylase (EROD) was analysedfluorimetrically (Prough et al., 1978). The reaction mixturescontainedHepes buffer (25mM, pH 7.8) with dicumarol (1mM),supernatant and 7-ethoxyresorufin (10 μM), whichwas used asa substrate. The reactionwas started by the addition of 0.2mMNADPH followed by incubation at 37°C for 20 min. The
excitation and emission wavelengths were set at 530 and585 nm, respectively. Enzymeactivity results are given as pmolresorufin permilligramprotein perminute. The GENiosmicro-plate reader (TecanGroup, Switzerland)was used formeasure-ment of absorbance in all spectrophotometric assays and thePOLARstar OPTIMA (BMG LABTECH, Germany) was used formeasurement of fluorescence.
2.6. Statistical evaluation
Statistical analyses were performed with Statistica for Win-dows® 7.0 (StatSoft, Tulsa, OK, USA). Results from differenttreatment groups were compared by one-way analysis ofvariance (ANOVA) and post-hoc analysis of means using theLSD test. Homogeneity of variances was tested by Levene'stest. Parameters that were not normally distributed as deter-mined by Shapiro–Wilk's test and/or for which the variancewas not homogeneous as determined by Levene's test werelog-transformed prior to analysis. In case of nonhomogeneousvariances, nonparametric Kruskall–Wallis test was used forcomparison of the treatment groups. Spearman rank ordercorrelations were used to characterize the relationshipsamong parameters. Variation of the biochemical parameterswas further summarized in the principal component analysis(PCA) as a tool for simplifying the information from inter-correlated variables through linear transformation of the orig-inal variables into a few principal components. PCA based oncorrelation matrix enabled to reduce the dimensions of mea-sured variables to the representative principal components.The results are presented in the component score and com-ponent weight plots showing the relationships among theparameters and their role in the evaluation of the samples aswell as the potential differences among various treatmentgroups. The length and direction of the lines represent thesignificance of the associated variables for the plotted compo-nents and for the discrimination of the samples based oncomponent scores. All statistical tests were performed withthe probability of type I error (α) set to be less than 0.05.
3. Results
In this study, four treatment groups of quails were fed 10 mLdaily of 3×105–3×108 cells/mL of a natural biomass with themajority content of M. aeruginosa for 10 and 30 days. The totalMC concentration in the biomass ranged from 4.5 μg/L (E1) to4600 μg/L (E4), consisting of about 40% each of MC-LR and MC-RR, 7% of MC-YR and 13% of unidentified MC-like compound(Table 1, Skocovska et al., 2007). The average weight ofexperimental animals was 205 g, Table 1 shows the dailydoses recalculated for the body weight. ELISA measurementsof MCs concentration in the liver andmuscles of experimentalbirds showed cyanotoxin accumulation in both acute and sub-chronic test (Fig. 1). The background values measured incontrols are caused by the unspecific matrix influence inELISA (Orr et al., 2003; Ernst et al., 2005). Many studies useELISA for determination of microcystins; LC-MC method isrecommended for better understanding of detoxification sinceit can distinguish the amount of MC-LR-GSH conjugate in
37S C I E N C E O F T H E T O T A L E N V I R O N M E N T 3 9 8 ( 2 0 0 8 ) 3 4 – 4 7
tissue (Dai et al., 2008). Low accumulation of MCs was ob-served in the muscles. There has been about 40% increase ofMCs concentration in the highest treatment group relative tothe background values in control in the muscles of both acuteand sub-chronic test, but this difference was statistically sig-nificant only for the acute test. On the other hand, significantaccumulation in dependence on exposure concentration wasobserved in the liver for both test durations. There wasgreatest accumulation in the liver from the acute test, wherethe average concentration reached 43.7 ng MCs/g FW in thehighest treatment group, while no significant MC accumula-tion was found for any of the other treatments in acute test. Inthe sub-chronic test, there has been significant accumulationof MC in the two highest treatment groups (E3, E4) withaverage concentration 2.7 and 7.5 ng MCs/g FW, respectively.However, there has been relatively great variability in theconcentrations among individuals within the greatest expo-sure groups reflecting interindividual differences.
Activities of cytochrome P-450-dependent 7-ethoxyresor-ufinO-deethylase (EROD) in the studied tissueswere increasedafter exposure to cyanobacterial biomass namely in the acutetest (Fig. 2). There was a significant increase from 0.7 to1.15 pmol resorufin/min/mg protein in the heart from acutetest, the increase of EROD activity from 3.1 to 4 pmol resorufin/min/mg protein in the brain from acute test (both starting atthe second lowest exposure group E2); similar increase wasfound in the brain from sub-chronic test.
The levels of GST activity in tissues from sub-chronic testshowed more distinct changes in comparison with acute test(Fig. 3). There was a non-significant increase in the liver GSTactivities from acute test (270 nmol/min/mg protein in controlto 320 nmol/min/mg protein in the highest exposed group E4).
On the other hand, the GST activities have been significantlyincreased in the liver of birds from all cyanobacteria-exposedgroups in sub-chronic exposure compared to control. The sub-chronic exposure to higher cyanobacterial concentration (E3)leads to the increase of GST activity also in the heart (36 to44 nmol/min/mg protein) and brain of the birds, while therehas been no effect in these two tissues after acute exposure.
The cyanobacterial exposure caused an increase in GSHlevel inmost tissues in both acute and sub-chronic test (Fig. 4).There was a significant dose-dependent increase of GSH levelin the liver and brain from sub-chronic test. A morepronounced effect was observed in the liver from acute testwhere the GSH level increased fivefold already in the lowestbiomass concentration (E1). Significant increase of GSH level (8to 14 nmol/mg protein) was also detected in the heart fromacute test. On the other hand, there was a decrease of GSHcompared to control (45 to 35 nmol/mg protein in E2) in thebrain of E1 and E2 groups of the acute test.
Both glutathione peroxidase and reductase activitiesslightly increased in the liver of the lowest exposure groupin the acute test (data not shown). GR activity was elevatedafter acute exposure also in the heart in the highest exposuregroup. The sub-chronic exposure caused increase of GPXactivity in the group E3 in the brain. On the other hand,significant decrease of GPX activity was measured in the liverfrom sub-chronic test scheme.
Dose-dependent increase in lipid peroxidation was ob-served in the heart from acute test (from 0.55 to 1.2 nmolTBARS/mg protein) (Fig. 5). The lowest tested concentrationhas induced lipid peroxidation also in the liver and brain ofbirds from the acute exposure. In sub-chronic test, therewas an increase of lipid peroxidation in the heart (E1) and
Fig. 1 –Concentration of MCs (ng/g FW tissue) in pectoral muscles and liver from acute and sub-chronic test. The results areexpressed as mean±standard error. Asterisks indicate the statistically significant difference from the control group [LSD test;*=Pb0.05; **=Pb0.01; ***=Pb0.001].
38 S C I E N C E O F T H E T O T A L E N V I R O N M E N T 3 9 8 ( 2 0 0 8 ) 3 4 – 4 7
brain (E3). The greatest basal values of lipid peroxides (rangingfrom 2.5 to 3.2 nmol TBARS/mg protein) were withrespect to high content of unsaturated lipids found in theavian brain.
Significant correlations of responses in biochemical para-meters within all tested organs were found (pb0.05). Correlationof GSH and GST was found in both the liver from acute and sub-chronic test. Also significant were the correlations of TBARSwithGSHandGPx in the liver fromacute test, aswell as the correlation
of GPx with GSH and EROD in the liver from sub-chronic test.Similar correlationsandalsocorrelationofGRtoERODandTBARSwere found in the heart tissue. With respect to the brain tissue,more significant correlations were found in sub-chronic test,showing interrelation of GSH with GST and EROD at low p-level(b0.001). Interestingly, there were some inter-tissue correlationsof biochemical parameters, for example GST and also GSH in theliver and heart, EROD in the heart and brain and finally TBARS inthe liver and brain.
Fig. 2 –Activity of 7-ethoxyresorufin-O-deethylase (EROD; pmol resorufin/min/mg protein) in tested tissues. Activity of7-ethoxyresorufin-O-deethylase (EROD; pmol resorufin/min/mg protein) in tested tissues. Box includes 50% values, middlepoint is a median and whiskers show non-outlier range. Asterisks indicate the statistically significant difference from control[LSD test; *=Pb0.05; **=Pb0.01].
39S C I E N C E O F T H E T O T A L E N V I R O N M E N T 3 9 8 ( 2 0 0 8 ) 3 4 – 4 7
The PCA (Fig. 6) clearly separated the control group from allthe exposed groups. Also, the lowest exposure group could beclearly distinguished from the other treatments. On the otherhand, the threegreatest exposure groups (E2, E3, E4) couldnot beclearly separated, showing thus similar biochemical responses(Fig. 6A). Fig. 6B shows the component weights of individualbiochemical parameters used for the PCA and documents thatthe separation was driven namely by the modification of
glutathione-related parameters in the liver and also by changesin EROD activities and lipid peroxidation (TBARS) in the heart.
4. Discussion
Cyanobacterial metabolites are known to cause adverseeffects in diverse organisms including plants, mammals, fish
Fig. 3 –Activity of glutathione-S-transferase (nmol/min/mg protein) in tested tissues. Box plot parameters as in Fig. 2 [LSD test;*=Pb0.05; **=Pb0.01; ***=Pb0.001].
40 S C I E N C E O F T H E T O T A L E N V I R O N M E N T 3 9 8 ( 2 0 0 8 ) 3 4 – 4 7
and other aquatic organisms (Figueiredo et al., 2004). Theyhave also been linked with unnatural bird deaths (Alonso-Andicoberry et al., 2002; Ballot et al., 2005; Ndetei andMuhandiki, 2005; Lugomela et al., 2006). In the previouslypublished part of our study, we have demonstrated histo-pathological hepatic changes including swelling of hepato-cytes, vacuolar dystrophy, steatosis, hyperplasia of lymphaticcenters and shrunken nuclei of hepatocytes, cristolysis withinmitochondria and vacuoles with pseudomyelin structures onsub-cellular level after exposure to Microcystis biomass (Sko-covska et al., 2007). Apart from hepatic changes on both thecellular and sub-cellular level, there were increased activities
of lactate dehydrogenase and a drop in the blood glucose inthe group receiving the highest dose of cyanobacteria for10 days.
Our data clearly document bioaccumulation ofMCs namelyin the bird liver. Most studies on bioaccumulation of MCs areconcerned with fish, but there are also some data available forother animal species, including zooplankton, mollusks, snails,shrimps, livestock or mice (Nishiwaki et al., 1994; Amorim andVasconcelos, 1999; Beattie et al., 2003; Orr et al., 2003;Adamovsky et al., 2007; Chen and Xie, 2007; Xie et al., 2007).High levels of accumulated MC were found in the liver offlamingos dissected in case of mass deaths in Spain in 2002
Fig. 4 –Level of glutathione (nmol/mgprotein) in tested tissues. Boxplot parameters as in Fig. 2 [*=Pb0.05; **=Pb0.01; ***=Pb0.001].
41S C I E N C E O F T H E T O T A L E N V I R O N M E N T 3 9 8 ( 2 0 0 8 ) 3 4 – 4 7
(Alonso-Andicoberry et al., 2002). Themeasured concentrationwas three orders of magnitude higher (81 μg MC equivalent/gliver) than in our study and cyanobacteria have beensuggested as an important agent in the high mortality offlamingos. No lethal effects were observed in quails exposedto cyanobacterial biomass (Skocovska et al., 2007), eventhough the dose administered in the highest exposure group(230 ng MCs/g/day) was close to the previously published LD50
of about 250 ng/g/day MC-RR for quail after intraperitonealinjection (Takahashi and Kaya, 1993). This difference isprobably related to both the oral way of exposure and thecomplex biomass as exposure material. Previous studies with
mammals indicate that the damage of tissues caused by MC-LR is possible and the route of exposure via oral ingestion is30–100 times less toxic than via intraperitoneal injection(Fawell et al., 1999). Moreover, it has been suggested that themechanisms of the incorporation ofMC-LR into the liver by i.p.and p.o. administrations are greatly different (Nishiwaki et al.,1994). In our experiment, group E4 in sub-chronic exposureingested overall 1381 μg total MCs/205 g (i.e. 6737 ng/g) in thirtydays and the tissue concentration was 7.5 ng MCs/g FW liver,which represents 11‰ of total ingested MCs. This observationcorresponds to study with beef cattle (Orr et al., 2003), wherethe MC-LR equivalents in the liver represented about 12‰ of
Fig. 5–Level of lipid peroxidation (nmol TBARS/mg protein) in tested tissues. Box plot parameters as in Fig. 2 [*=Pb0.05; **=Pb0.01].
42 S C I E N C E O F T H E T O T A L E N V I R O N M E N T 3 9 8 ( 2 0 0 8 ) 3 4 – 4 7
ingested cyanotoxins. On the other hand, the uptake in theacute 10 day test reached 191‰ of the total dosed MC amount(43.7 ng MCs/g FW liver from the overall dose 460.44 μg totalMCs/205 g, i.e. 2246 ng/g).
The concentration of MCs in the hepatopancreas of variousfish species after p.o. exposure ranged from 0.22 μg/g FW to17.8 μg/g DW (Xie et al., 2004; Li et al., 2004; Soares et al., 2004;Zhao et al., 2006; Adamovsky et al., 2007), depending on thedose and duration of exposure. In the muscle of fish, theconcentration ranged 0.014 μg/g FW to 1.77 μg/g DW (Xie et al.,2004; Magalhaes et al., 2003; Soares et al., 2004; Zhao et al.,2006; Adamovsky et al., 2007). The reported MC accumulationfrom experimental studies with various fish species is some-what higher than the levels detected in birds in our study.
In our experiment, daily doses were within the range of0.23–225 ng MCs/g body weight and the maximal mean con-centration of MCs was 0.94 and 2.3 ngMCs/g in themuscle and7.5 and 43.7 ng MCs/g in the liver in sub-chronic and acutestudy, respectively. This difference in accumulation rate couldbe caused by the species differences. On the other hand, theratio (liver/muscle) of MCs concentration in group E4 rangingfrom 8.3 to 20.5 is close to the ratio observed in various fishspecies (liver/muscle ratio 10 to 20) (Williams et al., 1997; Li et al.,2004; Malbrouck and Kestemont, 2006; Adamovsky et al., 2007).
Next to the bioaccumulation this paper documents sig-nificant modulations of sub-lethal parameters in the exposedindividuals. To our knowledge, this is the first study focusedon both detoxification and antioxidant parameters in birds(see summary Table 2) after exposure to natural cyanobacter-ial biomass in a controlled experiment. Most of the studiedparameters have shown stronger modulation at the shortertime of exposure (acute test) than in the prolonged exposure.Also the blood hematological and biochemical parameters
have shown greater changes (stronger effects) on day 10 thanon day 30 as shown in our previous report (Skocovska et al.,2007).
Many enzymes are involved in the first biotransformationsteps by cytochrome P450 enzyme family. P450 induction hasbeen shown as a sensitive parameter reflecting the exposureof birds to various contaminants (Walker and Ronis, 1989;Barron et al., 1995). The EROD activity studied in this work isonly one representative of this large enzyme family and doesnot completely reflect the detoxification capacity. A goodagreement between EROD levels in our experiment andplateau assessed in study of five different bird species wasfound (Liukkonen-Anttila et al., 2003). Our study documentsan increase in EROD activity in the heart and brain aftercyanobacterial exposure. However, there was no significantincrease of this enzyme activity in the liver. Correspondingly,Wang et al. (2006) did not observe significant modulations ofcytochrome P450 1A mRNA levels in the liver of tilapiaexposed to MC-LR, whereas gene expression of GPx andsGST was increased significantly. The elevated EROD activityin the heart and brain could be thus probably linked to othercyanobacterial components than MC.
Conjugation of MC-LR with GSH catalyzed by GST is acrucial part of its detoxification pathway (Pflugmacher et al.,1998; Fu and Xie, 2005). Moreover, GSH might be responsiblefor the higher resistance toMCs (Qiu et al., 2007). GST activitiesincreased in all studied organs, but namely in the liver, only inthe sub-chronic exposure, while GSH in the liver wasincreased in both the acute and sub-chronic exposure. Theseresults agree with another report pointing out the increase ofGST activity in the early stages of zebra fish embryos after5 days of exposure to MC-LR (Wiegand et al., 1999). Contra-riwise, exposure to MCs caused no effect on GST activity in
Fig. 6 –Component score (A) and component weight (B) plots from principal component analysis: distribution of samples fromdifferent treatment groups (A) based on the pattern of biochemical parameters (B) in the liver, heart and brain.Abbreviations: C—control, E1 to E4 —exposure groups, L—liver, H—heart, B—brain, EROD: 7-ethoxyresorufin-O-deethylase,GST: glutathione-S-transferase, GSH: glutathione content, TBARS: lipid peroxidation measured as TBARS.
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experiments with rats (brain), mice (liver) and hepatocytes ofcommon carp (Cyprinus carpio) (Li et al., 2003; Gehringer et al.,2004; Maidana et al., 2006) and other experiments with fisheggs exposed to cyanobacterial extracts and fish exposed toMC-RR and MC-LR showed inhibition of GST activity (Pietschet al., 2001; Malbrouck et al., 2003; Cazenave et al., 2006).However, the quails have been exposed to the complex cyano-bacterial biomass, which contains many other componentsthan just MCs. Modulations of MC-effects by co-exposure tocyanobacterial lipopolysaccharides or other cyanobacterialmetabolites were reported (Pietsch et al., 2001; Best et al., 2002;Dvorakova et al., 2002; Wang et al., 2006) and therefore effectsobserved in any study with complex cyanobacterial biomassshould not be simply linked to the MCs (Falconer, 2007). Incorrespondence with the enhanced activity of GST, there wasan increased level of GSH in the liver, heart and brain fromsub-chronic test confirming the importance of these twobiomolecules in protection from the harmful effects of MC.The non-enzymatic compound GSH is considered the majorintercellular antioxidant, which serves also as a substrate forGPX to reduce peroxides and directly acts as a free radicalscavenger. A significant rise of GSH level was identically de-tected after exposure of rat hepatocytes to MC-LR (Bouaichaand Maatouk, 2004) and in the hepatopancreas of the silvercarp (Blaha et al., 2004; Li et al., 2007) exposed toMC-producingcyanobacterial water bloom. On the contrary, decreased levelof GSH in fish hepatocytes or no modulations of GSH level inCarassius auratus p.o. exposed to MC-LR were shown (Li et al.,2003; Malbrouck et al., 2004).
The response of GPX activities differed in the acute andsub-chronic exposure scheme. The increased GPX activityobserved in the liver from acute test corresponds with resultsof some other studies, including the induction of GPX activityin tilapia (Oreochromis sp.) and loach (Misgurnus mizolepis) p.o.exposed to MCs (Jos et al., 2005; Li et al., 2005) as well as GPXinduction in the hepatopancreas and intestines of Corydoraspaleatus exposed to 2 μg/L MC-RR (Cazenave et al., 2006) andenhanced GPX activity in mice liver after 32-hour study withMC-LR (Gehringer et al., 2004). However, no changes or de-crease in GPX activity were observed in tilapia (Oreochromis sp.)after acute exposure to MCs and cyanobacterial cells contain-ing microcystins (Prieto et al., 2006, 2007).
Furthermore, the increase in the liver and heart in acuteexposure and the decrease in the liver in sub-chronic exposurewas observed for GR activities. Enhanced GR activity wasfound also in the hepatopancreas and brain of MC-LR or MC-
RR exposed fish (Cazenave et al., 2006; Prieto et al., 2006),whereas depletion of GR activity was found in the liver andkidney from rats exposed to MC-LR and in tilapia exposed tocyanobacterial cells containing microcystins (Moreno et al.,2005; Prieto et al., 2007).
GPX has been shown to play an important role in protectionagainst lipid peroxidation via removal of lipid hydroperoxides(Wang et al., 2001). Lipid peroxidation, mostly measured asTBARS, is commonly understood as oxyradical production byperoxidation of cellular lipids and it is known to be induced bycyanobacterial toxins (Halliwell and Gutterdige, 1999). In ourstudy cyanobacterial biomass containing predominantly MC-LR and MC-RR induced significant increase of TBARS level inall studied organs, mostly at the lowest exposure concentra-tion (4.5 μg/L MCs). TBARS levels were increased also in thehepatopancreas, kidneys and gills of tilapia exposed to MCs(Prieto et al., 2006, 2007) and to crushed cyanobacterial cells(Jos et al., 2005), or in the hepatopancreas of silver carpexposed to cyanobacterial bloom dominated byM. ichthyoblabeand M. aeruginosa (Blaha et al., 2004). Significant increase ofTBARS level was detected in mice exposed to 75% LD50 dose ofpure MC-LR (Gehringer et al., 2004) and also in rat hippocam-pus after injection of MC-LR (Maidana et al., 2006). Anotherstudy, however, reported a decrease in TBARS level in thehepatopancreas and gills of MC-RR exposed fish (Cazenaveet al., 2006) or no changes of lipid peroxidation in fish fed withcyanobacterial biomass (Li et al., 2005).
Correlations among the oxidative stress parameters anddetoxification enzymes activities illustrate the complex char-acter of the response and interdependence among parameters.The interrelation was demonstrated in the liver as the mostimportant place of detoxification of xenobiotics in birds (Riviereet al., 1985),with strongpotential for impact fromMCsknownasstronghepatotoxic agents.Moreover, significantmodulationsofdetoxification and antioxidative compounds and their relationswere also found in the heart and brain, indicating that theseorgans are also highly affected, since they showed increasedlipid peroxidation after cyanobacterial exposure. The observedeffects of MC-containing cyanobacterial biomass in the brainwell correspond with recent findings that MC-LR inducesoxidative stress in rat brains along with behavioral changes(Maidana et al., 2006).Moreover, organic anion transport proteinOATP1A2 expressed in human liver and brain has beendemonstrated tomediate intracellular uptake ofMC-LR (Fischeret al., 2005), which further indicates the brain as another targetof MC toxicity.
Birds coming to contact with eutrophicated aquatic eco-systems seem to react to the secondary metabolites ofcyanobacterial blooms as to xenobiotics. The overall patternof detoxification and oxidative stress responses clearly sepa-rates the control and the lowest exposure group from all thehigher exposed groups documenting the shift in the detox-ification and antioxidative balance after cyanobacterial expo-sure. General activation of the antioxidant enzymatic systemin quails after the exposure to natural cyanobacterial biomassdocuments the occurrence of oxidative stress in the studiedorgans and their ability to produce antioxidative moleculesprotecting cells against adverse oxidation processes.
Interesting differences were found between acute and sub-chronic exposure, both in the biochemical and accumulation
Table 2 – Summary of the effects of cyanobacterialbiomass with majority content of MC-LR and MC-RR onbird antioxidative and detoxification system
GSH GST GPX GR TBARS EROD
Acute test Liver ↑ – ↑ ↑ ↑ –Heart ↑ – – ↑ ↑ ↑Brain ↓ – – – ↑ ↑
Sub-chronic test Liver ↑ ↑ ↓ – – –Heart – ↑ – – ↑ –Brain ↑ – ↑ – ↑ ↑
Statistically significant increase ↑ and decrease ↓; Pb0.05.
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parameters, which may indicate the potential adaptation ofthe detoxification and antioxidative system to the exposurewith cyanobacterial biomasswith increasing time of exposure.They are characterized namely by dose-dependent increase ofGSH content as well as significant increase of GST activity inall three organs in the sub-chronic exposure. The increase ofGST activity and GSH content corresponds with their role indetoxification pathway of MCs (Pflugmacher et al., 1998; Fuand Xie, 2005), which implies that with increasing concentra-tion of MCs there is increasing need of GST and GSH providingthe conjugation of MCs to less toxic compound. On the otherhand, the higher accumulation in acute test has been linkedwith stronger changes of other detoxification (EROD) andoxidative stress parameters (GPx, GR). Taken together, thesedata document that increased activities of detoxification en-zymes could lead to greater biotransformation and elimina-tion of the MCs from both liver and muscle and thus loweraccumulation at the longer exposure time. This inferencecorresponds to the six times lower accumulation of MC in theliver of the highest exposure group of the sub-chronic testcompared to the acute one. These results support the hypo-thesis of the potential adaptation of the avian detoxificationsystem to the sub-chronic exposure.
5. Conclusions
The exposure of model birds to natural cyanobacterial biomasscaused significant changes in levels and activities of antiox-idative and detoxification compounds and accumulation ofcyanotoxins mainly in the liver and little accumulation in themuscles. Cyanobacteria are thus capable to induce oxidativestress responses in birds linked with activation or inhibition ofdetoxification compounds. The generation of oxidative stresscombined with insufficiency of defense mechanisms could insensitive species at prolonged exposure potentially result ineffects on the health status, especially if other stressors areinvolved at the same time, which is often the case in theenvironment.
Acknowledgements
Supported by project No. 1 M6798593901 of the programme“Research Centres PP2 — DP01”(1 M), project AVOZ60050516and project MSMT No. 6215712402.
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Paper IV.
Pašková, V., Paskerová, H., Pikula, J., Ban ouchová, H., Sedlá ková, J. and Hilscherová, K. (2011).
Combined exposure of Japanese quails to cyanotoxins, Newcastle virus and lead: Oxidative
stress responses.
Ecotoxicology and Environmental Safety 74 (7): 2082-2090.
Author's personal copy
Combined exposure of Japanese quails to cyanotoxins, Newcastle virusand lead: Oxidative stress responses
Paskova Veronika a, Paskerova Hana a, Pikula Jiri b, Bandouchova Hana b, Sedlackova Jana b,Hilscherova Klara a,n
a Research Centre for Toxic Compounds in the Environment (RECETOX), Faculty of Science, Masaryk University, Kamenice 126/3, 625 00 Brno, Czech Republicb Department of Veterinary Ecology and Environmental Protection, Faculty of Veterinary Hygiene and Ecology, University of Veterinary and Pharmaceutical Sciences Brno,
Palackeho 1/3, 612 42 Brno, Czech Republic
a r t i c l e i n f o
Article history:
Received 29 January 2011
Received in revised form
30 May 2011
Accepted 17 July 2011
Keywords:
Multiple exposure
Coturnix coturnix japonica
Cyanobacteria
Lead
Vaccine strain
Detoxification
a b s t r a c t
Wild birds are continually exposed to many anthropogenic and natural stressors in their habitats. Over
the last decades, mass mortalities of wild birds constitute a serious problem and may possibly have
more causations such as natural toxins including cyanotoxins, parasitic diseases, industrial chemicals
and other anthropogenic contaminants. This study brings new knowledge on the effects of controlled
exposure to multiple stressors in birds. The aim was to test the hypothesis that influence of
cyanobacterial biomass, lead and antigenic load may combine to enhance the effects on birds, including
modulation of antioxidative and detoxification responses. Eight treatment groups of model species
Japanese quail (Coturnix coturnix japonica) were exposed to various combinations of these stressors. The
parameters of detoxification and oxidative stress were studied in liver and heart after 30 days of
exposure. The antioxidative enzymatic defense in birds seems to be activated quite efficiently, which
was documented by the elevated levels and activities of antioxidative and detoxification compounds
and by the low incidence of damage to lipid membranes. The greatest modulations of glutathione level
and activities of glutathione-S-transferase, glutathione peroxidase, glutathione reductase, superoxide
dismutase, catalase and lipid peroxidation were shown mostly in the groups with combined multiple
exposures. The results indicate that the antioxidative system plays an important role in the protective
response of the tissues to applied stressors and that its greater induction helps to protect the birds from
more serious damage. Most significant changes of these ‘‘defense’’ parameters in case of multiple
stressors suggest activation of this universal mechanism in situation with complex exposure and its
crucial role in protection of the bird health in the environment.
& 2011 Elsevier Inc. All rights reserved.
1. Introduction
Combined exposure to both natural and anthropogenic stres-sors is a common problem in the polluted environment, wherethe organisms have to face multiple stressors including pathogens(Norris and Evans, 2000; Sagerup et al., 2009).
Mass development of cyanobacteria has become a seriousproblem in water bodies in many parts of the world. Moreover,their secondary metabolites, especially cyanotoxins, have beenshown to cause adverse effects in various organisms includingbirds (de Figueiredo et al., 2004; Wiegand and Pflugmacher, 2005;Skocovska et al., 2007; Paskova et al., 2008; Peckova et al., 2009).Some cases of mass mortality of wild birds in various parts of theworld have been suggested to be associated with the exposure to
toxic cyanobacteria since high concentrations of cyanotoxins havebeen reported in the water or in the stomach or intestine of deadbirds (Krienitz et al., 2003; Ballot et al., 2004, 2005; Ndetei andMuhandiki, 2005; Lugomela et al., 2006).
This association of cyanobacteria with bird mortalities has beenbased on field observations in wild birds under natural conditions.There are very few laboratory experiments focused on effects ofcyanobacteria in birds (Damkova et al., 2009, in press; Paskovaet al., 2008; Skocovska et al., 2007). Our previous laboratory studieswith model bird (Japanese quail, Coturnix coturnix japonica) exposedto cyanobacterial biomass showed accumulation of cyanotoxinsmicrocystins. Negative effects included significant induction ofoxidative stress parameters in liver, heart and brain and biochem-ical, hematological, subcellular and histopathological hepaticchanges (Paskova et al., 2008; Skocovska et al., 2007). Despite thesechanges no mortality has been reported. The above mentionedstudies focused on the effects of cyanobacterial biomass, but wildbirds can be certainly exposed to many other stressors. The mass
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Ecotoxicology and Environmental Safety
0147-6513/$ - see front matter & 2011 Elsevier Inc. All rights reserved.
doi:10.1016/j.ecoenv.2011.07.014
n Corresponding author. Fax: þ420 54949 2840.
E-mail address: hilscherova@recetox.muni.cz (H. Klara).
Ecotoxicology and Environmental Safety 74 (2011) 2082–2090
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mortalities may possibly have more causations such as othernatural toxins, pathogens, industrial chemicals and other anthro-pogenic contaminants.
Important widespread anthropogenic contaminants are heavymetals (HMs; Blus et al., 1995), which may constitute a seriousenvironmental problem because of their high concentrations andbioavailability for biota (Sanchez-Chardi et al., 2007). HMs canenter the environment from various sources such as wasteincineration, production of energy, use of lead shots for huntingand from former mining sites (Scheuhammer et al., 1998; Pacynaet al., 2009). Accumulated HMs were found in biota, includingbirds (Savinov et al., 2003; Bonilla-Valverde et al., 2004; Sanchez-Chardi et al., 2007). Some cases of lead-poisonings are knownmostly for aquatic birds as for example tundra and whooperswans (Cygnus columbianus; Cygnus cygnus) (Sileo et al., 2001;Degernes et al., 2006; O’Connell et al., 2008), Canada geese(Branta canadensis) and mallards (Anas platyrhynchos) (Hennyet al., 2000; Sileo et al., 2001) and other birds belonging to groupsAnatidae, Charadriidae, Scolopacidae and Rallidae (Pain, 1990).They have been linked with contaminated sediments in regionswith frequent birds-hunting (Guillemain et al., 2007) or formermining sites (Blus et al., 1991,1995). Both terrestrial and aquaticbirds have been shown to be exposed to lead by oral ingestion ofspent lead shot or bullet fragments (Fisher et al., 2006). Lead shotswere found in stomach and increased lead concentrations also inliver of some dead birds (Blus et al., 1995; Beyer et al., 1998).
Infectious diseases can be another frequent stressors in naturalconditions linked with mass mortalities. The most commoncauses of mass mortalities are bacterial (Parmelee et al., 1979;Delisle et al., 1990; Kwon and Kang, 2003; Waller and Underhill,2007), fungal (Stone and Okoniewski, 2001; Sotero-Santos et al.,2006) or viral infections. Well known are for example the WestNile virus or Newcastle disease virus, which caused mortalities ofaquatic birds in USA and China (Liu et al., 2008; Sovada et al.,2008). Pathogens induce immune responses of the defensesystem in birds, which can become more susceptible to otherstressors when investing energy to avoid pathological side-effectscaused by an elevated immune response (Costantini and Moller,2009). These stressors are able to influence defense system oforganisms and their negative effects may combine in naturalconditions. Moreover, chemicals may modify the effect of othersby altering their kinetics and/or dynamics (Costantini and Moller,2009; Naraharisetti et al., 2009).
Among many negative effects of cyanobacteria and heavymetals in the environment, there is at least one shared mechanismof action—their ability to increase the generation of reactiveoxygen species (ROS) such as the superoxide anion radical (O2� .),hydrogen peroxide (H2O2) and hydroxyl radicals ( dOH) (Stohs andBagchi, 1995; Ding et al., 2000; Li et al., 2003). Formation of ROSand oxidative stress are also associated with the development ofmany pathological states and damage, including immuno-pathol-ogy (Costantini and Moller, 2009), which may result from highdose of ROS released during the immune response. Oxidative stressmay occur either due to the decrease of the cellular antioxidantlevel or to the overproduction of ROS (Halliwell and Gutterdige,1999).
Exposure to cyanobacterial biomass or purified cyanotoxinsmicrocystins (MCs) has been shown to cause oxidative stress invarious organisms, including birds (Ding and Ong, 2003; Wiegandand Pflugmacher, 2005; Adamovsky et al., 2007; Paskova et al.,2008). Similarly, oxidative stress was observed in birds also afterexposure to lead (Mateo et al., 2003; Douglas-Stroebel et al., 2004).Induction of an immune response by e.g. bacterial or parasiticinfections linked with oxidative stress has been also shown in adiverse group of organisms, including some bird species (Georgievaet al., 2006; Liu et al., 2008). The modulation of oxidative stress
markers in birds after infection by some viruses has been demon-strated in a few studies as for example in a study with chickeninfected by Marek’s disease (Keles et al., 2010).
Liver as the general detoxifying organ is considered the mainregion of ROS generation in mammals and birds (Prieto et al.,2006). Endogenous enzymatic and non-enzymatic antioxidantdefenses are critical for the control of ROS-mediated oxidativedamage of biomolecules including proteins, RNA, DNA and mem-brane polyunsaturated lipids (Halliwell and Gutterdige, 1999).The main defense mechanisms against ROS and their toxic by-products include enzymes, glutathione-S-transferases (GSTs),glutathione reductase (GR), glutathione peroxidase (GPx), catalase(CAT) and superoxide dismutases (SOD), in particular, and alsonon-enzymatic compounds such as glutathione (GSH).
Significant modulations of the antioxidative and detoxificationsystem together with increased production of lipid peroxideswere observed in Japanese quails exposed to cyanobacterialbiomass (Paskova et al., 2008). That 10- and 30-day exposurealso resulted in accumulation of microcystins, but no mortality.
In this study, the aim was to test the hypothesis that combinedexposure to cyanobacteria, lead and immunological challengeenhances effects on birds including modulation of antioxidativeand detoxification responses. For this purpose we evaluated theeffects of single and combined exposures to cyanobacterialbiomass, lead shots and Newcastle disease vaccination in stan-dard model bird Japanese quail. We investigated the effects ofthese three stressors and their combinations on the hepatic andcardiac levels and activities of biotransformation and antioxida-tive compound GSH and enzyme GST. Further, we also studied theactivities of other antioxidative enzymes GR, GPx, SOD and CATand evaluated lipid peroxidation as a measure of oxidativedamage in the exposed birds.
2. Materials and methods
2.1. Bioassay
The study employed 30-day single and combined exposures of Japanese quails
to cyanobacterial biomass, lead and immunologic challenge of a live Newcastle
vaccination strain performed according to OECD Guideline for the testing of
chemicals 205—Avian Dietary Toxicity Test (OECD, 1984). Experiments were
conducted in compliance with laws for the protection of animals against cruelty
and were approved by the Ethical Committee of the University of Veterinary and
Pharmaceutical Sciences Brno, Czech Republic. Permit No. 9221/2009-30 was
issued by the Ministry of Education, Youth and Sports of the Czech Republic.
2.2. Experimental design
The experiment was performed with four months old male Japanese quails
(average weight 219 g) held individually in standard laboratory cages for birds
(floor area 1500 cm2/bird). Controlled conditions were maintained throughout the
test, i.e. temperature 25 1C, 12 h of light per day, light intensity 10 lx, relative
humidity 60%, ventilation 8 air changes per hour.
All birds were provided with commercial feeds and drinking water ad libitum
during the experiment. A total of 40 birds were divided on a random basis into
8 groups of 5 individuals (cf. Table 1 for labeling and description of treatments in
groups). Briefly, C¼control group, B¼cyanobacterial biomass-exposed group,
V¼Newcastle-vaccinated group, Pb¼ lead-exposed group, BPb¼cyanobacterial
biomassþ lead-exposed group, BV¼cyanobacterial biomass-exposedþNewcastle-
vaccinated group, PbV¼ lead-exposedþNewcastle-vaccinated group and BPbV¼cyanobacterial biomass-exposedþ lead-exposedþNewcastle-vaccinated birds.
The design of the cyanobacterial exposure and the preparation of the
cyanobacterial biomass have been described in our previous papers (Skocovska
et al., 2007; Pikula et al., 2010). Briefly, microcystin content in biomass was
analyzed using HPLC-DAD (Agilent 1100 Series) on Supelcosil ABZþPlus column,
150�4.6 mm, 5 mm according to Babica et al. (2006). Birds from groups B, BPb, BV
and BPbV were fed twice a day using a crop probe to reach the daily dose of 10 mL
of cyanobacterial biomass (1.92�109 cells, 83.46 mg of dry weight; microcystin
structural variants: 15.36 mg MC-RR, 12.70 mg MC-YR, 17.98 mg MC-LR,
46 mg sum of MCs) for 30 days. The birds not receiving the daily dose of 10 mL
of cyanobacterial biomass (i.e. groups C, Pb, V and PbV) were administered twice a
P. Veronika et al. / Ecotoxicology and Environmental Safety 74 (2011) 2082–2090 2083
Author's personal copy
day with 5 mL of control water to imitate the intake of experimental biomass by
the crop probe. Birds from groups V, BV, PbV and BPbV were vaccinated into
nostrils with live Newcastle disease vaccination strain (Avipest lyof. a.u.v. contain-
ing Paramyxovirus pseudopestis avium phyl. La Sota min. 106.0 EID50 per dose,
Mevak a.s., Nitra, Slovakia) at the beginning of the experiment to induce antigenic
stress and immune response. The intranasal administration of 0.05 mL of the
vaccine reconstituted in physiological saline solution was done as recommended
by the producer. Each bird from groups Pb, BPb, PbV and BPbV was given six
3.5 mm in diameter lead shots (containing in total 1.38 to 1.59 g lead per bird) into
the crop on day 0 of the experiment in order to induce lead toxicosis (Pikula et al.,
2010). Lead shots were produced by the ammunition company Sellier & Bellot
(Vlasim, Czech Republic). After the 30 days lasting exposure period, birds were
sacrificed by decapitation. Selected organs (liver and heart) were dissected and
stored at �80 1C for measurement of biochemical oxidative stress parameters.
2.3. Biochemical methods
All biochemicals and enzymes were purchased from Sigma-Aldrich (Prague,
CZ); other chemicals used for preparation of buffers were of the highest
commercial grade available. The biochemical analyses were performed in liver
and heart tissues. The tissues were homogenized on ice using mechanical
homogenizer (100 mg of tissue in 1 mL) in potassium phosphate buffer (50 mM
KH2PO4 with 1 mM EDTA, pH 7.4) for assessment of CAT and SOD and in
phosphate buffer saline (PBS, pH 7.2) for the other parameters. The postmitochon-
drial supernatant was collected after centrifugation (30 min at 30,000 g at 4 1C for
CAT and SOD and 15 min at 10,000 g at 4 1C for the other parameters) and stored
frozen at �80 1C until biochemical analyses.
The methods for assessment of most biochemical markers measured have
been described earlier (Paskova et al., 2008). Briefly, the glutathione-S-transferase
(GST) activity was measured spectrophotometrically using 1-chloro-2,4-dinitro-
benzene (Habig et al., 1974). The concentration of reduced glutathione (GSH) was
determined using 5,50-dithiobis-2-nitrobenzoic acid (DTNB) as a substrate
(Ellmann, 1959). Activities of glutathione peroxidase (GPx) and glutathione
reductase (GR) were determined from the rate of NADPH oxidation (Flohe and
Gunzler, 1984). The level of lipid peroxidation and also stimulated lipid peroxida-
tion in avian tissues was assessed as total thiobarbituric acid reactive species
(TBARS; Uchiama and Mihara, 1978; Livingstone et al., 1990). Activity of super-
oxide dismutase (SOD) was determined spectrophotometrically at 560 nm accord-
ing to the method using nitroblue tetrazolium (NBT) as a substrate (Ewing and
Janero, 1995). The reaction mixture contained 60 mM NBT, 100 mM NADH and
35 mM phenazine methosulfonate in 50 mM potassium phosphate/1 mM EDTA
buffer. Activity of catalase (CAT) was evaluated spectrophotometrically at 240 nm
in cuvettes as a rate of hydrogen peroxide break down (Aebi, 1984) in the mixture
containing 0.09% hydrogen peroxide in 50 mM TRIS/0.1 mM EDTA buffer. The
protein concentrations were determined by the method using the Folin-Ciocalteu
phenol reagent (Lowry et al. 1951). The GENios spectrophotometric reader (Tecan
Group, Switzerland) was used to measure the absorbance in microplates and
spectrophotometer VARIAN CARY 50 Bio (Varian, USA) was used for the measure-
ment of absorbances of solutions in cuvettes.
Methods for the assessment of the content of lead in liver and in the shots as
well as of the biochemical, hematological, serological and toxicological parameters
in blood samples and histological changes in tissues of exposed quails have been
described previously (Pikula et al., 2010). Briefly, for the metal analysis (according
to ISO 11466) samples (1 g dry wt) were leached with 2.3 mL HNO3 and 7 mL HCl
overnight followed by 2 h heating under reflux and analyzed by ICP-MS (Agilent
7500ce, Agilent Technologies, Japan).
Microcystin concentrations in liver of exposed birds were determined as
previously described (Babica et al., 2006; Skocovska et al., 2007). Element analysis
of cyanobacterial biomass was also performed by inductively coupled plasma-
mass spectrometry (ICP-MS; Agilent 7500ce, Agilent Technologies, Japan) and
resulted in finding 1.45 g/kg Na, 8.7 g/kg P, 11.0 g/kg K, 11.5 g/kg Ca, 0.43 g/kg Fe,
3.1 mg/kg As, 1.28 mg/kg Se and 1.7 mg/kg Pb.
2.4. Statistical evaluation
Statistical analyses were performed with Statistica for Windowss 7.0 (StatSoft,
Tulsa, OK, USA). Results from different treatment groups were compared by one-
way analysis of variance (ANOVA) and post-hoc analysis of means using the LSD
test. Homogeneity of variances was tested by Levene’s test. Parameters that were
not homogenous were log-transformed prior to analysis. In case of non-homo-
genous variances (GSH hepatic level and GST cardiac activity) the nonparametric
Kruskal–Wallis test was used for comparison of treatment groups. Spearman
rank order correlations were used to assess relationship among the measured
parameters.
Variation of the biochemical parameters was further summarized in the
principal component analysis (PCA) as a tool for simplifying the information from
inter-correlated variables through linear transformation of the original variables
into a few principal components. PCA based on a correlation matrix enabled to
reduce the dimensions of measured variables to the representative principal
components. The results are presented in the component score and component
weight plots showing the relationships among the parameters and their role in the
evaluation of the samples as well as the potential differences among various
treatment groups. The length and direction of the lines represent the significance
of the associated variables for the plotted components and for the discrimination
of the samples based on component scores. All statistical tests were performed
with the probability of type I error (a) set to be less than 0.05.
3. Results
This study showed differences in effects of the individual andcombined exposure to three environmentally relevant stressors(cyanobacterial biomass, lead and immunological challenge) onmodel bird Japanese quail. There was no mortality among controlJapanese quails (C), vaccinated controls (V), groups of single(B) and Newcastle-vaccinated (BV) cyanobacterial biomass-expo-sure and in group BPbV exposed to cyanobacterial biomass andlead and vaccination. One bird died in the single lead-exposure(Pb) and the combined exposures to cyanobacterial biomass andlead (BPb) and lead and Newcastle vaccination (PbV) resulted inthe death of two out of five birds.
The analysis of microcystin and lead concentrations in expo-sure material and in tissues have been described earlier (Pikulaet al., 2010). Weight of lead shots (prevailing content of Pb, withtraces of As and Sb) found in birds from lead (co)exposed groupsafter 30 days of exposure was 0.507–1.49 g lower than uponadministration. Table 2 summarizes the liver lead and microcys-tins concentrations at the end of exposure. Microcystins weredetected only in groups B, BV, BPb and BPbV. In groups BPb andBPbV the mean levels were higher than in other groups, but thedifferences were not statistically significant also due to the highvariation among individuals. There were also no differences inliver lead concentrations among lead-exposed groups (mean7.275.16 mg/g f.w.). The livers of birds from groups without leadexposure contained low background lead concentrations withmean around 0.07 mg/g, which could be originating from somematerials used during breeding.
Table 1Labeling and characterization of experimental groups.
Abbreviation Exposure Dosing
C Control 10 mL of control water/day
B Cyanobacterial biomass 10 mL of cyanobacterial biomass/day
Pb Lead 6 lead shots at the beginning of experimentþ10 mL of control water/day
V Vaccination vaccination at the beginning of experimentþ10 mL of control water/day
BPb Cyanobacterial biomassþ lead 6 lead shots at the beginning of experimentþ10 mL of cyanobacterial biomass/day
BV Cyanobacterial biomassþvaccination vaccination at the beginning of experimentþ10 mL of cyanobacterial biomass/day
PbV Leadþvaccination 6 lead shots at the beginning of experimentþvaccination at the beginning of
experimentþ10 mL of control water/day
BPbV Cyanobacterial biomassþ leadþvaccination 6 lead shotsþvaccination at the beginning of experimentþ10 mL of
cyanobacterial biomass/day
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The studied parameters of biotransformation, antioxidativeresponses and oxidative stress measured in liver and heart tissuesafter 30-day exposure differed significantly among the eighttreatment groups (Table 3). The level of the non-enzymaticantioxidant glutathione (GSH) was about threefold greater inthe co-exposures to cyanobacteriaþ lead (BPb) and cyanobacter-iaþ leadþvaccination (BPbV) groups compared to controls andalso to all the individual exposures in the liver (groups C, B, Pband V). Another pattern was obvious for GSH in heart where asignificant elevation against control was found in most single andcombined exposed groups except V and BV groups (Fig. 1).
Significantly higher hepatic glutathione-S-transferase activ-ities than in control and all single exposures were observed after
exposure to the group BPbV. In case of heart there was asignificantly higher GST activity in PbV and BPbV groups whencompared with both control and lead-exposed birds (Fig. 2).
The glutathione peroxidase activity was significantly elevatedin the BPb group in comparison with the single cyanobacterialexposure B in the liver. There was also a significant increase in theBPbV group against control and all the single exposure groups inliver. GPx activities in both co-exposure groups BPb and BPbVwere significantly threefold higher than control and all the singleexposures in heart. Moreover, an elevated cardiac GPx activityagainst control C and single V and B exposure groups was found inthe combined BV group (Fig. 3).
Significant increases of the glutathione reductase activity weredetected after both single vaccine and cyanobacterial exposuresin the liver. On the other hand, the hepatic GR of group BPbV didnot differ from the control group and it was significantly lowerwhen compared with B and V groups. Other results were obtainedfor heart where the GR activity in co-exposures BPbV, PbV andsingle B groups were significantly elevated against control andsingle V and Pb groups (Fig. 4).
Other biomarkers of antioxidative protection SOD and CATactivity were measured in liver but only in case of SOD activitythere was a significant fivefold increase in the BPbV co-exposedgroup in comparison with the single cyanobacterial exposure B.There were no statistically significant differences in the CATactivity among different exposure group (data not shown).
There was a significantly higher level of lipid peroxides studiedas markers of unsaturated fatty acids damage in the liver of BPbVco-exposed birds than in control (from 0.8 to 2 nmol TBARS/mg
Table 2Liver lead and microcystins (MCs) concentrations evaluated after 30-day exposure
(C¼control group; exposure-B: cyanobacterial biomass, Pb: lead, V: vaccination,
f.w.¼ fresh weight).
Exposure group Lead concentration
(mg/g f.w. tissue)
MCs concentration
(ng/g f.w. tissue)
C 0.1170.15
B 0.0770.05 39.9717.7
V 0.0570.01
BV 0.0770.03 36.9711.4
Pb 8.5078.99
BPb 6.8275.91 61.1733.1
PbV 6.9972.74
BPbV 6.5470.81 48.4710.6
Table 3Statistically significant increase m and decrease . (po0.05) of parameter in specific group compared to control and/or relevant single exposure groups; ’ no statistically
significant effect, – not measured.
GSH GST GPx GR LP stim. LP SOD CAT
m m m m m m m
Liver BPbV BPbV BPb B PbV BPbV BPbV
m BPbV V BPbV ’
BPb .
BPbV
m m m m m
Heart Pb PbV BV B PbV
B BV BPb PbV BV ’ – –
PbV BPbV BPbV BPbV BPbV
BPb
BPbV
Abbreviations: GSH: glutathione content, GST: glutathione-S-transferase, GPx: glutathione peroxidase, GR: glutathione reductase, LP stim: stimulated lipid peroxidation,
LP: lipid peroxidation, SOD: superoxide dismutase, CAT: catalase; Treatment labels—B: cyanobacterial biomass, Pb: lead, V: vaccination
*C
16
20
24
28
32
36
40
44
4
8
12
16
20
24
28
GSH
nm
ol /m
g pr
otei
n
C,Pb,B,VPbV,BV
C,Pb,B,VPbV,BV C
C
C CC
C PbVPb BV BPbVVPb BV BPbVV
BPbB C PbV BPbB
Fig. 1. Level of glutathione (nmol/mg protein) in liver (A) and heart (B). Box includes 50% values, middle point is median and whiskers show non-outlier range. Letters
indicate the statistically significant difference from control (C) or relevant treatment groups (lead Pb, cyanobacterial biomass B, vaccination V) [LSD test].
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protein), vaccinated birds and lead co-exposed birds from groupsPbV and BPb. In case of heart there was no statistically significantdifference in TBARS among treatment groups (Fig. 5), butincreased susceptibility to lipid peroxidation was evidenced forgroups PbV, BV and BPbV compared to single cyanobacterialexposure (data not shown).
Significant correlations among responses of the biochemicalparameters were foundmostly in liver tissue. GSH and GST positivelycorrelated with stimulated and non-stimulated TBARS, and there was
a positive correlation between GSH and GST both in liver and heart.Liver GSH positively correlated with GPx as well. There was also asignificant correlation of TBARS with GPx and SOD.
The PCA analysis revealed the multivariate pattern ofresponses of biochemical parameters and allowed the variablesand samples to be projected onto a two dimensional space. Themultivariate analyses combining parameters measured in liverand heart showed that studied parameters and their mutualassociation clearly separated the biomass-exposed groups
30
40
50
60
70
C,Pb C,Pb
200
240
280
320
360
400
440
C,Pb,B,V
GST
nm
ol /
min
/ m
g pr
otei
n
C PbVPb BV BPbVVPb BV BPbVV
BPbB C PbV BPbB
Fig. 2. Activity of glutathione-S-transferase (nmol/min/mg protein) in liver (A) and heart (B). Box plot parameters as in Fig. 1.
4.5
4.0
3.5
3.0
2.5
2.0
1.5
1.0
0.5
1.0
0.8
0.6
0.4
0.2
0
C,Pb,B,V
C,B,V
C,Pb,B,VC,Pb,B,V
GPx
nm
ol N
AD
PH/ m
in /
mg
prot
ein
B
C PbVPb BV BPbVVPb BV BPbVV
BPbB C PbV BPbB
Fig. 3. Activity of glutathione peroxidase (nmol NADPH/min/mg protein) in liver (A) and heart (B). Box plot parameters as in Fig. 1.
1.6
1.4
1.2
1.0
0.8
0.6
0.4
0.2
0.0
-0.25
6
7
8
9
10
C,Pb,VC,Pb,VC,Pb,VCC
B,V
C PbVPb BV BPbV
GR
nm
ol N
AD
PH/ m
in/ m
g pr
otei
n
VPb BV BPbVVBPbB C PbV BPbB
Fig. 4. Activity of glutathione reductase (nmol NADPH/min/mg protein) in liver (A) and heart (B). Box plot parameters as in Fig. 1.
P. Veronika et al. / Ecotoxicology and Environmental Safety 74 (2011) 2082–20902086
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(B, BPb and BPbV) from controls (Fig. 6A). Most parametersassessed in liver formed the primary trajectory, which explainedthe greatest proportion of the variability and was associated with
the 1st principal component (explaining 27.6% of variation). Onthe other hand, the parameters measured in heart contributedmore to the 2nd and 3rd principal components (Fig. 6B), whichexplained 15.7% and 12.9% of variation, respectively. Interestingly,the biomass-exposed groups were separated from each otheralong the PC1 axes (driven by liver glutathione parameters andTBARS), with BPb group separated in the same direction as BPbV.The control samples were separated from the other ones along thePC2 axis.
4. Discussion
The water pollution is a serious problem, which may inextreme concentrations possibly lead to deaths of wild birds asin case of mass mortalities associated with toxic cyanobacterialwater blooms (Krienitz et al., 2003; Ballot et al., 2004, 2005;Ndetei and Muhandiki, 2005; Lugomela et al., 2006), lead-con-tamination of bird habitats (Pain, 1990; Henny et al., 2000; Sileoet al., 2001; Degernes et al., 2006; O’Connell et al., 2008) orinfectious diseases (Sotero-Santos et al., 2006; Waller andUnderhill, 2007; Liu et al., 2008; Sovada et al., 2008).
While our previous studies concerned with the adverse effectsof cyanobacterial biomass containing MC on birds after 10- and30-day exposures (Skocovska et al., 2007; Paskova et al., 2008;Damkova et al., 2009; Peckova et al., 2009) documented nomortality, there was mortality observed in the present study ingroups exposed to lead (Pb), to cyanobacterial biomass and lead(BPb) and lead and Newcastle vaccination (PbV) as discussed indetail in our recently published paper (Pikula et al., 2010).Interestingly, there was no mortality in the group co-exposed toall three stressors (BPbV).
The research presented in this paper was focused on the studyof the modulations of antioxidative and detoxification parametersin experimental birds after single and combined exposures toboth natural and anthropogenic stressors. This design shouldsimulate possible processes under environmental conditionswhere the organism has to face multiple pressures. This novelapproach revealed significant changes of most studied parameters(summarized in Table 3) along with oxidative damage in the formof lipid peroxidation after 30-day single/combined exposures tocyanobacterial biomass, Newcastle disease vaccination and leadin the Japanese quail.
The study showed general stimulation of the antioxidativesystem with the greatest modulations of sublethal parameters inthe individuals from the groups with combined exposures. Theseresults support the hypothesis of higher energy demand to
2.4
2.0
1.6
1.2
0.8
0.4
TBA
RS
nmol
/ m
g pr
otei
n
AC,Pb
V
0
1
2
3
4
5
6
B
C PbVPb BV BPbV
BPbBC PbV BPbBVPb BV BPbVV
Fig. 5. Level of lipid peroxidation (nmol TBARS/mg protein) in liver (A) and heart (B). Box plot parameters as in Fig. 1.
C
CC
C
B
B
BB
B
PP
P BP
BP
BP
V VV
V
V
BV
BV
BVBV
BV
PVPV
PV
PV
BPV
BPV
BPV
BPVBPV
-5 -4 -3 -2 -1 0 1 2 3 4 5-4
-3
-2
-1
0
1
2
3
4
PC axis 1 : 27.6%
PC
axi
s 2
: 15.
7%
L-TBARS
L-TBARS stim
L-GSHL-GST
L-GPx
L-GR
H-TBARS
H-TBARS stim
H-GSHH-GST
H-GPx
H-GR
L-SODL-CAT
-1
1
0.5
0
-0.5
-1
PC
axi
s 2
: 15.
7%
PC axis 1 : 27.6% -0.5 0 0.5 1
Fig. 6. Component score (A.) and component weight (B.) plots from principal
component analysis. Treatment labels: control (C), lead (P), cyanobacterial
biomass (B), vaccination (V). Abbreviations: L: liver, H: heart, GSH: glutathione
content, GST: glutathione-S-transferase, GPx: glutathione peroxidase, GR: glu-
tathione reductase, TBARS stim: stimulated lipid peroxidation, TBARS: lipid
peroxidation, SOD: superoxide dismutase and CAT: catalase.
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counteract adverse effects of multiple exposures. In particular,when there is a risk of possible damage of tissues resulting frominsufficient antioxidative and detoxification protection and thus aneed of antioxidants synthesis.
Oxidative stress as a rather unspecific biochemical process isknown to be involved in toxic action of many stressors includingcyanobacterial biomass (Paskova et al., 2008; Peckova et al., 2009)and lead also in various bird species (Mateo et al., 2001, 2003;Douglas-Stroebel et al., 2004). Moreover, there is a clear associationbetween oxidative stress and immune responses of birds toinfectious agents, where ROS have an important role in killing thepathogens, but can possibly have adverse effects on the host tissues(Costantini and Moller, 2009). Induction of oxidative stress has beendocumented in birds after bacterial, parasitic or viral infections(Georgieva et al., 2006; Liu et al., 2008; Keles et al., 2010).
Significant modulations were shown for the glutathionerelated parameters. GSH levels increased in both cyanobacteriaand lead co-exposed groups (BPb, BPbV) in liver and in almost allexperimental groups in heart tissue. GST activities increased inthe cyanobacteria, lead and vaccination co-exposed group in liver(BPbV) and in two co-exposure groups in heart. Exposure tocyanobacteria with equivalent MC content for 30 days had noeffect on GST, and significantly increased GSH levels in the liver,but not in heart, in our previous study (Paskova et al., 2008).Increased GSH in several tissues has been reported after leadexposure of mallards and Canada geese (Mateo et al., 2003;Douglas-Stroebel et al., 2004). Contrariwise, exposure to leaddid not change the total content of GSH or GST activity amongpied flycatchers (Ficedula hypoleuca) from various metal contami-nated sites (Berglund et al., 2007). Moreover, no effects on GSTactivity or decreased hepatic GST activities were observed in lead-treated mallards and Canada goslings (Mateo and Hoffman, 2001;Mateo et al., 2003). The comparison with these studies and alsowithin our results clearly documents greater stimulation of theGSH and GST in the co-exposed groups.
Increased GR activities in single exposures to cyanobacteriaand vaccine groups were observed in liver and overall increases ofthis enzyme were detected in heart tissue of birds from allexposed groups. Increased enzymatic GPx activities were detectedin both avian liver and heart from most co-exposed (cyanobac-teriaþother stressors) experimental groups. Our previousresearch with cyanobacteria alone fed to quails showed thatacute 10-day exposure lead to increased GR and GPx activity inliver and GR also in heart, while there was no significantstimulation in subchronic 30-day exposure (Paskova et al.,2008). A slightly increased GPx activity was observed in geeseand mallards exposed to lead-contaminated sediments (Mateoand Hoffman, 2001). Inhibited GPx activities were contrariwisedetected in lead-exposed mallards (Mateo et al., 2003).
GPx is able to balance the normal rate of H2O2 production, butduring enhanced H2O2 formation CAT becomes more important(Halliwell and Gutterdige, 1999) and decomposes H2O2 veryefficiently (Berglund et al., 2007). Increased CAT activities wereshown for example in birds from HM-polluted sites when com-pared with reference sites (Berglund et al., 2007). However, nosignificant changes in CAT activity were detected in our study. Onthe other hand, increased activity of hepatic SOD, enzyme capableto metabolize superoxide to hydrogen peroxide, was documentedin the BPbV co-exposure group.
Significantly increased TBARS levels as a parameter of damage tomembrane lipids were shown only in the liver of birds from thecyanobacteria/lead/vaccination co-exposure group, but not in thesingle exposures, when compared with control. Increased lipidperoxidation was observed in heart but not in liver in previousexperiment with quails exposed to cyanobacteria alone for 30 days(Paskova et al., 2008). Significantly increased lipid peroxidation was
shown also in liver and brain of mallards feeding on diet with lead(Mateo et al., 2003) and in mallards and geese exposed to lead-contaminated sediments (Mateo and Hoffman, 2001). On the otherhand, no lipid peroxidation was observed in mallards receiving leadacetate in diet (Douglas-Stroebel et al., 2004).
Berglund et al. (2007) suggested that the antioxidant defenseresponds differently depending on pollution situation and species.When comparing our study with related studies dealing withoxidative stress in birds exposed to environmental stressors (e.g.Mateo and Hoffman, 2001; Mateo et al., 2003; Douglas-Stroebelet al., 2004; Berglund et al., 2007; Paskova et al., 2008), theresponses of antioxidative and detoxification system stronglydepend also on the experimental design and sensitivity of themodel species and/or the population. Moreover, even though thecyanobacterial biomasses used for the exposures in the currentexperiment and in our previous studies (Paskova et al., 2008)were always predominated by the Microcystis aerigunosa speciesand had the same MC content, they could significantly differ inother important effective cyanobacterial metabolites, which couldinfluence some differences in the responses of the studied para-meters. Important is also the ability of experimental animals tobalance the immune response. It has been namely shown that theinduction of immune response causes oxidative stress and affectsthe oxidative stress markers in birds.
Principal component analysis documents the greatest modula-tion of the studied parameters in the co-exposure to all threestressors (BPbV) and also in co-exposure to lead and biomass(BPb), by separating these two groups from the other samplesalong the PC1 driven by liver TBARS and liver glutathioneparameters (Fig. 6). Interestingly, single exposure to biomass isset apart from these two groups along PC1 axis documenting thatthe addition of lead to the exposure mix contributes strongly tothe modification of the studied parameters. The PC2 and PC3 axeswere driven mostly by the glutathione parameters in oppositedirection to TBARS in heart, which documents the significant roleof glutathione and related enzymes in detoxification and antiox-idative protection also in heart tissue. The heart parametersstrongly contributed to the separation of the control group frommost other treatments.
Correlations among the oxidative stress parameters and detox-ification enzymes activities documented also by the results ofprinciple component analysis illustrate the complex character ofthe response and interdependence among parameters. Significantcorrelations among studied parameters were shown especially inliver confirming thus the major role of liver in detoxification ofxenobiotics in birds (Riviere et al., 1985). Positive correlationswere found among most liver parameters including TBARS, whichsuggest similar pattern of stimulation of antioxidative and detox-ification protection, despite which, however, there was still someoxidative damage. The correspondence among the activity of GST,GPx and the increased level of GSH confirms the cooperation ofthe enzymes and previously reported crucial role of glutathione indetoxification after exposure to lead (Mateo and Hoffman, 2001;Berglund et al., 2007) and cyanobacterial biomass (Paskova et al.,2008) and the significance of these biomolecules in the protectionfrom harmful effects. In case of exposure to microcystin-contain-ing biomass, conjugation of microcystin with GSH catalyzed byGST poses an important part of its detoxification pathway(Pflugmacher et al., 1998). Similarly, the role of GST activity inlead binding to GSH and subsequent biliary excretion has beenshown within bird hepatic metabolism (Mateo and Hoffman,2001). It is also known that the enzyme g-glutamylcysteinesynthetase, involved in GSH synthesis, can be induced by heavymetals and oxidative stress (Griffith, 1999).
Analyses of lead concentration in the liver tissues have shownhigh variation among individual birds and no differences among
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lead-exposed groups, which suggest that cyanobacterial biomassor vaccination did not influence the accumulation of lead in testedbirds (Table 2). Concentration of microcystin in the liver tissuehave been more than 20% greater in groups of combined lead-exposures BPb and BPbV than in non-Pb-groups B and BV (Pikulaet al., 2010), which may indicate somewhat greater uptake of thecyanobacterial metabolites in birds weakened by the lead expo-sure. These higher levels of MC along with toxic effects of lead(and effects of immunological challenge) could contribute to thegreatest modulations of almost all examined antioxidative anddetoxification parameters in these groups. The most significantchanges after multiple stressors exposure confirm our hypothesisthat effects of cyanobacterial biomass, lead and immunologicalchallenge may combine to enhance their influence.
The antioxidative protection system was clearly the mostactivated in the three stressors co-exposure (BPbV), which couldcoincide with the fact that there was no mortality in this group.There were several mortalities in other groups exposed to lead,where the stimulation of antioxidative protection was lower. Theseresults suggest that effects of multiple stressors may actuallycombine to stimulate the antioxidative protective responses inthe tissues. In case that these protective mechanisms are insuffi-cient, there can occur some damaging effects. Despite the increasedantioxidative protection there was still some oxidative damage tothe lipidic membranes in the BPbV group, but no mortality.
The responses detected in our study reflect complex situationof the bird detoxification system, which fights in parallel withmore causations of oxidative stress (Costantini and Moller, 2009).
The antioxidative enzymatic defense in birds coming intocontact with both natural and anthropogenic stressors seems tobe activated quite efficiently, which was documented by theelevated levels and activities of antioxidative and detoxificationcompounds and by the low incidence of damage to lipid mem-branes. Most of the significant changes of these ‘‘defense’’ para-meters were detected in case of multiple stressors suggestingactivation of this universal mechanism in case of complexexposure and its crucial role in protection of the bird health inthe environment.
5. Conclusions
Under real environmental situation, complex of various stres-sors can affect the wild organisms, including birds, which makesthe assessment of potential causes and effects rather complicated.This study brings unique information on the effects of combinedavian exposure on important sublethal parameters. Generalactivation of the antioxidant enzymatic system in exposed quailsdocuments the greater need of antioxidative protection in thestudied organs and their ability to produce molecules protectingcells against adverse oxidation processes. The results indicate thatthe antioxidative system plays an important role in the protectiveresponse of the tissues to multiple stressors and that its greaterinduction could actually help to protect the birds from moreserious damage. However, a better understanding of processesand pathways involved in the toxic action of combined stressorsis necessary and would require further studies.
Acknowledgments
This research was supported by projects MSM 6215712402and 1M0571, INCHEMBIOL framework project MSM0021622412,and project CETOCOEN (CZ.1.05/2.1.00/01.0001) granted by theEuropean Union and administered by the Ministry of Education,Youth and Sports of the Czech Republic.
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P. Veronika et al. / Ecotoxicology and Environmental Safety 74 (2011) 2082–20902090
Paper V.
Pašková, V., Hilscherová, K. and Bláha, L. (2011).
Teratogenicity and embryotoxicity in aquatic organisms after pesticide exposure and the role of oxidative stress.
Reviews of Environmental Contamination and Toxicology, Vol. 211, pp. 25-61.
Teratogenicity and Embryotoxicity in AquaticOrganisms After Pesticide Exposureand the Role of Oxidative Stress
Veronika Pašková, Klára Hilscherová, and Ludek Bláha
Contents
1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25
2 Pesticides and Teratogenicity in Fish . . . . . . . . . . . . . . . . . . . . . 26
3 Pesticides and Teratogenicity in Amphibians . . . . . . . . . . . . . . . . . . 31
4 Pesticides as Possible Teratogens in Invertebrates . . . . . . . . . . . . . . . . 36
5 Role of Oxygen and Antioxidant Defenses in Embryogenesis . . . . . . . . . . 39
6 Oxidative Stress in Embryotoxicity and Teratogenicity . . . . . . . . . . . . . 40
7 Pesticides and Oxidative Damage During Early Development
in Aquatic Organisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . 40
8 Further Evidence – Pesticides and Antioxidative Defense in Adult Aquatic Biota . 45
9 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 51
References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 51
1 Introduction
Complex factors have contributed to the decline of aquatic populations worldwide.Among these factors are intensification of agriculture, including the application offertilizers and agents of crop protection, and loss of habitat. Various developmentalabnormalities in natural populations of aquatic vertebrates have been documented,and agricultural pesticides are considered by many to be one of the importantfactors that cause such abnormalities. Amphibians may potentially be a target ofenvironmental stressors and toxicants as a result of their biphasic life cycles andskin permeability. In this chapter, the role of oxidative stress in the teratogenicaction of pesticides is reviewed and addressed, with special attention given to non-target aquatic organisms such as amphibians, fish, and invertebrates. The review of
L. Bláha (B)Faculty of Science, Research Centre for Toxic Compounds in the Environment (RECETOX),Masaryk University; Kamenice 126/3, 625 00 Brno, Czech Republice-mail: blaha@recetox.muni.cz
25D.M. Whitacre (ed.), Reviews of Environmental Contamination and Toxicology,Reviews of Environmental Contamination and Toxicology 211,DOI 10.1007/978-1-4419-8011-3_2, C© Springer Science+Business Media, LLC 2011
26 V. Pašková et al.
available literature indicates that many pesticides enhance oxidative stress in aquaticorganisms, and such stress may be linked to developmental alterations, includingreproductive effects, embryotoxicity, and/or teratogenicity.
Any external factor affecting cellular proliferation, differentiation, or apoptosiscan produce embryotoxic or teratogenic effects, and such factors include chemicalexposures at high concentrations; such effects may result in permanent congeni-tal malformations, functional abnormalities, or even embryo death (Gilbert 2006).Several external factors may result in embryotoxicity and teratogenicity in theaquatic environment. These factors include ultraviolet radiation, extremes in pH,thermal and ionic conditions, infections, parasites, as well as chemicals such as phar-maceuticals, retinoid and aromatic compounds, and pesticides (Ankley et al. 2004;Bilski et al. 2003; Blaustein and Johnson 2003; Finn 2007; Hayes et al. 2006).
One mechanism by which chemicals induce toxicity is through oxidative stress,and it has been shown that several widely used pesticides are capable of pro-ducing pro-oxidants in cells (Tellez-Banuelos et al. 2009; Vismara et al. 2001a).Furthermore, oxidative stress is the major mechanism by which some pesticidesexert their effects; a prime example is the bipyridyl herbicides (Ruiz-Leal andGeorge 2004; Sewalk et al. 2000).
Pesticides are regularly applied onto agricultural land worldwide, and the resul-tant timing of exposure often parallels the appearance of the early developmentalstages of aquatic organisms (Greulich and Pflugmacher 2003). Although the variousside effects that pesticides have on biota have been documented in many studies,to our knowledge, no consistent overview of pesticide embryotoxicity in aquaticinvertebrates and vertebrates is available. Hence, in this review, we summarize theexisting knowledge on this topic. Moreover, we present an overview of availableinformation on the general teratogenic and embryotoxic effects of pesticides inaquatic biota such as fish, amphibia, and invertebrates. In this review we empha-size toxic effects that are related to oxidative stress and draw lines of evidence tosupport the view that it is a possible toxicity mechanism behind the induction ofteratogenicity.
2 Pesticides and Teratogenicity in Fish
In fish, developmental malformations have been linked to the presence of severalenvironmental pollutants such as persistent organochlorines, pesticides, or heavymetals (Westernhagen von 1988). In several studies, direct embryotoxicity hasresulted from the presence of complex matrices such as oil (Heintz et al. 1999),and recently, tests for embryonic malformations in fish have been used as generalwater quality indicators (Klumpp et al. 2002).
The array of effects that pesticides have had on embryonic development in fish issummarized in Table 1.
Zebra fish (Danio rerio; family Cyprinidae) constitutes the most common modelof test fish species, and it has been used in many pesticide studies; the results fromStrmac and Braunbeck (1999) and Osterauer and Köhler (2008) are examples. Otherspecies used in pesticide testing schemes include the Japanese medaka (Oryziaslatipes; family Adrianichthyidae; Villalobos et al. 2000) and various salmonids
Teratogenicity and Embryotoxicity in Aquatic Organisms 27
Tabl
e1
The
effe
cts
ofse
lect
edpe
stic
ides
onde
velo
pmen
tand
repr
oduc
tion
infis
h
Pest
icid
eO
rgan
ism
Toxi
cef
fect
Ref
eren
ces
Car
bam
ates
Thi
oben
carb
O.l
atip
esB
rady
card
ia,p
eric
ardi
aled
ema,
hem
osta
sis,
poor
yolk
reso
rptio
n,ce
phal
ican
dsp
inal
defo
rmiti
es,a
bnor
mal
hatc
hing
Vill
alob
oset
al.(
2000
)
Dith
ioca
rbam
ate
D.r
erio
Twis
ted
noto
chor
d,re
duce
dha
tchi
ngno
toch
ord
dist
ortio
nsD
isto
rted
noto
chor
dde
velo
pmen
tand
shor
tene
dan
teri
orto
post
erio
rax
is
Hae
ndel
etal
.(20
04),
Tilt
onet
al.(
2006
,200
8)va
nB
oxte
leta
l.(2
010)
Thi
uram
D.r
erio
Wav
yno
toch
ords
,dis
orga
nize
dso
mite
s,sh
orte
ned
yolk
sac
exte
nsio
nTe
raok
aet
al.(
2006
)
Car
bary
lD
.rer
ioR
edbl
ood
cell
accu
mul
atio
n,de
laye
dha
tchi
ngan
dpe
rica
rdia
lede
ma,
brad
ycar
dia
Lin
etal
.(20
07)
Synt
heti
cpy
reth
roid
sE
sfen
vale
rate
O.l
atip
es
Onc
orhy
nchu
sts
haw
ytsc
ha
Del
eter
ious
repr
oduc
tive
effe
cts,
redu
ced
hatc
hing
succ
ess,
and
larv
alvi
abili
tyM
yosk
elet
alab
norm
ality
,lor
dosi
s
Wer
ner
etal
.(20
02)
Via
ntet
al.(
2006
b)
Del
tam
ethr
inC
ypri
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io
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ykis
sB
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ydan
iore
rio
D.r
erio
Del
eter
ious
repr
oduc
tive
effe
cts,
decr
ease
dha
tchi
ngsu
cces
s,la
rvae
leth
ality
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red
deve
lopm
ent
Fry
leth
ality
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sof
equi
libri
umen
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edla
rvae
mor
talit
y,re
duce
dha
tchi
ngra
teE
mbr
yole
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ity,n
euro
beha
vior
alef
fect
s–
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ticm
ovem
ents
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icar
dial
edem
a,cr
anio
faci
alab
norm
aliti
es
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rücü
and
Ayd
in(2
004)
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land
Sagl
am(2
005)
,
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gean
dN
agel
(199
0),a
ndD
eMic
coet
al.(
2010
)C
yper
met
hrin
C.c
arpi
oD
.rer
ioE
mbr
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thal
ity,r
educ
edha
tchi
ngsu
cces
sE
mbr
yole
thal
ity,n
euro
beha
vior
alef
fect
s–
spas
ticm
ovem
ents
,per
icar
dial
edem
a
Ayd
inet
al.(
2005
)D
eMic
coet
al.(
2010
)
Bif
enth
rin
D.r
erio
Em
bryo
leth
ality
,cur
vatu
reof
the
body
axis
,ne
urob
ehav
iora
leff
ects
–sp
astic
mov
emen
tsD
eMic
coet
al.(
2010
)
�-C
yhal
othr
inD
.rer
ioE
mbr
yole
thal
ity,n
euro
beha
vior
alef
fect
s–
spas
ticm
ovem
ents
,per
icar
dial
edem
aD
eMic
coet
al.(
2010
)
28 V. Pašková et al.
Tabl
e1
(con
tinue
d)
Pest
icid
eO
rgan
ism
Toxi
cef
fect
Ref
eren
ces
Res
met
hrin
D.r
erio
Neu
robe
havi
oral
effe
cts
–sp
astic
mov
emen
tsD
eMic
coet
al.(
2010
)Pe
rmet
hrin
O.l
atip
es
D.r
erio
Del
ayed
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blad
der
infla
tion,
inab
ility
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tchl
ing
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spon
dto
stim
ulus
;unc
oord
inat
edm
ovem
ents
,m
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elet
alde
fect
s,an
dtr
ansi
ente
nlar
gem
ento
fga
llbl
adde
rof
larv
aeE
mbr
yole
thal
ity,c
urva
ture
ofth
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dyax
is,
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obeh
avio
rale
ffec
ts–
spas
ticm
ovem
ents
,cr
anio
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aliti
es
Gon
zále
z-D
once
leta
l.(2
003)
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icco
etal
.(20
10)
Org
anop
hosp
hate
sM
alat
hion
D.r
erio
Cla
rias
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us
Scia
enop
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Mel
anot
aeni
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viat
ilis
Red
uced
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urvi
val,
and
eye
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eter
sD
efor
med
noto
chor
dan
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rica
rdia
lede
ma,
larv
aew
ithbe
ntbo
dy,a
ndsw
olle
nyo
lksa
cD
ecre
ases
ingr
owth
inw
eigh
tof
larv
ae,i
ncre
ased
prot
ein
synt
hesi
sM
oder
ate
larv
alle
thal
ity
Coo
ket
al.(
2005
)L
ien
etal
.(19
97)
McC
arth
yan
dFu
iman
(200
8)
Rei
det
al.(
1995
)
Dia
zino
nC
.car
pio
O.l
atip
es
D.r
erio
Del
eter
ious
repr
oduc
tive
effe
cts,
redu
ced
hatc
hing
succ
ess,
and
larv
alvi
abili
tyE
dem
asbo
thal
ong
vite
lline
vein
san
dw
ithin
the
peri
card
ial
sac,
dela
yin
hatc
hing
,dec
reas
ein
swim
blad
der
infla
tion,
decr
ease
dle
ngth
Em
bryo
leth
ality
,dec
reas
edhe
artr
ate,
yolk
sac
and
hear
tsa
ced
ema,
spin
ede
form
atio
ns,a
ltere
dha
tchi
ngda
te
Ayd
inan
dK
öprü
cü(2
005)
Ham
man
dH
into
n(2
000)
Ost
erau
eran
dK
öhle
r(2
008)
Chl
orpy
rifo
sD
.rer
ioIn
crea
sed
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mot
orac
tivity
,pau
sed
jerk
ym
ovem
ents
,he
arte
dem
a,sp
inal
defo
rmity
Kie
nle
etal
.(20
09)
Teratogenicity and Embryotoxicity in Aquatic Organisms 29
Tabl
e1
(con
tinue
d)
Pest
icid
eO
rgan
ism
Toxi
cef
fect
Ref
eren
ces
Org
anoc
hlor
ines
End
osul
fan
O.l
atip
esD
.rer
ioD
elay
edha
tchi
ng,s
mal
ler
fry,
alte
red
mob
ility
Mild
trun
kcu
rvat
ure,
abno
rmal
beha
vior
,ede
ma,
mic
roce
phal
y,an
dim
pair
edm
ovem
ent,
alon
gw
ithin
crea
sed
deat
hra
te
Gor
mle
yan
dTe
athe
r(2
003)
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eyan
dK
rone
(200
1)
Lin
dane
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rata
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ykis
s
B.r
erio
Myo
skel
etal
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cts,
skin
opac
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xoph
thal
mia
,wea
ksw
imm
ing,
depi
gmen
tatio
n,be
havi
oral
chan
ges
Hep
atoc
ytic
alte
ratio
ns(g
lyco
geni
cde
plet
ion,
RE
Ran
ddi
ctyo
som
ech
ange
s,se
cond
ary
lyso
som
eac
cum
ulat
ion)
Enh
ance
dla
rvae
mor
talit
y,de
crea
sed
grow
th
Oliv
aet
al.(
2008
)
Sylv
ieet
al.(
1996
)
Gör
gean
dN
agel
(199
0)T
hiac
lopr
idD
.rer
ioH
eart
rate
affe
ctio
nO
ster
auer
and
Köh
ler
(200
8)
Org
anos
ulfu
rsM
ethy
liso
thio
cyan
ate
D.r
erio
Not
ocho
rddi
stor
tions
Tilt
onet
al.(
2006
)Te
rbut
ryn
(atr
iazi
ne)
+tr
iasu
lfur
onS.
aura
taC
urva
ture
sof
the
vert
ebra
lcol
umn,
the
hepa
tocy
tes
form
ing
slac
kly
arra
nged
cord
s,lo
ssof
cellu
lar
shap
eof
hepa
tocy
tes,
lipid
incl
usio
ns,n
ucle
arpy
knos
is
Aru
feet
al.(
2004
b)
Phe
nylp
yraz
ole
Fipr
onil
D.r
erio
Not
ocho
rdde
gene
ratio
n,sh
orte
ning
alon
gth
ero
stra
l–ca
udal
body
axis
,ine
ffec
tive
tail
flips
Steh
ret
al.(
2006
)
Tria
zine
sA
traz
ine
S.oc
ella
tus
B.r
erio
Alte
red
grow
th,h
yper
activ
ity,a
ndfa
ster
activ
esw
imm
ing
spee
d(e
leva
ted
rate
ofen
ergy
utili
zatio
n)D
eclin
esin
grow
thin
wet
wei
ghta
ndpr
otei
nco
nten
tin
larv
ae,i
ncre
ase
inra
tes
ofpr
otei
nde
grad
atio
nE
nhan
ced
larv
aem
orta
lity,
incr
ease
dnu
mbe
rof
defo
rmat
ions
and
edem
a
delC
arm
enA
lvar
ezan
dFu
iman
(200
5),M
cCar
thy
and
Fuim
an(2
008)
Gör
gean
dN
agel
(199
0)
Sim
azin
eS.
aura
taR
educ
edla
rvae
surv
ival
,hep
atic
lesi
ons,
loss
ofce
llula
rsh
ape
inhe
pato
cyte
s,lip
idin
clus
ions
,foc
alne
cros
is,
abun
dant
nucl
ear
pykn
osis
Aru
feet
al.(
2004
a)
30 V. Pašková et al.
Tabl
e1
(con
tinue
d)
Pest
icid
eO
rgan
ism
Toxi
cef
fect
Ref
eren
ces
Cya
nazi
neM
.fluv
iati
lis
Mod
erat
em
orta
lity
with
decr
easi
ngtr
end
from
the
day
ofha
tchi
ngR
eid
etal
.(19
95)
Org
anot
ins
Tri
phen
yltin
acet
ate
D.r
erio
Lar
valm
orta
lity,
dela
yed
hatc
hing
,ske
leta
lmal
form
atio
n,re
tard
edyo
lksa
cre
sorp
tion,
and
edem
ain
the
hear
tand
yolk
sac
regi
ons,
hist
o-an
dcy
to-p
atho
logi
cala
ltera
tions
ofla
rval
liver
incl
udin
gch
ange
sin
nucl
eian
dm
itoch
ondr
iaas
wel
las
glyc
ogen
depl
etio
n
Strm
acan
dB
raun
beck
(199
9)
Din
itro
phen
olD
inos
ebO
.lat
ipes
Lar
valm
orta
lity,
redu
ced
eye
grow
th,d
imin
ishe
dhe
artr
ate,
faile
dha
tchi
ng,d
evel
opm
enta
lret
arda
tion,
peri
card
ial
edem
a,an
dre
duce
dgr
owth
,red
uctio
nsof
eye
area
and
wid
th
Via
ntet
al.(
2006
a,b)
RE
R,r
ough
endo
plas
mic
retic
ulum
Teratogenicity and Embryotoxicity in Aquatic Organisms 31
such as Oncorhynchus sp. (family Salmonidae; Sylvie et al. 1996). The testingthat has utilized these species has involved a wide array of pesticides and pesti-cide classes. For example, the toxicity of the following classes has been tested inthese fish species: organophosphates, triazines, synthetic pyrethroids, carbamates,organochlorines, some studies with organosulfur pesticides (Arufe et al. 2004b;Tilton et al. 2006), phenylpyrazoles (Stehr et al. 2006), organotins (Strmac andBraunbeck 1999), and dinitrophenols (Viant et al. 2006a, b).
The array of pesticide effects that have been observed on fish embryonic develop-ment has included malformations in myoskeletal development (such as notochordabnormalities of degeneration), defects along the rostral–caudal body axis, curva-ture of the vertebral column, and reduced growth (McCarthy and Fuiman 2008;DeMicco et al. 2010; van Boxtel et al. 2010). In other studies, pesticides affectedvarious visceral organs in ways that led to defects in the hepatic, cephalic, and eyeregion, and various edemas in pericardial area or yolk sac (Strmac and Braunbeck1999; Hamm and Hinton 2000; Willey and Krone 2001). Besides these morpho-logical alterations, embryonic and larval exposures to pesticides have also resultedin decreased hatching success and larval mortality (Görge and Nagel 1990; Aydinand Köprücü 2005; Viant et al. 2006a, b) or behavioral alterations such as uncoor-dinated movements and loss of balance (Ural and Saglam 2005; González-Doncelet al. 2003; Kienle et al. 2009).
As is clearly shown in Table 1, various pesticides have produced significant detri-mental effects on developmental processes in different fish species. We address theseeffects and the role of oxidative stress in developmental toxicity in more detail below(Sections 5 and 6).
3 Pesticides and Teratogenicity in Amphibians
Amphibians are known to be highly sensitive organisms and can be affected bychemical, physical, and habitat factors; moreover, it is believed that pesticidesare one of the causes of the worldwide decline in amphibian populations (Muthset al. 2006). Most amphibian species breed during the spring when pesticidesare being applied onto the land for weed, fungal, insect, or other pest control,which makes amphibians highly vulnerable to pesticide toxicity (Greulich andPflugmacher 2003).
In natural frog populations, morphologically malformed individuals usually con-stitute a small fraction of less than 2% (Ouellet 2000). However, a much higherincidence (up to 60%) of malformed specimens was documented to occur incontaminated ponds (Meteyer 2000). In agricultural ecosystems, developmentalmalformations resulting from pesticide exposure were documented to have occurredin several amphibian species in India (Gurushankara et al. 2007) or in Canada(Ouellet et al. 1997).
Embryotoxicity and teratogenicity of various pesticide classes have been docu-mented to have occurred in amphibians in laboratory studies and in field observationstudies (Table 2). Most of the studies used the prototypical model organism Xenopuslaevis from the Pipidae family, but species from other families that includedRanidae, Bufonidae, Microhylidae, and others were also used.
32 V. Pašková et al.
Tabl
e2
The
effe
cts
ofse
lect
edpe
stic
ides
onde
velo
pmen
tand
repr
oduc
tion
inam
phib
ians
Pest
icid
eO
rgan
ism
Toxi
cef
fect
Ref
eren
ces
Mix
ture
ofpe
stic
ides
Ran
api
pien
sA
ltere
dde
velo
pmen
tand
grow
th,s
exdi
ffer
entia
tion,
beha
vior
,tim
ing
ofin
itiat
ion,
and
com
plet
ion
ofm
etam
orph
osis
Hay
eset
al.(
2006
)
Car
bam
ates
Car
bofu
ran
Mic
rohy
laor
nata
Blis
teri
ng,d
iste
ntio
nof
body
cavi
ties,
curv
atur
eof
the
body
axis
,poo
rbl
ood
circ
ulat
ion,
reta
rded
grow
th,
abno
rmal
beha
vior
,poo
rpi
gmen
tatio
n
Paw
aran
dK
atda
re(1
984)
Car
bam
ate
ZZ
-Aph
oxR
ana
pere
ziH
isto
logi
cald
amag
eof
gill,
liver
,gal
lbla
dder
,hea
rt,
and
noto
chor
dH
onru
bia
etal
.(19
93)
Car
bary
lA
mby
stom
aba
rbou
ri
X.l
aevi
s
Ran
asp
heno
ceph
ala
Del
ayed
hatc
hing
,red
uced
larv
alsu
rviv
al,l
ower
grow
thra
tes,
resp
irat
ory
dist
ress
,lim
bde
form
ities
Abn
orm
alta
ilfle
xure
,ske
leta
lmus
cle
lesi
ons,
wav
yor
bent
noto
chor
dIn
crea
sed
leng
thof
tadp
oles
,lar
ger
mas
sof
met
amor
phos
is
Roh
ret
al.(
2003
)
Bac
chet
taet
al.(
2008
)
Bri
dges
and
Boo
ne(2
003)
Org
anop
hosp
hate
sFe
nitr
othi
onM
.orn
ata
R.p
ipie
ns,
Ran
acl
amit
ans,
Ran
aca
tesb
eian
a
Blis
teri
ng,d
iste
ntio
nof
body
cavi
ties,
curv
atur
eof
the
body
axis
,poo
rbl
ood
circ
ulat
ion,
reta
rded
grow
th,
abno
rmal
beha
vior
,poo
rpi
gmen
tatio
nE
mbr
yoto
xici
ty,b
ehav
ior
alte
ratio
ns,p
aral
ysis
Paw
aran
dK
atda
re(1
984)
Ber
rill
etal
.(19
94)
Gut
hion
and
guth
ion
2SX
.lae
vis
Em
bryo
leth
ality
,dec
reas
edbo
dyle
ngth
,dev
elop
men
tal
alte
ratio
nsSc
huyt
ema
etal
.(19
94)
Mal
athi
onX
.lae
vis
Am
byst
oma
mex
ican
um
Def
ects
ofne
urom
uscu
lar
activ
itysu
chas
spas
ms,
trem
ors,
and
affe
cted
swim
min
g,ab
norm
alta
ilfle
xure
,dis
tort
edm
yocy
tes
Em
bryo
nic
mor
talit
y,de
laye
dor
stop
ped
deve
lopm
ent,
notc
ompl
eted
neur
ulat
ion,
thin
noto
chor
d,no
tfus
edne
ural
fold
s,em
bryo
sw
ithou
tnot
ocho
rdan
dne
ural
cana
land
loca
ted
edem
a,er
ratic
swim
min
g
Bon
fant
ieta
l.(2
004)
Rob
les-
Men
doza
etal
.(20
09)
Teratogenicity and Embryotoxicity in Aquatic Organisms 33
Tabl
e2
(con
tinue
d)
Pest
icid
eO
rgan
ism
Toxi
cef
fect
Ref
eren
ces
Chl
orpy
rifo
sX
.lae
vis
Hyl
ach
ryso
scel
isR
ana
sphe
noce
phal
a,A
cris
crep
itan
s
Gas
trop
hryn
eol
ivac
eaB
ufo
bufo
garg
ariz
ans
A.m
exic
anum
Ran
abo
ylii
Red
uced
myo
tom
esi
zean
dhy
pert
roph
ies;
defe
cts
ofne
urom
uscu
lar
activ
itysu
chas
spas
ms,
trem
ors,
and
affe
cted
swim
min
g,no
toch
ord
flexu
re,
dist
orte
dm
yocy
tes
tadp
ole
mor
talit
y,sw
imsp
eed
affe
ctio
n,lo
wer
mas
sof
tadp
oles
Shru
nken
fins,
tail
defo
rmiti
es,a
ndhe
aded
ema,
beha
vior
alch
ange
s,m
icro
nucl
eus
indu
ctio
n,ta
dpol
ele
thal
ityth
inno
toch
ord
and
neur
alca
nal,
late
ralt
ail
flexu
re,c
onvu
lsio
ns,s
pasm
san
dtr
emor
s,la
rval
mor
talit
y
Col
ombo
etal
.(20
05)
Bon
fant
ieta
l.(2
004)
Wid
der
and
Bid
wel
l(20
08)
Yin
etal
.(20
09)
Rob
les-
Men
doza
etal
.(20
09)
Spar
ling
and
Felle
rs(2
007)
Dia
zino
nB
ufo
mel
anos
tict
usPo
lype
date
scr
ucig
erR
.boy
lii
Lar
valm
orta
lity,
alte
red
activ
ity,g
row
thre
tard
atio
n
Lar
valm
orta
lity
Sum
anad
asa
etal
.(20
08)
Spar
ling
and
Felle
rs(2
007)
Org
anoc
hlor
ines
Die
ldri
nL
imno
dyna
stes
tasm
anie
nsis
Abn
orm
alot
olith
,otic
caps
ule,
and
ceph
alic
pigm
enta
tion
Bro
oks
(198
1)
Ben
zene
hexa
chlo
ride
M.o
rnat
aB
liste
ring
,dis
tent
ion
ofbo
dyca
vitie
s,cu
rvat
ure
ofth
ebo
dyax
is,p
oor
bloo
dci
rcul
atio
n,re
tard
edgr
owth
,ab
norm
albe
havi
or,p
oor
pigm
enta
tion
Paw
aran
dK
atda
re(1
984)
Met
hoxy
chlo
rX
.lae
vis
Xen
opus
trop
ican
a
Am
byst
oma
mac
roda
ctyl
um
Alte
red
hind
limb
diff
eren
tiatio
nan
dta
ilre
sorp
tion,
inte
rfer
ence
with
norm
alre
prod
uctiv
epr
oces
ses
Alte
red
rate
ofla
rval
deve
lopm
ent,
dela
yed
deve
lopm
ent,
chro
nic
repr
oduc
tive
effe
cts
Alte
red
star
tlere
spon
ses
and
ash
orte
rdi
stan
ceof
trav
elfo
llow
ing
appl
icat
ion
ofth
est
artle
stim
ulus
,in
crea
sed
pred
atio
n,m
orta
lity
ofla
rvae
Fort
etal
.(20
04a,
b)
Ero
sche
nko
etal
.(20
02)
End
osul
fan
Lit
oria
citr
opa
Red
uced
tadp
ole
surv
ival
,inc
reas
edvu
lner
abili
tyto
pred
atio
nB
room
hall
(200
2)
34 V. Pašková et al.
Tabl
e2
(con
tinue
d)
Pest
icid
eO
rgan
ism
Toxi
cef
fect
Ref
eren
ces
End
osul
fan
and
octy
lphe
nol
A.b
arbo
uri
Del
ayed
hatc
hing
,red
uced
larv
alsu
rviv
al,l
ower
grow
thra
tes,
resp
irat
ory
dist
ress
,lim
bde
form
ities
Roh
ret
al.(
2003
)
Hep
tach
lor
Ran
akl
.Esc
ulen
taD
ecre
ased
surv
ival
rate
inta
dpol
es,a
ltera
tions
inth
eep
ider
mis
ofta
dpol
esco
ntai
ning
dila
ted
and
irre
gula
rve
sicl
es,d
amag
edm
itoch
ondr
iash
owin
gal
tere
dcr
ista
e
Feno
glio
etal
.(20
09)
Pyr
idin
eT
ricl
opyr
R.p
ipie
ns,
R.c
lam
itan
s,R
.cat
esbe
iana
R.p
ipie
ns
Em
bryo
toxi
city
,beh
avio
ral
tera
tions
,par
alys
is
Red
uced
surv
ival
ofta
dpol
es
Ber
rill
etal
.(19
94)
Che
net
al.(
2008
)B
ipyr
idyl
Para
quat
X.l
aevi
sE
mbr
yole
thal
ity,g
row
thre
tard
atio
n,ve
ntra
ltai
lflex
ure,
abno
rmal
som
ites;
mito
seal
tera
tions
,ne
crot
icm
yocy
tes,
mal
form
edin
ters
omiti
cbo
unda
ries
,alte
red
swim
min
gac
tivity
,em
bryo
leth
ality
orem
bryo
sun
able
tosw
im,g
ener
alre
duct
ion
ofle
ngth
,med
ialfl
exur
esof
the
noto
chor
dan
dst
untin
g
Vis
mar
aet
al.(
2000
,200
1a,b
),M
ante
cca
etal
.(20
06),
Osa
noet
al.(
2002
)
Phe
noxy
carb
oxyl
icac
id4-
Chl
oro-
2-m
ethy
lphe
noxy
acet
icac
id
X.l
aevi
sG
row
thre
tard
atio
nB
erna
rdin
ieta
l.(1
996)
Chl
oroa
ceta
nili
deA
lach
lor
Buf
oam
eric
anus
R.p
ipie
nsE
mbr
yole
thal
ity,d
evel
opm
enta
lalte
ratio
nsH
owe
etal
.(19
98)
Teratogenicity and Embryotoxicity in Aquatic Organisms 35
Tabl
e2
(con
tinue
d)
Pest
icid
eO
rgan
ism
Toxi
cef
fect
Ref
eren
ces
Tria
zine
Atr
azin
eB
.am
eric
anus
R.p
ipie
nsX
.lae
vis
Rhi
nell
aar
enar
um
Em
bryo
leth
ality
,ede
mas
Indu
ced
inte
rsex
anim
als
No
effe
cton
mor
talit
y,gr
owth
,tim
eto
met
amor
phos
is,
gona
dan
dla
ryng
eald
evel
opm
ent,
orar
omat
ase
activ
ity,s
exua
ldif
fere
ntia
tion
affe
ctio
n–
incr
ease
sin
fem
ale
ratio
s,no
effe
cts
onre
prod
uctio
n,ha
tchi
ngsu
cces
s,or
deve
lopm
ent
Em
bryo
leth
ality
,gre
ater
mor
talit
yof
larv
ae,a
ltera
tions
inth
etim
ing
ofm
etam
orph
osis
How
eet
al.(
1998
)C
arr
etal
.(20
03)
Coa
dyet
al.(
2005
),O
kaet
al.(
2008
),D
uPr
eez
etal
.(20
08)
Bro
deur
etal
.(20
09)
Terp
enoi
dsM
etho
pren
eR
.pip
iens
Seve
rede
velo
pmen
tale
ffec
ts,d
ysm
orph
ogen
esis
,hig
hm
orta
lity
Ank
ley
etal
.(19
98)
Met
hopr
ene
degr
adat
ion
prod
ucts
X.l
aevi
sE
ye,c
rani
al,f
acia
ldef
ects
,spi
nalc
urva
ture
and
hear
tan
dgu
tmal
form
atio
ns,
dysm
orph
ogen
esis
ofcr
anio
faci
alre
gion
,ede
mas
,m
icro
phth
alm
ia,
redu
ctio
nsin
the
pros
ence
phal
onan
dm
esen
ceph
alon
,dev
elop
men
tald
elay
La
Cla
iret
al.(
1998
),an
dD
egitz
etal
.(2
003)
Am
idin
eA
mitr
azX
.lae
vis
Gro
wth
reta
rdat
ion,
edem
asof
the
face
,hea
rt,a
nd/o
rab
dom
enan
dax
ialfl
exur
es(c
urva
ture
ofth
eno
toch
ord
orbe
ndin
gof
the
tail)
Osa
noet
al.(
2002
)
2,4-
Dim
ethy
lani
line
X.l
aevi
sSu
btox
icst
imul
atio
nof
grow
th,l
oss
ofpi
gmen
ttog
ethe
rw
ithen
ceph
alom
egal
y,w
ellin
gof
the
brai
nO
sano
etal
.(20
02)
36 V. Pašková et al.
Similar to what occurs in fish, amphibians are known to be highly sensitive toseveral developmental effects; myoskeletal system, abnormal tail formation, andlimb differentiation are among the most often reported effects caused by pesticideexposure (Fort et al. 2004a, b; Bacchetta et al. 2008). Further alterations includeincomplete neurulation, edemas, epidermal defects, or gut malformations (Degitzet al. 2003; Robles-Mendoza et al. 2009), as well as severe dysmorphogenesis,embryonic and larval lethalities, delayed hatching, growth retardations, or alteredmetamorphosis (Vismara et al. 2000, 2001a, b; Brodeur et al. 2009).
In addition to investigations that have been performed with frogs and toads, a fewstudies were also performed with salamanders (family Ambystomatidae) exposed topesticides; effects observed included larval mortality, limb deformities, and behav-ioral changes (Eroschenko et al. 2002; Robles-Mendoza et al. 2009; Rohr et al.2003).
Similar to observations that have been made in fish studies (Table 1), test-ing results with amphibians (Table 2) indicate that significant embryotoxicity/teratogenicity and developmental toxicity result from pesticide exposure, andevidence suggests that oxidative stress may play a role in producing such effects.
4 Pesticides as Possible Teratogens in Invertebrates
Pesticides may not only alter development and reproduction in vertebrates butalso affect various aquatic invertebrates. Data from studies that have documenteddevelopmental effects in invertebrates are presented in Table 3. Most invertebratereproduction or developmental studies were performed with organophosphate andorganochlorine insecticides (Key et al. 2007; Lee and Oshima 1998), although theeffects of other pesticide classes (e.g., synthetic pyrethroids, chloroacetanilides andterpenoids; triazines, carbamates, azoles, and phenylpyrazoles) were also studied.
Among the effects found in gastropods, bivalve mollusks, echinoids, and deca-pod crustaceans were embryonic and larval lethality (Key et al. 2007; Harper et al.2008), decreased hatching success or delayed hatching times (Lee and Oshima 1998;Sawasdee and Köhler 2009), as well as alterations in embryolarval development, andlarval deformities (Bhide et al. 2006; Buznikov et al. 2007).
Model aquatic invertebrate organisms such as the cladoceran Daphnia were alsoinvestigated (Palma et al. 2009), but in this species, embryolethality was oftenmasked by changes in other parameters, such as adult immobilization or number ofoffspring (Abe et al. 2001). Embryos and larvae of ascidian Phallusia mammillatawere found to be a sensitive model, and the azole pesticides imazalil and triadime-fon inhibited sperm viability and fertilization rate, and deregulated organogenesisof the nervous system in this species (Pennati et al. 2006).
Although oxidative stress possibly plays a role in the toxicity of pesticides toinvertebrates as described in Table 3, to our knowledge no specific studies exist thatlink embryotoxicity directly to oxidative stress.
Teratogenicity and Embryotoxicity in Aquatic Organisms 37
Tabl
e3
The
effe
cts
ofse
lect
edpe
stic
ides
onde
velo
pmen
tand
repr
oduc
tion
inaq
uatic
inve
rteb
rate
s
Pest
icid
eO
rgan
ism
Toxi
cef
fect
Ref
eren
ce
Car
bam
ate
Prop
oxur
Lym
naea
stag
nali
sIn
crea
sed
mor
talit
y,de
crea
sed
hatc
habi
lity,
and
larv
alde
form
ities
linke
dw
ithde
viat
ion
inpr
otei
nfr
actio
nsB
hide
etal
.(20
06)
Org
anop
hosp
hate
sC
hlor
pyri
fos
Stro
ngyl
ocen
trot
usdr
oeba
chie
nsis
M.g
allo
prov
inci
alis
Pala
emon
etes
pugi
oC
alli
nect
essa
pidu
sD
aphn
iam
agna
Neu
roto
xici
ty,s
peci
ficm
alfo
rmat
ion
–“m
ushr
oom
”-sh
aped
larv
afo
rmat
ion
Red
uctio
nof
embr
yoge
nesi
ssu
cces
sL
etha
lity
ofne
wly
hatc
hed
larv
aean
dpo
stla
rvae
Inhi
bite
dha
tchi
ngU
nder
deve
lope
dse
cond
ante
nnae
,cur
ved
and
incu
rved
shel
lspi
ne,a
rres
ted
eggs
,red
uced
num
ber
ofof
fspr
ing
per
fem
ale,
embr
yos
rem
aini
ngat
cert
ain
stag
eE
arly
larv
aele
thal
ity,h
atch
ing
time
Buz
niko
vet
al.(
2007
)
Bei
ras
and
Bel
las
(200
8)K
eyet
al.(
2007
);L
eean
dO
shim
a(1
998)
Palm
aet
al.(
2009
)
Dic
hlor
vos
L.s
tagn
alis
Incr
ease
dm
orta
lity,
decr
ease
dha
tcha
bilit
y,an
dla
rval
defo
rmiti
esB
hide
etal
.(20
06)
Dia
zino
nPa
race
ntro
tus
livi
dus
Dec
reas
edle
ngth
ofpr
imar
ym
esen
chym
albr
anch
esan
dpl
utei
,alte
red
spee
dof
deve
lopm
ent
Mor
ale
etal
.(19
98)
Mal
athi
onP.
pugi
oL
etha
lity
ofla
rvae
,low
ernu
mbe
rof
inst
ars
topo
st-l
arva
e,le
thal
ityof
new
lyha
tche
dla
rvae
Key
etal
.(19
98,2
007)
Azi
npho
sm
ethy
lP.
pugi
oL
etha
lity
ofne
wly
hatc
hed
larv
aeK
eyet
al.(
2007
)
Org
anoc
hlor
ines
Met
hoxy
chlo
rSt
rong
yloc
entr
otus
purp
urat
usD
isru
ptio
nof
gast
rula
tion,
abno
rmal
clea
vage
,and
gut
deve
lopm
ent
Gre
enet
al.(
1997
)
Met
hoxy
chlo
r,di
eldr
in,
linda
neP.
livi
dus
Dec
reas
edra
teof
fert
iliza
tion,
incr
ease
dpo
lysp
erm
y,m
itotic
alte
ratio
ns,a
ltere
dde
velo
pmen
t,in
trac
ellu
lar
Ca
hom
eost
asis
Pesa
ndo
etal
.(20
04)
38 V. Pašková et al.
Tabl
e3
(con
tinue
d)
Pest
icid
eO
rgan
ism
Toxi
cef
fect
Ref
eren
ce
Lin
dane
M.g
allo
prov
inci
alis
Red
uctio
nof
embr
yoge
nesi
ssu
cces
sB
eira
san
dB
ella
s(2
008)
End
osul
fan
C.s
apid
usC
rass
ostr
eagi
gas
P.pu
gio
Inhi
bite
dha
tchi
ngin
high
erte
sted
conc
entr
atio
nsIn
crea
sed
abno
rmal
D-l
arva
ean
dth
ele
velo
fD
NA
stra
ndbr
eaks
Incr
ease
dha
tchi
ngtim
e
Lee
and
Osh
ima
(199
8)W
esse
leta
l.(2
007)
Wir
thet
al.(
2001
)Sy
nthe
tic
pyre
thro
ids
Fenv
aler
ate,
cype
rmet
hrin
C.s
apid
usIn
hibi
ted
hatc
hing
Lee
and
Osh
ima
(199
8)
Bif
enth
rin
P.pu
gio
Lar
vall
etha
lity
Har
per
etal
.(20
08)
Azo
les
Imaz
alil,
tria
dim
efon
P.m
amm
illa
taD
ecre
ased
rate
offe
rtili
zatio
n,al
tera
tions
ofth
ean
teri
orst
ruct
ures
,inc
orre
ctly
diff
eren
tiate
dpa
pilla
ryne
rves
and
nerv
ous
syst
em
Penn
atie
tal.
(200
6)
Phe
nylp
yraz
oles
Fipr
onil
P.pu
gio
Hig
hla
rvae
leth
ality
Key
etal
.(20
07)
Tria
zine
sA
traz
ine
P.pu
gio
Mer
cena
ria
mer
cena
ria
Mar
isa
corn
uari
etis
Non
-tox
icto
larv
aeat
test
edco
ncen
trat
ions
Incr
ease
ddr
ym
ass,
high
erco
nditi
onin
dex,
dose
-dep
ende
ntm
orta
lity,
high
ersh
ellm
ajor
axis
leng
thD
elay
inha
tchi
ng
Key
etal
.(20
07)
Law
ton
etal
.(20
06)
Saw
asde
ean
dK
öhle
r(2
009)
Ben
zoyl
phen
ylur
eas
Difl
uben
zuro
nC
.sap
idus
Inhi
bite
dha
tchi
ngL
eean
dO
shim
a(1
998)
Terp
enoi
dsM
etho
pren
eC
.sap
idus
P.pu
gio
Inhi
bite
dha
tchi
ngin
high
erte
sted
conc
entr
atio
nsN
oem
bryo
nic
mor
talit
yat
test
edco
ncen
trat
ions
Lee
and
Osh
ima
(199
8),
Wir
thet
al.(
2001
)N
itro
guan
idin
eni
coti
noid
sIm
idac
lopr
idP.
pugi
oM
.cor
nuar
ieti
sL
arva
ele
thal
ityD
ecre
ased
hear
trat
eK
eyet
al.(
2007
)Sa
was
dee
and
Köh
ler
(200
9)
Teratogenicity and Embryotoxicity in Aquatic Organisms 39
5 Role of Oxygen and Antioxidant Defenses in Embryogenesis
Despite intensive research, the mechanisms involved in teratogenesis are still notsufficiently understood, but it is known that they differ among various compounds.The following mechanisms by which xenobiotics may induce developmental effects,however, are known and include alterations of DNA (i.e., mutations, chromoso-mal aberrations, or nucleic acid metabolism effects), inhibition of specific enzymes,membrane alterations, modulation of cellular energy supplies, as well as disrup-tion of retinoic acid signaling or oxidative stress (Beckman and Brent 1984; Wellset al. 2005). Most often, a complex set of factors and/or the accumulation of errorsproduces morphological malformations or embryo lethality (Meteyer 2000).
Oxygen plays a key role in metabolism and is critical to the early developmentalstages of organisms. Several oxygen derivatives, known as reactive oxygen species(ROS), are known to have signaling functions and may affect several physiologicaland pathological processes in an organism (Covarrubias et al. 2008; de Lamirandeand Gagnon 1995). At the level of embryogenesis, sensitive regulation of ROShas been linked to control of oocyte cleavage (Allen and Balin 1989), as well asoocyte maturation, ovarian steroidogenesis, ovulation, implantation, and formationof blastocysts (Guerin et al. 2001).
However, ROS levels must be continuously controlled to prevent them frombecoming highly toxic to biological macromolecules (e.g., proteins, DNA, andmembrane lipids) (Agarwal et al. 2003). The teratogenic potential of xenobioticsthus depends on embryoprotective pathways and on detoxification and macro-molecule repair (Wells et al. 2005). General antioxidant defenses were recentlyshown to play an important role in protecting both early aquatic larval stages (Mariaet al. 2009; Tilton et al. 2008) and later developmental phases, as well as themetamorphosis process (Dandapat et al. 2003).
To protect against ROS, cells contain both non-enzymatic antioxidant molecules(the ubiquitous thiol-containing tripeptide glutathione, vitamin E, and metalloth-ioneins) (Tilton et al. 2008; Wiegand et al. 2001) and enzymes that can detoxifyROS. These enzymes include the following: superoxide dismutase (SOD; EC1.15.1.1) that exists in mitochondria and cytosol, catalase (CAT; EC 1.11.1.6) thatcatalyzes removal of hydrogen peroxide in peroxisomes, and glutathione peroxidase(GPx, EC 1.11.1.9) present in the nucleus, mitochondria, and cytosol (Wang et al.2002). Other important enzymes that are known to protect the embryo against reac-tive molecules are glutathione-S-transferase (GST; EC 2.5.1.18) (Anguiano et al.2001; Pena-Llopis et al. 2003) and glucose-6-phosphate dehydrogenase (G6PD; EC1.1.1.49) (Wells et al. 1997).
Antioxidants play an important role in protecting early larval stages against theeffects of ambient oxygen levels in water (Arun and Subramanian 1998; Dandapatet al. 2003). Dandapat et al. (2003) showed that glutathione content increased duringthe metamorphic progression of the giant prawn larvae Macrobrachium rosenbergii.A similar increase was also observed in developing grass shrimp, Palaemonetespugio (Winston et al. 2004), as well as in embryos of a toad, Bufo arenarum(Anguiano et al. 2001). In the trout, Salmo iridaeus, high CAT activities were
40 V. Pašková et al.
observed during early development (in contrast to relatively lower levels of GPx),thus documenting the prime role of CAT over GPx in the removal of toxic hydro-gen peroxide (Aceto et al. 1994). A gradual increase in CAT and GPx activities wasobserved during the progressive growth of the marine fish Dentex dentex from eggto larva (Mourente et al. 1999). On the other hand, in the same study, Mourenteet al. (1999) showed that the titers of two other detoxification enzymes (GST andSOD) reached their highest levels in eggs, compared to later developmental stages.A gradual increase of antioxidant enzyme activities during embryogenesis, fol-lowed by a sudden rise in freshly hatched larvae, was also observed for the prawnMacrobrachium malcolmsonii (Arun and Subramanian 1998).
As is apparent, the results of available studies confirm that temporal changes ofantioxidant agents are carefully regulated during the early development of aquaticorganisms. Hence, such antioxidants play an important role. Notwithstanding, moreresearch is needed to fully elucidate the physiological roles of ROS and antioxidantenzymes in maintaining homeostasis during early development.
6 Oxidative Stress in Embryotoxicity and Teratogenicity
While the previous section briefly described some physiological functions of ROSand the associated antioxidant defenses, the following paragraphs focus on theembryotoxicity that results from oxidative stress. Most research that has been con-ducted on this topic has included studies with model compounds such as hydrogenperoxide and has employed laboratory rodents or human embryos. In these studies,pro-oxidants induced severe oxidative stress damage to oocytes, mitochondrial alter-ations, ATP depletion, DNA damage and lipid peroxidation, apoptosis, or delaysin whole embryo development (Aitken and Krausz 2001; Duru et al. 2000; Nasr-Esfahani et al. 1990; Ozolins and Hales 1999). The importance of oxidative stressin causing embryotoxicity or teratogenicity was also indirectly confirmed in mam-malian and human studies, in which external additions of antioxidants preventeddamage to embryos (Feugang et al. 2004; Fraga et al. 1991).
Similar to what occurs in mammals, studies with model pro-oxidants have alsodemonstrated detrimental effects in fish embryos and larvae (Westernhagen von1988; Dietrich et al. 2005; Regoli et al. 2005), as well as in the larvae of the giantprawn M. rosenbergii (Dandapat et al. 2003). Moreover, the addition of antioxi-dants protected fish embryonal development against the effects of oxidative stress(Ciereszko et al. 1999; Toyokuni and Sagripanti 1992; Tilton et al. 2008).
7 Pesticides and Oxidative Damage During Early Developmentin Aquatic Organisms
Although pesticides may disrupt reproduction and development in many aquaticorganisms (see Tables 1, 2, and 3), our search of the literature disclosed only afew studies that experimentally documented the role of oxidative stress in pesticide-induced teratogenicity (see Table 4).
Teratogenicity and Embryotoxicity in Aquatic Organisms 41
Tabl
e4
Tera
toge
nic
effe
cts
ofso
me
pest
icid
esth
atar
elin
ked
toef
fect
son
deto
xific
atio
n,an
tioxi
dativ
epa
ram
eter
s,ox
idat
ive
stre
ss,
orm
odul
atio
nof
antio
xida
tive
proc
esse
sin
the
earl
ylif
est
ages
ofva
riou
sfis
hsp
ecie
s,am
phib
ians
,and
inve
rteb
rate
s
Pest
icid
eO
rgan
ism
Bio
chem
ical
effe
ctD
evel
opm
enta
leff
ect
Ref
eren
ces
Fis
hA
traz
ine
D.r
erio
Bre
akdo
wn
ofth
eG
STis
oenz
ymes
,al
tere
dm
icro
som
alan
dso
lubl
eG
STac
tivity
,abn
orm
alde
velo
pmen
t,at
razi
ne–G
SHco
njug
ate
form
atio
n
Del
ayin
embr
yoni
cde
velo
pmen
t,un
finis
hed
epib
oly
ored
ema,
decr
ease
ofhe
artr
ate
and
dysf
unct
ion
ofci
rcul
ator
ysy
stem
,red
uced
deve
lopm
ent,
redu
ced
gene
sis
ofey
es,s
omite
s,ot
olith
es,a
ndm
elan
opho
res
Wie
gand
etal
.(20
00,
2001
)
Ald
icar
b,al
dica
rbsu
lfox
ide
D.r
erio
Inhi
bite
dca
rbox
yles
tera
sehe
artr
ate
affe
ctio
n(e
mbr
yo)
Küs
ter
and
Alte
nbur
ger
(200
7)Pa
raqu
atO
.myk
iss
Ele
vate
dG
6PD
and
GR
activ
ity(j
uven
ile)
Lke
rman
etal
.(20
03)
Azi
npho
sm
ethy
lO
.myk
iss
Red
uced
GSH
leve
l,C
AT,
and
carb
oxyl
este
rase
activ
ity(j
uven
ile)
Ferr
arie
tal.
(200
7)
Car
bary
lO
.myk
iss
Inhi
bite
dca
rbox
yles
tera
sean
dG
SHle
vel,
alte
red
CA
T,in
duce
dG
STac
tivity
and
CY
P1A
leve
l(ju
veni
le)
Ferr
arie
tal.
(200
7)
Dic
hlor
vos
S.au
rata
Incr
ease
dL
Ple
vel,
decr
ease
dR
NA
/DN
Ara
tio(j
uven
iles)
Var
óet
al.(
2007
)
End
osul
fan
O.n
ilot
icus
RO
Spr
oduc
tion,
lipid
pero
xida
tion
(juv
enile
s)Te
llez-
Ban
uelo
set
al.
(200
9)H
exac
hlor
oben
zene
C.c
arpi
oA
ltere
dG
SHco
nten
tand
SOD
and
GPx
,G
R,G
SSG
activ
ity,R
OS
gene
ratio
n,lip
idpe
roxi
datio
n(j
uven
iles)
Song
etal
.(20
06)
42 V. Pašková et al.
Tabl
e4
(con
tinue
d)
Pest
icid
eO
rgan
ism
Bio
chem
ical
effe
ctD
evel
opm
enta
leff
ect
Ref
eren
ces
Am
phib
iaC
arba
ryl
B.a
rena
rum
Alte
red
GSH
cont
ent,
SOD
,GST
,and
GR
activ
ity,i
ncre
ased
CA
Tan
dG
PXac
tivity
Prog
ress
ive
drop
sy,b
ody
bend
ing,
and
para
lysi
sFe
rrar
ieta
l.(2
009)
Azi
npho
sm
ethy
lB
.are
naru
mD
ecre
ased
GSH
leve
l,al
tere
dG
R,G
PX,
GST
,and
CA
T,de
crea
sed
SOD
activ
ityG
illat
roph
y,no
toch
ord
curv
atur
e,fo
lded
tail
fin,g
ener
aliz
edde
lay
inth
ede
velo
pmen
t,hy
pera
ctiv
ity
Ferr
arie
tal.
(200
9)
Para
thio
nB
.are
naru
mA
ltera
tion
ofG
STac
tivity
Dec
reas
edra
teof
gast
rula
s,cu
rvat
ure
ofth
ean
tero
-pos
teri
orax
is,t
ailf
oldi
ng,
circ
le-s
wim
min
gm
ovem
ent,
freq
uent
drop
sy,a
nded
ema
Ang
uian
oet
al.(
2001
)
Mal
athi
onB
.are
naru
m
R.b
oyli
i
Red
uced
GSH
cont
enti
nbo
them
bryo
san
dla
rvae
,inc
reas
edG
STac
tivity
Dec
reas
edG
Ran
dC
AT
activ
ities
and
the
GSH
pool
(em
bryo
larv
al)
Inhi
bite
dca
rbox
yles
tera
seac
tivity
,in
duce
dm
ixed
func
tion
oxid
ase
(lar
vae)
Dep
lete
dac
id-s
olub
leth
iols
(lar
vae)
Lar
valm
orta
lity
Dec
reas
edra
teof
gast
rula
s,cu
rvat
ure
ofth
ean
tero
-pos
teri
orax
is,t
ailf
oldi
ng,
circ
le-s
wim
min
gm
ovem
ent,
freq
uent
drop
sy,a
nded
ema
Ang
uian
oet
al.(
2001
)
Ferr
arie
tal.
(200
8)
Ven
turi
noet
al.(
2001
a,b)
Spar
ling
and
Felle
rs(2
007)
Lin
dane
B.a
rena
rum
Dec
reas
edG
SHin
embr
yo,i
ncre
ased
GST
activ
ityD
ecre
ased
rate
ofga
stru
las,
irre
gula
rbl
asto
mer
es,a
xis
curv
atur
e,ta
ilfo
ldin
g,ed
ema,
orga
ndi
spla
cem
ent,
hem
orrh
age,
hype
ract
ivity
,pr
ofus
esc
alin
g,dr
opsy
,org
andi
spla
cem
ent,
and
bent
tail
Ang
uian
oet
al.(
2001
)
Teratogenicity and Embryotoxicity in Aquatic Organisms 43
Tabl
e4
(con
tinue
d)
Pest
icid
eO
rgan
ism
Bio
chem
ical
effe
ctD
evel
opm
enta
leff
ect
Ref
eren
ces
Die
ldri
nB
.are
naru
mIn
crea
sed
GST
activ
ityE
xoga
stru
latio
n,hy
pera
ctiv
ity,
hem
orrh
agia
Ang
uian
oet
al.(
2001
)
Ace
toch
lor
Buf
ora
ddei
Enh
ance
dL
Ple
vela
ndD
NA
sing
le-s
tran
dbr
eak
inliv
er(j
uven
ile)
Liu
etal
.(20
06)
Inve
rteb
rate
sM
alat
hion
C.g
igas
Incr
ease
dC
AT
activ
ity(l
arva
e)D
amie
nset
al.(
2004
)H
epta
chlo
rH
omar
usam
eric
anus
Ele
vate
dC
YP4
5an
dH
SP70
leve
lsL
arva
em
orta
lity,
dela
ysin
ecdy
sis
Snyd
eran
dM
ulde
r(2
001)
Car
bofu
ran
C.g
igas
Incr
ease
dL
Ple
vel,
mod
ulat
ion
ofC
AT
activ
ity(l
arva
e)D
amie
nset
al.(
2004
)
Perm
ethr
inP.
pugi
oIn
crea
sed
time
toha
tch,
leth
argy
ofla
rvae
,al
tere
dG
SHan
dL
Ple
vels
(lar
vae)
DeL
oren
zoet
al.(
2006
)
CA
T,ca
tala
se;C
oA,c
oenz
yme
A;C
YP1
A,c
ytoc
hrom
eP
450
1Ais
ozym
e;E
RO
D,e
thox
yres
orufi
n-O
-dee
thyl
ase;
G6P
D,g
luco
se-6
-pho
spha
tede
hydr
ogen
ase;
GSH
,gl
utat
hion
e;G
SSG
,ox
idiz
edgl
utat
hion
e;G
R,
glut
athi
one
redu
ctas
e;G
ST,
glut
athi
one-
S-tr
ansf
eras
e;G
Px,
glut
athi
one
pero
xida
se;
HSP
,he
at-s
hock
prot
ein;
LP,
lipid
pero
xida
tion;
RO
S,re
activ
eox
ygen
spec
ies;
SOD
,sup
erox
ide
dism
utas
e
44 V. Pašková et al.
Among the few pesticides on which oxidative stress effects were studied werethe following: the triazine herbicide atrazine, the organophosphate insecticidesparathion and azinphos methyl, and the organochlorine insecticides dieldrin andlindane.
Exposures of D. rerio embryos to atrazine lead to retardation of organogenesis(especially eyes, somites, otolithes, and melanophores), dysfunctions of the circu-latory system, edemas, and a delay in embryonic development; in addition, theseeffects occurred in parallel with alterations of GST activities (Wiegand et al. 2000,2001).
Mortality in embryos and developmental abnormalities, along with oxidativestress markers, were also observed in two studies with embryos of the toadB. arenarum. Anguiano et al. (2001) discovered that the organochlorine insecticidelindane caused abnormal segmentation of furrows, along with irregular blastomeres,profuse scaling, dropsy, organ displacements, and bent tail. Interestingly, only mod-erate alterations of embryonic morphology and hemorrhagia were observed afterexposure to another organochlorine insecticide – dieldrin (Anguiano et al. 2001). Inthe same study, Anguiano et al. (2001) also showed that the organophosphate insec-ticides malathion and parathion were highly embryotoxic and caused a pathologicalcurvature of the antero-posterior axis, tail folding edema, frequent dropsy, and alsoinduced circle-swimming movements. Ferrari et al. (2009) studied the effects ofcarbaryl and azinphos methyl on the embryos of B. arenarum and demonstratedprogressive dropsy, notochord malformations, gill atrophy, paralysis, and delayeddevelopment. The above-described effects were also correlated with modulationsof glutathione levels and elevated activities of GST, SOD, CAT, and glutathionereductase (GR; EC 1.8.1.7) (Anguiano et al. 2001; Ferrari et al. 2009).
In studies with invertebrates, Snyder and Mulder (2001) demonstrated oxida-tive stress and pesticide toxicity after exposure to heptachlor or disruption ofgrass shrimp development by permethrin (DeLorenzo et al. 2006). Damiens et al.(2004) also showed larval toxicity and modulation of antioxidant and detoxificationparameters after exposures to complex media contaminated with pesticides.
Direct toxic effects of pesticides on developing fish embryos were not found inother studies, but signs of oxidative stress and variable modulation of the antiox-idative system were observed (Lkerman et al. 2003; Song et al. 2006; Ferrari et al.2007; Varó et al. 2007; Küster and Altenburger 2007; Tellez-Banuelos et al. 2009).
The bipyridyl herbicides paraquat and diquat are of special interest. The majormechanism by which they produce their toxic action in target organisms, whetheranimals or plants, is through lipid peroxidation. Disturbances of normal early devel-opmental processes, after exposure to paraquat, were clearly documented to haveoccurred in X. laevis embryos (Vismara et al. 2000, 2001a; see Table 2). These toxiceffects were prevented after the addition of the water-soluble antioxidant ascorbicacid to the test medium (Vismara et al. 2001b, 2006). Antioxidant protection byascorbic acid was also confirmed in our studies (unpublished results), in which wecompared the embryotoxicity of diquat and paraquat to X. laevis.
Few studies exist in which the oxidative stress damage caused by pesti-cides (herbicides atrazine, paraquat, and diquat; the organophosphate insecticides
Teratogenicity and Embryotoxicity in Aquatic Organisms 45
parathion and azinphos methyl; organochlorines dieldrin and lindane) to developingaquatic organisms has been described. Therefore, further research is needed in thisarea to better understand the levels of embryotoxicity that may result from other,less-explored pesticides.
8 Further Evidence – Pesticides and AntioxidativeDefense in Adult Aquatic Biota
Although there are only a limited number of studies that link oxidative stress causedby pesticides with embryotoxicity, other evidence with adult aquatic organismsexists that supports the importance of this mechanism. Any comprehensive treat-ment of the topic of pesticide-induced oxidative stress in adults is beyond the scopeof this chapter, and there are several credible recent reviews that address this topic(Valavanidis et al. 2006; Monserrat et al. 2006; Slaninova et al. 2009; Debenest et al.2010). Nevertheless, to provide supporting data, representative studies that addressselected pesticides (e.g., organophosphates, organochlorines, and bipyridyl herbi-cides) are presented in Tables 5, 6, and 7 (these tables address studies with fish,amphibians, and invertebrates, respectively).
In general, exposures of adult specimens to different pesticides induced antiox-idative defenses, such as increases in titers of ethoxyresorufin O-deethylase(EROD), SOD, GST, and GR or G6PD, along with declines in glutathione con-centrations and oxidative damage to lipids, DNA, proteins, and tissues (hepaticalterations, necrosis, etc.). For example, diazinon and glyphosate induced tissue-specific alterations of CAT and GPx activities, together with enhanced SOD activityand lipid peroxidation, in fish species (Oreochromis niloticus and Prochilodus lin-eatus; Durmaz et al. 2006; Langiano et al. 2008). Similarly, lipid peroxidationand detoxification responses were induced in methyl parathion-exposed Bryconcephalus (Monteiro et al. 2006), as well as in the mosquito fish Gambusia affi-nis that was exposed to the organophosphorus insecticides monocrotophos andchlorpyrifos (Kavitha and Rao 2008; Kavitha and Rao 2007, 2008). Glutathioneredox cycle and CAT were shown to protect against endosulfan-induced toxicity introut Oncorhynchus mykiss cells (Dorval et al. 2003; Dorval and Hontela 2003).Endosulfan also induced lipid peroxidation and altered various enzymatic activitiesin Jenynsia multidentata (Ballesteros et al. 2009). Paraquat, a herbicide that acts viaROS production, induced lipid peroxides and modulated SOD, GR, and GST in var-ious fish, such as Sparus aurata (Pedrajas et al. 1995; Rodríguez-Ariza et al. 1999)or Nile tilapia O. niloticus (Figueiredo-Fernandes et al. 2006a, b). Paraquat alsoinduced protein carbonylation in liver, kidney, and gills and modulated glutathioneand ascorbic acid levels in Channa punctata (Parvez and Raisuddin 2005, 2006).
Comparable results were obtained in experiments with the amphibian (frog)Rana ridibunda (Table 6). Herein, mixtures of propamocarb and mancozeb causedelevated lipid and protein peroxidation and suppressed SOD activity (Falfushinskaet al. 2008). Oxidative stress caused by pesticides, pyrethroid insecticides forexample, was also observed in several experiments performed with invertebrates
46 V. Pašková et al.
Tabl
e5
The
effe
cts
ofse
lect
edpe
stic
ides
onde
toxi
ficat
ion,
antio
xida
tive,
and
othe
rim
port
antb
ioch
emic
alpa
ram
eter
sin
fish
Pest
icid
eO
rgan
ism
Bio
chem
ical
effe
ctR
efer
ence
s
Car
bam
ates
Mol
inat
eA
ngui
lla
angu
illa
Incr
ease
dhe
patic
GSH
,GR
and
GSH
:GSS
Gra
tio,
decr
ease
dm
uscu
lar
GSH
Pena
-Llo
pis
etal
.(20
01)
Car
bary
lO
.nil
otic
usD
ecre
ased
SOD
,GR
,GST
,and
CA
Tac
tivity
,he
pato
cellu
lar
baso
phili
a,ne
crot
icfo
ciM
atos
etal
.(20
07)
Org
anop
hosp
hate
sM
onoc
roto
phos
G.a
ffini
sSO
D,C
AT,
GR
,and
lipid
pero
xida
tion
indu
ctio
n,re
cove
ryre
spon
seof
antio
xida
nten
zym
esK
avith
aan
dR
ao(2
007)
Chl
orpy
rifo
sG
.affi
nis
SOD
,CA
T,an
dG
Rin
hibi
tion,
reco
very
resp
onse
ofan
tioxi
dant
enzy
mes
,lip
idpe
roxi
datio
nK
avith
aan
dR
ao(2
008)
Azi
npho
sm
ethy
lO
.nil
otic
us
C.c
arpi
o
Incr
ease
dac
tivity
ofG
6PD
,GPx
,and
GR
,SO
Dde
crea
seIn
crea
seof
SOD
and
GST
activ
ities
,ele
vatio
nin
CA
Tan
dG
Pxac
tiviti
esin
carp
decr
ease
ofG
Pxin
tilap
ia
Oru
çand
Üne
r(2
000)
Oru
çet
al.(
2004
)
Gly
phos
ate
P.li
neat
us
Car
assi
usau
ratu
s
Plas
ma
gluc
ose
and
CA
Tac
tivity
incr
ease
,his
tolo
gica
lal
tera
tions
impa
irin
gno
rmal
orga
nfu
nctio
nsR
educ
edSO
Dan
dG
Pxac
tiviti
es,i
ncre
ased
GST
activ
ityan
dG
SHle
vel,
lipid
pero
xida
tion
Dec
reas
edG
SHco
nten
tand
SOD
,GR
,G6P
HD
activ
ityan
dhe
patic
GST
,inc
reas
edC
AT
activ
ity
Lan
gian
oan
dM
artin
ez(2
008)
,Mod
esto
and
Mar
tinez
(201
0)
Lus
hcha
ket
al.(
2009
)
Dia
zino
nO
.nil
otic
us
O.m
ykis
s
C.c
arpi
o
SOD
incr
ease
,CA
Tan
dG
Pxal
tera
tion,
lipid
pero
xida
tion
Incr
ease
dL
Ple
vel,
GSH
depl
etio
n,m
odul
atio
nof
SOD
,G
R,G
ST,G
PXac
tiviti
esIn
crea
sed
SOD
,dec
reas
edC
AT,
and
alte
red
GPx
activ
ityan
dpr
otei
nca
rbon
ylle
vel,
lipid
pero
xida
tion
Dur
maz
etal
.(20
06)
Isik
and
Cel
ik(2
008)
Oru
çand
Ust
a(2
007)
Teratogenicity and Embryotoxicity in Aquatic Organisms 47
Tabl
e5
(con
tinue
d)
Pest
icid
eO
rgan
ism
Bio
chem
ical
effe
ctR
efer
ence
s
Met
hylp
arat
hion
Bry
con
ceph
alus
O.m
ykis
s
SOD
,CA
T,G
STin
duct
ion,
GPx
alte
ratio
n,lip
idpe
roxi
datio
nIn
crea
sed
LP
leve
l,G
SHde
plet
ion,
mod
ulat
ion
ofSO
D,
GR
,GST
,GPx
activ
ities
Mon
teir
oet
al.(
2006
)
Isik
and
Cel
ik(2
008)
Mal
athi
onS.
aura
taIn
crea
sed
LP
and
GST
Pedr
ajas
etal
.(19
95)
Fent
hion
O.n
ilot
icus
Incr
ease
dG
SHan
dG
SSG
cont
enta
ndG
Pxac
tivity
Pine
ret
al.(
2007
)
Org
anoc
hlor
ines
End
osul
fan
O.m
ykis
sJ.
mul
tide
ntat
a
C.p
unct
ata
CA
T,G
Px,G
SHal
tera
tions
,lip
idpe
roxi
datio
nR
educ
edG
SHan
dG
Px,i
nduc
edG
STan
dC
AT,
lipid
pero
xida
tion
GST
,GR
,GPx
,and
CA
Tal
tera
tions
,lip
idpe
roxi
datio
nIn
crea
sed
prot
ein
carb
onyl
sin
liver
,kid
ney,
and
gills
Dor
vala
ndH
onte
la(2
003)
Dor
vale
tal.
(200
3)
Bal
lest
eros
etal
.(20
09)
Parv
ezan
dR
aisu
ddin
(200
5)D
ield
rin
S.au
rata
CA
T,SO
D,a
ndpa
lmito
yl-C
oA-o
xida
sein
duct
ion,
incr
ease
dpr
otei
nco
ncen
trat
ion
inpe
roxi
som
alfr
actio
n;in
crea
sed
LP
and
GST
;incr
ease
d8-
oxoG
mar
ker
Pedr
ajas
etal
.(19
95,1
996)
,R
odrí
guez
-Ari
zaet
al.(
1999
)
Chl
orot
halo
nil
Mor
one
saxa
tili
sR
OS
prod
uctio
n,al
tere
dG
SHle
vel
Bai
er-A
nder
son
and
And
erso
n(2
000)
DD
TH
opli
asm
alab
aric
usIn
crea
sed
intr
acel
lula
rR
OS,
incr
ease
dC
AT
and
G6P
DH
activ
ities
,GSH
cont
ent,
lipid
pero
xida
tion
and
prot
ein
carb
onyl
leve
l,de
crea
sed
SOD
,GST
,and
GR
activ
ities
;dec
reas
edce
llvi
abili
ty
Filip
akN
eto
etal
.(20
08)
Tria
zine
sA
traz
ine
D.r
erio
Lep
omis
mac
roch
irus
Indu
ced
cyto
chro
me
P450
cont
ent,
incr
ease
dN
AD
PH-P
450
redu
ctas
e,er
ythr
omyc
inN
-dem
ethy
lase
,and
amin
opyr
ine
N-d
emet
hyla
seIn
crea
sed
hepa
ticG
SHan
dG
SSG
leve
lsan
dG
ST,S
OD
activ
ities
,lip
idpe
roxi
datio
n,al
tere
dG
Pxan
dG
Rac
tiviti
es
Don
get
al.(
2009
)
Elia
etal
.(20
02)
48 V. Pašková et al.
Tabl
e5
(con
tinue
d)
Pest
icid
eO
rgan
ism
Bio
chem
ical
effe
ctR
efer
ence
s
Mix
ture
ofat
razi
ne,
sim
azin
e,di
uron
,is
opro
turo
n
C.a
urat
usE
nhan
ced
prod
uctio
nof
supe
roxi
de,i
nduc
edSO
Dac
tivity
inliv
er,r
educ
edC
AT
activ
ityFa
tima
etal
.(20
07)
Sim
azin
eC
.car
pio
Lip
idpe
roxi
datio
n,in
crea
sed
GSH
leve
lO
rope
saet
al.(
2009
)
Bip
yrid
yls
Para
quat
O.n
ilot
icus
C.p
unct
ata
S.au
rata
Hig
hhe
pato
-som
atic
and
gona
do-s
omat
icin
dex,
incr
ease
dE
RO
Dac
tivity
,hep
atic
alte
ratio
nsof
pare
nchy
ma,
like
vacu
oliz
atio
n,ne
cros
is,a
ndan
incr
ease
ofm
acro
phag
eag
greg
ates
and
eosi
noph
ilic
gran
ular
cells
;hig
her
SOD
,GST
activ
ity,
gend
er-d
epen
dent
incr
ease
ofG
RIn
crea
sed
prot
ein
carb
onyl
sin
liver
,kid
ney,
and
gills
,re
duce
dG
SHin
liver
and
gills
,inc
reas
edas
corb
icac
idle
vel
Incr
ease
dL
P,G
ST,a
ndSO
DD
ecre
ased
GPX
,hig
her
G-6
PDH
activ
ity,i
ncre
ased
GR
Figu
eire
do-F
erna
ndes
etal
.(20
06a,
b)
Parv
ezan
dR
aisu
ddin
(200
5,20
06)
Pedr
ajas
etal
.(19
95),
Rod
rígu
ez-A
riza
etal
.(19
99)
2,4-
D-P
heno
xyca
rbox
ylic
acid
2,4-
Dic
hlor
ophe
noxy
acet
icac
iddi
met
hyla
min
eO
.nil
otic
usC
.car
pio
Incr
ease
dac
tivity
ofG
6PD
,GPx
,and
GR
Incr
ease
ofSO
Dan
dG
STac
tivity
,ele
vatio
nin
CA
Tan
dG
Pxac
tiviti
esin
carp
decr
ease
ofG
Pxin
tilap
ia
Oru
çan
dÜ
ner
(200
0)O
ruç
etal
.(20
04)
Synt
heti
cpy
reth
roid
sD
elta
met
hrin
C.p
unct
ata
Lip
idpe
roxi
datio
nan
dG
SHle
veli
ncre
ase,
alte
ratio
nsin
CA
T,SO
D,a
ndG
STac
tivity
and
asco
rbic
acid
leve
lIn
crea
sed
prot
ein
carb
onyl
sin
liver
,kid
ney,
and
gills
Saye
edet
al.(
2003
),Pa
rvez
and
Rai
sudd
in(2
005)
Cyp
erm
ethr
inC
.car
pio
O.n
ilot
icus
Incr
ease
dhe
patic
SOD
and
CA
Tac
tiviti
es,d
ecre
ased
GPx
,lip
idpe
roxi
datio
nIn
crea
sed
hepa
ticSO
D,G
Px,a
ndC
AT
activ
ities
,lip
idpe
roxi
datio
n
Une
ret
al.(
2001
)
Teratogenicity and Embryotoxicity in Aquatic Organisms 49
Tabl
e5
(con
tinue
d)
Pest
icid
eO
rgan
ism
Bio
chem
ical
effe
ctR
efer
ence
s
Bif
enth
rin
Cyp
rino
don
vari
egat
usIn
crea
sing
tren
din
GSH
leve
land
CA
Tac
tivity
with
toxi
cant
dose
Har
per
etal
.(20
08)
Chl
oroa
ceta
nili
des
But
achl
orC
.gar
iepi
nus
Lip
idpe
roxi
datio
n,al
tere
dSO
D,G
ST,a
ndC
AT
activ
ities
and
GSH
leve
lFa
rom
biet
al.(
2008
)
Org
anofl
uori
neE
toxa
zole
O.n
ilot
icus
Lip
idpe
roxi
datio
nSe
vgile
ret
al.(
2004
)
Pyr
azol
esFe
npyr
oxim
ate
Para
lich
thys
oliv
aceu
sA
ltere
dSO
D,C
AT,
GST
,GPx
,and
ER
OD
activ
ities
Na
etal
.(20
09)
Dic
hlor
oben
zene
s3,
4-D
ichl
oroa
nilin
eC
.aur
atus
Enh
ance
dSO
Dac
tivity
and
LP,
decr
ease
dN
Osy
ntha
sean
dG
SHle
veli
nliv
erL
ieta
l.(2
003)
CA
T,ca
tala
se;
ER
OD
,eth
oxyr
esor
ufin-
O-d
eeth
ylas
e;G
6PD
,glu
cose
-6-p
hosp
hate
dehy
drog
enas
e;G
SH,g
luta
thio
ne;
GSS
G,o
xidi
zed
glut
athi
one;
GR
,glu
-ta
thio
nere
duct
ase;
GST
,glu
tath
ione
-S-t
rans
fera
se;G
Px,g
luta
thio
nepe
roxi
dase
;NA
DPH
P450
,NA
DPH
–cyt
ochr
ome
P450
redu
ctas
e;N
O,n
itric
oxid
e;L
P,lip
idpe
roxi
datio
n;SO
D,s
uper
oxid
e
50 V. Pašková et al.
Table 6 Effects of carbamates on detoxification, antioxidative, and other important biochemicalparameters in amphibians
Pesticide Organism Biochemical effect Reference
Amphibians
CarbamateMixture of propamocarb
and mancozebR. ridibunda Decreased SOD; lipid and
protein peroxidation,neurotoxicity and endocrinedisruption
Falfushinska et al.(2008)
SOD, superoxide dismutase
(Table 7). Several studies were conducted with mollusks on the pyrethroid insec-ticides, for example, cypermethrin and alphamethrin; these insecticides inhibitedreproduction and induced oxidative stress in the freshwater snail Lymnaea acumi-nata (Tripathi and Singh 2004). In addition, lindane, an organochlorine insecticide,induced oxidative damage and stress responses in the mussel Mytilus galloprovin-cialis (Khessiba et al. 2005).
Table 7 The effects of selected pesticides on detoxification, antioxidative, and other importantbiochemical parameters in aquatic invertebrates
Pesticide Organism Biochemical effect References
OrganophosphatesFenitrothion M. galloprovincialis
Flexopectenflexuosus
GSH and GSSG depletion,reduction of GSH/GSSGratio, decreased survival
Pena-Llopis et al.(2002)
OrganochlorinesLindane M. galloprovincialis CAT activity induction Khessiba et al.
(2005)Endosulfan Penaeus monodon Increased HSP level in
musclesDorts et al. (2009)
Tetradifon D. magna Decreased protein, lipid,glycogen, and caloriccontent, decreased meanbody dry weight
Villarroel et al.(2009)
Synthetic pyrethroidsCypermethrinalphamethrin
L. acuminata Altered oxidative metabolismin hepatopancreas andovotestis tissues, reducedsurvival
Tripathi and Singh(2004)
Deltamethrin P. monodon Increased LP level and proteincarbonyls in gills
Dorts et al. (2009)
BipyridylsParaquat Biomphalaria
glabrataIncreased LP, decreased SOD Cochón et al. (2007)
CAT, catalase; GSH, glutathione; GSSG, oxidized glutathione; HSP, heat-shock protein; LP, lipidperoxidation; SOD, superoxide dismutase
Teratogenicity and Embryotoxicity in Aquatic Organisms 51
9 Summary
Many pesticides have been documented to induce embryotoxicity and teratogenic-ity in non-target aquatic biota such as fish, amphibians, and invertebrates. Ourreview of the existing literature shows that a broad range of pesticides, repre-senting several different chemical classes, induce variable toxic effects in aquaticspecies. The effects observed include diverse morphological malformations as wellas physiological and behavioral effects. When developmental malformations occur,the myoskeletal system is among the most highly sensitive of targets. Myoskeletaleffects that have been documented to result from pesticide exposures include com-mon notochord and vertebrate column degeneration and related abnormalities.Pesticides were also shown to interfere with the development of organ systemsincluding the eyes or the heart and are also known to often cause lethal or sub-lethal edema in exposed organisms. The physiological, behavioral, and populationendpoints affected by pesticides include low or delayed hatching, growth suppres-sion, as well as embryonal or larval mortality. The risks associated with pesticideexposure increase particularly during the spring. This is the period of time in whichmajor pesticide applications take place, and this period unfortunately also coincideswith many sensitive reproductive events such as spawning, egg laying, and earlydevelopment of many aquatic organisms.
Only few experimental studies with pesticides have directly linked developmen-tal toxicity with key oxidative stress endpoints, such as lipid peroxidation, oxidativeDNA damage, or modulation of antioxidant mechanisms. On the other hand, it hasbeen documented in many reports that pesticide-related oxidative damage occursin exposed adult fish, amphibians, and invertebrates. Moreover, the contribution ofoxidative stress to the toxicity of pesticides has been emphasized in several recentreview papers that have treated this topic.
In conclusion, the available experimental data, augmented by several indirectlines of evidence, provide support to the concept that oxidative stress is a highlyimportant mechanism in pesticide-induced reproductive or developmental toxicity.Other stressors may also act by oxidative mechanisms. This notwithstanding, thereis much yet to learn about the details of this phenomenon and further research isneeded to more fully elucidate the effects that pesticides have and the environmentalrisks they pose in the early development of aquatic organisms.
Acknowledgments Authors highly acknowledge all comments and recommendations of the edi-tor Dr. David M. Whitacre and two anonymous referees that significantly contributed to the qualityof the manuscript. Research Centre for Toxic Compounds in the Environment is supported bythe project CETOCOEN (no. CZ.1.05/2.1.00/01.0001) from the European Regional DevelopmentFund.
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Curriculum Vitae
Personal information
First name / Surname VERONIKA PAŠKOVÁ
Address POUCHOVSKÁ 179/3, 503 41 HRADEC KRÁLOVÉ, CZECH REPUBLIC
Telephone +420 724 532 969
E-mail paskova@recetox.muni.cz
Nationality Czech
Date of birth 26/07/1981
Gender FEMALE
Education and training
Dates
09/2005 – today PhD studies
Name and type of organisation providing education and training
Masaryk University, Faculty of Science, Research Centre for Toxic Compounds in the Environment, Brno, Czech Republic
Principal subjects/occupational skills covered
Dissertation thesis: “Biomarkers in experimental ecotoxicology” – Environmental Chemistry
Dates
11/2007 Rigorous exam
Name and type of organisation providing education and training
Faculty of Science, Masaryk University, Brno, Czech Republic
Principal subjects/occupational skills covered
Rigorous exam in Biology, specialization in Ecotoxicology, subject: „Plant oxidative stress responses after exposure to polycyclic aromatic compounds and thein N-heterocyclic derivates“
Level in national or international classification
Doctor of natural sciences
Title of qualification awarded RNDr.
Dates 09/2000 to 06/2005 Master study programme
Name and type of organisation providing education and training
Faculty of Science, Masaryk University, Brno, Czech Republic
Principal subjects/occupational skills covered
Five years Masters programme Ecotoxicology, finished by diploma work: “Ecotoxicity of polycyclic aromatic compounds and their derivatives in plants – assessment of biomarkers of exposure and effect” (graduate cum laude)
Title of qualification awarded M. Sc.
Dates 09/2000 to 06/2005 Master study programme
Name and type of organisation providing education and training
Faculty of Science, Masaryk University, Brno, Czech Republic
Principal subjects/occupational skills covered
Five years Masters programme Biology – Teacher Training of Biology for Secondary Schools (graduate cum laude)
Title of qualification awarded M.Sc.
Work experience
Dates July 2008 - today
Occupation or position held Project CETOCOEN administration; research assistant
Name and address of employer Research Centre for Toxic Compounds in the Environment, Masaryk University, Brno, Czech Republic
Dates 02/2009 – 04/2010
Occupation or position held Ph.D. student of the Academy of Science of the Czech Republic
Name and address of employer Institute of Botany, Czech Academy of Sciences, Czech Republic
Research project 2006 to 2007
Chancellor’s grant for support the students 2006 14 31 E 1232 (Masaryk University, Brno, Czech Republic) – Malformations and oxidative stress in the FETAX test together with the study of toxic effects of selected pollutants - principal investigator
Veronika Pašková – Publications Articles in international journals with impact factors Pašková, V., Hilscherová, K., Feldmannová, M. and Bláha, L. (2006). Toxic effects and oxidative stress in higher plants exposed to polycyclic
aromatic hydrocarbons and their N-heterocyclic derivatives. Environmental Toxicology and Chemistry 25/12, pp. 3238-3245
Skočovská, B. Hilscherová, K., Babica, P., Adamovský, O., Banďouchová, H., Horáková, J., Knotková, Z., Maršálek, B., Pašková, V. and Pikula, J. (2007). Effects of cyanobacterial biomass on the Japanese quail. Toxicon 49/1, pp. 493-803
Adamovský, O., Kopp, R., Hilscherová, K., Babica, P., Palíková, M., Pašková, V., Navrátil, S., Maršálek, B., Bláha, L. (2007). Microcystins kinetics
(bioaccumulation, elimination) and biochemical responses in common carp and silver carp exposed to toxic cyanobacterial blooms Environmental Toxicology and Chemistry 26/12, pp. 2687-2693 Smutná, M., Hilscherová, K., Pašková, V., Maršálek, B. (2008). Biochemical parameters in Tubifex tubifex as an integral part of complex sediment
toxicity assessment. Journal of Soils and sediments 8/3, pp. 154-164 Pašková, V., Adamovský, O., Pikula, J., Skočovská, B., Banďouchová, H., Horáková, J., Babica, P., Maršálek, B., Hilscherová, K. (2008).
Detoxification and oxidative stress responses along with microcystins accumulation in Japanese quail exposed to cyanobacterial biomass. Science of the Total Environment 398/1-3, pp. 34-47 Pikula J., Adamovský O., Banďouchová H., Horáková J., Machát J., Maršálek B., Pašková V. (2008). Effects of co-exposure to cyanobacterial
biomass, lead and immunological challenge in Japanese quails. Toxicology letters 180S1, pp. 197 Damková V., Sedláčková J., Banďouchová H., Pecková L., Vitula F., Hilscherová K., Pašková V., Kohoutek J.2, Pohanka M., Pikula J. (2009).
Effects of cyanotoxins on avian reproduction: a japanese quail model. Neuroendocrinology Letters, Sweden: Society of Integrated Sciences, 30, Suppl. 1, pp. 205 – 210.
Pecková L., Banďouchová H., Hilscherová K., Damková V., Sedláčková J., Vitula F., Pašková V., Pohanka M., Kohoutek J., Pikula J. (2009) Biochemical responses of juvenile Japanese quails to cyanobacterial biomass. Neuroendocrinology Letters, Sweden: Society of
Integrated Sciences, 30, Suppl.1, pp. 199 – 204. Pikula J., Banďouchová H., Hilscherová K., Pašková V., Sedláčková J., Adamovský O., Knotková Z., Laný P., Machát J., Maršálek B., Novotný L.,
Pohanka M., Vitula F. (2010) Combined exposure to cyanobacterial biomass, lead and the Newcastle virus enhances avian toxicity. Science of the Total Environment 408/21, pp. 4984-4992.
Pikula J., Damková V., Banďouchová H., Pašková V., Hilscherová K., Pohanka M., Ondráček K., Vitula F. (2011) Effects of cyanotoxins and lead on
avian reproduction. Toxicology letters 205/1, pp. S251 Damková V., Pašková V., Banďouchová H., Hilscherová K., Sedláčková J., Novotný L., Pecková L., Vitula F., Pohanka M., Pikula J. (2011)
Testicular toxicity of cyanobacterial biomass in Japanese quails. Harmful Algae 10/6, pp. 612-618 Pašková, V., Paskerová, H., Pikula, J., Banďouchová, H., Sedláčková, J. and Hilscherová, K. (2011). Combined exposure of Japanese quails to
cyanotoxins, Newcastle virus and lead: Oxidative stress responses. Ecotoxicology and Environmental Safety 74/7, pp. 2082-2090.
Pašková, V., Hilscherová, K. and Bláha, L. (2011). Teratogenicity and embryotoxicity in aquatic organisms after pesticide exposure and the role of oxidative stress. Reviews of Environmental Contamination and Toxicology, Vol. 211, pp. 25-61.
Czech peer-reviewed journal articles: Pašková V., Hilscherová K. (2007). Embryotoxicita a indukce oxidativního stresu po exposici herbicidu paraquatu na modelovém necílovém
organismu drápatce vodní (Xenopus laevis). Bulletin VÚRH Vodňany 43/3, s. 107-112 Palíková M., Mareš J., Pašková V., Kopp R., Adamovský O., Hilscherová K., Bláha L. & Navrátil S. (2007) Ovlivnění nutriční hodnoty svaloviny
kapra obecného (Cyprinus carpio) a tolstolobika bílého (Hypophthalmichtys molitrix) cyanobakteriemi. Bulletin VÚRH Vodňany 43/3, s. 99-106
Oral presentations at international and national conferences: Pašková V., Hilscherová K., Banďouchová H., Pikula J., Bláha L. (2010). Evidence of synergistic toxicity in birds: experimental co-exposures to
cyanobacteria and toxic metal (lead) cause immunosuppressions in Japanese quail. 8th International Conference on Toxic Cyanobacteria, Istanbul, Turkey
Pašková V., Hilscherová K.(2007) Embryotoxicita a indukce oxidativního stresu po expozici herbicide paraquati na modelovém necílovém
organismu drápatce vodní (Xenopus laevis). In Toxikologická konference VÚRH Vodňany.Czech Republic. Pašková V., Hilscherová K., Bláha L. (2007) Role of oxidative stress in embryotoxicity and teratogenesis in FETAX test. In SETAC Europe 17th
Annual Meeting, Porto, Portugal – poster spotlight presentation. Pašková V., Hilscherová K., Babica P., Pikula J., Skočovská B. (2006) Efekty expozice sinicové biomasy u křepelky japonské (Coturnix coturnix
japonica). In Študentská vedecká konferencia, Bratislava, Slovak Republic. Poster presentations at international conferences: Pašková V., Hilscherová K., Banďouchová Damková V., Pikula J. (2010). Toxicity of cyanobacterial biomass to birds - effects in testes including
detoxification parameters and histology. 14th International Conference of Harmful Algae, Herssonissos, Crete, Greek, pp. 254. Moosová Z., Pašková V., Hilscherová K., Bláha L. (2010). Effects of algal and cyanobacterial cultures and their fractions on the Xenopus laevis
development. 8th International Conference on Toxic Cyanobacteria, Istanbul, Turkey, pp. 177. Pašková V., Hilscherová K., Banďouchová H., Horáková J., Pikula J., Maršálek B., Bláha L. (2009). Mixture of natural and chemical compounds:
effects on detoxification and oxidative stress parameters in bird Coturnix coturnix japonica. In NORMAN Workshop - Mixtures and metabolites of chemicals of emerging concern.
Pašková V., Hilscherová K. (2009) The involvement of oxidative stress in teratogenity and embryotoxicity of bipyridyl-pesticides. In SETAC Europe
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