Biomarkers in experimental ecotoxicology

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

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



CHAPTER 5 – PAPER No.3 .........................................................................................................57

DETOXIFICATION AND OXIDATIVE STRESS RESPONSES IN JAPANESE QUAIL EXPOSED TO CYANOBACTERIAL BIOMASS



CHAPTER 6 – PAPER No.4 .........................................................................................................65

DETOXIFICATION AND OXIDATIVE STRESS RESPONSES IN JAPANESE QUAIL EXPOSED TO MULTIPLE STRESSORS



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)


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.).


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.-


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


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


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


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.


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


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.


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


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


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


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.


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


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


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


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).

PAPER No.4 - DETOXIFICATION IN BIRDS AFTER MULTIPLE-STRESSORS

EXPOSURE

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.


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

CHAPTER 6

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.


EXPOSURE

69

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

CHAPTER 6

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


EXPOSURE

71

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).

CHAPTER 6

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

CHAPTER 7

74

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.

PAPER No.5 - DETOXIFICATION IN AQUATIC ORGANISMS AFTER

PESTICIDES EXPOSURE

75

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

CHAPTER 7

76

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


PESTICIDES EXPOSURE

77

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

CHAPTER 7

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.


PESTICIDES EXPOSURE

79

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

CHAPTER 7

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


PESTICIDES EXPOSURE

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.

CHAPTER 7

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)).


PESTICIDES EXPOSURE

83

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

CHAPTER 7

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).



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



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



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.



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


2.5mg/LDQ


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

)

.

* *

* *



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



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

*

* * *

* *


Px

nmol

NA

DP

H/m

in/m

g pr

otei

n

control

24 hod48 hod

72 hod96 hod

10

20

30

40


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.



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


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


*

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


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].

CHAPTER 8

102

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.



103

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

CHAPTER 8

104

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



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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106

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.



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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108

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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112

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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116

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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118

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([email protected]).

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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———


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��

Benzo[h]quinoline T. aestivumS. albaP. vulgaris

—��

—

——

��

��—

��

��

��

—��

��

��

��

Phenanthridine T. aestivumS. albaP. vulgaris

—�—

�——

��—

��—

��—

��—

��——

Anthracene T. aestivum — — — — — — —S. albaP. vulgaris

—��

——

—��

—��

——

—��

——

Acridine T. aestivumS. albaP. vulgaris

��—

��—

��

��

—

��—

��

—

��——

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


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.


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

——

��

—��

——

��

—��

��

��1,10-Phenanthroline T. aestivum

S. albaP. vulgaris

———

———

—��

�

—��

��


———

———

——�

��—

��

��


��—

��

—��

��

��

��

Benzo[h]quinoline T. aestivumS. albaP. vulgaris

——

��

��—

��

��

��

��

��

Phenanthridine T. aestivumS. albaP. vulgaris

��

—

��

—

��

��

�

��

�Anthracene T. aestivum

S. alba——

��—

——

——

��—

P. vulgaris — �� — —Acridine T. aestivum

S. albaP. vulgaris

�——

—��

��

��——

��——

Fluorene T. aestivumS. albaP. vulgaris

��

——

��

��—

��

��——

��

—Carbazole T. aestivum

S. albaP. vulgaris

��

��

�—

��

—��

—

—��

—�—

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


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


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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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([email protected]).

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]).


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-


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).


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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44. Prieto AI, Jos A, Pichardo S, Moreno I, Camean AM. 2006.Differential oxidative stress responses to microcystins LR and RRin intraperitoneally exposed tilapia fish (Oreochromis sp.). AquatToxicol 77:314–321.

45. Cazenave J, Bistoni MDA, Pesce SF, Wunderlin DA. 2006. Dif-ferential detoxification and antioxidant response in diverse organsof Corydoras paleatus experimentally exposed to microcystin-RR. Aquat Toxicol 76:1–12.

46. Harada K-I, Tsuji K, Watanabe MF. 1996. Stability of microcys-tins from cyanobacteria—III. Effect of pH and temperature. Phy-cologia 35:83–88.

47. U.S. Environmental Protection Agency. 1989. Risk Assessment—Guidance for Superfund, Vol I—Human Health Manual (Part A),Interim final. EPA/540/1-89/002. Office of Emergency and Re-medial Response, Washington, DC.

48. Ernst B, Dietz L, Hoeger SJ, Dietrich DR. 2005. Recovery ofMC-LR in fish liver tissue. Environ Toxicol 20:449–458.

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

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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: [email protected] (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


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].


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].


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].


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].


(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].


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.


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.


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: [email protected] (H. Klara).

Ecotoxicology and Environmental Safety 74 (2011) 2082–2090


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


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

P. Veronika et al. / Ecotoxicology and Environmental Safety 74 (2011) 2082–20902084


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].



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


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


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.



(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.



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



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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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: [email protected]

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

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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

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roid

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sfen

vale

rate

O.l

atip

es

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sts

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ha

Del

eter

ious

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oduc

tive

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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)

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ntet

al.(

2006

b)

Del

tam

ethr

inC

ypri

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carp

io

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ykis

sB

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eter

ious

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cts,

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ease

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ent

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libri

umen

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mor

talit

y,re

duce

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ngra

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mbr

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fect

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004)

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land

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005)

,

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gean

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agel

(199

0),a

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al.(

2010

)C

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met

hrin

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arpi

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.rer

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mbr

yole

thal

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educ

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yole

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euro

beha

vior

alef

fect

s–

spas

ticm

ovem

ents

,per

icar

dial

edem

a

Ayd

inet

al.(

2005

)D

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al.(

2010

)

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enth

rin

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erio

Em

bryo

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ality

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reof

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body

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urob

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iora

leff

ects

–sp

astic

mov

emen

tsD

eMic

coet

al.(

2010

)

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othr

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.rer

ioE

mbr

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thal

ity,n

euro

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vior

alef

fect

s–

spas

ticm

ovem

ents

,per

icar

dial

edem

aD

eMic

coet

al.(

2010

)


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

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coet

al.(

2010

)Pe

rmet

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atip

es

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erio

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ayed

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ility

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ing

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ulus

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ents

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alde

fect

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dtr

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ente

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gem

ento

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rof

larv

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mbr

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thal

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003)

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.(20

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rias

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Red

uced

grow

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val,

and

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eter

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med

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dan

dpe

rica

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lede

ma,

larv

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ntbo

dy,a

ndsw

olle

nyo

lksa

cD

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ases

ingr

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eigh

tof

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ae,i

ncre

ased

prot

ein

synt

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sM

oder

ate

larv

alle

thal

ity

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ket

al.(

2005

)L

ien

etal

.(19

97)

McC

arth

yan

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iman

(200

8)

Rei

det

al.(

1995

)

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zino

nC

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pio

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atip

es

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erio

Del

eter

ious

repr

oduc

tive

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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

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ality

,dec

reas

edhe

artr

ate,

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sac

and

hear

tsa

ced

ema,

spin

ede

form

atio

ns,a

ltere

dha

tchi

ngda

te

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inan

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cü(2

005)

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man

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into

n(2

000)

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008)

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crea

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sed

jerk

ym

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ents

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arte

dem

a,sp

inal

defo

rmity

Kie

nle

etal

.(20

09)


Tabl

e1

(con

tinue

d)

Pest

icid

eO

rgan

ism

Toxi

cef

fect

Ref

eren

ces

Org

anoc

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ines

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osul

fan

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esD

.rer

ioD

elay

edha

tchi

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mal

ler

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alte

red

mob

ility

Mild

trun

kcu

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ure,

abno

rmal

beha

vior

,ede

ma,

mic

roce

phal

y,an

dim

pair

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ovem

ent,

alon

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ithin

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sed

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te

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mle

yan

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athe

r(2

003)

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eyan

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rone

(200

1)

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ykis

s

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erio

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skel

etal

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cts,

skin

opac

ity,e

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thal

mia

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ksw

imm

ing,

depi

gmen

tatio

n,be

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oral

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ges

Hep

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ytic

alte

ratio

ns(g

lyco

geni

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plet

ion,

RE

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ary

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cum

ulat

ion)

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ance

dla

rvae

mor

talit

y,de

crea

sed

grow

th

Oliv

aet

al.(

2008

)

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ieet

al.(

1996

)

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gean

dN

agel

(199

0)T

hiac

lopr

idD

.rer

ioH

eart

rate

affe

ctio

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and

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ler

(200

8)

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anos

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cyan

ate

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erio

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ocho

rddi

stor

tions

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onet

al.(

2006

)Te

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ryn

(atr

iazi

ne)

+tr

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lfur

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aura

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sof

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ebra

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umn,

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tes

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ing

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lar

shap

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tes,

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incl

usio

ns,n

ucle

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knos

is

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feet

al.(

2004

b)

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nylp

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ole

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onil

D.r

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Not

ocho

rdde

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ratio

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orte

ning

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l–ca

udal

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axis

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ffec

tive

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flips

Steh

ret

al.(

2006

)

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zine

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ine

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ella

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B.r

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Alte

red

grow

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grow

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tin

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ae,i

ncre

ase

inra

tes

ofpr

otei

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grad

atio

nE

nhan

ced

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orta

lity,

incr

ease

dnu

mbe

rof

defo

rmat

ions

and

edem

a

delC

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enA

lvar

ezan

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iman

(200

5),M

cCar

thy

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an(2

008)

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(199

0)

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azin

eS.

aura

taR

educ

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rvae

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ival

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atic

lesi

ons,

loss

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llula

rsh

ape

inhe

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idin

clus

ions

,foc

alne

cros

is,

abun

dant

nucl

ear

pykn

osis

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feet

al.(

2004

a)


Tabl

e1

(con

tinue

d)

Pest

icid

eO

rgan

ism

Toxi

cef

fect

Ref

eren

ces

Cya

nazi

neM

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iati

lis

Mod

erat

em

orta

lity

with

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easi

ngtr

end

from

the

day

ofha

tchi

ngR

eid

etal

.(19

95)

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ins

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phen

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ate

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lity,

dela

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hing

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tion,

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ain

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th

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a,b)

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retic

ulum


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.


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)


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)


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)


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)


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.


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)


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)


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


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).


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)


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

)


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


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


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


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)


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)


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

)


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


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


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 [email protected]

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


Faculty of Science, Masaryk University, Brno, Czech Republic


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




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




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

19th Annual Meeting. Goeteborg, Sweden, pp. 134-134. Pašková V., Adamovský O., Hilscherová K., Banďouchová H., Horáková J., Pikula J., Maršálek B. (2008) Effects of avian co-exposure to natural

and anthropogenic compounds on accumulation, detoxification and oxidative stress parameters. In SETAC Europe 18th Annual Meeting Warsaw, Poland, pp. 90-321.

Adamovský O., Hlávková J., Kopp R., Hilscherová K., Babica P., Palíková M., Pašková V., Navrátil S., Bláha L. (2008) Kinetics of biomarkers,

bioaccumulation and elimination of peptide toxins microcystins in different freshwater fish species. In SETAC Europe 18th Annual Meeting Warsaw, Poland, pp. 191-422.

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. pp. 279-297. Pašková V., Hilscherová K., Babica P., Skočovská B., Horáková J., Pikula J., Maršálek B. (2006) Toxic effects of cyanobacterial biomass in

Japanese quail. In SETAC Europe 16th Annual Meeting, Abstract Book. The Hague, Belgium. pp. 258-625.

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