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AVERTISSEMENT
Ce document est le fruit d'un long travail approuvé par le jury de soutenance et mis à disposition de l'ensemble de la communauté universitaire élargie. Il est soumis à la propriété intellectuelle de l'auteur. Ceci implique une obligation de citation et de référencement lors de l’utilisation de ce document. D'autre part, toute contrefaçon, plagiat, reproduction illicite encourt une poursuite pénale. Contact : [email protected]
LIENS
Code de la Propriété Intellectuelle. articles L 122. 4 Code de la Propriété Intellectuelle. articles L 335.2- L 335.10 http://www.cfcopies.com/V2/leg/leg_droi.php http://www.culture.gouv.fr/culture/infos-pratiques/droits/protection.htm
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Laboratoire Interdisciplinaire des Environnements Continentaux LIEC – UMR 7360 CNRS, Metz
THESE
Pour l’obtention du titre de:
DOCTEUR DE L’UNIVERSITE DE LORRAINE
Mention: Ecotoxicologie, Biodiversité, Ecosystèmes
Kahina MEHENNAOUI
Understanding the impact of engineered nanoparticles
Gammarus sp. as a valuable non-vertebrate model?
Soutenue publiquement Le 20 décembre 2017 devant la commission d’examen :
Laure Giamberini Université de Lorraine – UMR 7360 CNRS - LIEC
Arno Gutleb Luxembourg Institute of Science and Technology
Catherine Mouneyrac Université Catholique de l’Ouest – MMs - Angers
Mélanie Auffan CNRS – CEREGE – Aix en Provence
Erik Ropstad Norwegian University of Life Sciences
Elise David Université Reims Champagne-Ardenne – UMR-I 02 SEBIO
François Guérold Université de Lorraine – UMR 7360 CNRS - LIEC
Sebastien Cambier Luxembourg Institute of Science and Technology
Ecole Doctorale Sciences et Ingénierie Ressources Procèdes Produits Environnement,
RP2E, ED Nº 410, Nancy
Environmental Research and Innovation (ERIN)
Department,
Luxembourg Institute of Science and Technology,
(LIST), Luxembourg
Directeur de thèse
Directeur de thèse
Rapporteur
Rapporteur
Examinateur
Examinateur
Examinateur
Examinateur
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To my parents
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ABSTRACT The potential toxicity of nanomaterials is of high societal and scientific interest due to the promise of ground-breaking innovations for many technical applications. However, toxicity can often not be related to the actual size, mass or surface area of the single nanoparticles (NPs) or the NP agglomerates. Therefore, it can be proposed that the toxicity is greatly influenced by other inherent and non-understood properties of the particles to which ions dissolving from the particle, surface or molecules adhering to the surface interfering with the uptake of NPs into cells, may have important contributions. The PhD project “NANOGAM”, closely linked up to CORE2012 NANION project that aims to obtain knowledge to understand some of the processes and factors involved in NP uptake and toxicity as such understanding is a prerequisite for the development of nanomaterials following the safer-by-design philosophy. This PhD project aims to investigate, based on known characteristics of the key physico-chemical parameters; as size and surface functionalities, of a well-chosen list of silver and gold NPs, the uptake, and dependent biological effects of different complexity (mortality, behavioural effects, physiological effects, transcriptomic effects, etc.), on a sensitive species; Gammarus fossarum (Crustacea Amphipoda), in order to understand to which extent toxicity of nanomaterials is due to intrinsic material properties or ion leaching. Such understanding will contribute to the prediction of toxicity based on material properties rather than repetitive testing of an indefinite number of new nanomaterials. G. fossarum were exposed at low concentrations of AgNPs and AuNPs for 72h or 15 days in presence or absence of food. The obtained results showed that (i) surface coating is the main factor governing AgNPs and AuNPs uptake by G. fossarum, (ii) both released ions and NPs themselves play a role in the potency of the studied AgNPs and AuNPs and (iii) chemical composition led to different effects at the sub-individual levels (target genes expression) and different tissue distribution as AgNPs were found in G. fossarum gills while AuNPs were found in the intestinal caeca. Additionally, this work shows that Gammarus sp. are valuable models for the study of the effects of AgNPs and AuNPs.
Keywords: Gammarus sp.; silver nanoparticles; gold nanoparticles; ions release; multi-biomarker approach;
transcriptomic
RESUME La toxicité potentielle des nanomatériaux présente un intérêt sociétal et scientifique élevé en raison de la promesse d'innovations pour de nombreuses applications techniques. Cependant, elle n’est pas forcément liée à la taille réelle, à la masse, à la surface des nanoparticules (NP) ou à leurs agglomérats. La toxicité des NPs pourrait être fortement influencée par d'autres propriétés inhérentes et encore incomprises telles que le relargage d’ions, de la particule elle-même, sa surface, ou des molécules adhérentes à la surface, qui interfèreraient avec l'absorption cellulaires des NPs. Le projet « NANOGAM» étroitement lié au projet « FNR CORE2012 NANION », vise à définir certains processus et facteurs impliqués dans l'absorption des NPs et leur toxicité. Une telle compréhension est une condition préalable au développement des nanomatériaux, fondement de la philosophie « safer-by-design ». Les objectifs de ce projet de thèse sont multiples. En tenant compte des caractéristiques des principaux paramètres physico-chimiques tels que la taille et l’aspect de la surface, l’étude a porté sur l'absorption de NPs d'argent et d'or, et leurs effets biologiques via une approche multi-biomarqueurs (mortalité, effets comportementaux, effets physiologiques, effets transcriptomiques, etc.) sur une espèce sensible, Gammarus
fossarum (Crustacea Amphipoda). Le but de cette investigation est de comprendre si la toxicité des nanomatériaux est inhérente aux propriétés intrinsèques des NPs ou plutôt aux ions relargués, ce qui contribuera à la prédiction de la toxicité des NPs en rapport avec leurs propriétés physico-chimiques et ce afin de limiter le nombre d’essais répétitifs sur de nouveaux nanomatériaux. G. fossarum ont été exposés à de faibles concentrations d'AgNPs et AuNPs pendant 72h à jeun et 15 jours nourris. Les résultats obtenus ont montré que (i) la nature de l’enrobage de surface est le principal facteur responsable de l'absorption d'AgNPs et d'AuNPs par G. fossarum ; (ii) les ions libérés et les NPs elles-mêmes jouent un rôle dans la toxicité des AgNPs et AuNPs étudiées ; (iii) la composition chimique des NPs a conduit à des effets différents aux niveaux sub-individuels (transcriptomique), ainsi qu’à une distribution différente dans les tissues selon la nature métallique de la NP. Les AgNPs ont été localisées dans les branchies de G. fossarum tandis que les AuNPs ont été observées dans les caeca intestinaux. Cette étude a également révélé que Gammarus sp. est un excellent modèle pour l'étude de la toxicité et des effets des AgNPs et des AuNPs.
Mots-clé : Gammarus sp. Nanoparticules d’argent, nanoparticules d’or, relargage d’ions, approche multi-biomarquers, transcriptomiques
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AVANT PROPOS/ FOREWORD
The research performed within the project” NanoGAM” have been funded by the National Fund for Research
(FNR Luxembourg- AFR-PhD-9229040). This project was closely linked up to CORE2012 NANION and FP7
FUTURENANONEEDS (FNN). These two projects allowed the funding of ICP-MS and NanoSIMS analyses.
Additionally, NanoGAM project were closely linked to PhD projects at Université de Lorraine (ANR P2N
MESONNET and ANR nanoSALT) and PhD programs at Oslo University (NanoZebra).
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ACKNOWLEDGEMENTS
These last four years were such a rewarding and meaningful experience. I would never have succeeded in
this work without the support of colleagues, relatives and friends.
First of all, I would like to thank Arno Gutleb and Pr. Laure Giamberini. Thank you for giving me the
opportunity to make this PhD and research possible. Your advices, ideas and inputs helped me throughout
this process. You also taught me how to face the difficulties encountered during this work. Thank you, Laure,
for your trust and for accepting to be my supervisor. Thank you for your availability, your precious advices
and the good moments shared during the different conferences over Europe. Thank you, Arno, for giving me
the opportunity to make a “PRE-doc” in the TOX group that resulted in this present PhD project. Thanks a lot
for your availability, trust and optimism. Thank you also for keeping my ideas structured and realistic.
Working with you was such an amazing experience.
I also would like to thank Pr. François Guérold and Sebastien Cambier for accepting to be my co-supervisors.
Thank you, François, for your advices and our discussions about the Gammarus. Sebastien, thank you for
your availability, for teaching me molecular biology, for your help in the lab even on Sundays, for criticizing
my work, and helping me finding solutions. Thanks also for all the discussions we had during these last 4
years.
I would like to address my sincere acknowledgments to the opponents who accepted to evaluate my present
work: Catherine Mouneyrac, Mélanie Auffan, Elise David and Erik Ropstad. I am also deeply grateful to the
Luxembourg Institute of Science and Technology, Lucien Hoffmann, the Fond National de la Recherche
(Luxembourg) and the Laboratoire Interdisciplinaire des Environnements Continentaux at Lorraine University
for making this research possible.
This work would have been impossible without the help of colleagues. Tommaso, thank you for correcting
my works, for your help and support, especially during these last very stressful months. I learned a lot with
you. I am also grateful to Natasa, who supervised my work during the internship and who kept helping me
from NIVA Oslo where she is now. Working with you was a great experience. I will never be thankful enough
to Aline, Boris and Sylvain. Thank you all for the support. Sylvain thank you for your precious help and
advices, for all the discussions we had on the way back to Metz. Thanks also to Aurélie for her support during
difficult moments. Thanks, the both of you for your presence and support. Aline you’ve been a precious
colleague and friend. I’ll be always grateful to you. You’ve always been here and it was so nice to work with
you. Thank you very much for your support. I am also grateful to Boris for his good spirit, optimism, and
enthusiasm. You’ve always made the days funnier in the lab… there’re so many things to say that I do not
even know from where to start… Thank you for being here. I have been very lucky to share the office with
you girls; Anouk, Joanna and Blandine. I wish you all the best for your future careers. I also thank all my PhD
fellows: Marc, Marie B, Sebastien L, Rodolphe and Benoit, I wish you the best and I’m sure you will succeed.
I am also indebted to all the LISRA group and the nice moments shared at SETAC. Thank you, Enrico, for your
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availability. I am also grateful to all the ERIN and MRT colleagues: Gea, Xavier, Magda, Christelle, Elisa,
Vincent R, and Brahim.
This work would not be possible without the help of all the ERIN support. Thank you Servane, Marie F,
Delphine and Lionel for helping me in the lab and in the field for Gammarus sampling. Servane, it was so
nice to work with you. Thank you also Johanna, Sebastien P, Audrey L, Audrey J, Aude, Laurent, Francois,
Marine, Celine, Cecile, Cyril, Jeff, Cedric and Jenny. Thank you, Alain, for all your help in the beginning of my
PhD, for helping me finding the nice “spot” for the Gammarus sampling. I am also greatly thankful to Anaïs
C for your support during difficult moments and availability and for helping me in the lab. Thank you also
Jean-Nicolas, Nathalie, Esther, Gaëlle, Jean-Sebastien and Patrick for your help and fruitful collaboration. I
learned a lot with you concerning new techniques. I also thank Maryline and Vincent for your trust and for
the collaboration.
I also would like to thank my colleagues from LIEC; Maël Garaud, Jennifer Andreï, Vanessa Koehle-Divo and
Alice Gossiaux for your help during the lab work. Thank you, Carole, Sandrine, Vincent and Simon, for your
advices these last years.
I would like to dedicate this work to the late Pr. Stephen J Klaine who inspired me for many years. I am so
grateful to have known you.
I am also deeply grateful to my friends in France and Algeria. Thank you all for being here. I thank also Hanne
for all the great moments and very interesting discussions we had. I also thank Greg for your support, your
help your advices and kindness. Thank you for keeping me in the right behaviour to achieve my objectives.
Last but not least, I would like to thank my family; my parents, my brother and his wonderful wife; Mehdi
and Fairouz for your great support and all the good moments spent with you each time I came back home.
My parents, Smaïl and Fatima, thank you for allowing me to achieve this work. Thank you for giving me this
opportunity despite the very hard moments 7 years ago when I left home. I am deeply grateful to you, you
always supported me, helped me in the difficult moments. I learned so much with you and I hope that I make
you proud. I love you.
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THESIS ACHIEVEMENTS
Articles in the thesis frame
Gançalo Vale, Kahina Mehennaoui, Sébastien Cambier, Giovanni Libralato, Stéphane Jomini, Rute F. Domingos. Manufactured nanoparticles in the aquatic environment – biochemical responses on freshwater organisms: a critical overview. 2016. Aquatic toxicology 170, 162-174.
Kahina Mehennaoui, Anastasia Georgantzopoulou, Vincent Felten, Jennifer Andreï, Maël Garaud, Sébastien Cambier, Tommaso Serchi, Sandrine Pain-Devin, François Guérold, Jean-Nicolas Audinot, Laure Giamberini, Arno C. Gutleb. 2016. Gammarus fossarum (Crustacea, Amphipoda) as a model organism to study the effects of silver nanoparticles. Science of the Total Environment 566, 1649-1659.
Kahina Mehennaoui, Sébastien Cambier, Sylvain Legay, Tommaso Serchi, François Guérold, Laure Giamberini, Arno C. Gutleb. Identification of reference genes for RT-qPCR data normalization in Gammarus fossarum (Crustacea Amphipoda). Submitted to Scientific reports
Kahina Mehennaoui, Sébastien Cambier, Tommaso Serchi, François Guérold, Johanna Ziebel, Jean-Sebastien Thomann, Nathalie Valle, Laure Giambérini, Arno C. Gutleb. Influence of size and surface coating on silver and gold nanoparticles uptake and their molecular effects on Gammarus fossarum (Crustacea Amphipoda). Submitted to Science of the Total Environment
Kahina Mehennaoui, Sébastien Cambier, Tommaso Serchi, François Guérold, Laure Giambérini, Arno C. Gutleb. Sub-chronic effects of silver and gold nanoparticles on Gammarus fossarum (Crustacea Amphipoda): from molecular to behavioural responses. In prep.
Book chapter
Arno C. Gutleb, Sébastien Cambier, Teresa Fernandes, Anastasia Georganztopoulou, Thomas A.J. Kuhlbusch, Iseult Lynch, Ailbhe Macken, Kahina Mehennaoui, Ruth Moeller, Carmen Nickel, W. Peijnenburg, Tommaso Serchi. 2016. Chapter 4: Environmental Fate and Effects of Nanomaterials in aquatic Freshwater Environments. 96-114. Nanomaterials: A guide to fabrication and applications. Edited by Sivashankar
Krishnamoorthy. CRC Press, Taylor & Francis Group.
Articles in the nanotoxicology topic
Jennifer Andreï, Sandrine Pain-Devin, Vincent Felten, Simon Devin, Laure Gaimberini, Kahina Mehennaoui, Sébastien Cambier, Arno C. Gutleb, François Guérold. 2016. Silver nanoparticles impact the functional role of Gammarus roeseli (Crustacea Amphipoda). Environmental Pollution 208, part B, 608-618.
Articles out of the topic
Vincent Rogé, Anastasia Georgantzopoulou, Kahina Mehennaoui, Ioana Fechete, François Garin, Aziz Dinia, Arno C. Gutleb, Damien Lenoble. 2015. Tailoring the optical properties of ZnO nano-lazers and their effect on in vitro biocompatibility. RCS advances 5, 97635-97647.
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Oral presentations
Effets de différentes tailles de nanoaprticles d’argent sur le comportement et la physiologie de Gammarus sp. Mehennaoui K., Georgantzopoulou A., Felten V., Garaud M., Andreï J., Cambier S., Serchi T., Contal S., Balachandran Y.L., Pain-Devin S., Giamberini L., Gutleb A.C. Colloque ARET-SFTG, Paris France, 3-4.06.2014.
Influence of size and surface coating on silver and gold nanoparticles uptake by Gammarus fossarum. Mehennaoui K., Cambier S., Serchi T., Ziebel J., Chauvière A., Lentzen E., Valle N., Thomann J.S., Guérold F., Giamberini L., Gutleb A.C., 32nd international conference on environmental geochemistry and health, Brussels, Belgium. 4-8.07.2016.
Influence of size and surface coating on silver and gold nanoparticles uptake by Gammarus fossarum. Mehennaoui K., Cambier S., Serchi T., Ziebel J., Chauvière A., Lentzen E., Valle N., Thomann J.S., Guérold F., Giamberini L., Gutleb A.C., ES1205 final conference, Aveiro, Portugal. 7-8.02.2017.
Do size and surface coating of AgNPs and AuNPs influence their uptake and molecular effects on Gammarus
fossarum? Mehennaoui K., Cambier S., Serchi T., Ziebel J., Chauvière A., Lentzen E., Valle N., Thomann J.S., Guérold F., Giamberini L., Gutleb A.C., PRIMO19 international conference, Matsuyama, Ehime, Japan. 30.6-3.07.2017.
Posters
Effects of different sizes of silver nanoparticles on physiological and behavioural responses of Gammarus
fossarum. Mehennaoui K., Georgantzopoulou A., Felten V., Garaud M., Andreï J., Cambier S., Serchi T., Contal S., Balachandran Y.L., Pain-Devin S., Giamberini L., Gutleb A.C. Geel, Belgium. 4.12.2014.
Surface and Size-dependent effects of silver nanoparticles on Gammarus fossarum: Link between physiological and behavioural responses. Mehennaoui K., Georgantzopoulou A., Felten V., Garaud M., Andreï J., Cambier S., Serchi T., Contal S., Ziebel J., Guignard C., Balachandran Y.L., Pain-Devin S., Giamberini L., Gutleb A.C. Nanoposter Virtual conference. 04.2015.
Surface and Size-dependent effects of silver nanoparticles on Gammarus fossarum: Link between physiological and behavioural responses. Mehennaoui K., Georgantzopoulou A., Felten V., Garaud M., Andreï J., Cambier S., Serchi T., Contal S., Ziebel J., Guignard C., Balachandran Y.L., Pain-Devin S., Giamberini L., Gutleb A.C. 25th SETAC Europe Annual meeting, Barcelona Spain. 3-7.05.2015.
Effects of Ag nanoparticles on two aquatic invertebrates. S. Cambier, K. Mehennaoui, E.J. Keuzenkamp, A. Georgantzopoulou, T Serchi, L Giamberini, A.C. Gutleb. SOT Annual meeting, New-Orleans, Louisiana, United States. 13-17.03.2016.
Surface and size-dependent effects of silver nanoparticles on behavioural and physiological responses of Gammarus fossarum and uptake evaluation with NanoSIMS50. Mehennaoui K., Georgantzopoulou A., Felten V., Garaud M., Andreï J., Cambier S., Serchi T., Contal S., Balachandran Y.L., Pain-Devin S., Giamberini L., Gutleb A.C. SOT Annual meeting, New-Orleans, Louisiana, United States. 13-17.03.2016.
Influence of Size and Surface coating on silver nanoparticles uptake by Gammarus fossarum. K. Mehennaoui, S. Cambier, J. Ziebel, N. Valle, J.S. Thoamnn, F. Guérold, L. Giamberini, A.C. Gutleb. 26th SETAC Europe Annual meeting, Nantes, France. 22-26.05.2016.
Influence of Size and Surface coating on silver and gold uptake by Gammarus fossarum. K. Mehennaoui, S. Cambier, J. Ziebel, N. Valle, J.S. Thoamnn, F. Guérold, L. Giamberini, A.C. Gutleb. Beltox Annual meeting, Louvain-La-Neuve, Belgium, 12.2016.
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Viualisation of the assimilation of nanoaprticles in biological tissues by SIMS nano-analyses. E. Lentzen, J.N. Audinot, N. Valle, A. Georgantzopoulou, K. Mehennaoui, S. Cambier, A.C. Gutleb. 6th international NanoSIMS user meeting, Utrecht, The Netherlands. 26-27.09.2016.
Influence of Size and Surface coating on silver and gold uptake by Gammarus fossarum. K. Mehennaoui, S. Cambier, J. Ziebel, N. Valle, J.S. Thoamnn, F. Guérold, L. Giamberini, A.C. Gutleb. Nanosafety 2017, Saarbrucken, Germany, 10-13.10.2017.
Identification of reference genes for RT-qPCR data normalization in Gammarus fossarum (Crustacea Amphipoda). K. Mehennaoui, S. Legay, T. Serchi, F. Guérold, L. Giamberini, A.C. Gutleb, S. Cambier. Beltox Annual Meeting, Leuven, Belgium, 01.12.2017.
Lectures
Utilisation des biomarqueurs en écotoxicologie : effets de différentes tailles de nanoparticules d’argent sur le comportement et la physiologie de Gammarus sp. K. Mehennaoui. Mentouri University, Constantine, Algeria. 2.01.2015.
Utilisation des biomarqueurs en écotoxicologie : effets de différentes tailles de nanoparticules d’argent sur les organismes aquatiques. K. Mehennaoui. Mentouri University, Constantine, Algeria. 3.01.2016.
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TABLE DES MATIERES/ TABLE OF CONTENT
ABSTRACT .................................................................................................................................................... 4
RESUME ....................................................................................................................................................... 4
AVANT PROPOS/ FOREWORD....................................................................................................................... 6
ACKNOWLEDGEMENTS ................................................................................................................................ 7
THESIS ACHIEVEMENTS .............................................................................................................................. 10
ARTICLES IN THE THESIS FRAME ...................................................................................................................... 10
BOOK CHAPTER .......................................................................................................................................... 10
ARTICLES IN THE NANOTOXICOLOGY TOPIC ....................................................................................................... 10
ARTICLES OUT OF THE TOPIC .......................................................................................................................... 10
ORAL PRESENTATIONS ................................................................................................................................. 11
POSTERS ................................................................................................................................................... 11
KEYNOTE LECTURES ..................................................................................................................................... 12
LIST OF FIGURES......................................................................................................................................... 19
LIST OF TABLES .......................................................................................................................................... 23
ABBREVIATIONS ......................................................................................................................................... 25
GENERAL INTRODUCTION .......................................................................................................................... 27
CHAPTER 1 ................................................................................................................................................. 30
PART 1: STAT OF THE ART BASED ON THE PUBLISHED REVIEW: .................................................................. 30
ABSTRACT .................................................................................................................................................. 32
1. INTRODUCTION .................................................................................................................................. 32
2. NPS TRANSFORMATIONS IN AQUATIC SYSTEMS ................................................................................. 33
3. NANOTOXICITY TOWARD AQUATIC ORGANISMS ................................................................................ 36
3.1. GENERATION OF ROS ...................................................................................................................... 36
3.2. OMICS ENDPOINTS ......................................................................................................................... 38
4. NPS TOXICITY ON FRESHWATER ORGANISMS ..................................................................................... 40
4.1. SILVER NPS (AGNPS) ........................................................................................................................ 40
4.2. GOLD NANOPARTICLES (AUNPS) ..................................................................................................... 44
5. GENERAL REMARKS AND CONCLUSIONS ............................................................................................. 46
ACKNOWLEDGEMENTS .............................................................................................................................. 48
REFERENCE ................................................................................................................................................ 48
CHAPTER 1 ................................................................................................................................................. 57
PART 2: G. FOSSARUM AS A MODEL ORGANISM IN NANOTOXICOLOGY: .................................................... 57
1. SYSTEMATIC AND IDENTIFICATION OF GAMMARUS FOSSARUM ......................................................... 58
1.1. SYSTEMATIC POSITION .................................................................................................................... 58
1.2. IDENTIFICATION OF GAMMARUS FOSSARUM .................................................................................. 60
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2. MORPHOLOGY.................................................................................................................................... 62
2.1. PROSOMA ....................................................................................................................................... 63
2.2. MESOSOMA .................................................................................................................................... 63
2.3. METASOMA .................................................................................................................................... 64
2.4. UROSOMA ...................................................................................................................................... 64
3. ANATOMY OF GAMMARUS FOSSARUM .............................................................................................. 64
3.1. NERVOUS SYSTEM AND CIRCULATORY ORGANS .............................................................................. 64
3.2. INTESTINAL CAECA AND DIGESTION SYSTEM ................................................................................... 65
3.3. GAMMARUS GILLS AND RESPIRATION ............................................................................................. 66
3.4. REPRODUCTION ORGANS, LIFE CYCLE AND DEVELOPMENT ............................................................. 67
3.4.1. ANATOMY OF THE REPRODUCTION ORGANS ............................................................................... 67
3.4.2. REPRODUCTION AND LIFE CYCLE ................................................................................................. 67
3.4.3. DEVELOPMENT ............................................................................................................................ 68
4. GAMMARUS SP. ECOLOGY .................................................................................................................. 69
5. FORAGING PLASTICITY ........................................................................................................................ 69
6. GAMMARUS FOSSARUM AS A MODEL ORGANISM IN ECOTOXICOLOGY .............................................. 70
REFERENCES .............................................................................................................................................. 71
CHAPTER 2 ................................................................................................................................................. 77
GAMMARUS FOSSARUM (CRUSTACEA AMPHIPODA) AS A MODEL ORGANISM FOR NANO-ECOTOXICOLOGY
.................................................................................................................................................................. 77
AIM OF THE STUDY .................................................................................................................................... 78
EXPERIMENTAL DESIGN ............................................................................................................................. 78
KEY FINDINGS ............................................................................................................................................ 78
ARTICLE 1 ................................................................................................................................................... 79
COMPLEMENT 1......................................................................................................................................... 91
SUPPLEMENTARY MATERIAL ...................................................................................................................... 92
PARTICLE DISPERSION AND CHARACTERISATION IN VOLVIC® WATER ....................................................................... 92
ENERGY RESERVES, DETOXIFICATION AND LPO LEVEL MEASUREMENTS ................................................................... 92
CHAPTER 3 ................................................................................................................................................. 96
G. FOSSARUM MOLECULAR RESPONSES ..................................................................................................... 96
AIM OF THE STUDY .................................................................................................................................... 97
EXPERIMENTAL DESIGN ............................................................................................................................. 97
KEY FINDINGS ............................................................................................................................................ 97
ARTICLE 2 ................................................................................................................................................... 98
1. INTRODUCTION ................................................................................................................................ 100
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2. RESULTS ........................................................................................................................................... 101
2.1. STABILITY OF THE CANDIDATE REFERENCE GENES IN G. FOSSARUM .............................................. 101
2.2. OPTIMAL NUMBER OF REFERENCE GENE FOR DATA NORMALIZATION IN G. FOSSARUM USING
GENORM ................................................................................................................................................. 103
2.3. VALIDATION OF THE SELECTED REFERENCE GENES FOR G. FOSSARUM .......................................... 104
3. DISCUSSION ...................................................................................................................................... 105
4. MATERIALS AND METHODS .............................................................................................................. 107
4.1. ORGANISMS SAMPLING AND ACCLIMATION ................................................................................. 107
4.2. AGNO3, AGNPS AND AUNPS CONTAMINATION ............................................................................. 108
4.3. GENE IDENTIFICATION AND QPCR PRIMER DESIGN ....................................................................... 108
4.3.1. DNA EXTRACTION ...................................................................................................................... 108
4.3.2. LIBRARY PREPARATION AND SEQUENCING ................................................................................ 108
4.3.3. DE NOVO ASSEMBLY.................................................................................................................. 108
4.3.4. GENE IDENTIFICATION ............................................................................................................... 109
4.3.5. PRIMER DESIGN ......................................................................................................................... 110
4.4. RNA EXTRACTION, CDNA AND RT-QPCR ........................................................................................ 110
4.5. STABILITY OF THE CANDIDATES’ REFERENCE GENES ...................................................................... 110
4.6. STATISTICAL ANALYSES ................................................................................................................. 111
5. CONCLUSIONS .................................................................................................................................. 111
ABBREVIATIONS ....................................................................................................................................... 112
REFERENCES ............................................................................................................................................ 113
COMPLEMENT 2....................................................................................................................................... 117
CHAPTER 4 ............................................................................................................................................... 118
AIM OF THE STUDY .................................................................................................................................. 119
EXPERIMENTAL DESIGN ........................................................................................................................... 119
KEY FINDINGS .......................................................................................................................................... 119
ARTICLE 3 ................................................................................................................................................. 120
1 INTRODUCTION ................................................................................................................................ 123
2 MATERIALS AND METHODS .............................................................................................................. 124
2.1 PARTICLES AND CHEMICALS ............................................................................................................. 124
2.2 PARTICLE CHARACTERIZATION ........................................................................................................... 124
2.3 ORGANISM SAMPLING AND ACCLIMATION ........................................................................................... 125
2.4 ACUTE TOXICITY TEST ...................................................................................................................... 125
2.4.1 Total and dissolved silver and gold measurements ............................................................. 126
2.4.2 Silver and gold bioaccumulation ......................................................................................... 126
2.4.3 Nanoparticle uptake........................................................................................................... 127
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2.4.4 Molecular responses .......................................................................................................... 128
2.5 STATISTICAL ANALYSES .................................................................................................................... 132
3 RESULTS ........................................................................................................................................... 132
3.1 NANOPARTICLE CHARACTERIZATION ................................................................................................... 132
3.2 ACUTE TOXICITY TEST ...................................................................................................................... 133
3.2.1 Survival .............................................................................................................................. 133
3.2.2 AgNP exposure ................................................................................................................... 133
3.2.3 AuNP exposure ................................................................................................................... 136
3.2.4 Particle uptake ................................................................................................................... 137
3.2.5 Molecular responses .......................................................................................................... 140
4 DISCUSSION ...................................................................................................................................... 144
5 CONCLUSION .................................................................................................................................... 148
COMPLEMENT 3....................................................................................................................................... 157
SUPPLEMENTARY MATERIAL .................................................................................................................... 158
1- COATING OF AGNPS AND AUNPS BY WET CHEMISTRY ...................................................................... 158
2- PARTICLE CHARACTERIZATION .......................................................................................................... 159
3- PRELIMINARY TEST: DAPHNIA MAGNA STRAUS MOBILITY INHIBITION TEST (ISO 6341:1996) ............ 163
3.1- TEST ORGANISMS .............................................................................................................................. 163
3.2- BIOLOGICAL ASSAY ............................................................................................................................. 163
3.2.1- Reference test ......................................................................................................................... 163
3.2.2- Final test ................................................................................................................................. 164
3.2.3- Results .................................................................................................................................... 164
4- ACUTE TOXICITY OF AGNO3 ON GAMMARUS FOSSARUM .................................................................. 165
CHAPTER 5 ............................................................................................................................................... 167
AIM OF THE STUDY .................................................................................................................................. 168
EXPERIMENTAL DESIGN ........................................................................................................................... 168
KEY FINDINGS .......................................................................................................................................... 168
1. INTRODUCTION ................................................................................................................................ 171
2. MATERIALS AND METHODS .............................................................................................................. 172
2.1. PARTICLES AND CHEMICALS ............................................................................................................. 172
2.2. PARTICLE CHARACTERIZATION ........................................................................................................... 173
2.3. ALDER LEAVES CONDITIONING .......................................................................................................... 173
2.4. ORGANISMS SAMPLING AND ACCLIMATION.......................................................................................... 173
2.5. TROPHIC EXPOSURE ........................................................................................................................ 174
2.5.1. EXPERIMENTAL DESIGN ............................................................................................................... 174
2.5.2. TOTAL AND DISSOLVED SILVER AND GOLD MEASUREMENTS .................................................................. 174
2.5.3. SILVER AND GOLD BIOACCUMULATION ............................................................................................ 175
2.5.4. PARTICLES UPTAKE: CYTOVIVA ® ANALYSES ...................................................................................... 175
2.5.4.1. SAMPLE PREPARATION ............................................................................................................. 175
2.5.4.2. CYTOVIVA® DARK FIELD HYPERSPECTRAL IMAGING ......................................................................... 175
2.5.5. MOLECULAR EFFECTS .................................................................................................................. 176
18
2.5.5.1. RNA EXTRACTION AND CDNA SYNTHESIS .................................................................................... 176
2.5.5.2. PRIMER DESIGN AND QUANTITATIVE REAL-TIME PCR ...................................................................... 176
2.5.6. BEHAVIOURAL RESPONSES AND OSMOREGULATION ........................................................................... 177
2.5.6.1. LOCOMOTOR ACTIVITY ............................................................................................................. 177
2.5.6.2. VENTILATION ACTIVITY............................................................................................................. 177
2.5.6.3. OSMOREGULATION ................................................................................................................. 177
2.5.6.3.1. HAEMOLYMPH SAMPLING ..................................................................................................... 177
2.5.6.3.2. HAEMOLYMPH NA+, CL- AND CA2+ CONCENTRATIONS MEASUREMENTS ........................................... 178
2.6. STATISTICAL ANALYSES .................................................................................................................... 178
3. RESULTS ........................................................................................................................................... 178
3.1. PARTICLE CHARACTERIZATION ...................................................................................................... 178
3.2. SURVIVAL ...................................................................................................................................... 178
3.3. SILVER AND GOLD BIOACCUMULATION......................................................................................... 179
3.4. CYTOVIVA® DARK FIELD HYPERSPECTRAL IMAGING....................................................................... 181
3.5. MOLECULAR EFFECTS .................................................................................................................... 182
3.6. BEHAVIOURAL RESPONSES AND OSMOREGULATION .................................................................... 182
3.6.1. LOCOMOTION AND VENTILATION ACTIVITY ............................................................................... 182
3.6.2. OSMOREGULATION ................................................................................................................... 185
4. DISCUSSION ...................................................................................................................................... 185
4.1. PARTICLE CHARACTERIZATION ...................................................................................................... 186
4.2. BIOLOGICAL ENDPOINTS ............................................................................................................... 186
5. CONCLUSIONS .................................................................................................................................. 190
CHAPTER 6 ............................................................................................................................................... 198
1. NPS SIZE, SURFACE COATING AND CHEMICAL COMPOSITION INFLUENCE THE UPTAKE ..................... 203
2. SYNTHESIS METHODS AND CHEMICAL COMPOSITION INFLUENCE AGNPS AND AUNPS EFFECTS ....... 205
3. CONTRIBUTION OF SOLUBLE IONS TO THE EFFECTS .......................................................................... 206
4. LIMITS OF THE MULTI-BIOMARKERS APPROACH ............................................................................... 207
5. CONCLUDING REMARKS AND FUTURE PERSPECTIVES ....................................................................... 208
REFERENCES ............................................................................................................................................ 209
REFERENCE .............................................................................................................................................. 213
19
LIST OF FIGURES
Chapter 1 Part 1
Figure P1. 1. Representative chemical and physical transformations of NPs when entering in natural aquatic systems:
dissolution, phosphatization, sulfidation, homo- and hetero-aggregation, and sedimentation. Important constituents
with which NPs can interact governing their fate and transport includes hardness cations (e.g., Ca2+, Mg2+), alkalinity,
phosphate and sulfide anions, pH, dissolved organic carbon (DOC), organic matter (OM) and mineral surfaces (such as
iron and manganese oxides, and clays). Legend: blue circles: Engineered NPs; yellow circles: humic substances (HM);
brown circles: natural inorganic colloids; blue lines: rigid biopolymers; gray surroundings: representing sulfidation; Mz+:
free metal ion. Adapted from(Domingos et al., 2015b) ............................................................................................... 34
Figure P1. 2. Antioxidant defense system in an animal cell (Cossu et al., 1997; Sroda, 2011; Garaud, 2015). CAT:
catalase; G6PD: glucose-6-phospho dehydrogenase; SeGPx: selenium dependent glutathione peroxidase, GPx:
glutathione peroxidase; GR: glutathione reductase; GSH/GSSG: reduced/oxidised glutathione; NAPD+/NAPDH:
oxidised/reduced nicotinamide dinucleotide phosphate; NOS: nitric oxide synthase; SOD: superoxide dismutase......... 37
Figure P1. 3. Potential routes for the generation of ROS due to the presence of NPs. 1) Internalization of NPs ROS
generation could occur due to the NPs dissolution inside the cells and/or due to the NPs photocatalytic activity. 2)
Dissolution of the NPs leads to an increase concentration of metal ions in the media; some of these metals can also be
uptake by the organisms. 3) NPs and/or their surrounding coatings can adsorb/complex other metals present in the
media, being taken up by the cells. 4) Photocatalytic activity of the NPs in the presence of UV and/or natural light. ... 38
Chapter 1 Part 2
Figure P2. 1. (A) Tree of life of the order Amphipoda (Yellow arrow) and its constituents’ suborders Gammaridea,
Senticaudata and Hyperiidea. (B) Tree of life of Gammarus fossarum (Yellow arrow) (Lifemap)............................... 58
Figure P2. 2. Geographical distribution of Gammarus fossarum in (A) Europe and (B) Luxembourg (Dohet et al., 2008;
Fauna Europaea). ...................................................................................................................................................... 59
Figure P2. 3. Determination key [in French] of Gammarus fossarum (Felten, 2003) .................................................. 60
Figure P2. 4. Distribution of Gammarus fossarum cryptic species in Europe (Adapted from Westram et al., 2011). Grey
shaded zone indicates the contact zone between Type A and B. ............................................................................... 62
Figure P2. 5. General morphology of Gammarus sp. Adapted from Felten 2003. ...................................................... 63
Figure P2. 6. Lateral view of the anatomy of Gammarus sp. Illustrating the principal organs (Schmitz, 1992). ......... 65
Figure P2. 7. Cross section of (A) Gammarus fossarum illustrating the main organs. (B) midgut and intestinal caeca
and (B) gills of G. fossarum observed with optical microscope at 60x magnification. MG: midgut, IC: intestinal caeca
(Pictures: Chauvière A. and Mehennaoui K.) ............................................................................................................... 66
Figure P2. 8. Reproduction organs of a male Gammarus and a female (Trapp, 2015). cmu: mucus cells, GA: androgen
gland, ci: incubation chamber, mt: non-differentiated mesenchymal tissue, ov1: primer vitellogenese oocyte, ov2:
secondary vitellogenese oocyte, ovd: oviduct, ovd vst: vestigial oviduct, spc: spermatocyte, spg: spermatogonium, spz:
spermatozoid, vd: spermiduct .................................................................................................................................... 67
Figure P2. 9. Gammarus fossarum male and female forming a prepopulate pair (Picture: Untereiner B. and Mehennaoui
K.) ............................................................................................................................................................................. 68
20
Chapter 2 Article 1
Figure 1. Experimental design (adapted from(Arce Funck et al., 2013). CAT: Catalase; GPx: Glutathione
Peroxidase; TAC: Total Antioxidant Activity; ACP: Acid Phosphatase; GST: Glutathione S-Transferase; CHOL:
Cholesterol; TRIG: Triglyceride: PROT: Proteins: ETS: Electron Transport System; LDH: Lactate dehydrogenase; LPO:
Lipid peroxidation; CASP-3: Caspase 3………………………………………………………………………………………………………………..82
Figure 2. Elemental distribution of 12C14N- cluster and 109Ag- ion in 300 nm cuts of gills. G. fossarum were exposed
to AgNPs (A) 23 nm, (B) 27nm, (C) AgNO3 and (D) Control (Volvic® water). Scale bar is 5 µm…………………………..85
Figure 3. G. f2 survival rates (Mean ± SD) after 72h of exposure to Ag, AgNPs 23 nm, AgNPs 27 nm, AgNPs 20
nm. Data were arcsin root square transformed. Letters (a-c) illustrate significant differences (One-way ANOVA +
Fisher LSD post hoc test at P<0.05 level of significance, n=15). ……………………………………………………………………………86
Figure 4: G .f2 haemolymph osmolality (Mean ± SD) after 72h exposure to AgNO3, AgNPs 20 nm, AgNPs 23 nm
and AgNPs 27 nm. Different letters illustrate significant differences between different treatments (One-way
ANOVA + LSD Fisher post hoc test at P < 0.05 level of significance, n=15)……………………………………………………………87
Figure 5: Behavioural responses of G.f2 after 72h of exposure to AgNO3, AgNPs 20 nm, AgNPs 23 nm, AgNPs 27
nm. (A) Ventilation (Mean pleopod beat frequency ± SD). No significant differences were detected for any of
exposure conditions (One-way ANOVA + Fischer LSD post hoc test, P < 0.05, n=10). (B) Locomotor activities
(Mean percentage of moving G.f2 ± SD). Letters illustrate significant differences. (One-way ANOVA + Fisher LSD
post hoc test at P < 0.05 level of significance, n=10)…………………………………………………………………………………………….88
Chapter 3 Article 2
Figure 3. 1. Global ranking of candidate reference genes in G. fossarum. A number (from 1 to 6) was assigned to each
stability coefficient. A mean rank was generated and error bars .............................................................................. 103
Figure 3. 2 Determination of the optimal number of reference gens for data normalization in G. fossarum exposed to
AgNO3, AgNPs 40 nm and AuNPs 40 nm. The pairwise variation (Vn/Vn+1) was calculated between normalization factors
NF/NFn+1. The recommended cut-off threshold of 0.15 was applied in this study....................................................... 104
Figure 3. 3 HSP90 expression analysis using different normalization strategies. Error bars indicate the standard errors
of the means (n=4). Different letter (a-c) indicate significant differences at P < 0.05. ................................................ 105
Chapter 4 Article 3
Figure 4. 1. G. fossarum Ag bioconcentration (mean ± SD) after 72h exposure to CIT and PEG-AgNPs 20, 40 and 80 nm.
................................................................................................................................................................................ 135
Figure 4. 2. G. fossarum au bioconcentration (mean ± SD) after 72h exposure to CIT and PEG AuNPs 20, 40 and 80 nm.
................................................................................................................................................................................ 137
Figure 4. 3. A) 300 nm cross section of G. fossarum observed under optical microscope. Red square represents the area
observed with NanoSIMS 50. Scale bar is 250 µm. B) Elemental distribution of 12C14n- clusters. C) 109Ag- ions in 300 nm
cross sections of G. fossarum gills. Animals were exposed to 10 µg. L-1 of CIT-AgNPs 40 nm and PEG-AgNPs 40 nm in
Volvic water. Scale bar is 2 µm. ................................................................................................................................ 137
Figure 4. 4. A) 300 nm cross section of G. fossarum observed under optical microscope. Red square represents the area
observed with NanoSIMS 50. Scale bar is 250 µm. B) Elemental distribution of 12C14N- clusters. C) 109Au- ions in 300 nm
cross sections of G. fossarum intestinal caeca. Animals were exposed for 72h to 10 µg. L-1 of CIT-AuNPs 40 nm and PEG-
AuNPs 40 nm in Volvic water. Scale bar is 2 µm ........................................................................................................ 138
Figure 4. 5. A) 300 nm cross section of G. fossarum observed under optical microscope. Red square represents the area
observed with Cytoviva darkfield hyperspectral microscope. Scale bar is 250 µm. B) G. fossarum gills viewed with
Cytoviva (60x oil immersion magnification). C) AgNPs accumulation in G. fossarum gills (red spots and white arrows).
21
Animals were exposed for 72h to 10 µg. L-1 of CIT-AgNPs 40 nm and PEG-AgNPs 40 nm in Volvic water. Scale bar is 6 µm
................................................................................................................................................................................ 139
Figure 4. 6. A) 300 nm cross section of G. fossarum observed under optical microscope. Red square represents the area
observed with Cytoviva darkfield hyperspectral microscope. Scale bar is 250 µm. B) G. fossarum intestinal caeca viewed
with Cytoviva (60x oil immersion magnification). C) AuNPs accumulation in G. fossarum intestinal caeca (green spots
and white arrows). Animals were exposed for 72h to 10 µg. L-1 of CIT-AgNPs 40nm and PEG-AgNPs 40 nm in Volvic water.
Scale bar is 6 µm ...................................................................................................................................................... 140
Chapter 5 Article 4
Figure 5. 1. Survival rates (mean ± SD) of G. fossarum exposed for 15 days to 0.5 and 5 µg. L-1 of CIT and PEG-AgNPs
and PEG-AuNPs 40 nm ............................................................................................................................................ 179
Figure 5. 2 A) 300 nm cross section of G. fossarum observed under optical microscope. Red square represents the area
observed with cytoviva darkfield hyperspectral microscope. Scale bar is 250 µm. B) G. fossarum gills or intestinal caeca
viewed with Cytoviva (60x oil immersion magnification). C) AgNPs accumulation in G. fossarum gills (red spots and
white arrows). Animals were exposed for 72h to 10 µg. L-1 of CIT-AgNPs 40nm and PEG-AgNPs 40 nm in Volvic water.
Scale bar is 6µm. ...................................................................................................................................................... 181
Figure 5. 3. A) 300 nm cross section of G. fossarum observed under optical microscope. Red square represents the area
observed with CytoViva darkfield hyperspectral microscope. Scale bar is 250 µm. B) G. fossarum intestinal caeca viewed
with CytoViva (60x oil immersion magnification). C) AuNPs accumulation in G. fossarum caeca (red spots and white
arrows). Animals were exposed for 72h to 10 µg. L-1 of CIT-AgNPs 40nm and PEG-AgNPs 40 nm in Volvic water. Scale bar
is 6µm. .................................................................................................................................................................... 182
Figure 5. 4. Behavioural responses of G. fossarum exposed for 15 days to CIT and PEG-AgNPs and AuNPs 40 nm. A)
Locomotor activity (mean percentage of moving G. fossarum ± SD) and b) Ventilation (mean pleopods beat frequency
± SD). Different letters (a-f) indicates significant differences (one-way ANOVA + Tukey HSD post hoc test at P < 0.05 level
of significance, n = 10). ............................................................................................................................................ 184
Chapter 6
Figure 6. 1. Physiological and behavioural effects of synthetic and biological AgNPs on G. fossarum exposed for 72h:
summary of results presented in Chapter 2. G. fossarum were exposed for 72h to synthetic AgNPs 20 and 200 nm and
biologically synthetized from plant leaf extract AgNPs 23 and 27 nm. Effects on antioxidant responses (GPx, TAC, CAT),
defense mechanism (GST, ACP), cellular damage (LDH, CASP3, LOOH), energy reserves (Prot, Chol, Trig, ETS),
osmoregulation and behaviour (Locomotion and ventilation) were assessed............................................................. 200
Figure 6. 2. Influence of size (20, 40 and 80 nm) and surface coating (CIT and PEG) of AgNPs and AuNPs on their uptake,
tissue distribution and molecular effects of G. fossarum exposed for 72h: summary of the results presented in Chapter
4. G. fossarum were exposed for 72h to up to 50 µg. L-1 of AgNPs and AuNPs in absence of food. Influence of size and
surface coating on bioaccumulation was assessed using ICP-MS, internal distribution of AgNPs and AuNPs were
evaluated using NanoSIMS50 and Cytoviva, and molecular effects were assessed using RT-qPCR. A set of stress-related
genes expression including genes implied in cytoskeleton trafficking (Actin, TUB, UB), exoskeleton cuticle (Chitinase),
antioxidant defence (CAT, MnSOD, CuZnSOD, GPx7), general stress (GST, HSP90, GAPDH), DNA damage and repair
(Gadd45, NfkB, c-jun, P53), lysosomes (Cathepsin L), osmoregulation and respiration (Na+K+ATPase, HEM) was used.
................................................................................................................................................................................ 201
Figure 6. 3. Sub-chronic toxicity of CIT and PEG-AgNPs and CIT and PEG-AuNPs 40 nm on molecular, physiological and
behavioural responses of G. fossarum: summary of the results presented in Chapter 5. G. fossarum were exposed for
15 days to 0.5 and 5 µg.L-1 of CIT and PEG-AgNPs and AuNPs 40 nm in presence of food. Bioaccumulation was assessed
using ICP-MS, internal distribution of AgNPs and AuNPs were evaluated using Cytoviva and molecular effects were
assessed using a set of stress-related genes expression including genes implied in cytoskeleton trafficking (Actin, TUB,
22
UB), exoskeleton cuticle (Chitinase), antioxidant defence (CAT, MnSOD, CuZnSOD, GPx7), general stress (GST, HSP90,
GAPDH), DNA damage and repair (Gadd45, NfkB, c-jun, P53), lysosomes (Cathepsin L), osmoregulation and respiration
(Na+K+ATPase, HEM). ............................................................................................................................................... 202
Supplementary materials
Figure S 1. Size distribution of Ag NPs 23 nm (A), 27 nm (B), 20 nm (C) and 200 nm (D) in Volvic® water expressed as
particle concentration x 106.mL-1. The red error bars indicate the ± SD of the mean of triplicate measurements......... 96
Figure S 2 PEG coating on metal Nanoparticles (here silver NPs). SEM shows clearly the conformal silver coating with
the PEG layer. NTA analysis shows a small size increasing for the 40 and 80 nm silver nanoparticles as a consequence of
their coating with the PEG layer. The sensitivity of NTA was not enough to resolve the size increase for 20 nm
nanoparticles. .......................................................................................................................................................... 158
Figure S 3. Size distribution of CIT-AgNPs 20 nm (A), CIT-AgNPs 40 nm (B) and CIT-AgNPs 80 nm (C) in Volvic Water
(T0h) and CIT-AgNPs 20 nm (D), CIT-AgNPs 40 nm (E) and CIT-AgNPs 80 nm (F) in Volvic water after 24h of incubation.
Size distribution is expressed as particle concentration E6.mL-1. The red errors bars indicate ±SD of the mean of triplicate
measurements ......................................................................................................................................................... 159
Figure S 4. Size distribution of PEG-AgNPs 20 nm (A), PEG-AgNPs 40 nm (B) and PEG-AgNPs 80 nm (C) in Volvic Water
(T0h) and PEG-AgNPs 20 nm (D), PEG-AgNPs 40 nm(E) and PEG-AgNPs 80 nm (F) in Volvic water after 24h of
incubation. Size distribution is expressed as particle concentration E6.mL-1. The red errors bars indicate ±SD of the mean
of triplicate measurements ...................................................................................................................................... 160
Figure S 5. Size distribution of CIT-AuNPs 20 nm (A), CIT-AuNPs 40 nm (B) and CIT-AuNPs 80 nm (C) in Volvic Water
(T0h) and CIT-AuNPs 20 nm (D), CIT-AuNPs 40 nm (E) and CIT-AuNPs 80 nm (F) in Volvic water after 24h of incubation.
Size distribution is expressed as particle concentration E6.mL-1. The red errors bars indicate ±SD of the mean of triplicate
measurements ......................................................................................................................................................... 161
Figure S 6. Size distribution of PEG-AuNPs 20 nm (A), PEG-AuNPs 40 nm (B) and PEG-AuNPs 80 nm (C) in Volvic Water
(T0h) and PEG-AuNPs 20 nm (D), PEG-AuNPs 40 nm(E) and PEG-AuNPs 80 nm (F) in Volvic water after 24h of
incubation. Size distribution is expressed as particle concentration E6.mL-1. The red errors bars indicate ±SD of the mean
of triplicate measurements ...................................................................................................................................... 162
Figure S 7. Effects AgNO3 on survival of Gammarus fossarum collected in A) June, B) September and C) November after
72h of exposure....................................................................................................................................................... 165
23
LIST OF TABLES
Chapter 1 Part 2
Table P2. 1. Classification of the species Gammarus fossarum .................................................................................. 59
Chapter 2 Article 1
Table 2.1. Particles size distribution (mode ± SD, 3 replicates) and ζ potential of AgNPs in Volvic® water………..84
Table 2.2: LC50 values with 95% confidence intervals for G.f1 and G.f2 exposed to AgNO3 and AgNPs20 nm,
AgNPs 23 nm, AgNPs 27nm and AgNPs 200 nm for 72h……………………………………………………………………………………84
Table 2.3: Total and dissolved Ag concentrations (mean ± SD) and dissolution rates of AgNO3 and AgNPs after
72h of exposure in Volvic water. Different letters illustrate significant differences between different treatments
(Two-way ANOVA + Tukey HSD post-hoc test at P < 0.05 level of significance, n=3)…………………………………………….85
Table 2.4: G.f2 Ag Bioconcentration (Mean± SD) after 72h exposure to AgNO3, AgNPs 20 nm, AgNPs 23 nm,
AgNPs 27 nm. Different letters illustrate significant differences between different treatments (Kruskal-Wallis
ANOVA + Mann-Whitney U test at P < 0.05 level of significance, n=3)…………………………………………………………………86
Table 2.5: Mean values (±SD) of biomarkers measured in Gf2 exposed to AgNO3, AgNPs 20nm, AgNPs 23nm and
AgNPs 27 nm for 72h. Biomarkers in italic were analysed using One-way ANOVA and Tukey post hoc test, whereas
the others were analysed using Kruskal-Wallis ANOVA and Mann-Whitney U test. No significant differences were
detected for any of the exposure conditions. a: mg. g fresh weight-1; b: µmol O2. g proteins-1. h-1.; c: µmol p-
nitrophenol. g protein-1.h-1; d: µmol CDNB.min-1.g-1 proteins; e: µmol NADPH.g protein-1.min-1; f: mmol Trolox
equivalent.g protein-1; g: mmol H2O2.g proteins-1.min-1; µmol NADH. g proteins-1. h-1; h: µmolpNA. g proteins-1.h-1;
j: nmol TBH. g proteins-1……………………………………………………………………………………………………………………………………..87
Chapter 3 Article 2
Table 3. 1.Ranking of candidate reference genes according to the five algorithms used ......................................... 102
Table 3. 2 Identification of Gammarus fossarum gene sequences ........................................................................... 108
Table 3. 3 List of primers of the candidate reference genes and target gene HSP90 ................................................ 109
Chapter 4 Article 3
Table 4. 1. Specific primer pairs used for RT-qPCR analyses on Gammarus fossarum exposed for 72h to AgNPs and
AuNPs (coated with CIT or PEG, sizes: 20, 40 and 80 nm, F: forward sequence, R: reverse sequence) ...................... 130
Table 4. 2. Size distribution of particles (mode ± SD, 3 replicates) and ζ potential of AgNPs and AuNPs in Volvic water
(exposure medium) at T0h and T24h. ...................................................................................................................... 133
Table 4. 3. Total ag concentrations (mean ± SD) and recovery rates (mean ± SD) of AgNO3, CIT-AgNPs and PEG-AgNPs
in Volvic water (exposure medium). ........................................................................................................................ 134
Table 4. 4. Relative gene expression of G. fossarum exposed for 72h to CIT-AgNPs and PEG-AgNPs 20, 40 and 80 nm
................................................................................................................................................................................ 142
Table 4. 5. Relative gene expression of G. fossarum exposed for 72h to CIT-AuNPs and PEG-AuNPs 20, 40 and 80 nm
................................................................................................................................................................................ 143
24
Chapter 5 Article 4
Table 5. 1. G. fossarum Ag and Au uptake (mean ± SD) after 15 days of exposure to CIT and PEG-AgNPs and AuNPs 40
nm. ......................................................................................................................................................................... 180
Table 5. 2. Relative gene expression of G. fossarum exposed for 15 days to CIT- and PEG-AgNPs 40 nm and CIT and
PEG-AuNPs 40 nm ................................................................................................................................................... 183
Table 5. 3. Haemolymph [Cl-], [Na+] and [Ca2+] of G. fossarum exposed for 15 days to CIT and PEG-AgNPs 40 nm and
CIT and PEG-AuNPs 40 nm. ...................................................................................................................................... 185
Supplementary materials
Table S 1. Physico-chemical parameters of Volvic® water ......................................................................................... 95
Table S 2 D. magna reference test ........................................................................................................................... 163
Table S 3. EC50 values obtained after 48h exposure of D. magna to AgNPs and AuNPs .......................................... 164
25
ABBREVIATIONS
ATP Adenosine triphosphate
cDNA Complementary DNA
CYP450 Cytochrome P450
DTT Dithioreitol
EF1 Elongation factor 1 alpha
EMBL European Molecular Biology Laboratory
EST Expressed Sequence Tag
FC Fold change
FNR National Research Funds – Luxembourg
MIQe Minimum information for publication of qPCR experiments
mRNA Messenger ribonucleic acid
PCR Polymerase Chain Reaction
qPCR Quantitative real-time polymerase chain reaction
RIN RNA integrity number
RNA Ribonucleic acid
RNA-seq RNA sequencing
SI Supporting information
TEM Transmission electron microscope
TF Transcription factor
UV Ultraviolet
AgNPs Silver nanoparticles
AuNPs Gold nanoparticles
AgNO3 Silver nitrate
Ag2S Silver sulphide
DMSO Dimethyl Sulfoxide
NTA Nanoparticles tracking analysis
DLS Dynamic light scattering
HIM-SIMS Helium Ion microscopy – secondary ion mass spectrometer
EDTA Ethylenediaminetetraacetic acid
HNO3 Nitric acid
TAC Total antioxidant capacity
HCl Chloride acid
CIT Citrate
PEG Polyethylene glycol
NanoSIMS 50 Nano secondary ion mass spectroscopy
ICP-MS Inductively coupled plasma – mass spectrometer
ACP Acid phosphatase
CAT Catalase
OS Oxidative stress
NPs Nanoparticles
ENPs Engineered nanoparticles
NMs Nanomaterials
HM Humic acid
DOC Dissolved organic carbon
TOC Total organic carbon
OM Organic matter
FPOM Fine particle organic matter
CASP Caspase
PEC Predicted environmental concentration
L(E)C 50 Lethal or effective concentration that impact 50% of the exposed population
ETS Electron transport system
GR Glutathione reductase
GSH Reduced glutathione
GST Glutathione S-transferase
NOEC Non-observable effects concentrations
H2O2 Hydrogen peroxide
HSP Heat shock protein
MDA Malondialdehyde
ROS Reactive oxygen species
SOD Superoxide dismutase
LPO Lipid peroxidation
GPx Glutathione peroxidase
26
27
GENERAL INTRODUCTION
In the recent decades, nanotechnology has emerged as a fast-growing sector impacting key economical fields
and providing new engineered nano-enabled products, constituted by nanoparticles (NPs), with novel and
unique functions that reach the market every day. NPs are defined as compound presenting at least one
dimension less than 100 nm (Klaine et al., 2008).
The use of nanomaterials is increasingly and continuously growing as they are used in different areas such as
electronics, medicine, environmental technology, etc. Silver nanoparticles (AgNPs) are among the most
promising group of NPs. AgNPs are used in different kind of daily-life products such as textiles, food
packaging, healthcare products, etc., mostly for their antibacterial properties. AuNPs are being investigated
for their unique optical properties and are used as contrast agents in electron microscopy, optical sensors,
catalysts, and for therapeutic uses. The increasing use of AgNPs and AuNPs lead to their inevitable release in
the environments and may reach the aquatic ecosystems where they may represent a threat for aquatic
organisms. The potential toxicity of nanomaterials is of high societal and scientific interest and can be related
to many properties such as the size, the mass, surface area and characteristics, aggregation and
agglomeration (Georgantzopoulou et al., 2013). Their toxicity could also be linked to less known or less
understood parameters like released ions and adhering molecules which could have an influence on their
toxicity by interfering with their uptake and fate on living organisms.
The PhD project “NANOGAM” is a FNR funded project closely linked up to the “CORE2012 NANION” and to
“FP7 FUTURENANONEEDS (FNN)” projects that aim to obtain knowledge to understand some of the
processes and factors involved in NPs uptake and toxicity. The specific objectives of NANION are:
• Leaching ions from NPs have their own spectrum of toxicity and effects can be clearly separated
between free ions in the exposure media and the NPs (Georgantzopoulou, 2015).
• Within NANION, a set of AgNPs and AuNPs comprising different size classes and coated with various
surfaces (citrate and polyethylene glycol) are used. Prior to any testing, these NPs were carefully
characterized (zeta-potential, agglomeration, aggregation, etc.) in the relevant exposure media. The
ion release from uncoated and coated NPs (Au does not release ions) were studied in dependency of
exposure media and related to toxicity endpoints ranging from molecular to organism level.
• Biomolecules attach easily and quickly to the surface of NPs. This so-called corona influences uptake,
kinetics, distribution within the organism and thereby finally impact endpoints such as biochemical
biomarkers but also reproduction, behaviour of the organism, and other relevant whole organism
endpoints. The uptake kinetics, distribution of the NPs within the organism and toxicity for a range
of relevant endpoints were studied and related to physico-chemical properties of the NPs.
28
Generally, NANOGAM aims at contributing to the understanding in how far surface properties of NPs
(physical parameters, ion leaching, biological molecules attached, etc.) affect and interfere with the toxicity
of NPs in a very relevant freshwater model organism, the crustacean Gammarus fossarum. G. fossarum was
selected as test species due to its large distribution in Europe (Barnard and Barnard, 1983), high abundance
(Felten et al., 2008; Kunz et al., 2010), clear sexual dimorphism, easiness of identification to the species level,
collection and handling, the high sensitivity to a large range of toxicants and their major functional role in
ecosystems. Indeed, as shredders, they play a key role in the litter breakdown process and thus in freshwater
food chain and nutrient cycling (Forrow and Maltby, 2000; Vellinger et al., 2012a). Consequently, Gammarus
sp. could provide valuable information about the potential effects of studied contaminants on some other
taxa in aquatic ecosystem communities (Vellinger et al., 2012b) To our knowledge, only few studies have
been done on the ecotoxicity of AgNPs and AuNPs in these species (Andreï et al., 2016; Baudrimont et al.,
2017; Bundschuh et al., 2011; Park et al., 2015).
The “NanoGAM” project aims at investigating the characteristics of the key physico-chemical parameters and
surface functionalities of a well-chosen list of AgNPs and AuNPs that control uptake, and dependent
biological effects of different complexity (using a battery of biomarkers) on G. fossarum. Starting from a set
of commercially available NPs that will form the baseline for the studies proposed within NANOGAM.
The following questions have been raised within the present project:
• Is G. fossarum a valuable model to study the effects of nanoparticles?
• How do particle size, surface coating, synthesis method and chemical composition contribute to the
toxicity and effects of AgNPs and AuNPs?
• To what extent does leaching ions play a role in the toxicity of AgNPs?
• What are the acute effects of AgNPs and AuNPs, with different sizes (20, 40 and 80 nm) and two
different coatings (CIT and PEG), on G. fossarum?
• What are the sub-chronic effects of AgNPs and AuNPs on molecular, physiological and behavioural
responses of G. fossarum?
These research questions are addressed in the following chapters:
• Chapter 1 is a critical overview on biochemical effects of AgNPs and AuNPs on aquatic organism (Part
1) with a focus on G. fossarum, the model organism used in the present PhD project (Part 2)
• In Chapter 2, the effects of a well-characterized and well-studied set of AgNPs are used in order to
evaluate their effects on G. fossarum through a multi-biomarker approach. This study allows also the
assessment of the contribution of leaching ions on the toxicity and effects of AgNPs. Furthermore,
the internal distribution of AgNPs in G. fossarum is discussed.
• As G. fossarum is to our best knowledge a non-sequenced species, it was necessary to identify target
stress-related genes for RT-qPCR experiments. For accurate analyses, it was necessary to identify a
set of reference genes for RT-qPCR normalization and results are presented in Chapter 3.
29
• Chapter 4 provides evidence of the influence of AgNPs and AuNPs size and surface-coating on their
uptake and internal tissue distribution in G. fossarum. Furthermore, the acute effects on the
molecular responses are investigated.
• Chapter 5 deals with the sub-chronic effects of AgNPs and AuNPs on G. fossarum. A multi-biomarker
approach including molecular, physiological and behavioural responses was applied to assess the
effects of AgNPs and AuNPs. Additionally, tissues distribution of NPs is evaluated
• Finally, Chapter 6 discuss the knowledge acquired in the present work and the relevance of the
current findings. Conclusions and future perspectives are also addressed.
Chapter 1 Part1: NPs in the aquatic environment
30
CHAPTER 1
Part 1: Stat of the art based on the published review:
Gançalo Vale, Kahina Mehennaoui, Sébastien Cambier, Giovanni Libralato, Stéphane Jomini, Rute F.
Domingos. Manufactured nanoparticles in the aquatic environment – biochemical responses on
freshwater organisms: a critical overview. 2016. Aquatic toxicology 170, 162-174.
The published review is presented in Annexe 1
CHAPTER 1
Part 1: Stat of the art based on the published review:
Gançalo Vale, Kahina Mehennaoui, Sébastien Cambier, Giovanni Libralato, Stéphane Jomini,
Rute F. Domingos. Manufactured nanoparticles in the aquatic environment – biochemical
responses on freshwater organisms: a critical overview. 2016. Aquatic Toxicology 170, 162-174.
The published review is presented in Annexe 1
Chapter 1 Part1: NPs in the aquatic environment
31
Manufactured nanoparticles in the aquatic environment –Biochemical responses on freshwater
organisms: A critical overview1
Gonçalo Valea,b*, Kahina Mehennaouic,e, Sebastien Cambierc, Giovanni Libralatod, Stéphane
Jominie, Rute F. Domingosa,f
a) Centro de Química Estrutural, Instituto Superior Técnico, Universidade de Lisboa, Torre Sul, Av. Rovisco Pais,
1049-001 Lisboa, Portugal. Gonçalo Vale: [email protected]
b) Department of Molecular Genetics, University of Texas Southwestern Medical center, Harry Dallas, TX 75390,
USA (present address)
c) Luxembourg Institute of Science and Technology, Environmental Research and Innovation (ERIN) Department,
Belvaux, Luxembourg. Kahina Mehennaoui: [email protected]; Sebastien Cambier:
d) Department of Environmental Sciences, Informatics and Statistics, University Ca’ Foscari Venice, Via Torino, 155,
30172, Mestre, Venice, Italy. Giovanni Libralato: [email protected]
e) Laboratoire Interdisciplinaire des Environements Continentaux (LIEC), Université de Lorraine, UMR 7360,
Campus Bridoux rue du Général Delestraint, 57070 Metz, France. Stéphane Jomini:
f) Institut de Physique du Globe de Paris, Sorbonne Paris Cité, UMR CNRS 7154, Université Paris Diderot, 75205
Paris Cedex 05, France. Rute F. Domingos: [email protected] (present address)
*Corresponding author
Email: [email protected]
1 Vale, G., Mehennaoui, K., Cambier, S., Libralato, G., Jomini, S., Domingos, R.F., 2016. Manufactured
nanoparticles in the aquatic environment-biochemical responses on freshwater organisms: A critical overview.
Aquat. Toxicol. 170, 162–174. doi:10.1016/j.aquatox.2015.11.019
Chapter 1 Part1: NPs in the aquatic environment
32
ABSTRACT
The enormous investments in nanotechnology have led to an exponential increase of new manufactured
nano-enabled materials whose impact in the aquatic systems is still largely unknown. Ecotoxicity and
nanosafety studies mostly resulted in contradictory results and generally failed to clearly identify biological
patterns that could be related specifically to nanotoxicity. Generation of reactive oxygen species (ROS) is one
of the most discussed nanotoxicity mechanism in literature. ROS can induce oxidative stress (OS), resulting
in cyto- and genotoxicity. The ROS overproduction can trigger the induction of anti-oxidant enzymes such as
catalase (CAT), superoxide dismutase (SOD) and glutathione peroxidases (GPx), which are used as biomarkers
of response. A critical overview of the biochemical responses induced by the presence of NPs on freshwater
organisms is performed with a strong interest on indicators of ROS and general stress. A special focus will be
given to the NPs transformations, including aggregation, and dissolution, in the exposure media and the
produced biochemical endpoints.
Keywords: nanoparticles, transformations, freshwater organisms, nanotoxicity, reactive oxygen species
(ROS), anti-oxidant enzymes
1. INTRODUCTION
Nanotechnology has emerged as a fast growing sector impacting key economical fields and providing new
engineered nano-enabled products, constituted by nanoparticles (NPs), with novel and unique functions that
reach the market every day (Bour et al., 2015). NPs are defined as materials with a size between 1 and 100
nm on at least one dimension, having unique physico-chemical properties differing from their bulk forms due
to their greater surface area to volume ratio (Hood, 2004; The European Commision, 2011). The size related-
properties results in larger reactivity and higher mobility, leading to numerous applications in medical
diagnostics, electronics, computers, cosmetics and environmental remediation (Rauscher et al., 2015). The
worldwide consumption of NPs is expected to grow from 225,060 metric tons in 2014 to nearly 584,984
metric tons in 2019 representing an annual growth rate of 21.1 % (BCC RESEARCH, 2017). Although
impressive, these numbers are in fact “expected” values obtained by estimation or modelling. The lack of
legislation for nanotechnologies gives the manufacturers no onus to reveal the real figures, thus, indeed,
these predicted values are most probably significantly higher. The absence of real numbers hinders the
prediction of the NPs amount that are actually being released into the environment (Piccinno et al., 2012).
Even though several studies have been performed with the goal of modelling NPs environmental
concentrations (Gottschalk et al., 2009, 2013; Yang et al., 2016), they should only be considered as guidelines,
since they derive from uncertain data about the NPs production (often obtained by surveys to producers)
and extrapolations used to scale up regional to worldwide amounts (Gottschalk et al., 2013; Keller et al.,
2010; Piccinno et al., 2012).
Chapter 1 Part1: NPs in the aquatic environment
33
When released in natural media NPs will be subjected to a dynamic physical and chemical environment that
consequently results in different and unknown endpoints far from their pristine or as released state.
Therefore, environments and living organisms are not facing pristine manufactured NPs but rather
transformed nano-enabled products, which is factually accepted but so far neglected. In fact, the large
majority of the physico-chemical and toxicity data obtained so far was focused on simple nanoscale particles
and not on relevant nano-enabled products. This includes not only the NP embedded in the manufactured
matrix but also the materials resulting from the interaction with biotic and abiotic (bio)molecules composing
the natural systems (Nasser and Lynch, 2016). To further complicate the interpretation of the NPs studies,
there are two distinct mechanisms that should be considered but are not easily differentiated:
• Chemical toxicity by the release of possible ions and/or formation of reactive oxygen species (ROS)
(Fu et al., 2014)
• Physical stress or stimuli caused by NPs size, shape and surface properties (Libralato, 2014; Vale et
al., 2014).
These materials are generally associated with cellular perturbations such as ROS generation, gene expression
and proteome profiles alterations. For these reasons, the NPs escalating production and applications has
raised concerns about their environmental and human safety, which have led to large investments in
nanosafety-related projects resulting in a considerable amount of data assessing their potential hazard
(Savolainen et al., 2013). However, the establishment of relationships between bioavailable NP-containing
species and the specific bioadverse or biocompatible endpoints is still lacking, mainly since the effects are
NP-dependent and also specie-dependent (Burić et al., 2015).
This work provides an overview of the latest studies on the impact of NPs onto aquatic environment with a
focus on freshwater ecosystems, considered by many as the ultimate sink of these particles, with a special
focus:
• NPs transformations and characterization in the different test media,
• Toxicological effects such as generation of ROS, genotoxicity, transcriptomic and proteomic changes.
This survey is focused on metallic NPs including silver nanoparticles (AgNPs) and gold nanoparticles
(AuNPs), mostly due to the great number of studies dedicated to these particles.
2. NPs TRANSFORMATIONS IN AQUATIC SYSTEMS
NPs can enter in an aquatic compartment from:
• Wastewater treatment plants effluents,
• Direct use (e.g., application of NPs-containing paintings on boats),
• Deposition from the air compartment.
Chapter 1 Part1: NPs in the aquatic environment
34
When entering aquatic compartment, NPs will be exposed to a highly dynamic physical and chemical
environment that leads to several transformations that will change their pristine or as released physico-
chemical properties (Figure P1. 1). These transformations, including dissolution, aggregation and
sedimentation, are dependent on both physico-chemical properties of the NPs and those of the environment
into which they were released (Auffan et al., 2012).
Colloidal particles, including organic and inorganic matter, are ubiquitous in the aquatic environment and can
be originated from both natural and anthropic sources. These colloids can strongly interact with NPs, thereby
determining their forms over space and time (dynamic speciation), and greatly affecting their bioavailability.
Thus, the NPs will have a specific speciation in each environmental compartment, and this speciation is always
dynamic with reaction rates that depend upon the chemical nature and physical sizes of the engineered and
natural colloids (Levard et al., 2012). Although it is clear that dynamic speciation must be considered in order
to make relevant predictions of NPs fate, toxicity and risk, was mostly neglected (Domingos et al., 2015b).
Figure P1. 1. Representative chemical and physical transformations of NPs when entering in natural aquatic
systems: dissolution, phosphatization, sulfidation, homo- and hetero-aggregation, and sedimentation.
Important constituents with which NPs can interact governing their fate and transport includes hardness cations (e.g., Ca2+, Mg2+), alkalinity, phosphate and sulfide anions, pH, dissolved organic carbon (DOC), organic matter (OM) and mineral surfaces (such as iron and manganese oxides, and clays). Legend: blue circles: Engineered NPs; yellow circles: humic substances (HM); brown circles: natural inorganic colloids; blue lines: rigid biopolymers; gray surroundings: representing sulfidation; Mz+: free metal ion. Adapted from(Domingos et al., 2015b)
Chapter 1 Part1: NPs in the aquatic environment
35
Dissolution, which is one of the main transformations of metallic NPs such as AgNPs, is mainly due to:
• The formation of partially soluble metal-oxide (Domingos et al., 2013; Heinlaan et al., 2008; Wang et
al., 2015),
• The oxidation of the particle constituents (Dale et al., 2013; Ma et al., 2013; Wang et al., 2013),
• The complexation of the metal particle constituent by complexants present in the environmental
compartment or even in the NPs embedded matrix (including the manufactured stabilizers)
(Domingos et al., 2013, 2015a).
The sulfidation of the metallic NPs can retard their oxidation and, thus, their dissolution (Dale et al., 2013;
Ma et al., 2013; Wang et al., 2013). This dissolution mechanism results in the release of toxic cations, such
that their persistence is reduced but the toxicity is increased. Evidently, complete dissolution of the NPs
allows the prediction of their impact using already existing models for metal speciation and toxicity.
Aggregation is another critical transformation, which, mainly is occurring by interaction with naturally bio- or
geo-macromolecules, affect NPs size and surface chemistry. For example, organic matter (OM) provides both
charge and steric stabilization (Domingos et al., 2009) of the NPs, although they may also result in bridging
flocculation in presence of multiple charged cations and anions (Domingos et al., 2010). OM effects are
complex and difficult to predict; however, it is of extremely importance to explore these interactions since
OM concentrations are typically orders of magnitude higher in concentration than engineered NPs, and, thus,
likely to substantially modify their properties and behaviours.
Dissolution and aggregation are dynamic processes that can decrease the NPs available surface area, thereby
decreasing their reactivity. However, this decrease is dependent on the surface properties, particle number,
size distribution, and the fractal dimensions of the aggregate (Hotze et al., 2010). The NP size will affect its
bioavailability to the organisms; when aggregates become too large for direct transport across the cell wall
and/or membrane, uptake may be prevented, whereas partial dissolution, which will lead to smaller sizes,
would facilitate this cellular transport. Since these transformations are most often not in equilibrium, they
require real-time kinetic measurements, limiting the methodology to be used:
• The storage of whole unfractionated samples for ion analysis may not be possible since the
dissolution rate may be fast or not attaining the equilibrium during the experimental time,
• The aggregation rate can be fast or the aggregates size distribution may not reach equilibrium within
the experimental time window.
Despite the large number of studies focused on nanotoxicology, most of them disregard the particles kinetic
physicochemical characterization under the exposure conditions, hindering the establishment of crucial
predictive structure-activity relationships that can be used afterwards in the categorization and function for
Chapter 1 Part1: NPs in the aquatic environment
36
risk assessment studies. In absence of these realistic studies, two less constrain conditions were used to select
the literature studies that will be discussed in this critical overview:
• Studies where the NPs characterization was performed in the same media as the bioassays;
• Quantification of the NPs dissolution for studies using metallic NPs with propensity for dissolution
such as AgNPs.
3. NANOTOXICITY TOWARD AQUATIC ORGANISMS
The increasing use of engineered NPs in consumers’ products such as cosmetics, paints, food, computers,
medicine, etc. lead to their increasing release in aquatic environments and raise concerns about their
potential toxicity of aquatic organisms (Klaine et al., 2008; Manzo et al., 2013; Nowack et al., 2012).
3.1. Generation of ROS
Despite the large number of studies on NPs toxicity both in cell line systems and organisms, a complete
understanding about the mechanisms behind is still lacking (Bour et al., 2015; Fu et al., 2014; Manke et al.,
2013; Schultz et al., 2014). ROS generation, whose overproduction can lead to oxidative stress (OS) in the
organism tissues, is unquestionably the most studied nanotoxicity mechanism.
Molecular oxygen is used as an oxidizing agent for the production of adenosine triphosphate (ATP) in the
organism cells, being afterwards reduced to water. The non-reduced oxygen results in the formation of
superoxides (O"#.) that can be further converted to hydroxyl radicals(HO.), which has the highest reduction
potential of all the physiological relevant ROS. When under control, these species are easily scavenged by
• Antioxidant agents such as polyphenols (Fu et al., 2014; Lipinski, 2011) (e.g., elimination ofHO.),
• Enzymes such as superoxide dismutase (SOD), catalase (CAT) and glutathione peroxidase (GPx)
(Barata et al., 2005; Regoli et al., 2002; Vasseur and Leguille, 2004).
The SOD enzymes catalyse the dismutation of O"#. into oxygen or hydrogen peroxide (H2O2), which is
decomposed by CAT into water and oxygen. Even though H2O2 is less reactive than the radical species, is still
a strong oxidant that needs further elimination. The GPx, also plays a role in the detoxification of H2O2 by
using glutathione (GSH) as a reductant. During the process, GSH is oxidized and converted to glutathione
disulfide (GSSG) being latter reduced back to GSH by glutathione S-transferase enzymes (GST), thus
completing the cycle (Brigelius-Flohé and Maiorino, 2013; Deponte, 2013). The antioxidant-enzymes activity
is considered a reflection of the redox state of the cells and is frequently studied as a biomarker of OS (Figure
P1. 2).
Chapter 1 Part1: NPs in the aquatic environment
37
Figure P1. 2. Antioxidant defense system in an animal cell (Cossu et al., 1997; Sroda, 2011; Garaud, 2015). CAT: catalase; G6PD: glucose-6-phospho dehydrogenase; SeGPx: selenium dependent glutathione peroxidase, GPx: glutathione peroxidase; GR: glutathione reductase; GSH/GSSG: reduced/oxidised glutathione; NAPD+/NAPDH: oxidised/reduced nicotinamide dinucleotide phosphate; NOS: nitric oxide synthase; SOD: superoxide dismutase.
When ROS is overproduced beyond the organism antioxidant response capacity, it leads to several
deleterious effects on the cells components such as lipids, proteins and DNA, possibly resulting in lipid
peroxidation, apoptosis and/or cancer initiation processes, respectively. The production of ROS can be
enhanced by the presence of NPs, depending mainly on their size, aggregation, solubility and coating. It is
commonly accepted that smaller particles can easily penetrate cell membranes, and thus induce cytotoxicity
(Klaine et al., 2008). However, this correlation between size and toxicity is still controversial. For instance, a
size dependent toxicity of 5-10 nm AgNPs on Tetrahymena pyriformis compared to 15-25 nm AgNPs was
reported (Shi et al., 2013) while no size-dependent response on Danio rerio was obtained when exposed to
20, 50 and 110 nm (Bowman et al., 2012). Most studies show that toxicity increases with decreasing particle
size. However, others reported that either the size has no role on toxicity or that smaller NPs are less toxic
(Ivask et al., 2014). A consensus about the size effect is still lacking, and, most probably, will be unlikely to be
attained since the effects seem to be NP- and even specie-dependent (Burić et al., 2015).
The dissolution of metallic NPs such as AgNPs and CuNPs results in the release of Cu and Ag ions, which are
known to catalyse Fenton, Fenton-like and Haber-Weiss reactions, leading to the formation of ROS (Fu et al.,
O2
O2⦁ -
H2O2
OH⦁
H2O
R ⦁
R’
H2O 2GSH
GSSG
NADP+
NADPH
2GSH
GSSG
NADP+
NADPH
ROO⦁ /ROOH
ROH/ONOH
GPx Grd G6PD
GR G6PDSeGPx
SODO2
2H+
e-
CAT H2O+½O2
Men+
Me(n+1)+ +OH-
Fenton
reaction
NO⦁
ONOO-/ONOOH
NOSe-
nitrocompounds
Organicmolecules
Chapter 1 Part1: NPs in the aquatic environment
38
2014; Lipinski, 2011; Wang et al., 2013). Moreover, the highly reactive surface of NPs and the presence of
manufactured and/or natural coatings can lead to the adsorption/complexation of trace metals present in
the environmental compartment also acting as a catalyzer platform to the above-mentioned reactions, and
thus increasing the concentration of ROS in the system. Photoactive NPs such as nTiO2 and nZnO, can also
induce the formation of ROS. When exposed to visible or UV light, these NPs can be photo excited resulting
in the formation of electron-holes, which are powerful oxidants that can react with surface bounded
molecules forming radicals (Clemente et al., 2014). All these processes are schematized in Figure P1. 3.
Figure P1. 3. Potential routes for the generation of ROS due to the presence of NPs. 1) Internalization of NPs ROS generation could occur due to the NPs dissolution inside the cells and/or due to the NPs photocatalytic activity. 2) Dissolution of the NPs leads to an increase concentration of metal ions in the media; some of these metals can also be uptake by the organisms. 3) NPs and/or their surrounding coatings can adsorb/complex other metals present in the media, being taken up by the cells. 4) Photocatalytic activity of the NPs in the presence of UV and/or natural light.
3.2. Omics endpoints
The omics tools, such as toxicogenomic and proteomic, are very useful on the establishment of toxic
endpoints. A toxicogenomic approach allows the identification of gene and protein activities in organisms’
cells induced when in the presence of a certain xenobiotic. A central assumption is that chemicals generating
toxicity by the same mechanism will produce similar gene expression responses under a given set of
conditions, bringing new insights about their mode of action that can be linked to their specific
physicochemical properties. A metallomic and proteomic approach will allow a complete analysis on the
metal and metalloid species composition within a cell or tissue and the establishment of metallo-proteins
profiles leading to the identification of new biomarkers (e.g., proteins expressed by the NP itself (Shepard et
Chapter 1 Part1: NPs in the aquatic environment
39
al., 2000). This will allow a better understanding and profiling of NPs toxic mechanisms discriminating them
from their bulk components.
Despite the large number of nanotoxicology studies, few of them have reported the use of omics tools to
evaluate NPs toxic effects at the molecular level on freshwater organisms (Kuznetsova et al., 2014; Rainville
et al., 2014; Revel et al., 2017). The “omics” approach generates a huge amount of data whose interpretation
is not always straightforward. The large amount of data together with unappropriated physico-chemical
characterization prior to the biological assays results in an escalating number of unknown variables impeding
a comprehensible understanding of the biochemical responses.
Despite the lack of genetic information needed for applying proteomic techniques to freshwater organisms
(Larkin et al., 2003), these approaches are interesting in nanotoxicology as they allow a better understanding
and profiling of toxic mechanisms of nanoparticles and discriminate them from their bulk components
(Rainville et al., 2014). In addition, it has been proved that, to use biomarker measurements, it is better to
have a deep knowledge of mode of actions of pollutants (Vioque-Fernández et al., 2009) and to avoid the
influence of confounding factors such as gender (Sornom et al., 2010), parasitism (Gismondi et al., 2012a,
2012b) and seasonal variations (Sroda and Cossu-Leguille, 2011). Therefore, the use of proteomic approaches
can allow the identification of unbiased responses for nanoparticle exposures (Shepard et al., 2000).
Even though the number of studies concerning nanoparticles-induced toxicity in aquatic organisms under
laboratory conditions continues to increase, the mode of action behind NPs toxicity in freshwater organisms
need further clarification (Canesi et al., 2017; Moore, 2006; Scown et al., 2010). Some studies on metallic
nanoparticles reported, for instance, the effects of copper nanoparticles on the proteomes of aquatic
organisms by considering the effects of CuNPs and Cu2+ on two marine bivalves M. galloprovincialis (Gomes
et al., 2014) and M. edulis (Hu et al., 2014). M. galloprovincialis were exposed for 15 days to CuNPs and Cu2+
(10 μg. L-1). CuNPs showed a high tendency to up-regulate proteins in the gills tissues and down-regulate
them in the digestive gland, while Cu2+ showed the opposite tendency. The effects were metal-dependent
with either common or distinctive response mechanisms induced by CuNPs and Cu2+. Both CuNPs and Cu2+
affected cytoskeleton and cell structure (actin, α-tubulin, paramyosin), stress response (heat shock cognate
71, putative c1q domain containing protein), transcription regulation (zinc finger BED domain-containing
protein 1, nuclear receptor subfamily 1G), and energy metabolism (ATP synthase FO subunit 6). CuNPs
specific mechanisms were also identified with oxidative stress (GST), proteolysis (Cathepsin L) and apoptosis
(caspase3/7-1) while Cu2+ impacted percollagen-D, an adhesion and mobility protein associated with the
detoxification mechanism of Cu2+. This study demonstrated that CuNPs effects are not only due to ion release
but are also due to mitochondrial and nucleus stress induced cell-signalling cascade that can lead to apoptosis
(Gomes et al., 2012, 2013a). M. edulis were exposed to 100 nm CuNPs (1000 ppb) for 1h. CuNPs accumulated
Chapter 1 Part1: NPs in the aquatic environment
40
in gills and caused an alteration of the expression of six proteins: α and β-tubulin, actin, tropomyosin,
triosephosphate isomerase and CuZnSOD. Actin and triosephosphate showed a decrease in proteins thiol
levels while alpha-tubulin, tropomyosin and CuZnSOD showed an increase in carbonylation indicating protein
oxidation of cytoskeleton and antioxidant enzymes in response to CuNPs. This study showed that CuNPs
affect first cytoskeletal protein and disrupt lysosomal membrane stability (Hu et al., 2014). Other studies
reported the effects of ZnO NPs at the transcriptomic level of D. magna (Poynton et al., 2012). Exposure of
adult D. magna to very high concentration (9 mg.L-1) of 20 nm ZnO-NPs for 24h and to 0.52 mg.L-1 of 30 nm
of ZnO-NPs for 96h led to induction of cytoskeletal transport proteins and a repression of the expression of
genes linked to the reproduction and antioxidant responses (Adam et al., 2015; Poynton et al., 2012). Within
metallic NPs, TiO2 NPs are among the most studied ones. Molecular effects of TiO2 NPs on zebrafish were
reported. TiO2 NPs 20 nm were showed to cause a down-regulation of genes involved in ribosomal functions
of D. rerio gills (Griffitt et al., 2009). Other studies reported that microinjection of TiO2 NPs (25 nm; anatase;
8.5 ng.g-1 bw; 48h at 28.5°C) in zebrafish embryo significantly modified the expression of genes involved in
circadian rhythm, cell signalling, exocytosis and vesicular trafficking (Jovanović et al., 2011). However, no
induction of ROS induction was detected in these two studies.
4. NPS TOXICITY ON FRESHWATER ORGANISMS
4.1. Silver nanopartic les (AgNPs)
AgNPs are known for their antifungal and antimicrobial properties, being extensively used in several products
such as clothing, cosmetics, medical devices, paints, etc. (Fabrega et al., 2011; McGillicuddy et al., 2017;
Vance et al., 2015). Several databases have been created in order to collect information of nano-
functionalised products. Two reviews compiled lists of these inventories (Hansen et al., 2016; Vance et al.,
2015). According to the Woodrow Wilson database, more than 400 consumer products contain AgNPs which
make them among the most used NPs in consumer products (Hansen et al., 2016; Vance et al., 2015). The
global annual production of silver has been estimated around 27,000 tons (US Geological Survey, 2016) with
an annual consumption of 55 tons/year to 450 tons/year as AgNPs (Zhang et al., 2016). The increasing
application of AgNPs will lead to their release in the environment through different routes from synthesis,
manufacturing, distribution, end product use and end of life disposal of every day consumer products
containing AgNPs (Blaser et al., 2008; Mueller and Nowack, 2008). Release of AgNPs in the environment may
occur via wastewater treatment plants (WWTP) even though AgNPs may be retained in sewage sludge (Kaegi
et al., 2013). In wastewater, it was suggested that AgNPs are in contact with sulfides leading to the formation
of an important amount of silver sulfide (Ag2S) that are kept in sewage sludge (Kaegi et al., 2013; Völker et
al., 2015) leading to the release of about 15% of AgNPs in surface water (Keller et al., 2013). Thus, the
Chapter 1 Part1: NPs in the aquatic environment
41
predicted environmental concentrations in surface water are usually in the low ng. L-1 to µg. L-1 (Mueller and
Nowack, 2008) and range from 0.088 to 2.16 ng. L-1 in European and North American surface waters
(Gottschalk et al., 2009). Recent studies reported predicted environmental concentrations, in surface waters,
in the range from 0.01 and 0.32 µg. L-1 (Tiede et al., 2009; Yang et al., 2016).
Many questions have been raised in order to determine the fate of AgNPs once they reach the environment
(Klaine et al., 2008). Several factors including size, structure and environmental physico-chemical parameters
may affect the behaviour of AgNPs in aquatic environment and thus influence their toxic effects. The
speciation and concomitant interference is of particular note to researchers attempting to risk assess AgNPs
as their toxic mechanisms towards aquatic organisms are still not fully clear (McGillicuddy et al., 2017). It was
possible to identify two distinct routes that could induce biochemical responses: (Auffan et al., 2009; Lowry
et al., 2012; Schultz et al., 2014):
• Presence of the NPs per se
• Presence of both AgNPs and dissolved Ag+.
The identification of responses that are uniquely due to the presence of AgNPs and do not occur in matched
Ag+ exposures are crucial. This is possible by:
• Using AgNPs with low dissolution rates, so, that the leached Ag+ in the media is insufficient to induce
toxicity to the organisms,
• Identify endpoints specific to AgNPs, such as internalization of NPs, cytotoxicity and genotoxicity.
The work performed by Kumar et al., (2014) is an example of the application of these strategies. They reported
a significant increase of ROS and SOD in addition to morphological alterations on freshwater bacteria exposed
to polyvinyl-pyrrolidone PVP-AgNPs. Since the NPs dissolution was very low (leached Ag+ < 1 µg. L-1), it was
concluded that the results were related to AgNPs form. Morphological changes and alteration in genes
profiles, related to the presence of AgNPs itself, were also observed in carp (Cyprinus carpio) (Lee et al., 2012),
zebrafish (Danio rerio) (Choi et al., 2010; Griffitt et al., 2009), D. magna (Poynton et al., 2012), medaka
(Oryzias latipes) (Pham et al., 2012) and rainbow trout (Oncorhynchus mykiss) (Gagné et al., 2012). The
exposure of 7 days old D. magna to citrate-coated AgNPs (size 10 nm, 30 μg L-1) and AgNO3 (2.5 μg L-1) during
24h showed that both AgNPs and Ag+ increased proteins thiol content, while only particles increased proteins
carbonyl levels (Gündel et al., 2007). Similarly, to D. magna, AgNPs and Ag+ also impacted different proteins
in the mussel M. gallopronvinciallis. Although M. galloprovincialis is a marine organism, the observed effects
on this species exposed to AgNPs was highly significant. M. galloprovincialis were exposed for 15 days to
AgNPs and Ag+ (10 μg. L-1). AgNPs affected similar cellular pathways than Ag+ with common responses
mechanisms in cytoskeleton and cell structure (catchin, myosin heavy chain), stress responses (heat shock
protein 70), oxidative stress (GST), transcriptional regulation (nuclear receptor subfamily 1G), adhesion and
Chapter 1 Part1: NPs in the aquatic environment
42
mobility (percollagen P), and energy metabolism (ATP synthase FO subunit 6 and NADH dehydrogenase
subunit 2). Exposure to AgNPs altered the expression of protein associated with stress response (major vault
protein and ras partial) and proteins involved in cytoskeleton and cell structure (paramyosin) while Ag+ had
a strong effect in proteins related to stress response (putative c1q domain containing protein) and proteins
involved in cytoskeleton and cell structure (actin and a-tubulin). The identification of altered protein
suggested that AgNPs toxicity is mediated by oxidative stress-induced cell signalling cascade (including
mitochondria and nucleus) that can lead to cell death. In fact, it has been described that oxidative stress
caused by AgNPs and Ag+ induce cytoskeleton disorganization (Gomes et al., 2013b). Transcriptomic effects
of two kinds of AgNPs (CIT-AgNPs 40 nm and PVP-AgNPs 35 nm) were investigated on 10 days old D. magna
after 24 h of exposure to different concentrations corresponding to 1/10 of the LC50 of each AgNPs (CIT-
AgNPs: 0.43 µg.L-1; PVP-Ag NPs: 1.1 µg.L-1) and also at the LC25 (CIT-AgNPs: 3.5 µg.L-1; PVP-AgNPs: 8.1 µg.L-1).
This work revealed common effects of these two types of AgNPs with an induction of the ubiquitination and
proteolysis pathways and a repression of ribosomal genes suggesting a general effect on the protein
metabolism. This work also revealed an effect of the PVP coated AgNPs on the expression of immune and
circadian genes of D. magna (Poynton et al., 2012). Other studies reported a significant decrease in lysosomal
membrane stability and the presence of vacuolization and necrosis in D. rerio exposed through the diet to
100 µg.L-1 of PVP/PEI-AgNPs 5 nm was reported (Lacave et al., 2017). Additionally, DNA damage, total
degeneration of hepatocytes and down regulation of stress related genes in L. rohita exposed to PVP-AgNPs
20 and 30 nm were observed. These studies suggested that the observed effects were caused by the AgNP
itself.
It is well known that Ag+ has a great propensity to bioaccumulate in tissues leading to ROS generation,
genotoxicity and inhibition of Na+/ K+-ATPase activity by blocking the Na+ uptake by the cells (Arce Funck et
al., 2013; Luoma et al., 2014; Morgan et al., 1997). For instance, toxic effects were observed in juvenile
Oncorhynchus mykiss exposed to CIT-AgNPs, with claimed low propensity for dissolution (Schultz et al., 2012).
However, it was not possible to confidently establish if the observed effects were caused by the AgNPs per
se or by dissolved Ag+. In fact, it is very difficult to distinguish between the toxic effects induced by particulate
or ionic Ag, and, thus, is crucial not only to evaluate the size distribution but also to quantify the AgNPs
dissolution in the medium.
Both forms, NPs and ionic Ag, can induce OS and genotoxicity being the distinction between these effects a
truly challenge. For example, in the algae Chlorella vulgaris a positive correlation between ROS production
and LPO on the tissues after 24h of exposure to uncoated AgNPs (1 and 10 mg L-1) (Oukarroum et al., 2012)
was found, but with no possibility to establish a positive correlation with solely the particulate or ionic Ag.
Similar difficulties were also observed in the snail Lymnaea luteola (Ali et al., 2014), as exposure to AgNPs
Chapter 1 Part1: NPs in the aquatic environment
43
resulted in DNA damage and induction of OS but without the possibility to identify which Ag forms has the
dominant role on the observed effects. In these situations, a complete physicochemical characterization of
the NP in the exposure media, localisation within biological tissues along with the use of controls containing
ionic Ag are crucial to understand which Ag form (or both) is responsible for the observed biochemical
responses. Ions release form AgNPs can directly bind to cell structure or interact with sulfhydryl groups of
proteins (such as Ca2+-ATPase) and disrupt calcium homeostasis through ROS formation (AshaRani et al.,
2009; AshaRani et al., 2009; Gomes et al., 2013b). Mechanical injury caused by AgNPs in mitochondria may
be another reason of important changes in membrane permeability, disruption of the respiratory chain and
contributes to oxidative stress (Asharani et al., 2009b). The up-regulation on the nuclear receptor sub-family
1G in the gills were reported after exposure to Ag+ indicates the ability of silver ions (released from AgNPs)
to interfere with signal transduction in DNA-related functions and induce genotoxicity (Gomes et al., 2013b).
Localisation of NPs in biological tissues is a step forward in understanding uptake route and internal
distribution. For instance, PVP-AgNPs 50 nm were observed in the brain of the fish Piaractus mesopotamicus
(Bacchetta et al., 2017) and poly-N-vynil-2-pirrolidone/polyethylenimine (PVP/PEI) AgNPs 5 nm were found
in liver and intestine of D. rerio (Lacave et al., 2017) as CIT-AgNPs 20 nm were observed in liver and gills of
rainbow trout (Bruneau et al., 2016). Other studies reported a dose and time-dependent uptake of PVP-
AgNPs 20 and 30 nm in the fish L. rohita exposed to up to 800 µg.L-1 for 168h (Sharma et al., 2016). Usually,
localisation of NPs is performed using transmission electron microscopy (Heinlaan et al., 2011; Lovern et al.,
2008). This method allows the detection of NPs without identifying the chemical nature of the observed
element. Thus, more sensitive techniques have been used in order to better define AgNPs internal
distribution in aquatic organisms. For example, a secondary ion mass spectrometry instruments, NanoSIMS
50, was used to define AgNPs internal distribution in D. magna exposed for 48h to up to 10 mg.L-1 of
chemically synthesised AgNPs 20 and 200 nm and biologically synthesised from plant leaf extract AgNPs 23
and 27nm (Georgantzopoulou et al., 2013). AgNPs were observed as big aggregates in the gut lumen of D.
magna with AgNPs 20 nm being able to cross the gastrointestinal tract (GIT), AgNPs 200 nm passing the
peritrophic membrane and possibly being present within the epithelial layer of the cells and AgNPs 23 nm
being observed around developing oocytes. However, AgNPs 27 nm were not found to have cross D. magna’s
GIT (Georgantzopoulou et al., 2013). These results were linked to surface properties of the studied AgNPs as
the biological AgNPs 23 and 27 nm were synthesised from to different plants, Ocinum sanctum and
Azadirachta indica, respectively. AgNPs 23 nm appeared as the one to release the highest amount of Ag ions.
Thus, although NanoSIMS 50 allowed the internal distribution of AgNPs in D. magna it did not allow
distinction between the AgNPs per se and the released ions.
Chapter 1 Part1: NPs in the aquatic environment
44
4.2. Gold nanoparticles (AuNPs)
Compared to AgNPs, AuNPs are considered as chemically stable NPs as they are not expected to release ions
(Auffan et al., 2012). AuNPs are being investigated for their unique optical properties and are used as contrast
agents in electron microscopy, optical sensors, catalysts, and for therapeutic uses. AuNPs have one of the
broadest range of synthetic strategies (Thompson et al., 2017) resulting in cubes, spheres, pyramids, stars
octopods and other variants (Xia et al., 2008). Au metal is extremely resistant to oxidation and essentially
insoluble under ambient conditions. Thus, using AgNPs and AuNPs as model NPs in the present work allowed
us to integrate the effects of ions and NPs on one hand and effects of NPs only on the other hand.
A major challenge in nano-ecotoxicology is finding suitable methods to determine the uptake and localisation
of NPs on a whole organism level (Skjolding et al., 2017). For nanoparticles in order to cause toxicity, it is
assumed that they are taken up and distributed within an organism, therefore a better understanding of how
they behave in an in vivo system and what factors influence their uptake by living organisms is necessary
(Wepener et al., 2011). Some studies reported a size and coating dependent bioaccumulation of AuNPs. CIT-
AuNPs and mercaptoundecanoic acid (MUDA)-AuNPs 10 and 30 nm were reported to accumulate in Daphnia
magna with CIT-AgNPs being more taken up then MUDA-AuNPs and MUDA-AuNPs 10 nm being more
bioconcentrated than MUDA-AuNPs 30 nm (Skjolding et al., 2014). One aspect on NPs structure that may
influence toxicity is the identity and charge of ligand molecules used for functionalization of NPs surface.
Thus, three different types of capping agent of 4-5 nm AuNPs were tested on D. magna. Animals were
exposed to CIT-AuNPs, MPA-AuNPs and PAH-AuNPs, and acute and chronic toxicity tests were performed.
The negatively charged AuNPs appeared to be less toxic while after chronic exposure, both positively and
negatively charged AuNPs affected reproduction (Bozich et al., 2014). However, the mechanisms underlying
the observed effects could not be clearly determined even though the surface chemistry appeared as a critical
factor influencing NPs toxicity on D. magna (Bozich et al., 2014). Uptake of different sizes and surface-coated
AuNPs was modelled using D. magna as model organism (Wray and Klaine, 2015). Size and surface charges
resulted as the main influencing factors controlling AuNPs uptake by D. magna and elimination, whereas
shape had no significant effects (Wray and Klaine, 2015). Examination of intestinal microvilli indicate no
assimilation of AuNPs by D. magna as AuNPs remained in the gut lumen and the carapace (Wray and Klaine,
2015). One study used an innovative technique that allows the localisation of NPs but not of ions to localise
and determine the internal distribution of AuNPs in D. magna. Animals were exposed for 14 days to up to 20
mg. L-1 of AuNPs. Cytoviva darkfield imaging microscope was used to localise AuNPs in D. magna. AuNP
aggregates were observed in the gut of D. magna when animals were fed as no AuNPs were observed in the
body cavity. However, a direct exposure to AuNPs led to the adsorption of AuNPs on the carapace of D.
Chapter 1 Part1: NPs in the aquatic environment
45
magna and led also to their ingestion and uptake into the gut. This significant uptake had no significant effect
on reproduction and moulting pattern of D. magna (Botha et al., 2016).
Few other studies investigated the effects of AuNPs on freshwater invertebrates. For instance, asiatic clams
Corbicula fulminea were exposed to algae contaminated with 10 nm of amino-acid AuNPs for 7 days. A
concentration-dependent mortality of algae was observed after 24h of exposure. Trophic exposure led to a
significant bioaccumulation with amino-acid AuNPs were observed in the gills and digestive gland. AuNPs
were also observed in the lysosome and led to an oxidative stress in C. fulminea (Renault et al., 2008). Very
few studies have investigated the effects of AuNPs on Gammarus sp. G. pulex were exposed to differently
functionalized AuNPs, namely CIT-AuNPs, MUDA-AuNPs, NH2-AuNPs and PEG-AuNPs. A surface-coating-
dependent uptake of AuNPs was observed in G. pulex with CIT-AuNPs and MUDA-AuNPs being the most
bioaccumulated by G. pulex while no size effects were observed (Park et al., 2015). To our best knowledge
there is only one study investigating effects of AuNPs on G. fossarum. Individuals were exposed through a
contaminated biofilm to up to 45 mg.L-1 of 10 nm amino-acid AuNPs (Baudrimont et al., 2017). Cellular
damage linked to an oxidative stress and significant effect on mitochondrial respiration was observed after 7
days of exposure with a daily renewal of contaminated food. Additionally, modulation of digestive enzymes
was also observed suggestion a modification of digestive functions (Baudrimont et al., 2017).
Within freshwater vertebrates, the model organism zebrafish is one of the most studied species. Zebrafish
embryo were exposed to different metal-based nanoparticles (Ag, Au, CuO and ZnO). AuNPs appeared as the
most bioconcentrated ones in the eggs followed by Ag, ZnO and CuO. However, when the whole organism
was assessed, Ag appeared as the highest bioaccumulated NPs followed by Zn (Böhme et al., 2017). A
previous study also reported presence of AuNPs in ovaries of female zebrafish exposed to AuNPs 15 nm and
47 nm (Dayal et al., 2016). Zebrafish was exposed to fluorescent AuNPs via aqueous or dietary route (Skjolding
et al., 2017). Internal distribution was assessed using light sheet microscopy (LSM) at different time point (1,
3 and 7 days). AuNPs were observed within the gut of zebrafish after trophic exposure. AuNPs were not
observed in the gut epithelia indicating the absence or limited uptake of AuNPs through the intestinal villi.
Direct exposure of zebrafish to AuNPs led an increase in relative swimming distance. This study showed the
importence of exposure routes and subsequent localisation of AuNPs in zebrafish (Skjolding et al., 2017).
Deleterious effects were reported as AuNPs caused modulation of genes linked to oxidative stress,
mitochondrial metabolism, detoxification and DNA repair in zebrafish exposed to AuNPs 14 nm (Dedeh et al.,
2015). Other studies using estuarine and marine organisms reported further toxic effects of AuNPs. Estuarine
copepod Eurytemora affinis were exposed for 30 min to 11.4 µg. L-1 of AuNPs and a decrease in swimming
behaviour and a lower velocity and acceleration were observed in females whereas males appeared to be
less sensitive. However, the mechanisms underlying these effects were not clearly determined as neither
Chapter 1 Part1: NPs in the aquatic environment
46
adsorbed nor ingested AuNPs could be localised in E. affinis (Michalec et al., 2017). AuNPs showed a higher
uptake in the digestive gland than in the gills of the clam R. philippinarum. Changes in oxidative stress and
inflammatory response markers as phase II antioxidant enzymes and qPCR changes were observed (Volland
et al., 2015). However, no strong oxidative damage was reported and a significant depuration from digestive
tract after 7 days in cleans medium was observed (Volland et al., 2015). The oligochaetes Tubifex tubifex were
exposed for 5 days to 10 and 60 µg. g-1 dry weight of sediments of 5 nm AuNPs. A concentration dependent
uptake of Au was observed but the internal concentration did not cause significant mortality of T. tubifex
(Zhang et al., 2017). Marine blue mussels Mytillus edulis were exposed to 750 ppb of 5 nm AuNPs for 24h
which mainly accumulated in the digestive gland. A significant decrease in lysosomal membrane stability in
addition to a significant increase in lipid peroxidation was observed together with a decrease in thiol-
containing proteins. All these results suggest that the observed effects are linked to an increase in ROS leading
to oxidative stress (Tedesco et al., 2010).
5. GENERAL REMARKS AND CONCLUSIONS
Most of the biochemical responses reported are related to the organism’s ROS defence mechanisms, mainly
through gene expression or changes on anti-oxidant enzyme activities. As mentioned before (section 3.2),
data related with changes in freshwater organism genome and proteome due to the presence of NPs is very
scarce. This is indeed surprisingly since -omic techniques have already proved a great potential on the
recognition of signatures related to specific stress, eventually leading to the discover of new biomarkers
(Revel et al., 2017). The few data available suggest that the interaction with the organisms is NP-specific
regarding the diversity of effects of AgNPs and AuNPs on aquatic organisms and raise questions about clear
and defined mechanism of action.
NPs composition can play an important role on ROS generation, since some metals constituting the NP can
instigate Fenton and Weiss-type reactions releasing ROS in the intra or extracellular media. Evidently, and as
usually performed, the bioassays should contain a control group exposed to the salt form of the metal
constituting the NP, allowing the distinction between the effects provoked by the NP per se and/or by the NP
dissolution products.
Despite the large number of studies dealing with ecotoxicology of NPs, it is evident that is still not possible to
establish crucial predictive structure-activity relationships. This is mainly due to the fact that the majority of
the available studies have critical deficiencies on their experimental designs; a comprehensive physico-
chemical characterization of the particles under the exposure conditions is mostly miscarried or restricted to
a secondary task. Despite the scientific community is already aware about the importance of the NPs
physicochemical characterization prior and during the bioassay, this is still frequently neglected in the most
recent studies giving rise to more confusing and contradictory data. Clearly, this pushes back the possibility
Chapter 1 Part1: NPs in the aquatic environment
47
to establish a proper environmental risk assessment plan for these current early generations on ENPs (1st and
2nd generation passive and active nanostructures, respectively), while advanced generations of ENPs (3rd and
4th generation nano-systems and molecular nano-systems) may not be far away, bringing additional
challenges that require further novel approaches.
The dynamic speciation of the NPs should be assessed in the same exposure media of the bioassay by
following key NPs transformations:
• Dissolution,
• Homo- and hetero-aggregation,
• Sedimentation.
Several analytical tools are nowadays available for the quantification of these physicochemical
transformations, each of them having their specific advantages and limitations being able to provide different
information on ENPs properties (see reviews (Domingos et al., 2009; Tiede et al., 2009). Several analytical
tools are nowadays available for the quantification of these physicochemical transformations, each of them
having their specific advantages and limitations being able to provide different information on ENPs
properties (see reviews Domingos et al., 2009; Tiede et al., 2009). This physicochemical characterization
approach allows:
• To assess the bioavailable NP-containing species to which the organisms will be exposed,
• To relate the biocompatible or bioadverse effects with the NP-containing species permitting a NP
categorization and function.
Nanotoxicology is indeed a multidisciplinary field where the study of the NPs physic, chemistry and biological
impacts is crucial for a complete toxicological assessment. Unfortunately, there is a lack of legislation
controlling the production, use and release of these materials to the environment, and new NPs are
commercialized every day without an appropriate assessment about their impacts in environment and
human health. The establishment of national and international laws regulating the production of these
materials is mandatory. Furthermore, it is also urgent to increase the number of comprehensive
nano(eco)toxicology studies under natural more environmentally-realistic conditions implying the co-
presence of ENPs (at low and environmentally-realistic doses) and environmental constituents such as natural
organic and inorganic dissolved and colloidal matter. Only with these approaches a comprehensive risk
assessment will be possible with production of environmentally safe-by-design ENPs.
Chapter 1 Part1: NPs in the aquatic environment
48
Acknowledgements
All the authors acknowledge to the European COST action ES1205, Engineered Nanomaterials - From
Wastewater Treatment & Stormwater to Rivers (ENTER). G. Vale and R.F. Domingos acknowledges to
Fundação para a Ciência e Tecnologia (FCT, Portugal) the post - doctoral grant SFRH/BPD/73117/2010 and
the project PTDC/AAC-AMB/111998/2009, respectively. The contribution of K. Mehennaoui and S. Cambier
were in part possible within NanoGAM (AFP-PhD-9229040) and NANION (FNR/12/SR/4009651) respectively,
both founded by Fonds National de la Recherche Luxembourg.
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CHAPTER 1
Part 2: G. fossarum as a model organism in nanotoxicology
CHAPTER 1
Part 2: G. fossarum as a model organism in nanotoxicology
Chapter 1 Part 2: G. fossarum as a model organism
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1. SYSTEMATIC AND IDENTIFICATION OF Gammarus fossarum
1.1. Systematic position
More than 4500 species belong to the crustacean super-family Gammaridea and represent more than 85%
of amphipods (Bousfield, 1973; Kunz et al., 2010). Gammarids are one of the most widespread group
geographically, with more than 300 species composing the Gammaridae family (Väinölä et al., 2008) (Figure
P2. 1 A, B). They are found throughout a range of marine, freshwater and terrestrial habitats (Bousfield, 1973;
Kunz et al., 2010) whereas the three other amphipod families (Hyperiida, Ingolfiellidea and Caprellida) are
highly specialized and ecologically restricted (Kunz et al., 2010).
Figure P2. 1. (A) Tree of life of the order Amphipoda (Yellow arrow) and its constituents’ suborders Gammaridea,
Senticaudata and Hyperiidea. (B) Tree of life of Gammarus fossarum (Yellow arrow) (Lifemap)
The genus Gammarus, the most important component of the Gammaridae family, is one of the most frequent
in freshwater ecosystems (Barnard and Barnard, 1983; Karaman, 1977). The genus Gammarus counts more
than 100 known species (MacNeil et al., 1997). The classification of Gammarus fossarum, the model organism
used in the present work, is presented below:
AB
Chapter 1 Part 2: G. fossarum as a model organism
59
Table P2. 1. Classification of the species Gammarus fossarum
Gammarus are epigean species widely distributed all over the northern hemisphere (Karaman, 1977; Kunz et
al., 2010; Živić and Marković, 2007) (Figure P2.2 A). The most frequent species in Europe are: Gammarus
fossarum (Koch, 1835), Gammarus pulex (Linnaeus, 1758) and Gammarus roeseli (Gervais, 1835) with G.
fossarum being one with the most common and the most diverse species complex in Europe (Weiss et al.,
2014). A species complex represents a group of species being morphologically close but very distinct from a
molecular biology point of view (Brown et al., 1995). The highest diversity of G. fossarum complex is located
in the southern Carpathian and the Balkan peninsula (Copilaş-Ciocianu et al., 2017). G. fossarum is also the
most abundant species in Luxembourg (Figure P2.2 B).
Figure P2. 2. Geographical distribution of Gammarus fossarum in (A) Europe and (B) Luxembourg (Dohet et al.,
2008; Fauna Europaea).
Kingdom Metazoa
Phylum Arthropoda
Subphylum Crustacea
Class Malacostraca
sub-class Eumalacostraca
Super-order Peracarida
Order Amphipoda
Sub-order Senticaudata
Infra-order Gammarida
Parv-order Gammaridira
Super-family Gammaroidea
Family Gammaridae
Genus Gammarus
Species fossarum
FaunaEuropea
A B
Chapter 1 Part 2: G. fossarum as a model organism
60
1.2. Identification of Gammarus fossarum
Figure P2. 3. Determination key [in French] of Gammarus fossarum (Felten, 2003).
Chapter 1 Part 2: G. fossarum as a model organism
61
The identification of the Gammarus species in question is generally easy and is performed on mature adult
males as they present morphological characteristics more easily discernible compared to females (Piscart and
Bollache, 2012). The identification key (Figure P2. 3, [in French]) describes the species mainly found in France
and Luxembourg. G. fossarum are characterised by the absence of dorsal carinae (hull) and strips. G. fossarum
are described to have few spines and bristles on the antennae and the extremity of urosoma. The internal
oar of the third uropod is 1/3 the size of the external one. This parameter allows morphological distinction
between G. fossarum and G. pulex (Felten 2003, Figure P2. 3). However, the recent advances in molecular
biology techniques allow a better identification of Gammarus species. Indeed, different cryptic species have
been identified within the G. fossarum complex (Müller, 2000).
The concept of cryptic species is known for more than 300 years (Bickford et al., 2007; Lagrue et al., 2014). It
describes species that are morphologically similar but with a significant genetic differentiation that led to
their wrong classification under the same name (Bickford et al., 2007). The genetic differentiation between
cryptic species is linked to a reproductive isolation that occurred potentially already millions years ago (Webb
and Bartlein, 1992). Three G. fossarum cryptic species were genetically identified, namely Type A, B and C
(Müller, 2000). This split occurred during or before the ice age of the Pleistocene (Webb and Bartlein, 1992)
leading to biological (drift in mitochondrial DNA) rather that morphological differentiation (Feckler et al.,
2012; Müller, 1998; Westram et al., 2010). G. fossarum cryptic species have been reported to have different
geographical distributions (Figure P2. 4). Type A and Type B have been observed in central Europe
(Scheepmaker and van Dalfsen, 1989) and can coexist (Müller, 1998). Type C, in addition to Type B have been
found in Luxembourg (Müller, 1998) with type B being the most abundant one (Annexe 2).
An accurate identification of the cryptic species is crucial in ecotoxicological studies (Feckler et al., 2014;
Lagrue et al., 2014; Sattler et al., 2007; Westram et al., 2010). Differences in sensitivity between species have
already been reported. For instance, G. fossarum were found to be more sensitive to the fungicide
Tebuconazole than G. pulex (Adam et al., 2009). Therefore, it can be assumed that different cryptic species
have different sensitivities to chemicals and deviations among cryptic lineages regarding their physiological
and behavioural characteristics are possible and have been reported (Bickford et al., 2007). A higher
sensitivity to pesticides was observed in G. fossarum Type A compared to Type B (Feckler et al., 2012).
Different sensitivities to parasitism was also highlighted as a higher infection rate by acanthocephalan
parasites was observed in Type B compared to Type A (Westram et al., 2011). All this could explain the
observed species differences in sensitivity to chemicals (Gismondi et al., 2012a, 2012b).
Chapter 1 Part 2: G. fossarum as a model organism
62
Figure P2. 4. Distribution of Gammarus fossarum cryptic species in Europe (Adapted from Westram et al., 2011).
Grey shaded zone indicates the contact zone between Type A and B.
2. MORPHOLOGY
Gammarus are Crustacea malacostraca characterized by a segmented body. They present a bilateral
symmetry, a laterally flattened body, sessile eyes, a pair of antennae with homogenous size and a total length
of 20 mm maximum (Tachet, 2000). Usually, females are smaller than males at the same developmental stage
(Gagné et al., 2005; Tachet, 2000). Compared to shrimps, Gammarus are characterized by the absence of a
carapace which is replaced by a cuticle composed of chitin. The first thoracic segment is fused to the head
and their body is divided in four different parts: prosoma, mesosoma, metasoma and urosoma (Chevreux et
al., 1970, Figure P2. 5). Gammarids usually have large coxal plates and a large abdomen with six pairs of
appendages (Figure P2. 5).
Chapter 1 Part 2: G. fossarum as a model organism
63
Figure P2. 5. General morphology of Gammarus sp. Adapted from Felten 2003.
2.1. Prosoma
Prosoma represents the cephalic region of Gammarus sp. Antennae (2 pairs of sensorial appendages), eyes,
mouth (4 pairs of masticatory appendages), and cephalon compose this region. The eyes are kidney-shaped
and sessile and the antennae have similar sizes and present a flagellum allowing distinction of Gammarus
from Niphargidae (Tachet, 2000). The mouth present mandibles located lateral to the mouth, in conjunction
with the upper and lower lips and surround the mouth opening (MacNeil et al., 1997). They consist of a strong
chitinized incisor, a small accessory plate (the lacinia mobilis), a large medial molar and a spine row between
the molar and the lacinia. In Gammarus, the molar is tough and ridged for crushing and grinding (Lincoln,
1979). Hence, these characteristics allow a wide variety of food to be used by Gammarus sp. (MacNeil et al.,
1997).
2.2. Mesosoma
The mesosoma represents the thoracic region. Seven pereon plates compose this region on which coxal
plates are sealed. Each coxal plate carry a pair of periopods. The first pair represent the gnatopods and they
are used for handling and catching females and for locomotor activity. Males usually present bigger
gnatopods than females. Moreover, these appendages are used for feeding, grooming, burrowing and for
agonistic encounters between males (Borowsky, 1984; MacNeil et al., 1997). Gills and genital organs are also
present in this region fixed under the coxal plates and are in direct contact with the external environment.
Antennae
Peduncle
eyes
Head
GnathopodsPereopods
Uropods
Telson
Pereopods
Pleopods
Pe: Pereo PC: Coxae Pl: Pleon Ur: Urosome
Chapter 1 Part 2: G. fossarum as a model organism
64
2.3. Metasoma
The metasoma represents the abdominal region of a Gammarus body. It carries three pairs of pleopods. In
addition to locomotion, these appendages are used for renewal of water flux on the gills surface. Their
continuous movement is directly linked to oxygen absorption. Therefore, the pleopod beats frequency is
increasingly used as a behavioural biomarker in ecotoxicology studies. This response indicates the
oxygenation rate (Andreï et al., 2016; Arce Funck et al., 2013; Felten et al., 2008; Sornom et al., 2010; Vellinger
et al., 2013).
2.4. Urosoma
The urosoma presents three different segments called uromeres. Each of them carries a uropod pair allowing
the animal to move. Each uropod presents a peduncle, an endopod and an expopod. The shape and size of
the third uropod is different from the two others. This parameter is used for the species determination and
is used as a characteristic for discriminating Gammarus from Echinogammarus. The last segment of the
urosoma present a telson at its extremity.
3. ANATOMY OF Gammarus fossarum
3.1. Nervous system and c irculatory organs
The nervous system consists of a protocerebrum or brain with a long nerval chain covering the entire body
of G. fossarum (Figure P2.6). There are ganglions in the ventral part of the body all the way to the urosoma
and they are all linked with intersegmental nerve connections. All appendages are connected to nervous nets
(Chevreux et al., 1970; Felten, 2003; Charron, 2014). The eyes communicate directly with the ganglia in the
head by large optic nerves. The antenna are sensory and large nerve run out to them as well (Shimek, 2008).
Haemolymph circulation is based on the heart located close to the head (Figure P2.6). This organ has his two
extremities open allowing circulation of haemolymph in all the body. The heart is surrounded by an
heamocoele located on the dorsal side of G. fossarum body and is found in the mesosoma (Felten, 2003;
Charron 2014).
All the organs are in direct contact with haemolymph. Haemolymph plays an important role is transporting
nutrients, hormones, excretions products, ions and oxygen (Nebeker et al., 1992). Haemolymph has an
osmolality around 300 mOsmol allowing a good equilibrium of water and ions in the cells (Charron, 2014).
Chapter 1 Part 2: G. fossarum as a model organism
65
Figure P2. 6. Lateral view of the anatomy of Gammarus sp. Illustrating the principal organs (Schmitz, 1992).
aa: anterior aorta, an: anus, ap: posterior aorta, ba: arterial bulb, cc: circumoesophagial connexion, cd: digestives caeca, cda: anterior digestive caeca, chc: heart chamber, ci: intersegmental connexion, co: heart, cp: posterior caeca, cpy: pyloric chamber, gpe: pereionic ganglion gpl: pereionic ganglion 1, im: midgut, ip: intestinal tract, lds: suspended dorsal ligament, mu: urosomic mass, ms: suboesophagial mass, nan: antennule nerve, nat: antennae nerve, oe: oesophagus, pr: protocerebrum
3.2. Intestina l caeca and digestion system
In Gammarids the mouth is located near the base of the head. It leads to a short oesophagus located before
the stomach where the food is grinded by chitinized plates (Shimek, 2008; Figure P2.6). Behind the stomach
is the midgut located which is as long as G. fossarum body. The midgut is surrounded by two pairs of pouches
or intestinal caeca. Two caeca are located on the side and the two other are located at the bottom of the
midgut (Shimek, 2008). These organs are the central site of metabolism were both digestion and secretion of
digestive enzymes occur. Intestinal caeca or hepatopancreatic caeca are the central organs of digestions in
amphipods. The caeca play a role in nutrient uptake, excretion, moulting cycle, storage of inorganic reserves
and lipid and carbohydrates metabolism (Chevreux et al., 1970; Correia et al., 2002; Grassé, 1961; Schmitz
and Scherrey, 1983; Charron, 2014). They present a tubular structure surrounded with epithelial cells and
muscles covered with connective tissues (Schmitz and Scherrey, 1983,
Figure P2. 7 B). There are four different types of cells forming the epithelial tissues in amphipods (cells E, F,
R and B). The epithelial cells form a barrier between the lumen of the gut and the haemolymph (Kutlu et al.,
2002). The plant material and the fungi are digested in the midgut using cellulase (Chamier and Willoughby,
1986; MacNeil et al., 1997).
Chapter 1 Part 2: G. fossarum as a model organism
66
3.3. Gammarus g il ls and respiration
Gammarids have 6 pairs of gills in direct contact with the external environment. They are present under the
thoracic appendages (Felten and Guérold, 2006; Sutcliffe, 1984). Gills are thin flattened oval plates called
lamellae formed from the epipodits of the pereopods (Sutcliffe, 1984, Figure P2. 7 A, C).
Crustacean gills are lined by a single layer of epithelial cells whose basement membrane at the serosal side is
directly bathed by the haemolymph (Figure P2. 7 C). The mucosal side is covered by a chitinous cuticle and
faces the external medium. The thickness of the epithelium varies depending on its physiological and
biochemical function (Dunel-Erb et al., 1997; Martinez et al., 2005). Thinner regions of the epithelium (1-5
µm) are associated with gaseous exchanges and passive diffusion of ions while thicker epithelium (10-20 µm)
are mainly concerned with ions transport mechanisms such as sodium potassium-adenosine triphosphate
(Na+K+-ATPase) (Henry et al., 2012). It has been described that the thick hydrophobic cuticle on the surface
body limit the gaseous exchanges and electrolytes. They play a crucial role in gaseous exchanges, respiration,
ions absorption, osmoregulation, acid-base balance, calcium homeostasis, extraction of ammonia and
intracellular pH regulation (Mantel and Farmer, 1983; Pequeux, 1995).
In addition to the gills, the pleopods present in the metasoma area are in continuous movement for the
renewal of the water influx on the surface of the gills leading to the improvement of oxygen absorption
(Sutcliffe, 1984). The diffusion of the oxygen is therefore dependent on the surface and thickness of the gills,
the thickness of the water layer around the gills and the difference in the internal and external pressure on
oxygen (Sutcliffe, 1984). Once the oxygen is taken up, it is trapped by hemocyanin which allow its transport
from the haemolymph to the different organs.
Figure P2. 7. Cross section of (A) Gammarus fossarum illustrating the main organs. (B) midgut and intestinal
caeca and (B) gills of G. fossarum observed with optical microscope at 60x magnification. MG: midgut, IC:
intestinal caeca (Pictures: Chauvière A. and Mehennaoui K.)
Chapter 1 Part 2: G. fossarum as a model organism
67
3.4. Reproduction organs, life cyc le and development
3.4.1. Anatomy of the reproduction organs
Two long tubules compose the reproduction apparel in males and are located above the midgut. They are
divided in three different parts: testicle, seminal vesicle and the different canal. Each canal lead to the
periomere 7 where the genital papillae are located (Figure P2. 8). The female reproductive organs consist of
two ovaries and are also located above the midgut. In periomere 5, an oviduct stands out from each ovary
and reach the ventral side of the G. fossarum body. Each oviduct leads to the osteitis where the embryo
development until hatching of eggs takes place (Figure P2. 8) (Charron, 2014; Felten 2003, Trapp 2015).
Figure P2. 8. Reproduction organs of a male
Gammarus and a female (Trapp, 2015). cmu: mucus
cells, GA: androgen gland, ci: incubation chamber,
mt: non-differentiated mesenchymal tissue, ov1:
primer vitellogenese oocyte, ov2: secondary
vitellogenese oocyte, ovd: oviduct, ovd vst: vestigial
oviduct, spc: spermatocyte, spg: spermatogonium,
spz: spermatozoid, vd: spermiduct
3.4.2. Reproduction and l i fe cycle
Gammarus sp. have a complex life cycle, which is of value in ecotoxicological studies (Kunz et al., 2010).
Generally, Gammarids reproduce throughout most of the year with a peak in spring and early summer (Pöckl
et al., 2003). Females are available for mating only during brief periods directly after moulting while males
are available for mating during almost the entire moult cycle (Sutcliffe, 1993). This results in a male-biased
operational sex-ratio. Therefore, males are usually engaged in precopula mate guarding when meeting a
female close to the moult stage (Ridely 1983). When the females are sexually mature, they release
pheromones in the environment. Males detect these pheromones with the chemoreceptors present on their
antennae (Sutcliffe, 1992). Males grab different females before choosing the one that is likely to produce
more eggs than the others (Kunz et al., 2010). Pairing in Gammarus is size-dependent as bigger males grab
bigger females (Kunz et al., 2010). Males will hold the female under and parallel to their body using the first
Chapter 1 Part 2: G. fossarum as a model organism
68
pair of gnatopodes (Borowsky, 1984; Chevreux et al., 1970). Hence, males present bigger gnatopodes
compared to females. This characteristic is used for gender differentiation in sorting (Figure P2. 9). Female
may produce 6 to 10 broods of juveniles and three generations may be reproducing at the same time in
midsummer (Pöckl et al., 2003; Sutcliffe, 1992). Once the female sheds her skin, male can mate with her and
release her. Female carries the developing eggs in her brooding pouch. Juveniles hatches after 1 to 3 weeks
and remains in the brooding pouch until the next moulting stage. After 4 to 6 weeks juveniles swim out of
the brooding pouch. They feed first by coprophagy and then their diet starts to include conditioned leaves
when they reach 1 month age (McCahon and Pascoe, 1988). Gammarids can live for 1 to 2.5 years (Kunz et
al., 2010).
Figure P2. 9. Gammarus fossarum male and female forming a prepopulate pair (Picture: Untereiner B. and Mehennaoui K.)
3.4.3. Development
Gammarids growing is effective through a successive moult process as the presence of the cuticle does not
allow a continuous growth (Sutcliffe, 1992). A cycle of moulting and inter-moulting steps can take place for
many weeks (~4 weeks) and is dependent on the water temperature (12 °C). One cycle is divided in six steps
• Two pre-moult phases (1-3 days, each) lead to a thickening of the existing cuticle
• Two inter-moult phases (9 days each) lead to a total change of the existing cuticle and they are
characterised by an increase in Gammarus permeability for water (Lockwood and Inman, 1973)
• Two post-moult phases (3 to 5 days each) lead to the formation of the new cuticle
Whereas many crustaceans’ juvenile present larvae stages during their life cycle, Gammarus juvenile are
directly similar to adults. Genital papillae appear in males and osteitis appear in females after the fifth moult
stage. Then, the gnatopodes are differentiated and finally the reproduction organs are well differentiated
0.5mm
Chapter 1 Part 2: G. fossarum as a model organism
69
after 9 to 10 moult stages (Gross et al., 2001). Generally, males have a longer life time than females. Females
can leave to up to 23 months while male can live for up to 30 months (Leroy et al., 2010; Pöckl et al., 2003).
Moulting is a very critical process during Gammarus lifetime. Indeed, cuticle is described to be impermeable
to water and pollutants (Rainbow, 1997, 1995), the inter-moult stage may represent a very critical moment
for the animals. Hence, it had been suggested that Gammarids use more energy for osmoregulation processes
and if needed for defence mechanisms (Lockwood and Inman, 1973).
4. Gammarus sp. ECOLOGY
Abiotic factors such as temperature, salinity, oxygen, acidity and pollution play an important role in the
distribution of Gammarus species (Kunz et al., 2010; Meijering, 1991; Whitehurst and Lindsey, 1990). As
benthic macroinvertebrates and shredders, gammarids are usually found in high abundance under rocks,
gravels, in coarse substrates and among living and dead vegetation (Kunz et al., 2010). These substrates offer
shelter from predators and a supply of organic detritus leading to a high presence of amphipods species,
making them one of the most important component of the aquatic macroinvertebrate assemblage in terms
of density and biomass (Kunz et al., 2010; MacNeil et al., 1997). Gammarus fossarum are usually found in
central and eastern areas in Europe headwaters. They are therefore considered as a typical woodland-brook
element in respect to their feeding habits and their resistance to currents (Meijering, 1991). Hence, it had
been described that G. fossarum could be the first species to be impacted by souring water and pollution as
breeding conditions in running waters can be impacted (Meijering, 1991).
5. FORAGING PLASTICITY
Gammarus play an important role as shredders in aquatic ecosystems and present a very diversified trophic
repertoire. They have been described as herbivores, detritivores and predators (Dangles and Guérold, 2001;
Kelly et al., 2002). Their diet is composed of stream conditioned leaves, biofilms that grow on the leaves,
dead chironomids, live juvenile isopods and even juvenile and wounded/trapped fish (Fielding et al., 2003;
Kelly et al., 2002; MacNeil et al., 1997). Gammarus have also been described as cannibals (MacNeil et al.,
1997). Their success in colonizing very diverse habitats is linked to their foraging plasticity making them of
high ecological value (Kunz et al., 2010; MacNeil et al., 1997). Generally, Gammarus feed of conditioned alder,
elm or maple leaves. The presence of microorganisms will influence the palatability of leaves (MacNeil et al.,
1997). Gammarus will prefer fresh conditioned leaves than sterile old ones in addition to leaves with low
level of tannin and lignin (MacNeil et al., 1997). The fresh conditioned leaves with high presence of fungi
leads to an increase of 30% of digestibility as leaves are partially degraded by microorganisms (Chamier and
Willoughby, 1986). Moreover, Gammarus were described to differentiate conditioned leaves regarding the
species of fungi present on their surface (Graça et al., 1994).
Chapter 1 Part 2: G. fossarum as a model organism
70
6. Gammarus fossarum AS A MODEL ORGANISM IN ECOTOXICOLOGY
Bioindicators have been defined as species that are sensitive to different kind of stressors, give significant
biochemical and physiological responses to anthropogenic pollutants and allow the predictions of the effects
of contaminants on higher level of biological organisations (population, community, ecosystems) (Vellinger,
2012). Gammarids fulfil all these criteria and are therefore increasingly used as bioindicators of the quality of
aquatic ecosystems (Farkas et al., 2003). Additionally, Gammarids have a well-known ecology and sensitivity
to different pollutants (Kunz et al., 2010) and are increasingly used as model organisms in ecotoxicology
studies (Andreï et al., 2016; Arce Funck et al., 2013; Felten et al., 2006; Gismondi et al., 2012a; Vellinger et
al., 2013).
As described above, G. fossarum is considered as an ubiquitous species with a wide distribution all over
Europe (Janetzky, 1994; Karaman, 1977; Živić and Marković, 2007). They are found in high abundance and
high density in many freshwaters. Their wide foraging plasticity offer them a central position in the food web
chain. G. fossarum play also a major functional role in litter breakdown process and nutrient cycling (Forrow
and Maltby, 2000; Kelly et al., 2002; Lacaze et al., 2010; MacNeil et al., 1997). They also play an important
structural role as they represent an important food source for other aquatic species like fish, amphibians and
other invertebrates (Kelly et al., 2002; MacNeil et al., 1997) and they can also be prey for birds like ducks.
Many individuals are parasitized with acanthocephalans which modify their behavior and make them swim
close to the water surface and be more available for the birds the final host of the parasite (Gismondi et al.,
2012a).
In contrast with other macroinvertebrates, Gammarids have a long life cycle, as it could reach 2.5 years.
Therefore, they are able to integrate the different contamination fluxes that occur in the aquatic ecosystems.
Gammarids have also been described as good accumulators of metal contaminations as the internal
concentrations measured within the animal were usually proportional and dependent on the concentrations
present in the environment (Arce Funck et al., 2013; Kunz et al., 2010; Lebrun et al., 2012; Vellinger et al.,
2012).
Gammarids are easy to collect, handle and manipulate in the laboratory and it is possible to collect them in
the field all year long. They are very easy to identify and genders are easy to differentiate. They can be used
for laboratory experiment as well as in situ experiments (Andreï et al., 2016; Besse et al., 2013; Coulaud et
al., 2011; Dedourge-Geffard et al., 2009). G. fossarum is described as a sensitive species to contaminants.
Thus, they are frequently used as bioindicators as their presence or absence in aquatic ecosystems, their
density, population dynamic are good indicators of the quality of the aquatic environment. Furthermore, they
are considered as a good model for integrative studies by enhancing the link between responses observed at
Chapter 1 Part 2: G. fossarum as a model organism
71
the molecular level (omics) to responses observed at the individual and functional level (Andreï et al., 2016;
Lagadic et al., 1994).
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Chapter 2 G. fossarum as a model organism for nano-ecotoxicology
77
CHAPTER 2
Gammarus fossarum (Crustacea Amphipoda) as a model organism for
nano-ecotoxicology
CHAPTER 2
Gammarus fossarum (Crustacea Amphipoda)
as a model organism for nano-ecotoxicology
Chapter 2 G. fossarum as a model organism for nano-ecotoxicology
78
KEY FINDINGS
Aim of the study
G. fossarum was selected to assess effects of well-characterised and well-studied set of AgNPs. Synthetically
produced AgNPs of different sizes (20 and 200 nm) as well as AgNPs synthetized by a biological method (using
plant leaf extracts of Ocinum sanctum and Azadirachta indica, AgNPs 23 and 27 nm, respectively) were used
in order to elucidate the relation between size, synthesis method, NPs surface properties, ions dissolution
and toxicity. A multibiomarker approach was used to investigate the acute effects of these AgNPs on
physiological and behavioural responses of G. fossarum.
Experimental design
Key findings
• Size-dependent effects
o AgNPs 20 nm > AgNPs 200 nm
• Coating-dependent effects
o AgNPs 23 nm > AgNP 27 nm
• High ion release from AgNPs 23 nm compared to AgNPs 20, 27 and 200 nm
• Presence of Ag in gills of G. fossarum exposed AgNPs 23 nm and AgNPs 27 nm
• Significant Ag uptake from AgNO3 and AgNPs 23 nm
• AgNO3 and AgNPs 23 nm led to an osmoregulation impairment and a decrease in G. fossarum
locomotion
• The observed effects seemed to be linked to Ag ions released from AgNPs 23 nm
Chapter 2 G. fossarum as a model organism for nano-ecotoxicology
79
ARTICLE 1
Science of the Total Environment 566, 1649-1659.
Gammarus fossarum (Crustacea, Amphipoda) as a model organism to study silver nanoparticles effects
Kahina Mehennaoui a,b, Anastasia Georgantzopoulou a,d, Vincent Felten b, Jennifer Andreï b, Maël Garaud b,
Sébastien Cambier a, Tommaso Serchi a, Sandrine Pain-Devin b, François Guérold b, Jean-Nicolas Audinot c,
Laure Giambérini b, Arno C. Gutleb a *
a Environmental Research and Innovation (ERIN) Department, Luxembourg Institute of Science and
Technology (LIST), 5 avenue des Hauts-Forneaux, Esch-sur-Alzette, Luxembourg.
b Laboratoire Interdisciplinaire des Environnements Continentaux (LIEC), CNRS UMR 7360, Université de
Lorraine – Metz, France.
c Materials Research and Technology (MRT) Department, Luxembourg Institute of Science and Technology,
(LIST), 5 avenue des Hauts-Forneaux, Esch-sur-Alzette, Luxembourg.
d Norwegian Institute for Water Research (NIVA), Gaustadalléen 21, NO-0349 Oslo, Norway.
We wish to dedicate this manuscript to the late Stephen J. Klaine who inspired the work of our group for many
years.
Gammarus fossarum (Crustacea, Amphipoda) as a model organism tostudy the effects of silver nanoparticles☆
Kahina Mehennaoui a,b, Anastasia Georgantzopoulou a,d, Vincent Felten b, Jennifer Andreï b, Maël Garaud b,Sébastien Cambier a, Tommaso Serchi a, Sandrine Pain-Devin b, François Guérold b, Jean-Nicolas Audinot c,Laure Giambérini b, Arno C. Gutleb a,⁎
a Environmental Research and Innovation (ERIN) Department, Luxembourg Institute of Science and Technology (LIST), 5 avenue des Hauts-Fourneaux, Esch-sur-Alzette, Luxembourgb Laboratoire Interdisciplinaire des Environnements Continentaux (LIEC), CNRS UMR 7360, Université de Lorraine, Metz, Francec Materials Research and Technology (MRT) Department, Luxembourg Institute of Science and Technology, (LIST), 5 avenue des Hauts-Fourneaux, Esch-sur-Alzette, Luxembourgd Norwegian Institute for Water Research (NIVA), Gaustadalléen 21, NO-0349 Oslo, Norway
H I G H L I G H T S
• Gammarus fossarum was used as modelorganism to study silver nanoparticleseffects.
• Four AgNPs were used to link ions re-lease and surface chemistry to the ef-fects.
• AgNPs sub-lethal effects were investi-gated using multibiomarker approach.
• Survival, heamolymph osmolality andlocomotor activity of G. fossarum wereimpacted.
• Ions released by AgNPs dissolutioncould be a main factor of toxicity.
G R A P H I C A L A B S T R A C T
a b s t r a c ta r t i c l e i n f o
Article history:
Received 9 March 2016Received in revised form 10 June 2016Accepted 10 June 2016Available online 18 June 2016
Editor: D. Barcelo
Amphipods are one of the most important components of freshwater ecosystems. Among them, gammarids arethemostwidespread group in Europe and are often used as bioindicators andmodel organisms in ecotoxicology.However, their use, especially of Gammarus fossarum for the study of the environmental impact of nanoparticles,has been rather limited so far.G. fossarum was selected to assess effects of well-characterized chemically synthesized silver nanoparticles(AgNPs 20 nm and 200 nm) and “green” laboratory synthetized (from plant leaf extracts) AgNPs (AgNPs23 nm and 27 nm). AgNO3 was used as a positive control to compare AgNPs effects and silver ions effects. Amultibiomarker approachwas used to investigate the sub-lethal effects of AgNPs on physiological and behaviour-al responses of G. fossarum.Two different experiments were carried out. In a preliminary experiment, two populations of G. fossarum (G.f1and G.f2) were tested for sensitivity differences and the most sensitive one was exposed, in a final experiment,to sub-lethal concentrations of AgNO3 and the most toxic AgNPs. AgNO3 and AgNPs 23 nm led to a significant
Keywords:
Gammarus fossarum
Silver nanoparticlesLocomotor activityOsmoregulationBiomarkers
Science of the Total Environment 566–567 (2016) 1649–1659
☆ We wish to dedicate this manuscript to the late Stephen J. Klaine who inspired the work of our group for many years.⁎ Corresponding author at: Environmental Research and Innovation (ERIN) Department, Luxembourg Institute of Science and Technology, 5, avenue des Hauts-Fourneaux, L-4362 Esch-
sur-Alzette, Luxembourg.E-mail address: [email protected] (A.C. Gutleb).
http://dx.doi.org/10.1016/j.scitotenv.2016.06.0680048-9697/© 2016 Elsevier B.V. All rights reserved.
Contents lists available at ScienceDirect
Science of the Total Environment
j ourna l homepage: www.e lsev ie r .com/ locate /sc i totenv
decrease in survival rates, osmoregulation and locomotor activity. Ag internalisation, performedwith SecondaryIon Mass Spectrometry (SIMS), showed the presence of silver in gills of G.f2 exposed to AgNPs 23 and 27 nm.This study highlighted the influence ofmethod of synthesis on ion release, uptake and toxic effects of AgNPs onG.fossarum. Osmoregulation appeared to be an effective biomarker indicating the physiological health status of G.fossarum. Locomotor activity, which was the most impacted response, reflects the potential effects of releasedions from AgNPs 23 nm at the population level as locomotion is necessary for foraging, finding mates and escap-ing from predators. Therefore, we propose G. fossarum as a suitable model for environmental nanotoxicology,providing information both at individual and population levels.
© 2016 Elsevier B.V. All rights reserved.
1. Introduction
The increasing production of nanomaterials and their ongoing use ina wide range of industries leads to their inevitable release into the envi-ronment (Bhatt and Tripathi, 2011; Klaine et al., 2008). According to theWoodrow Wilson Database, 1814 nanotech consumer products existand 435 of them contain silver nanoparticles (AgNPs) (Vance et al.,2015). AgNPs have distinctive chemical and physical properties suchas catalytic and electronic properties, thermal conductivity and non-lin-ear optical behaviour (Fabrega et al., 2011; Sharma et al., 2014), andthey are widely used as bactericides in a variety of products such as cos-metics, clothing, detergents, water filters, etc. The global annual silverproduction was estimated to be about 27,000 t (US Geological Survey,2016), and due to the increasing usage, large quantities of silver maybe potentially released into the environment and enter aquatic ecosys-tems where Ag and AgNPs persist and bioaccumulate (Fabrega et al.,2011). Sound data on the environmental concentration of AgNPs arestill missing and somemodels calculated predicted environmental con-centrations of AgNPs in water ranging from 0.6 ng·L−1 to 0.32 μg·L−1
(Batley et al., 2013; Sun et al., 2014; Gottschalk et al., 2013). AgNPswere reported to contribute to only 1–15% of the total Ag present inthe environment (Blaser et al., 2008), however, given their expandeduse, AgNPs concentrations are expected to increase in the future.
Despite the growing number of studies on the toxicity of AgNPs,their behaviour, their environmental fate in aquatic environments andthe consequent effects on living organisms are still not fully understood(Foldbjerg et al., 2015; Vale et al., 2016). A few studies investigated andclearly demonstrated that AgNPs are taken up in aquatic organisms suchas molluscs (Ali, 2014; Buffet et al., 2013; Canesi et al., 2012; Ringwoodet al., 2010), crustaceans (Andreï et al., 2016; Georgantzopoulou et al.,2013; Ulm et al., 2015) and fish (Ašmonaite et al., 2016; Griffitt et al.,2013; Jung et al., 2014; Kwok et al., 2012), and behave differently de-pending on their physico-chemical characteristics. In this way, size,coating and chemical composition were suggested as important con-tributing factors to the toxicity of AgNPs (Castranova, 2011; Vale et al.,2016). For instance, laboratory synthesised “coffee-coated” AgNPswere reported to be more toxic to Daphnia magna than commercialPVP-coated and non-coated AgNPs (Allen et al., 2010). Similarly, differ-ent kinds of colloidal AgNPs induced higher mortality in D.magna thanAgNPs powder dispersed by sonication, and their effects were size-de-pendent with the smaller particles being more toxic than the largerones (Asghari et al., 2012).
In freshwater ecosystems, amphipods are an important componentand can be the dominant part of benthic macro-invertebrate assem-blages (MacNeil et al., 1997). Gammarus fossarum is increasingly usedin ecotoxicology as test species due to its wide distribution (BarnardandBarnard, 1983) all over Europe and all over Luxembourg, high abun-dance (Felten et al., 2008a; Kunz et al., 2010), clear sexual dimorphism,easiness of collection and handling and its sensitivity to a large range oftoxicants (Issartel et al., 2005; Xuereb et al., 2011; Bundschuh et al.,2013; Arce Funck et al., 2013; Besse et al., 2013). G. fossarum plays amajor functional role in ecosystems in leaf litter breakdown processes,freshwater food chains and nutrient cycling (Forrow and Maltby,2000; Kelly et al., 2002; MacNeil et al., 1997). Consequently, G. fossarum
could provide valuable information, both in laboratory and in situ exper-iments (Besse et al., 2013; Coulaud et al., 2011, 2014), about the poten-tial effects of environmental contaminants including nanoparticles inaquatic ecosystems. To the best of our knowledge, at present, only onestudy used Gammarus to assess AgNPs effects (Andreï et al., 2016). Cit-rate coated AgNPs (10 nm, 0.5 μg·L−1) impacted the functional role ofG. roeseli after 72 h of exposure, disturbed the production of fine partic-ulate organic matter (FPOM) and significantly decreased locomotor ac-tivity of G. roeseli (Andreï et al., 2016), a species with a very restricteddistribution in Luxembourg. However, there is still a lack of informationabout mechanisms of actions underlying AgNPs effects on Gammarids.Therefore, multibiomarker approaches, including biochemical (e.g. de-fencemechanisms, antioxidant responses), physiological (e.g. osmoreg-ulation) and behavioural (e.g. locomotion) responses, are appropriatetools to investigate sub-lethal effects and determine mechanisms of ac-tions of AgNPs in aquatic invertebrates (Lagadic, 2002; Moore et al.,2004). These biomarkers are known to be early warning indicatorsused as a diagnostic or predictive tool for long-term effects that couldoccur at individual and/or population level (Garaud et al., 2015, 2016;Jemec et al., 2008; Lagadic, 2002; Ulm et al., 2015; Vasseur andLeguille, 2004).
The aim of this study was to investigate in a multibiomarker ap-proach whether G. fossarum could serve as a good model organism inthe field of environmental nanotoxicology, allowing the identificationof potential AgNPs effects from the molecular to the organism level.Chemically synthesised non-coated particles of different initial sizes(AgNPs 20 nm and AgNPs 200 nm) were used to better understandthe link between size and toxicity. In addition, AgNPs synthesised byOcimum sanctum (initial size: AgNPs 23 nm) and Azadirachta indica (ini-tial size: AgNPs 27 nm) plant leaf extract (Balachandran et al., 2012)were used to evaluate the influence of themethod of synthesis and sur-face chemistry on the toxicity of AgNPs. Out of two different populationsof G. fossarum exposed to the selected AgNPs in a first acute experiment,the most sensitive population was then used for further experimentswith lower concentrations. Effects on survival, bioaccumulation, osmo-regulation, antioxidant responses (catalase, glutathione peroxidase,total antioxidant activity), defencemechanisms (glutathione S-transfer-ase, acid phosphatase), cellular damage (lactate dehydrogenase, lipidperoxidation, apoptosis), energy reserves (cholesterol, triglycerides,proteins) and behaviour (locomotor activity and ventilation) wereinvestigated.
2. Organisms, material and methods
2.1. Particles and chemicals
Non-coated AgNPs (20 and 200 nm) were obtained fromPlasmaChem GmbH (Berlin, Germany) and were characterised(Lankoff et al., 2012). A. indica (AgNPs 27 nm) and O. sanctum (AgNPs23 nm) plant leaf extract synthesized AgNPs were provided andcharacterised by the department of Biotechnology, Bharathiar Universi-ty, India (Balachandran et al., 2012, 2013). Silver nitrate (AgNO3,) waspurchased fromVWR (Leuven, Belgium), osmium tetroxide (OsO4), glu-taraldehyde and the epoxy resin embedding kit from Sigma Aldrich
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(Bornem, Belgium). Phosphate buffer saline (PBS) was obtained fromInvitrogen (Merelbeke, Belgium).
2.2. Particle dispersion and characterisation in Volvic® water
2.2.1. AgNPs 20 and 200 nm
Two milligrams (2 mg) of particles were weighed and dispersed in1 mL of a solution containing 5% (v/v) dimethyl sulfoxide (DMSO) inMilli-Q water. The solution was sonicated (cycle 0.5; amplitude 30%)on ice for 3 min using a UP200S probe ultra sonicator with a low diam-eter tip of 2 mm (Hielscher, Germany) resulting in 6.56 J/s of deliveredacoustic power (See Supplementary material for calculations details).Stocks were always prepared freshly before each experiment(Georgantzopoulou et al., 2013; Lankoff et al., 2012). Total dilutions ofDMSO in exposure medium were always in excess of 100.000, leadingto minute DMSO exposure.
2.2.2. AgNPs 23 nm and 27 nm
The NPs were provided as stocks in Milli-Q water and stored at 4 °Cin glass vials. Before use, the vials were gently vortexed. No further son-ication steps were included (Georgantzopoulou et al., 2013).
2.2.3. AgNPs characterisation
AgNPs 20 and 200 nm were characterised using Scanning electronmicroscopy (SEM) type DSM942 (Zeiss, Germany) in the secondaryelectron (SE) mode and transmission electron microscopy (TEM)(JEOL 1200 EXII,JEOL, Japan) (Lankoff et al., 2012).
AgNPs 23 and 27 nmwere characterised usingHigh Resolution Scan-ning Electron Microscope (HRSEM). Images were obtained using SirionHRSEM (FEI Company). Additionally, characterisation and size distribu-tion using TEM were provided (Balachandran et al., 2012). The surfacechemistry of AgNPs 23 and 27 nm were characterised using ShimadzuFTIR-8400S (Fourier Transform Infrared) spectrometer (Balachandranet al., 2012). Results were summarized in our previous study(Georgantzopoulou et al., 2013).
Particle size distribution of the AgNPs in the exposure medium(commercialmineralwater Volvic®, France)was evaluatedusingNano-particle Tracking Analysis (NanoSIGHT, Malvern Instrument Ltd., UK).
2.3. Organism sampling and acclimation
Two different G. fossarum populations were sampled at two riversconsidered to be unpolluted (Arce Funck et al., 2013; Dohet et al.,2008). The first set of animals (G.f1) was collected in June 2013
(experiment 1) at La Maix (48°28′55.7″ N and 07°04′20.1 E,Vexaincourt, Vosges Mountains, North-East France). The second set ofanimals (G.f2)was collected in July 2013 (experiment 1) and September2013 (experiment 2) at Schwaarzbaach (49°48′24.9″N and 06°04′53.2″E, Attert River, Colmar-berg, Luxembourg). Specimens were collectedusing a hand net and they were quickly transported in plastic tankswith river water to the laboratory, where they were kept at 12 °C untilsorting. In order to avoid potential gender and between-life stage influ-ences on the organism sensitivity to the treatments, only adult maleswere selected for the experiments (Arce Funck et al., 2013; Sornom etal., 2010; Sroda and Cossu-Leguille, 2011; Vellinger et al., 2013). Malesfrom precopula pairs were directly identified (males are bigger than fe-males). Other individuals were sexed based on sexual dimorphism suchas gnathopode sizes (males present bigger gnathopodes compared tofemales). As parasitism is known to interferewith the studied responses(Gismondi et al., 2012a, 2012b, 2012c), Acanthocephala sp. parasitizedorganisms (parasites appear as orange spots visible through the cuticle)were excluded from sampling. After selection, non-parasitized maleswere acclimatised for at least 10 days in commercial mineral water(Volvic®, France) under controlled conditions (13.1 ± 0.2 °C with aphotoperiod of 16 h light and 8 h darkness). Volvic® water was gradu-ally added to river water for the first three days of acclimation. Gamma-ridswere then kept for at least 10 days in 100%Volvic®water (Andreï etal., 2016). Individuals were fed ad libitum with alder leaves (Alnus sp.)up to 24 h before experiment. Volvic® water in the tanks was aeratedand changed every 24 h during the acclimation period to avoid organicmatter accumulation and potential increase of ammonium, nitrite andnitrate. All relevant characterization parameters such as mineral con-tent, pH, conductivity, dissolved oxygen, hardness and alkalinity ofVolvic® water were controlled and are presented in Supplementarymaterials (Table S1).
2.4. Experiment 1: acute toxicity test
After the acclimation period, 5 males per treatment group weretransferred into plastic tanks (250 mL polypropylene tanks) containing70 mL of exposure medium (Volvic® water with or without AgNPs orAgNO3). A piece of mesh was added in each tank in order to provide aresting surface for the organisms and to reduce losses due to cannibal-ism. Males were exposed to various treatments: AgNO3 and AgNPs23 nm (0.5–1–2–4–8 μg·L−1), AgNPs 27 nm (1–3–10–30–100 μg·L−1), AgNPs 20 nm and AgNPs 200 nm (10–30–100–300 -1000 μg·L−1), and a control for 72 h at 12 °C with a photoperiod of16 h light and 8 h darkness in the absence of feeding (Fig. 1). The
Fig. 1. Experimental design (adapted from Arce Funck et al., 2013). CAT: Catalase; GPx: Glutathione Peroxidase; TAC: Total Antioxidant Activity; ACP: Acid Phosphatase; GST: GlutathioneS-Transferase; CHOL: Cholesterol; TRIG: Triglyceride: PROT: Proteins: ETS: Electron Transport System; LDH: Lactate dehydrogenase; LPO: Lipid peroxidation; CASP-3: Caspase 3.
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different tested range of concentrations were selected based on previ-ous results obtained after exposing D. magna to the same AgNPs(Georgantzopoulou et al., 2013). AgNO3 and AgNPs stock solutionswere diluted to the desired nominal concentrations in mineral water(Volvic®, France). Exposure medium was renewed daily (Volvic®water with or without AgNPs or AgNO3 at the same concentrationslisted above; no aeration). Dead animals were removed and survivorswere counted. Each treatmentwas performed in 4 replicates. Acute tox-icity of each AgNPs was assessed using LC50 values.
2.4.1. Total and dissolved Ag release
At the end of the exposure period, 5 mL of exposure medium(Volvic® water + AgNPs at the lowest and highest working concentra-tions where at least one animal survived)were kept for total silver con-centrations measurements.
In order to assess the ion release for each class of particles, 4 mL ofexposure medium (Volvic® water + AgNPs at the lowest and highestworking concentrations where at least one animal survived) were cen-trifuged at the end of the 72 h exposure period at 4000 g for 40 minusing centrifugal filter devices (Amicon ultra-4, Milli-pore, Ireland)with a 3 kDa cut-off level. Total (exposure medium) and dissolved(lower fraction) Ag were quantified using Inductively Coupled PlasmaMass Spectrometry (ICP-MS) (Elan DRC-e, Perkin-Elmer, Waltham,MA, USA) as previously described (Boscher et al., 2010;Georgantzopoulou et al., 2013). All measurements were performed inthe absence of animals. Percentages of dissolution were calculatedbased on total silver measured concentrations. The quantification limitof ICP-MS was around 0.051 μg·L−1.
2.4.2. Particle uptake in gills of G. fossarum
Gills of G. fossarum are one of the first organs that are directly ex-posed to contaminants in the water (Felten et al., 2008b; Henry andWheatly, 1992; Issartel et al., 2010; Lignot et al., 2000; Pequeux, 1995)and were therefore selected for AgNPs uptake evaluation.
At the end of the 72 h exposure period, gammarids' gills were har-vested and fixed overnight with 5% glutaraldehyde in Phosphate BufferSaline (PBS) at 4 °C. Glutaraldehyde was then removed and sampleswere washed in PBS. After that, gills were embedded in small 1% agarcubes (adapted from Georgantzopoulou et al., 2013). The gills werepost-fixed with 1% osmium tetroxide in Milli-Q water for 1 h at roomtemperature. After an additional washing step with PBS, samples weredehydrated with five increasing acetone concentrations (30, 50, 70and 100% v/v in Milli-Q water) and were embedded in epoxy resin(Epon 812 substitute). Samples were cut to 300 nm semi-thin sections(Leica ultracut UCT, LePeq Cedex, France) and finally placed on siliconwafers (Siltronix, Archamps, France) for Secondary Ion Mass Spectrom-etry (SIMS) analysis (Georgantzopoulou et al., 2013).
Sampleswere analysedwith a NanoSIMS 50 (Cameca, Gennevilliers,France) using Cs+ as primary ion source (8 KeV) sputtering the surfaceof the sample (−8 KeV) with a raster of 40 × 40 μm2 to generate sec-ondary negative ions. Images were recorded in a pixel format of256 × 256 image points with a counting time of 30ms per pixel. The in-strument was tuned for a mass resolution (M/ΔM) up to 5000 for theelimination of atomic or molecular isobar interference. The mass cali-bration of the silver ion was carried out using a silver foil (Goodfelow,Huntingdon, UK). The isotopic ratio between 107Ag− (m = 106.9051,51.8%) and 109Ag− (m = 108.9048, 48.2%) was measured and verified(ratio = 1074). The 12C14N− cluster was simultaneously detectedwith the silver ions to allow the recognition of the essential anatomicalfeatures (Eybe et al., 2009; Georgantzopoulou et al., 2013).
2.5. Experiment 2: physiological and behavioural responses
The G.f2 population, which was slightly more sensitive to AgNPs,was used for the second experiment in which the sub-lethal effects of
AgNPs that had showed a toxic effect in the first experiment (AgNPs20 nm, AgNPs 23 nm and AgNPs 27 nm) were assessed (Fig. 1).
Groups consisting of 15 individuals were housed in plastic tanks(one 500 mL polypropylene tank for each group) containing 210 mLof exposure medium (Volvic®, France). The same parameters used forexperiment 1 (exposure time, temperature and photoperiod) were ap-plied for the setup of experiment 2. Two concentrations of AgNO3 andfor each AgNPs (1 and 3 μg·L−1) were used. The group treated withonly Volvic® water represented the negative control group. The choiceof 1 and 3 μg·L−1 was based on the results obtained during the first ex-periment as they were lower than the LC5 of AgNPs 23 nm, which re-sulted as the most toxic AgNPs among those analysed in this work.Every 24 h, dead animals were removed from each tank and livingones were counted. Each treatment was performed in 3 replicates.
2.5.1. Ag bioconcentration
After 72 h of exposure, a pool of 4 gammarids per condition wererinsed with Milli-Q water and gently dried on filter paper. Animalswere weighed and stored at −20 °C until analysis. 3 mL of mass gradeMilli-Q water supplemented with 30% (v/v) of HNO3 and 14% (v/v) ofH2O2were added to each pool followed bymineralisation under a max-imum pressure of 35 bars andmaximum temperature of 200 °C in ami-crowave (Anton Paar Multiwave Pro). After mineralisation, total Agcontentwas quantified using ICP-MS following the procedure describedelsewhere (Georgantzopoulou et al., 2013). Bioconcentration factors(BCF) were calculated based on the internal Ag concentrations in gam-marids andmeasured total Ag concentration in the exposuremediumatthe end of exposure period (BCF= [Ag]organism in μg·kg−1 - [Ag]control inμg·kg−1/[Ag]water in μg·L−1).
2.5.2. Locomotor activity
For each treatment, measurements of locomotor activity, ventilationand haemolymph osmolalitywere performed on the same pool of 10 in-dividuals. At the end of the exposure period, locomotor activity wasfirstly assessed by counting the number of animals in movement in a80 mL glass tank containing 10 organisms with a piece of net added toprovide a resting surface (Felten et al., 2008b). Measurements weredone after 5 min of acclimation at the same time of day with similarlight conditions and in a quiet environment. Moving G.f2were countedfor a period of 2 s and this process was repeated 40 times.
2.5.3. Ventilatory activity
Ventilatory activity was recorded immediately after locomotor ac-tivity measurements on the same animals (G.f2), by measuring the fre-quency of pleopod beats (ventilator appendages of malacostraceancrustaceans). Ten gammarids from each treatment group were placedindividually in a glass tube containing Volvic® water and left for a 30 sacclimation period. Then, pleopod beats were visually counted threetimes for 10 s using a manual cell counter only when animals were atrest. Measurements were performed at the same period of the day toavoid possible effects of a circadian rhythm on respiration (Rosas etal., 1992).
2.5.4. Haemolymph osmolality
At the end of the exposure period and immediately after the mea-surement of locomotor and ventilator activities, the organisms wereused for the measurement of haemolymph osmolality. Prior to thehaemolymph extraction, animals (G.f2) were gently dried betweentwo pieces of filter paper to remove the excess of exposure media.Haemolymph samples were collected from the telson using a modifiedmicrocapillary (Felten et al., 2008b). A drop of haemolymph (≈20 nL)was transferred into mineral oil to avoid evaporation. Osmolality wasmeasured using an Otago nanolitre osmometer (Otago OsmometersLtd., Dunedin, New Zealand) using a 300 mOsm·kg−1 standard(Felten et al., 2008b).
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2.5.5. Energy reserves, detoxification and lipid peroxidation (LPO) level
measurements
Five pools of five gammarids each (G.f2) exposed to 1 and 3 μg·L−1
of AgNO3 and AgNPs, as well as the non-treated negative control, wereweighed and frozen in liquid nitrogen and stored at−80 °Cuntil furtheranalysis. Each pool was homogenised as described previously (Srodaand Cossu-Leguille, 2011). Briefly, each pool was grindedwith amanualpotter Elvejheim tissue grinder at 4 °C in homogenisation buffer(50 mM phosphate buffer) (pH 7.6) supplemented with 1 mM L-ser-ine-borate mixture and 1 mM phenylmethylsulfonyl fluoride (PMSF)as protease inhibitors, adjusted to the weight of the pooled tissues(ratio 4/1, v/w). Homogenates were centrifuged at 250 g for 5min to re-move cuticle fragments and 112 μL of the resulting supernatant (S1)were kept for biomarker measurements. The remaining supernatantwas centrifuged at 1000 g for 20min at 4 °C. The supernatantwas recov-ered and centrifuged again at 20,000 g for 50 min at 4 °C. The final su-pernatant, corresponding to the cytosolic fraction (S2), was used forenzyme activities measurement.
Acid phosphatase (ACP), lipid peroxidation (LOOH), lactate dehy-drogenase (LDH), electron transport system (ETS) and energy reserveswere measured on the S1 fractions, while total antioxidant capacity(TAC), catalase (CAT), glutathione peroxidase (GPxtot), and glutathioneS-transferase (GST), caspase-3 (CASP-3) weremeasured on the cytosol-ic fraction (S2) by an automated colorimetricmethods developed on theKonelab 20-XTi (Thermo Scientific) (Garaud et al., 2015, 2016). All tech-nical procedures are described in detail in the Supplementary material.
2.6. Statistical analysis
LC50 values were calculated on nominal concentrations based on alogistic curve-fitting procedure applying the method described earlier(Isnard et al., 2001; Vindimian et al., 1999). The Excel macro REGTOXallowed us to estimate both LC50 values using Hill model (Vindimianet al., 1983) and their 95% confidence intervals (Vellinger et al., 2012).95% confidence intervals were considered for identifying significant dif-ferences existing between Gammarus populations.
All other results are reported as mean ± SD. Total and dissolved Agconcentrations in exposure medium were compared using two-wayANOVA, with exposure conditions and Gammarus population as factors,followed by Tukey HSD post hoc test for multiple comparisons. Thecomparison of each biomarker response in tested conditions was per-formed using one-way ANOVA (if normality and homoscedasticitywere identified using Shapiro-Wilk and Levene tests, respectively)followed by appropriate post hoc test formultiple comparisons. If signif-icant heterogeneity among tested conditions (P b 0.05) had been iden-tified, non-parametric Kruskal-Wallis test for multiple independentgroups followed by Mann-Whitney U test were performed. An arcsinesquare root transformationwas doneon responses expressed in % (mor-tality and locomotion) before applying statistical tests to normalise thedata. All statistical analyses were performed using Statistica 7.0 soft-ware (Statsoft Inc).
3. Results
3.1. Particle characterisation
Nanoparticles tracking analysis (NTA) showed that the studiedAgNPs of the nominal size of around 20 nm agglomerated in Volvic®water (Table 1, Fig. S1). Zeta (ζ) potential ranged from −3.5 to −4.5,suggesting unstable dispersion of AgNPs in Volvic® water (Table 1).
3.2. Experiment 1: acute toxicity test and particle uptake
In general, AgNO3 and AgNPs 23 nm were the most toxic for bothpopulations with AgNO3 having the lowest LC50 values followed byAgNPs 23 nm, AgNPs 27 nm and AgNPs 20 nm (Table 2). No LC50
value could be recorded for AgNPs 200 nm within the tested range ofconcentrations (10–1000 μg·L−1). The LC50 calculated for G.f2was sig-nificantly lower than the LC50 determined for G.f1 (no overlap in 95%confidence intervals) when they were exposed to AgNO3, AgNPs23 nm and AgNPs 27 nm. G.f2 was thus considered to be the slightlymore sensitive population and was therefore selected for furtherexperiments.
Uptake study, performed by NanoSIMS 50 analysis, showed no pres-ence of silver in the gill tissues of control gammarids (Fig. 2D).Internalisation of Ag within gill tissues was observed in G. fossarum ex-posed to AgNPs 23 nmand AgNPs 27 nm (fig. 2A and B). AgNO3 seem tobe present only in the external membrane of G. fossarum gills (Fig. 2C).
Dissolution of ions from AgNPs was below the limit of quantificationat the lowest exposure concentrations tested, therefore only concentra-tions from the high exposure can be reported.
The dissolved Ag concentrations measured in Volvic® water at theend of exposure period for AgNPs 23 nm were at the same level as forAgNO3 (~0.5 μg·L−1). The percentage of dissolution of AgNPs 23 nmwas about 30% (G.f2), while lower levels were found for all the otherAgNPs (between 0.3 and 4.4%) (Table 3).
3.3. Experiment 2: physiological and behavioural responses
3.3.1. Survival
After 72 h of exposure, survival rate in controls remained high andabove 95% (Fig. 3). Survival rates of G.f2 were significantly reduced(one-wayANOVA, F=9.470, 8 df, P b 0.0001)when organismswere ex-posed to the highest concentration (3 μg·L−1) of AgNO3, reaching 28±30% compared to the untreated control. No significant effects on surviv-al were recorded for G. fossarum exposed to AgNPs 23 nm, AgNPs 27 nmand AgNPs 20 nm (P N 0.05).
3.3.2. Ag bioconcentration
ICP-MS analysis of internal Ag+ concentrations in control gamma-rids revealed that non-treated animals contained already0.274 μg·g−1 of Ag in their body (Table. 4). 72 h of direct exposure to1 μg·L−1 of AgNO3 led to a 4-fold increase of Ag concentration in thewhole body of gammarids compared to controls.
Internal Ag concentration in G. fossarum exposed to AgNPs 20 nmand 1 μg·L−1 of AgNPs 23 nm were not different from those measuredin negative control animals. Internal Ag concentration of gammarids ex-posed to 3 μg·L−1 of AgNPs 23 nm was increased by 24 and 12-fold
Table 1
Size distribution of particles (mode± SD, 3 replicates) and ζ potential of AgNPs in Volvic®water.
Particle Size (nm) Zeta potential (mV)
AgNPs 23 nm 56 ± 3.2 −3.49AgNPs 27 nm 63 ± 3.2 −3.55AgNPs 20 nm 69 ± 8.4 −3.49AgNPs 200 nm 135 ± 18.7 −4.55
Measurements were performed in Volvic® water at 12 °C, pH 7.7. Particle concentrationused was 100 μg·L−1 (detection limit of NTA).
Table 2
LC50 values with 95% confidence intervals for G.f1 and G.f2 exposed to AgNO3 and AgNPs20 nm, AgNPs 23 nm, AgNPs 27 nm and AgNPs 200 nm for 72 h.
Compound G.f1 G.f2
LC50-72 h (μg·L−1) LC50-72 h (μg·L−1)
AgNO3 3.9 (3.3–4.3) 2.3 (2.0–2.6)AgNPs 20 nm 835 (590–988) N1000AgNPs 23 nm 7.7 (6.7–8.6) 4.9 (4.2–5.5)AgNPs 27 nm N100 5.5 (3.9–8.2)AgNPs 200 nm N1000 N1000
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compared to control and animals exposed to 1 μg·L−1 of AgNPs 23, re-spectively. Ag concentration of gammarids exposed to 3 μg·L−1 ofAgNPs 27 nm were increased by 28 and 2.7-fold compared to controlanimals and those exposed to 1 μg·L−1 of AgNPs 27 nm, respectively.AgNPs 23 nm had the highest bioconcentration factor (BCF = 4238)in G.f2 followed by AgNPs 27 nm (BCF = 2382), AgNO3 (BCF = 2277)and AgNPs 20 nm (BCF = 147) (Table 4).
3.3.3. Haemolymph osmolality
The basal level of haemolymph osmolality in controls G.f2 reached295 ± 13 mOsm·kg−1 (Fig. 4). AgNO3 (1 and 3 μg·L−1) and AgNPs23 nm (3 μg·L−1) significantly decreased haemolymph osmolality ofG.f2 (one-way ANOVA, F = 3.466, 8 df, P b 0.001) reaching 258 ±31mOsm·kg−1, 261 ± 24 mOsm·kg−1 and 275 ± 53 mOsm·kg−1 re-spectively. Osmolality in G. fossarum exposed to AgNPs 20 nm, 27 nmand 1 μg·L−1 of AgNPs 23 nm was not affected.
3.3.4. Energy reserves, antioxidants, defence mechanisms, cellular damage
and behavioural responses
Due to the high mortality rate recorded for G. fossarum exposed to3 μg.L−1 of AgNO3, antioxidant responses, defencemechanisms, cellulardamage and energy reserve measurements could not be performed forthis condition.
Antioxidant responses, defence mechanisms, cellular damage, ener-gy reserves (Table 5) and ventilatory activity (Fig. 5a) were not impact-ed by 1 μg·L−1 AgNO3 and 1 and 3 μg·L−1 AgNPs treatments (one-wayANOVA, F = 0.921, 8 df, P N 0.05).
Locomotion was significantly decreased by 20.3% and 33.9% (one-way ANOVA, F = 18.705, 8 df, P b 0.0001) when gammarids were ex-posed to 1 μg·L−1 and 3 μg·L−1 of AgNO3, respectively and by 17.7%at 3 μg·L−1 of AgNPs 23 nm. Locomotor activity of G. fossarum exposedto AgNPs 20 nm and AgNPs 27 nm was not affected (Fig. 5b).
4. Discussion
4.1. Experiment 1: acute toxicity test
The observed mortality rates in G. fossarum exposed to AgNO3 andAgNPs indicates that the observed effects are concentration-dependent.AgNO3 appeared to be themost toxic form of silver for both G. fossarumpopulations. The recorded LC50-72 h values were in the same rangethan those previously reported for Gammarus pulex and G. fossarum(Bury et al., 2002; Arce Funck et al., 2013). The G. fossarum populationused in this study was the same one used in Arce Funck et al. (2013)study (collected at la Maix).
Generally, the order of toxicity of all AgNPs was similar for both G.fossarum populations with AgNPs 23 nm being the most toxic onefollowed by AgNPs 27 nmand AgNPs 20 nm. Nomortality was observedfor both G. fossarum populations exposed to AgNPs 200 nm. The sameorder of toxicity of AgNPs for Vibrio fischeri, Desmodesmus subspicatus,D. magna (Georgantzopoulou et al., 2013) and in-vitro co-culturemodel for gastrointestinal epithelium (Georgantzopoulou et al., 2016)was reported. Ag 200 nm appeared the be the least toxic particle forall themodels tested andmultivariate analysis had shown that this par-ticle always clusters togetherwith control groups (Georgantzopoulou etal., 2016).
G.f2 appeared to be the slightly more sensitive of the two selected G.fossarum populations. The different LC50s obtained between the twopopulations could be linked to the presence of different cryptic species(Feckler et al., 2014;Weiss et al., 2014), particular life-history traits spe-cific to each population (e.g. size, reproduction, feeding habits, etc.)
Table 3
Total and dissolved Ag concentrations (mean± SD) and dissolution rates (mean± SD) ofAgNO3 and AgNPs after 72 h of exposure in Volvic®water.
Compound Nominal[Ag]
(μg·L−1)
Gammarus
populationTotal [Ag](μg·L−1)
Dissolved[Ag]
(μg·L−1)
Dissolution(%)
Control 0 G.f1 0.08 0 0 a
G.f2 0 0 0 a
AgNO3 2 G.f1 0.6 ± 0.2 0.5 ± 0.1 DNAG.f2 1.1 ± 0.1 0.4 ± 0.1 DNA
AgNPs 23nm
4 G.f1 1.2 ± 0.1 0.3 ± 0.1 25.1 ± 2.6b
G.f2 1.8 ± 0.2 0.5 ± 0 29.1 ± 1.0b
AgNPs 27nm
100 G.f1 61.3 ± 1.2 2.7 ± 0.1 4.4 ± 0.2a
G.f2 Na Na Na
AgNPs 20nm
300 G.f1 42.2 ± 4.1 1.8 ± 0.1 3.2 ± 1.9a
G.f2 20.4 ± 2.5 0.6 ± 0.1 0.3 ± 0.5a
AgNPs 200nm
1000 G.f1 26.7 ± 3.7 3.7 ± 0.1 1.9 ± 0.1a
G.f2 20.4 ± 2.2 0.8 ± 0.1 3.9 ± 0.4a
Different letters illustrate significant differences between different treatments (Two-wayANOVA + Tukey HSD post hoc test at P b 0.05 level of significance, n = 3), Na: NotAnalysed due to a high mortality rate, DNA= Do not apply.
Fig. 2. Elemental distribution of 12C14N− cluster and 109Ag− ion in 300 nm cuts of gills. G.fossarum were exposed to AgNPs (A) 23 nm, (B) 27 nm, (C) AgNO3 and (D) Control(Volvic® water). Scale bar is 5 μm.
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(Lambrechts et al., 1997; Usseglio-Polatera et al., 2000) as well as phys-iological parameters (e.g. life cycle) (McCahon andPascoe, 1988; Alonsoet al., 2010; Vellinger et al., 2012; Arce Funck et al., 2013), samplingdates (June vs July), water quality in the field and food availability(Heugens et al., 2006).
The differences in toxicity observed for the different AgNPs (20 nmand 200 nm obtained by chemical synthesis and AgNPs 23 nm and27 nm obtained by “green” synthesis) could be partially explained bytheir different dissolution in exposuremedium. Indeed, ICP-MS analysisrevealed that AgNPs 23 nm dissolved more in exposure medium thanAgNPs 27 nm, AgNPs 20 nm and AgNPs 200 nm. These results are in ac-cordance with those reported previously after exposingD.magna to thesame particles (Georgantzopoulou, 2015). In this previous study theamount of dissolved Ag from AgNPs 23 nm was about 30% which fitsour results exactly. However, the observed effects could not only belinked to the effects of Ag ions as the observed effects for AgNPs23 nmwere lower than those observed for AgNO3with comparable dis-solved Ag concentrations in Volvic® water. The differences in effectscould be explained by different factors such as the size, surface chargeand specific molecules present on the surface of AgNPs. As previouslydescribed, AgNPs 23 nm and AgNPs 27 nm are synthetized using plantleaf extract. Therefore, bio-molecules (e.g proteins, flavonones, terpe-noids and sugars) are present on their surface (Balachandran et al.,2013; Georgantzopoulou et al., 2013). However, an additional com-pound was detected at the surface of AgNPs 27 nm (germinal methylor symmetric stretching of carbonyl group) making them more stablein exposuremedium (Georgantzopoulou et al., 2013) thereby potential-ly reducing their toxicity (Kroll et al., 2014). Uptake evaluation per-formed with NanoSIMS 50 showed the presence of Ag inside the gillsof gammarids exposed to AgNPs 23 nm and AgNPs 27 nm, whereas Agin AgNO3 exposed animals was only present in the external membraneof the gammarids' gills. Thus, it can be hypothesised that AgNPs 23 nmand AgNPs 27 nm have a different or an additional uptake route com-pared to AgNO3 and AgNPs 23 nm may readily release the ions in thecells after uptake (Georgantzopoulou, 2015).
4.2. Experiment 2: physiological and behavioural responses
4.2.1. Survival rates and Ag bioconcentration
This experiment showed that exposing G. fossarum to AgNPs and toAgNO3 led to a significant uptake of Ag. Increased internal concentra-tions of Ag are associated to decrease of survival rates of G.f2 withAgNO3 being the most toxic, followed by AgNPs 23 nm, AgNPs 27 nm
and AgNPs 20 nm. Additionally, AgNPs 23 nm also showed the highestpotential for bioaccumulationwith the highest bioconcentration factors,followed by AgNPs 27 nm and AgNPs 20 nm. These results are in accor-dance with previous studies that reported an accumulation of water-borne Ag in crustaceans, mainly in hepatopancreas and gills (Bury etal., 2002; Grosell et al., 2002; Arce Funck et al., 2013), which increasedthe mortality rate. However, despite a high amount of Agbioaccumulated in G. fossarum exposed to AgNPs 27 nm, this did notcause an increase inmortality rate or any other disturbances. The differ-ences in toxicity observed between the similar sized AgNPs 23 nm andAgNPs 27 nm can thus be linked to AgNPs dissolution and silver ions re-lease from AgNPs 23 nm (Georgantzopoulou, 2015). It could also besuggested that proteins released by G. fossarum in exposure mediummight increase the toxicity of AgNPs 23 nm by the formation of aneco-corona (Nasser and Lynch, 2016) due to observed differences inthe surface molecules of the native Ag particles (Georgantzopoulou etal., 2013). Indeed, it had been shown that the presence of proteins ofD.magna in exposure medium increased AgNPs uptake by D.magna ne-onates thus leading to a higher retaining dose of AgNPs and thus highertoxicity (Nasser and Lynch, 2016).
Fig. 3.G.f2 survival rates (Mean±SD) after 72 h of exposure to Ag, AgNPs 23nm, AgNPs 27nm, AgNPs 20nm. Datawere arcsine square root transformed. Letters (a–c) illustrate significantdifferences (One-way ANOVA + Fisher LSD post hoc test at P b 0.05 level of significance, n = 15).
Table 4
G.f2AgBioconcentration (Mean±SD) after 72 h exposure to AgNO3, AgNPs 20 nm, AgNPs23 nm, AgNPs 27 nm.a
Conditions Nominal[Ag]
(μg·L−1)
[Ag] inexposuremedium(μg·L−1)
Gammarusinternal [Ag]
(μg·ggammarids−1)
BCF (μg·kggammarids−1/μg·L−1)
Control 0 bLOQ 0.27 ± 0.02a NDAgNO3 1 0.36 ± 0.03 1.09 ± 0.010a 2277
AgNPs 20nm
1 0.23 ± 0.03 0.70 ± 0.003c 18693 0.68 ± 0.03 0.37 ± 0.001a 147
AgNPs 23nm
1 0.44 ± 0.05 0.54 ± 0.005a 6133 1.47 ± 0.01 6.50 ± 0.003b 4238
AgNPs 27nm
1 0.96 ± 0.04 2.83 ± 0.003a 26663 3.16 ± 0.10 7.80 ± 0.009b 2382
Different letters illustrate significant differences between different treatments (Kruskal-Wallis ANOVA + Mann-Whitney U test at P b 0.05 level of significance, n = 3). LOQ:limit of quantification.
a Due to an unexpectedmortality during sorting and acclimation period in addition to ahigh mortality rate during exposure, Ag bioconcentration in Gammarus exposed to 3μg·L−1 of AgNO3 could not be measured.
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4.2.2. Haemolymph osmolality
Freshwater organisms are dependent on osmoregulation. Absorptionand excretion of ions is needed to maintain their ion homeostasis. Inthis study, heamolymph osmolality of G.f2 exposed for 72 h to 1 and3 μg·L−1 of AgNO3 and 3 μg·L−1 of AgNPs 23 nm was significantly re-duced in addition to a significant higher mortality rate. This result wasin accordance with a previous study, in which a significant decrease ofheamolymph [Na+] and [Cl−], due to gill iono/osmoregulation failure,was observed in G. fossarum exposed for 96 h to 0.5 μg·L−1 of AgNO3
(Arce Funck et al., 2013). Ag+was described to act by disturbing the bran-chial ion transport pathways infish, leading to an ion influx inhibition anda significant loss of Na+ and Cl− across the gill (Morgan et al., 1997). Con-sequently, osmoregulation failure observed in gammarids exposed toAgNO3 and AgNPs 23 nm might be explained by the loss of Na+ andCl− ions through the gills. In crustaceans, Ag+was described as a compet-itor of Na+ for the same transport site, attributable to a similar radius be-tween the two ions (Grosell et al., 2002; Bianchini and Wood, 2003;Völker et al., 2013). Therefore, it could be assumed that the observedAg+ toxicity is related to the inhibition of Na+ uptake.
4.2.3. Defence mechanisms and energy reserves
In general, one of the most widely accepted toxicity mechanismsproposed for AgNPs is through the generation of ROS (Klaine et al.,2008; Vale et al., 2016), which can lead to lipid peroxidation inducing,finally, oxidative stress (Ulm et al., 2015). In this study, despite a
significant bioconcentration of Ag in gammarids exposed to AgNO3,AgNPs 23 nm and AgNPs 27 nm, and their presence within G. fossarumgills tissue, no significant alterations in antioxidant responses, defencemechanisms, cellular damage and energy reserves could be detected.These results are consistent with those reported by Arce Funck et al.(2013) who observed no effects at 0.5 and 1 μg·L−1 of AgNO3 on cata-lase and selenium-dependent GPx activities of G. fossarum. However,unlike our study, significant lipid peroxidation was observed and wasdiscussed to potentially result from an overwhelming of the defencemechanisms activities (Arce Funck et al., 2013). In our study, the ab-sence of effect could be explained by the fact that the basal levels of en-zyme activities were maybe effective enough to cope with potentialincreased ROS levels generated by the AgNPs.
4.2.4. Behavioural responses
In the current study, we could observe a significant decrease in loco-motor activity after 72 h of exposure to AgNO3 and AgNPs 23 nm (1 and3 μg·L−1) while no observable effect on locomotor activity of G.fossarum treated with AgNPs 20 nm and AgNPs 27 nm could be detect-ed. These results were in accordance with a study which showed that0.5 μg·L−1 of citrate-coated AgNPs significantly decreased locomotoractivity of G. roeseli after 72 h of exposure (Andreï et al., 2016). The ef-fects were size-dependent with the 10 nm citrate coated AgNPs beingmore potent in inhibiting locomotor activity than 60 nm citrate coatedAgNPs (Andreï et al., 2016). In our study, the results indicate that theimpact of AgNPs on behavioural responses of G. fossarummay be linked
Fig. 4. G.f2 haemolymph osmolality (Mean ± SD) after 72 h exposure to AgNO3, AgNPs 20 nm, AgNPs 23 nm and AgNPs 27 nm. Different letters illustrate significant differences betweendifferent treatments (One-way ANOVA + LSD Fisher post hoc test at P b 0.05 level of significance, n = 15).
Table 5
Mean values (±SD) of biomarkers measured in Gf2 exposed for 72 h to AgNO3, AgNPs 20 nm, AgNPs 23 nm and AgNPs 27 nm.
Biomarkers Energy reserves Defence mechanisms Antioxidant responses Cellular damage
Conditions(μg·L−1)
Prota Chola Tria ETSb ACPc GSTd GPxe TACf CATg LDHh CSP3i LOOHj
Control 0 21.7 ± 3.6 0.6 ± 0.1 1.8 ± 0.4 1.3 ± 0.1 89.7 ± 15.2 190.1 ± 31.8 68.3 ± 6.3 36.5 ± 2.2 17.5 ± 7.9 1047.4 ± 411.1 1.5 ± 0.7 1511.9 ± 350.9AgNO3 1 24 ± 6.5 0.7 ± 0.2 2 ± 1.1 1.2 ± 0.2 85 ± 7.1 156.1 ± 51.4 62.3 ± 12.2 29.9 ± 6.9 15 ± 6.1 1093.3 ± 327.5 1.2 ± 0.3 1208.2 ± 445.1AgNPs20 nm
1 22.8 ± 5.9 0.7 ± 0.2 1.9 ± 0.7 1.3 ± 0.2 87.3 ± 10.5 154 ± 58.8 61.1 ± 12.5 32.6 ± 7.5 16.2 ± 8.4 1093.4 ± 276.1 1.4 ± 0.9 1577.1 ± 6203 24.4 ± 1.4 0.8 ± 0.1 2.1 ± 0.5 1.3 ± 0.1 89.4 ± 13.7 147.5 ± 36.8 74.9 ± 2.5 33.6 ± 4.7 19.1 ± 8.5 1238 ± 167.5 1.4 ± 0.4 1308.8 ± 225.8
AgNPs23 nm
1 24.6 ± 2.6 0.7 ± 0.1 1.2 ± 0.3 1.1 ± 0.1 85 ± 7.9 154.1 ± 28.9 66.6 ± 9.3 31.1 ± 6.5 20 ± 4.6 1090.5 ± 306.6 1.3 ± 0.1 1110.6 ± 178.33 23.6 ± 4.1 0.7 ± 0.2 1.7 ± 0.5 1.2 ± 0.1 84.7 ± 3.3 146.2 ± 39.8 59.1 ± 12.3 27.9 ± 3.6 14.7 ± 5.7 963.9 ± 278.4 1.1 ± 0.3 1513.3 ± 566.5
AgNPs27 nm
1 24.1 ± 2.8 0.7 ± 0.1 1.7 ± 0.7 1.2 ± 0.1 87.8 ± 6 146.8 ± 28.1 65.8 ± 14.8 34.6 ± 9.7 19.6 ± 9.4 1056.8 ± 213.3 1.3 ± 0.4 1199.2 ± 204.33 24.6 ± 3.3 0.7 ± 0.1 2.4 ± 1.1 1.24± 0.1 82.9 ± 11.4 170.6 ± 30.5 71.7 ± 8 31.7 ± 6.3 14.6 ± 3.2 1336.1 ± 200.7 1.3 ± 0.2 1241.2 ± 156.6
Biomarkers in italic were analysed using One-way ANOVA and Tukey post hoc test, whereas the others were analysed using Kruskal-Wallis ANOVA andMann-Whitney U test. No signif-icant differences were detected for any of the exposure conditions.a:mg·g freshweight−1; b: μmol O2·g proteins
−1·h−1; c: μmol p-nitrophenol·g protein−1·h−1; d: μmol CDNB·min−1·g proteins−1; e: μmol NADPH·g protein−1·min−1; f: mmol Troloxequivalent·g protein−1; g: mmol H2O2·g proteins−1·min−1; μmol NADH·g proteins−1·h−1; h: μmolpNA·g proteins−1·h−1; j: nmol TBH·g proteins−1.
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to the surface chemistry of AgNPs and eco-corona (Nasser and Lynch,2016). Indeed, authors reported that the formation of eco-corona onAgNPs affected the feeding ability of D. magna and could thereforelead to multi-trophic disturbances in aquatic ecosystems (Nasser andLynch, 2016). Besides the potential direct effect of AgNPs, the observeddecrease in locomotor activity could also be linked to an energy reallo-cation or preservation in favour of maintenance mechanisms. Otherstudies showed that the impairment of iono/osmoregulation of animalstreated with Cd or under acidic stress, led to a reduction of energy allo-cation to locomotor activity of G. pulex (Felten et al., 2008a, 2008b).Therefore, osmoregulation impairment of G.f2 exposed to AgNO3 andAgNPs 23 nm could explain the observed lower locomotion activity.Measurement of this parameter could allow the use of G. fossarum not
only at individual level, but also at population level, thus allowing a bet-ter and more realistic evaluation of the potency of a compound, whichgoes beyond the mere evaluation of the EC50s. Indeed, movement is ahighly ecologically relevant biomarker as locomotion is necessary forforaging, finding a mate and escaping from predators. Consequently,by interfering with locomotion, AgNO3 and AgNPs 23 nm could reducethe fitness of organisms and have a potential effect at population level,resulting in “ecological death” (Scott and Sloman, 2004).
5. Conclusion
This study allowed the obtaining of first information on the toxicity ofAgNPs on G. fossarum. AgNPs 23 nm, which was more toxic than AgNPs
Fig. 5. Behavioural responses of G.f2 after 72 h of exposure to AgNO3, AgNPs 20 nm, AgNPs 23 nm, AgNPs 27 nm. (A) Ventilation (Mean pleopod beat frequency ± SD). No significantdifferences were detected for any of exposure conditions (One-way ANOVA + Fischer LSD post hoc test, P b 0.05, n = 10). (B) Locomotor activities (Mean percentage of movingG.f2± SD). Letters illustrate significant differences. (One-way ANOVA + Fisher LSD post hoc test at P b 0.05 level of significance, n = 10).
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27nm, appeared to be the one that releasedmore Ag ions andwas readilytaken up byG. fossarum. AgNPs 23 nm led to significant iono/osmoregula-tion impairment, decrease in locomotor activity and survival rates.Haemolymphosmolality appeared to be aneffectivemarker as it indicatesthe physiological health status of G. fossarum. Altered locomotor activity,which was the most impacted response, may indicate potential effectsof released silver ions from AgNPs 23 nm at population level.
G. fossarum is proposed in this work as model organism fornanotoxicology as it allows studies at individual levels and at populationlevel. Our results indicate the potential usefulness of G. fossarum as amodel organism for future nanotoxicology studies at population level.
Acknowledgments
This work was in part supported by the Fonds National de laRecherche (NanoGAM AFR-PhD-9229040), the ResEAU LorLux projectand the NanEau II project (C10/SR/799842). The authors are gratefulto S. Contal andD. Collard for technical support, C. Guignard and J. Ziebelfor chemical analyses, P. Grysan andA. Chauvière for NanoSIMS analysis.V. Peardon and S. Pitt are thanked for thorough proofreading and lan-guage corrections. The contribution of YL. Balachandran in the develop-ment of plant-synthesised AgNPs in gratefully acknowledged. M.Kruszewski is thanked for providing chemically synthesised AgNPs.
Appendix A. Supplementary data
Supplementary data to this article can be found online at http://dx.doi.org/10.1016/j.scitotenv.2016.06.068.
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COMPLEMENT 1
Science of the Total Environment 566, 1649-1659.
Gammarus fossarum (Crustacea, Amphipoda) as a model organism to study silver nanoparticles
effects
Kahina Mehennaoui a,b, Anastasia Georgantzopoulou a,d, Vincent Felten b, Jennifer Andreï b, Maël
Garaud b, Sébastien Cambier a, Tommaso Serchi a, Sandrine Pain-Devin b, François Guérold b, Jean-
Nicolas Audinot c, Laure Giambérini b, Arno C. Gutleb a *
a Environmental Research and Innovation (ERIN) Department, Luxembourg Institute of Science and
Technology (LIST), 5 avenue des Hauts-Forneaux, Esch-sur-Alzette, Luxembourg.
b Laboratoire Interdisciplinaire des Environnements Continentaux (LIEC), CNRS UMR 7360,
Université de Lorraine – Metz, France.
c Materials Research and Technology (MRT) Department, Luxembourg Institute of Science and
Technology, (LIST), 5 avenue des Hauts-Forneaux, Esch-sur-Alzette, Luxembourg.
d Norwegian Institute for Water Research (NIVA), Gaustadalléen 21, NO-0349 Oslo, Norway.
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SUPPLEMENTARY MATERIAL
Particle dispersion and characterisation in Volvic® water
Sonication calorimetric calibration
The delivered acoustic power of the UP200S probe ultra sonicator was determined. Temperature (K°) of
medium during sonication was measured for 30 minutes and recorded as a function of time. Then, the
delivered acoustic power P (W= J/s) was calculated as the following:
P = (dT/dt)MCp
Where (dT/dt) is the slope of temperature (K) vs time (s), M is the mass of water and Cp is the specific heat of
water (4.186J/gK°).
Energy reserves, detoxification and LPO level measurements
Detoxification mechanisms and antioxidant responses
TAC was measured based on a method previously published (Erel, 2004). The loss of colour of the radical 2,2'-
azino-bis(3-ethylbenzothiazoline-6-sulphonic acid) (ABTS*+) upon reduction by antioxidants in the cytosolic
fraction was measured by measuring absorbance at 660nm after a 342s incubation time. TAC is calculated
against a calibration curve designed using Trolox (6-hydroxy-2,5,7,8-tetramethylchroman-2-carboxylic acid)
as standard, and expressed in mmol Trolox equivalent. g protein-1 (mmolTeq.gprot-1).
GPxtot activity was measured based on a method (Paglia and Valentine, 1967) adapted for its automation on
the Konelab (Garaud et al., 2015). The cytosolic fraction was diluted to 1/32 (v/v) to a total volume of 160 µL
with a solution containing 50 mM phosphate buffer (pH = 7.6), reduced glutathione (GSH, 2 mM), NADPH
(0.36 mM), glutathione reductase (1 U/mL) and cumene hydroperoxyde (1.125 mM), being the last substrate
for the reduction. NADPH consumption was followed for 126s by monitoring absorbance at 340 nm and
GPxtot activities were expressed in µmol NADPH.g protein-1.min-1 (ε NADPH = 6220 M-1.cm-1).
GST activity was measured on the Konelab automat by monitoring the absorbance at 340 nm for 4 min. The
measurements were performed in a phosphate buffer (0.1 M; pH 6.5) supplemented with 1-chloro2,4-
dinitobenzene (CDNB; 1 mM), GSH (1 mM) and 13/200 (v/v) of diluted cytosolic fraction. GST activity is
expressed in µmol CDNB.min-1.g-1 proteins.
ACP activity was measured by measuring the hydrolysis of 4-nitrophenyl phosphate (1.67 g/L) by acid
phosphatase. Total homogenate was diluted (1/144 v/v) in a total volume of 90 µL of citrate buffer (41.7 mM,
pH=4.8) supplemented with 0.16% Triton X-100. Reaction was stopped and the p-nitrophenol resulting from
Chapter 2 G. fossarum as a model organism for nano-ecotoxicology
93
the hydrolysis of 4-nitrophenyl was revealed after 15 min by the addition of 125 µL NaOH (138.7 mM). A p-
nitrophenol standard calibration curve was constructed. ACP activity was expressed in µmol p-nitrophenol. g
protein-1.h-1 (Garaud et al., 2015).
Catalase activity was measured by spectrophotometry (SPECORD® 205; analytikjena) according to a published
method (Beers and Sizer, 1951). Catalase catalyses the decomposition of hydrogen peroxide into water and
oxygen. Briefly, cytosolic fraction (S2) was first diluted 1 to 4 in 50 mM phosphate buffer solution (pH=7.6).
Then, the diluted S2 was incubated for 30 seconds with H2O2 (30 mM) in the reactive mixture (1/10 v/v) and
the consumption of H2O2 was followed at 240 nm. Catalase enzymatic activity is expressed in mmol H2O2.g
proteins-1.min-1 (ε H2O2 = 40 M-1.cm-1).
Cellular damage and oxidative stress assessment
Lipid hydroperoxide (LOOH) concentration was measured using the Konelab following the automated method
previously described (Arab and Steghens, 2004). LOOH were quantified based on the oxidation of Fe2+ to Fe3+,
under acidic conditions, followed by a complexation of Fe3+ by xynelol orange whose absorbance is finally
measured at 620 nm after 20 min incubation time. The final xylenol orange (XO) concentration used was 240
mM as described (Garaud et al., 2015). A tert-butylhydroperoxide (TBH) standard calibration curve (0.125-4
µmol/L) was established and LOOH contents were expressed in nmol TBH. g proteins-1.
Lactate Dehydrogenase (LDH) is an enzyme involved in glycogen metabolism and is used as a biomarker of
metabolic disruption. LDH activity was measured using Thermo-Scientifc Konelab ready-to-use reagents and
methods (LDH IFCC) adapted to gammarids (Garaud et al., 2015). LDH activity was expressed µmol NADH. g
proteins-1.h-1.
Apoptosis was measured based on Caspase-3 (CASP-3) activity. CASP-3 was measured in the cytosolic fraction
following an automated method developed on the Konelab. Cytosolic fraction was diluted 1 to 7 (v/v) with a
solution containing Tris buffer (10 mM, pH=7.4), DTT (5 mM) and EDTA (2 mM). The absorbance of p-
Nitroaniline (pNA) released by the cleavage of caspase—3–substrate z-DEVD-pNA (25 µM) by CASP-3 was
followed at 405 nm for 1h. CASP-3 activity was expressed in µmolpNA. g proteins-1.h-1 (µmolpNa/gprot/h,
εpNA=10500 M-1.cm-1 at 405 nm) (Garaud et al., 2015).
Energy reserves
Protein, triglyceride and cholesterol concentrations were measured using Thermo-Scientifc Konelab ready-
to-use reagents (following manufacturer’s instructions) and methods adapted to gammarid (Garaud et al.,
2015). Scal bovine-based serum (Thermo-Scientific) was used as standard to create calibration curves and
energetic reserve contents were expressed in mg.g-1 fresh weight.
Chapter 2 G. fossarum as a model organism for nano-ecotoxicology
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Total protein concentration in the cytosolic fraction was determined by the pyrogallol red colorimetric
method using ready-to-use U/CSF protein (following manufacturer’s instructions) reagents and Scal bovine-
based serum as reference (Thermo-Scientific).
Electron Transport System (ETS) mitochondrial activity was measured following the original method (Owens
and King, 1975) modified by (De Coen and Janssen, 1997) and automated on the Konelab (Garaud et al.,
2015). ETS activity was measured by monitoring for 12 min at 480 nm, the production of formazan by the
reduction of p-iodonitrotetrazolium (133 mg. L-1), which works as final electron acceptor of the mitochondrial
respiratory chain. This reaction was measured in 80 mM phosphate buffer (pH=8.5) containing MgSO4 (15
µM), Triton X-100 (0.16%), PVP (300 mg. L-1), NADH (1 mM) and NADPH (0.15 mM) with 1/35 (v/v) of diluted
whole homogenate. ETS activity was calculated according to (De Coen and Janssen, 1997) and expressed in
µmol O2. g proteins-1.h-1.
REFERENCES
Arab, K., Steghens, J.-P., 2004. Plasma lipid hydroperoxides measurement by an automated xylenol orange
method. Anal. Biochem. 325, 158–163. doi:10.1016/j.ab.2003.10.022
Beers, R.F., Sizer, I.W., 1951. A Spectrophotometric Method for Measuring the Breakdown of Hydrogen Peroxide
by Catalase. J. Biol. Chem. 195, 133–140.
De Coen, W.M.D., Janssen, C.R., 1997. The use of biomarkers in Daphnia magna toxicity testing. IV. Cellular Energy
Allocation: a new methodology to assess the energy budget of toxicant-stressed Daphnia populations. J. Aquat.
Ecosyst. Stress Recovery 6, 43–55. doi:10.1023/A:1008228517955
Erel, O., 2004. A novel automated direct measurement method for total antioxidant capacity using a new
generation, more stable ABTS radical cation. Clin. Biochem. 37, 277–285. doi:10.1016/j.clinbiochem.2003.11.015
Garaud, Auffan, M., Devin, S., Felten, V., Pagnout, C., Pain-Devin, S., Proux, O., Rodius, F., Sohm, B., Giambérini, L.,
2015. Fate and integrated assessment of ceria nanoparticle impacts on the freshwater bivalve Dreissena
polymorpha: a mesocosm approach. Nanotoxicology.
Owens, T.G., King, F.D., 1975. The measurement of respiratory electron-transport-system activity in marine
zooplankton. Mar. Biol. 30, 27–36. doi:10.1007/BF00393750
Paglia, D.E., Valentine, W.N., 1967. Studies on the quantitative and qualitative characterization of erythrocyte
glutathione peroxidase. J. Lab. Clin. Med. 70, 158–169.
Chapter 2 G. fossarum as a model organism for nano-ecotoxicology
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TABLE S 1. PHYSICO-CHEMICAL PARAMETERS OF VOLVIC® WATER
Mineral content and physic chemical Parameters Volvic water
Ca2+ (mg.L-1) 11.5
Mg2+ (mg.L-1) 8
Na+ (mg.L-1) 11.6
K+ (mg.L-1) 6.2
SO42- (mg.L-1) 8.1
HCO3-(mg.L-1) 71
NO3- (mg.L-1) 6.3
Cl- (mg.L-1) 13.5
SiO2 (mg.L-1) 31.7
F (mg.L-1) 0.22
pH 7.75 ± 0.07
Conductivity (µS.cm-1) 213.67 ± 1.53
Dissolved oxygen (mg.L-1) 8.31 ± 0.02
Alkalinity (meq.L-1) 1.214 ± 0.1
Hardness (mg CaCO3.L-1) 63.3
Chapter 2 G. fossarum as a model organism for nano-ecotoxicology
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FIGURE S 1. SIZE DISTRIBUTION OF AG NPS 23 NM (A), 27 NM (B), 20 NM (C) AND 200 NM (D) IN VOLVIC® WATER
EXPRESSED AS PARTICLE CONCENTRATION X 106.ML-1. THE RED ERROR BARS INDICATE THE ± SD OF THE MEAN OF TRIPLICATE
MEASUREMENTS.
97
Chapter 3 Reference genes of G. fossarum
96
CHAPTER 3
G. fossarum molecular responses
CHAPTER 3
G. fossarum molecular responses
Chapter 3 Reference genes of G. fossarum
97
KEY FINDINGS
Aim of the study
The aim of the study is the determination of a suitable set of reference genes for data normalization of RT-
qPCR experiments using G. fossarum. Six common candidate reference genes, tested and validated with other
species, were selected and their reliability for RT-qPCR were tested in different exposure conditions using
AgNO3, silver nanoparticles (AgNPs) and gold nanoparticles (AuNPs) contamination.
Experimental design
Key findings
• Set of six reference genes was tested
o Actin, TUB, UB, GAPDH, SDH, Clathrin
• One target gene was used to validated the stability of reference genes
o HSP90
• Stability of reference genes:
o Clathrin > SDH > GAPDH > TUB > Actin > UB
• RT-qPCR data normalization
o SDH + clathrin: down-regulation of HSP90 in G. fossarum exposed to AuNPs 40 nm
o SDH + clathrin + GAPDH: same results when data were normalized only with the two most
stable genes
o All genes: up-regulation of HSP90 in G. fossarum exposed to AgNPs 40 nm
• These results highlight the importance of selection of reference genes to obtain robust RT-qPCR data
RNA
Ribonucleic acid
Nucleobases
Base pair
Guanine
Uracil
Guanine
Adeninehelix of
sugar-phosphates
Nucleobases
of RNA
DNAoxyribonucleic a
Nucleobases
Base pair
Cytosi
Guanine
Adenine
Thymine
elix ofosphates
RNAextraction
Reversetranscriptase
cDNAsynthesis
qPCR
Acclimation
D-10
Exposure
D-7 D-1
Temperature : 12 oC
Photoperiod : 16h/8h
Daily renewal of Volvic water
Stalling
D15D-30 D-20
Leave
conditioning
Temperature : 12 oC
Photoperiod : 16h/8h
Exposure medium renewal: every 3 days Addition of food: every 2 days
D0
Chapter 3 Reference genes of G. fossarum
98
Article 2
Article submitted to Scientific Reports
Identification of reference genes for RT-qPCR data normalization in Gammarus fossarum (Crustacea
Amphipoda)
Kahina Mehennaoui 1,2, Sylvain Legay 1, Tommaso Serchi 1, François Guérold 2, Laure Giambérini 2, Arno C.
Gutleb 1, Sébastien Cambier 1*
1 Environmental Research and Innovation (ERIN) Department, Luxembourg Institute of Science and
Technology (LIST), 5, avenue des Hauts-Fourneaux, Esch-sur-Alzette, Luxembourg
2 Laboratoire Interdisciplinaire des Environnements Continentaux (LIEC), CNRS UMR 7360, Université de
Lorraine – Metz, France
Chapter 3 Reference genes of G. fossarum
99
Identification of reference genes for RT-qPCR data normalization in Gammarus fossarum
(Crustacea Amphipoda)
Kahina Mehennaoui 1,2, Sylvain Legay 1, Tommaso Serchi 1, François Guérold 2, Laure Giambérini
2, Arno C. Gutleb 1, Sébastien Cambier 1*
1 Environmental Research and Innovation (ERIN) Department, Luxembourg Institute of Science
and Technology (LIST), 5, avenue des Hauts-Fourneaux, Esch-sur-Alzette, Luxembourg;
[email protected] (K.M.); [email protected] (T.S.); [email protected] (A.C.G.)
2 Laboratoire Interdisciplinaire des Environnements Continentaux (LIEC), CNRS UMR 7360,
Université de Lorraine – Metz, France ; [email protected] (F.G.) ;
[email protected] (L.G.)
* Correspondence: [email protected]; Tel.: +352-275-888-5018
Chapter 3 Reference genes of G. fossarum
100
Abstract: Gene expression profiling via RT-qPCR is a robust technique increasingly used in ecotoxicology.
Determination and validation of optimal reference genes is a requirement for initiating RT-qPCR experiments.
To our best knowledge, this study is the first attempt of identifying a set of reference genes for the freshwater
crustacean Gammarus fossarum. Six candidate genes (actin, TUB, UB, SDH, clathrin and GAPDH) were tested
in order to determine the most stable ones in different stress conditions and to increase the robustness of
RT-qPCR data. SDH and clathrin appeared as the most stable ones. A validation was performed using G.
fossarum samples exposed for 15 days to AgNO3, AgNPs 40 nm and AuNPs 40 nm. Effects on HSP90 were
evaluated and data normalized using clathrin and SDH. HSP90 was down-regulated when G. fossarum were
exposed to both AuNPs 40 nm whereas no effects were observed when G. fossarum were exposed to AgNPs
40 nm. This study highlights the importance of the preliminary determination of suitable reference genes for
RT-qPCR experiments. Additionally, this study allowed the determination of a set of valuable genes that can
be used in other RT-qPCR studies using G. fossarum as model organism
Keywords: Gammarus fossarum; Reference genes; RT-qPCR; Gene expression; Environmental toxicology
1. INTRODUCTION
Understanding the mechanisms underlying the effects of stressors on organisms needs sensitive analytical
techniques that can cover and link responses observed at different biological levels (from molecular to
individual responses). One of the most reliable technique is the measurement of changes or alterations in
gene expression in response to an external stimuli [1]. Recent advances in “omics” and bioinformatics
methodologies applied in ecotoxicological studies provided a new-angle of study of non-model organisms,
opening new ways in determining new molecular biomarkers (genes) as an alteration of their regulation may
influence the fitness of organisms [2,3]. Reverse transcription quantitative polymerase chain reaction (RT-
qPCR) is currently described as one of the most reliable techniques to assess these changes due to its
effectiveness, sensitivity and reproducibility [1,2,4]. This method, which allows studying the expression of a
set of selected genes in an organism, requires multiple critical quality controls in order to obtain robust
results. This includes RNA purity and integrity control, genomic DNA contamination assessment, evaluation
of PCR primer efficiency and specificity and, in case of relative quantification of gene expression, the
identification of suitable reference genes for data normalization [1,5,6].
Reference genes are described to be stable regardless the exposure conditions making them suitable for data
normalization of genes of interest. Therefore, their determination is crucial for a good analyze of RT-qPCR
results [2,5,7,8]. Furthermore, combination of a set of the most stable reference genes is fundamental to
avoid misinterpretation of the results that may occur when using a unique reference gene [1]. Consequently,
Chapter 3 Reference genes of G. fossarum
101
a preliminary experimental determination of the optimal reference genes is necessary before each new
experimental design [5,8]. For this purpose, many software and tools have been developed in order to
determine the best reference genes. This selection is performed following a ranking method based on their
stability across different stress conditions [5,7,9]. Unfortunately, many ecotoxicological studies are still using
only one, or a set of reference genes selected from previous studies, applying different experimental designs,
making them sometimes improper for data normalization. Classically known housekeeping genes such as
actin, tubulin (TUB), glyceraldehyde-3-phosphate dehydrogenase (GAPDH), and 18S ribosomal RNA are used
as reference genes [2,4,6,8]. However, recent studies demonstrated that some of these genes might not be
suitable as they are not always stable depending on the experimental design, studied stress and organisms
leading sometimes to the misinterpretation of the results [8,10].
Among the aquatic macroinvertebrates assemblage, amphipods of the genus Gammarus, and more
specifically Gammarus fossarum, represent a major part of it [11,12]. They are used in different
ecotoxicological studies for their high ecological relevance linked to their wide distribution all over Europe
and their major functional role in litter breakdown process and nutrient cycling [13,14]. Their well-
documented sensitivity to different kind of pollutants make them good model organisms for ecotoxicological
studies [15–18]. However, to our knowledge, few studies investigated their responses to exogenous stress at
molecular level, such as the vitellogenin expression following an estrogenic stress (Xuereb et al., 2011), the
antioxidant responses via catalase and MnSOD gene expression following exposure to gold nanoparticles
(AuNPs) (Baudrimont et al., 2017) and the identification of proteins expression profiles during
spermatogenesis [19–21] . Nevertheless, there is still no reference genes properly characterized and
dedicated to the amphipod G. fossarum. Reference genes are mandatory for an accurate data normalization
of gene expression [1]. Therefore, the aim of the present study is to fill this gap and to determine a suitable
set of reference genes for data normalization of RT-qPCR experiments using G. fossarum. Six common
candidate reference genes, tested and validated with other species [4,8,9], were selected and their reliability
for RT-qPCR were tested in different exposure conditions using AgNO3, silver nanoparticles (AgNPs) and gold
nanoparticles (AuNPs) contamination.
2. RESULTS
2.1. Stability of the candidate reference genes in G. fossarum
To identify the most appropriate set of reference genes for G. fossarum, six candidates were tested. The
genes included some frequently used ones (actin, TUB and GAPDH) and less common ones, namely, succinate
dehydrogenase (SDH), ubiquitin (UB) and clathrin. Five different methods were compared in order to
Chapter 3 Reference genes of G. fossarum
102
determine the most stable gene sub-set: GeNormPlus [22], NormFinder [23], BestKeeper [7], RefFinder [24]
and the comparative delta-CT method [25].
The ranking of the studied genes varied dependently on the algorithms used (Table 3. 1). Interestingly, the
same ranking was obtained using the most commonly used softwares, GenormPlus and NormFinder, with
clathrin and SDH as the most stable genes and UB as the least stable one. In the same way, RefFinder and the
comparative delta-CT method identified clathrin and SDH as the most stable genes. GAPDH was the third
most stable gene according to GenormPlus, NormFinder, and RefFinder, whereas, the rankings were different
for the least stable ones (Table 3. 1) as Actin was assigned the highest score by BestKeeper, the comparative
delta-CT method and RefFinder (Table 3. 1).
As BestKeepper’s ranking is based on correlation factor (R, the closer to 1, the better) and standard deviation
(the bigger, the worse) [7, 42], the obtained results differed from the other algorithms as Clathrin and Actin appeared
as the most stable genes whereas GAPDH and TUB were the least stable ones. According to BestKeeper, SDH was the
third most stable gene (Table 3.1)
Table 3. 1.Ranking of candidate reference genes according to the five algorithms used
Ranking
GeNorm NormFinder BestKeeper Comparative delta
Ct RefFinder
Gene Stability
coeff. Gene
Stability
coeff. Gene
Stability
coeff. Gene
Stability
coeff. Gene
Stability
coeff.
1 Clathrin 0.433 Clathrin 0.105 Clathrin 0.670 Clathrin 1.035 Clathrin 1.000
2 SDH 0.470 SDH 0.198 Actin 0.624 SDH 1.067 SDH 1.682
3 GAPDH 0.550 GAPDH 0.265 SDH 0.571 TUB 1.188 GAPDH 3.464
4 Actin 0.593 Actin 0.292 UB 0.575 GAPDH 1.190 TUB 3.464
5 TUB 0.652 TUB 0.465 GAPDH 0.459 UB 1.232 UB 5.000
6 UB 0.703 UB 0.520 TUB 0.395 Actin 3.217 Actin 6.000
In order to determine the stability of each gene, a global ranking was generated by assigning a number (from
1 to 6 where 1 is the most stable gene) to each stability coefficient presented in Table 3. 1 and by averaging
them [1,9]. This allowed the confirmation of the high stability of clathrin and SDH (Figure 3. 1). UB appeared
as the least stable gene followed by actin and TUB (Figure 3. 1).
Chapter 3 Reference genes of G. fossarum
103
Figure 3. 1. Global ranking of candidate reference genes in G. fossarum. A number (from 1 to 6) was assigned
to each stability coefficient. A mean rank was generated and error bars
2.2. Optimal number of reference gene for data normalization in G. fossarum using GeNorm
In order to calculate the optimal number of reference genes for data normalization in G. fossarum,
GeNormplus was used to generate the pairwise variation (Vn/Vn+1) between two normalization factors
(NF/NFn+1). Indeed, only GenormPlus allows an estimation of the optimal number of reference genes to use
in a specific experimental design.
The analysis conducted on 31 samples of G. fossarum exposed to AgNO3, AgNPs 40 nm and AuNPs 40 nm,
showed that the optimal number of reference genes is 2 as the V value is below the cut-off threshold of 0.15
making the addition of a third gene unnecessary (Figure 3. 2). As previously described, GeNormPlus identified
clathrin and SDH as the best combination for data normalization. In the same way, NormFinder determined
clathrin/SDH as the best pair as these two genes showed the highest stability coefficient (Table 3. 1).
0
1
2
3
4
5
6
7
clathrin SDH GAPDH TUB Actin UB
Meanrankvalue
Candidates referencegenes
Chapter 3 Reference genes of G. fossarum
104
Figure 3. 2 Determination of the optimal number of reference gens for data normalization in G. fossarum
exposed to AgNO3, AgNPs 40 nm and AuNPs 40 nm. The pairwise variation (Vn/Vn+1) was calculated between normalization factors NF/NFn+1. The recommended cut-off threshold of 0.15 was applied in this study.
2.3. Validation of the selected reference genes for G. fossarum
In order to validate G. fossarum reference genes, the expression of a general stress-related gene, the heat
shock protein 90 (HSP90), was evaluated. HSP90 expression profiles were studied on G. fossarum exposed
for 15 days to 0.5 µg. L-1 of AgNO3, CIT-AgNPs 40 nm, PEG-AgNPs 40 nm, CIT-AuNPs 40 nm and PEG-AuNPs
40 nm. Data were analysed using the Biogazelle qbase+ software and normalized using clathrin and SDH. As
shown in Figure 3. 3, a significant decrease in HSP90 expression was observed when G. fossarum were
exposed to 0.5 µg. L-1 of CIT-AuNPs 40 nm and PEG-AuNPs 40 nm (One-way ANOVA, P <0.001) while none of
the tested AgNPs or AgNO3 impacted HSP90 expression (Figure 3. 3, One-way ANOVA, P>0.05). However,
when data were normalized using all the six reference genes, the statistical analysis showed a significant
induction of the expression of HSP90 after treatment with CIT-AgNPs 40 nm (Figure 3. 3, One-way ANOVA, P
<0.05). Data were also normalized using the three best reference genes clathrin, SDH and GAPDH. No
differences in HSP90 expression were observed (Figure 3. 3, One-way ANOVA, P <0.05) enhancing that adding
a third gene for data normalization is unnecessary for our experimental design (Figure 3. 3).
0
0.02
0.04
0.06
0.08
0.1
0.12
0.14
V2/3 V3/4 V4/5 V5/6
Pairwisevariation(V)
Chapter 3 Reference genes of G. fossarum
105
Figure 3. 3 HSP90 expression analysis using different normalization strategies. Error bars indicate the standard
errors of the means (n=4). Different letter (a-c) indicate significant differences at P < 0.05.
3. DISCUSSION
Despite the rapid advances in the “omics” field marked by the development of next generation sequencing
(NGS) methods, RT-qPCR remains one of the most accurate and reliable techniques for targeted gene
expression and NGS data validation. This method is recognized for its capacity to highlight sensitive changes
in gene transcription levels [1,26]. However, for an accurate interpretation of RT-qPCR results, in term of
relative gene expression, one of the most important criteria is the selection of suitable reference genes,
which, to the best of our knowledge, are lacking for G. fossarum species. Previous studies have already
investigated molecular responses in Gammarus sp., exposed to AuNPs, temperature and ammonia, using
actin, or GAPDH as reference genes [19,27] on the assumption that these genes were stable and without
prior experimental verifications. Such approaches might conduct to misinterpretation of RT-qPCR results
[5,28,29] since the expression of these genes might be influenced by biotic or abiotic stress as well as
developmental stage and tissue type [4,8,30]. It is well established that a single reference gene could not be
applied to all the experimental designs [1,5]. Therefore, a case-specific choice of the best reference genes for
RT-qPCR is mandatory. In the present study, six candidate genes have been tested in adult G. fossarum males,
regarding a contamination with AgNO3, AgNPs and AuNPs, for their appropriateness to be used as reference
genes in transcriptomic studies [4]. The efficiency and specificity of the designed primers have been checked.
All the tested primers showed correct PCR efficiency and specificity during the PCR ensuring robust and
precise results interpretation [5], and confirmed that our RT-qPCR experiments comply with the known
0.00
0.50
1.00
1.50
2.00
2.50
Citrate PEG Citrate PEG
AgNPs40nm AuNPs40nm AgNO3
Norm
alizedrelativeexpression
Exposureconditions
Clathrin+SDH allgenes clathrin+SDH+GAPDH
b b
c
ac
a aa
aa
ab
a
Chapter 3 Reference genes of G. fossarum
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recommendations [4,5]. Five different algorithms were used to determine optimal reference genes to
normalize our data. The five used softwares identified clathrin and SDH as the most stable genes. These
results are in accordance with previous studies that described SDH as a suitable reference gene in Rhodnius
prolixus [10] and in red abalone Haliotis rufescens’ gonad and digestive tissues [31]. Clathrin was also
identified as one of the most stable genes in Lilium formolongi [32].
It is important to notice that our ranking of the candidate reference genes shows that the most commonly
used genes in transcriptomics studies are the least stable ones in this study. This observation is in agreement
with other recent studies which showed that GAPDH, 18s rRNA, actin and TUB were not good candidates as
reference genes [4,8,26,33]. GAPDH, commonly considered as a housekeeping gene, plays an important role
in energetic metabolism and its expression was described as significantly impacted in bivalves, like Mytilus
spp. or Crassostrea gigas, exposed to different environmental stress (harbor pollution), and in Haliotis discus
hannai under heavy metal stress conditions [4,33]. On the contrary, other studies showed that GAPDH was
one of the most stable gene in two flatworms species, Macrostomum lignano and Schmidtea mediterranea,
exposed to cadmium [26] and in R. prolixus in different physiological conditions and feeding status [10].These
observation indicate that GAPDH should not be considered as suitable for data normalization without prior
validation. Other studies demonstrated that selection of unstable genes like actin and 18S rRNA as reference
genes greatly altered interpretation of data for Cathepsin D in Ruditapes philippinarum exposed to copper
[8]. Similar observations were done for TUB and UB in Haliotis discus hannai [4]. These results are in
agreement with our observations in G. fossarum where classical reference genes like actin, TUB, UB and
GAPDH appeared among the least stable genes. The present study shows that it is crucial to experimentally
assess the stability of reference genes for each species tested in each experimental design applied before
they can be selected as housekeeping genes [8].
In addition, the stability of the present reference genes was verified by exposing G. fossarum to AgNO3, AgNPs
40 nm and AuNPs 40 nm for 15 days and studying the expression of HSP90. HSP90 is involved in the regulation
of proteostasis under both physiological and stress conditions. It plays many roles such as protein folding as
a chaperone, DNA repair and immune responses [34]. HSP90 is also known to play a role, with HSP70, in
keeping inactive the heat shock factor (HSF1) which induce the expression of HSP90 in stress condition
[34,35]. Therefore, it was used as a target gene in the present study. Data were normalized using, firstly
clathrin and SDH, which were indicated by GeNormPlus to be the most stable genes, then, clathrin, SDH and
GPDH and finally with the whole set of six candidate reference genes. When clathrin and SDH are used for
normalization, a significant decrease in HSP90 expression was observed in G. fossarum exposed to CIT-AuNPs
40 nm and PEG-AuNPs 40 nm. However, when data were normalized using clathrin, SDH and GAPDH, no
differences in HSP90 expression profile was observed making the addition of a third gene unnecessary.
Chapter 3 Reference genes of G. fossarum
107
Furthermore, data normalization using all the six-studied reference genes, led to the detection of an
upregulation of HSP90 in G. fossarum exposed to CIT-AgNPs 40 nm, which in this case is to be considered as
an overestimation of the response as “incorrect references” were used. This is in accordance with what has
been previously reported [26]. Normalization of HSP90 expression in M. lignano exposed to Cd using all the
nine reference genes tested led to a high variability between replicates. Authors stated that this observation
could be linked to an important variability between replicates that lower the resolution of detection of
differences between their different conditions [26]. However, when data were normalized using the three
most stable genes, no significant differences in HSP90 expression were observed [26]. Other study showed
that a non-optimal selection of the best combination of reference genes may lead to statistical
misinterpretation [10]. Authors showed that data normalization using an unstable gene such eiF-1a in R.
prolixus led to a clear but false increase in an olfactory gene, RproIR76b expression while no statistical
difference were observed when data were normalized using the most stable genes [10]. Moreover, other
study underlined the importance of selection of the appropriate reference genes with the highest stability
coefficient as data normalization of metallothionein expression in abalone exposed to copper let to an
underestimation or overestimation of the effects when data were normalized using unstable reference genes
[4]. These results highlight once more the importance of an experimental validation of reference gene in
addition to the selection of the optimal number and the appropriate genes for data normalization [5,10,26].
4. MATERIALS AND METHODS
4.1. Organisms sampling and acclimation
G. fossarum were collected at an unpolluted stream (49°48’24.9’’ N and 06°04’53.2’’ E, Schwaarzbaach,
Colmar-berg, Luxembourg) [18,36]. Animals were collected using a hand net and were sorted in the field.
They were immediately brought to the laboratory in river water, where they were kept at 12 °C. In order to
avoid influence of gender on the studied parameters, only adult males were kept for the experiment [16,37].
They were selected from precopula pairs or based on sexual dimorphism like gnatopode size [18]. Adult males
were then acclimated to laboratory conditions [18,38]. The acclimation was conducted on 2 steps. First,
Gammarids were acclimated for 72h to mineral water (Volvic, France) by progressively changing field water
to Volvic water (30% v/v, 50% v/v, 100% v/v) Then, a stalling period of 10 days was conducted in 100% Volvic
water [18,38]. The acclimation was performed under controlled conditions at 12 °C with a 16h light 8h dark
photoperiod. Volvic water was aerated and changed every 24h to avoid organic matter accumulation and
potential increase of ammonium, nitrite and nitrate. Gammarids were fed ad libitum with alder leaves (Alnus
glutinosa) up to 24h before experiment.
Chapter 3 Reference genes of G. fossarum
108
4.2. AgNO3, AgNPs and AuNPs contamination
At the end of the acclimation period, four pools of 3 males were exposed to 0.5 µg.L-1 of AgNO3, AgNPs 40
nm and AuNPs 40 nm, either stabilized with citrate (CIT-AgNPs 40 nm and CIT-AuNPs) or coated with
polyethylene-glycol (PEG-AgNPs 40 nm and PEG-AuNPs 40 nm), for 15 days at 12 ºC with a photoperiod of
16h light and 8h darkness [18]. Exposure medium (Volvic water) was changed every 72h. Food (Alnus
glutinosa disk leaves) was added every 48h. At the end of the exposure period, Gammarids were gently dried
and stored at -80 ºC in RLT buffer supplemented with 1% β-mercaptoethanol until RNA extraction.
4.3. Gene identification and qPCR primer design
4.3.1. DNA extraction
Pools of 50 eggs of G. fossarum were placed in ATL Buffer supplemented with 10% (v/v) of proteinase K,
placed on thermomixer at 56 ºC at 800 rpm for 1h. Genomic DNA was extracted using DNeasy blood and
tissue kit following the manufacturer’s instructions (Qiagen, Leusden, The Netherlands). Quantification and
purity were performed using Nanodrop ND-1000 (Thermo scientific, Waltham, MA, USA). The resulting gDNA
was stored at -20ºC until use.
4.3.2. Library preparation and sequencing
Libraries were prepared from 50 ng of genomic DNA using the Nextera DNA library preparations kit
according to the manufacturer’s instructions (Illumina, San Diego, CA, USA). After labelling of gDNA
fragments, clean up and PCR amplification, libraries were controlled using the Agilent 2100
bioanalyzer and DNA high sensitivity kit. Libraries were quantified using the Library quantification
kit – Illumina Low ROX (Kapa Biosystems, Boston, MA, USA) on a Viia7 96-well real-time-qPCR
instrument (Life Technologies). The libraries were sequenced using an Illumina Miseq by means of
3 consecutive runs (Illumina Miseq reagent V3) to generate 76 base-pairs paired-end reads.
4.3.3. De novo assembly
Miseq generated FASTQ files were imported in CLC genomics workbench v9 discarding reads with low quality
(<Q30). Illumina adapters and indexes were trimmed using the Nextera adaptor sequence and filtered
(%filtered). Reads with nucleotide ambiguity (N) and CLC quality index above 0.05 were then filtered from
the remaining sequences. Two trimming procedures were applied. First, the Nextera adaptor sequences were
filtered in order to avoid the presence of technical sequences (adaptors and indexes) in the reads. Conserved
sequences generated by the hybridization and amplification of random hexamers (commonly found in the
Chapter 3 Reference genes of G. fossarum
109
stranded libraries), were trimmed using an arbitrary “hard trim” of 16 base pairs at the 5’ end of the reads
and 5 base pairs at the 3’ end of the reads. These values were determined from the QC report analysis
obtained from CLC genomics. Finally, the remaining reads were assembled using a De novo assembly
following the upcoming criteria: minimum contig length of 200bp, mismatch, insertion and deletion costs of
3, length and similarity fraction of 0.8 and maximum distance of 1000 bp.
4.3.4. Gene identification
In order to identify and amplify putative genes from G. fossarum, an initial data mining was performed on
the raw reads previously sequenced by [20] in addition to the reads obtained as described above (see 4.3.3).
Reads were mapped to the sequences of Hyalella azteca (https://www.hgsc.bcm.edu/arthropods/hyalella-
azteca-genome-project), a closely related species for which the transcriptome is available, with the following
criteria: the mapped reads must have multiple hits lower than 10, a minimum of 80% identity and 80%
coverage with the reference. Mismatch costs was set at 2 (medium) and deletion/insertion cost at 3 (highest
stringency). A consensus sequence was generated from the mapped reads for each gene (Table 3. 2). In order
to verify whether the obtained sequences are coding for a protein, a translation of the obtained nucleotide
sequences was performed using ExPASy translation tool (http://web.expasy.org/translate/). Finally, a Blastx
search was performed against non-redundant protein databases from the National Centre for Biotechnology
(NCBI) to check the identity of the selected sequences. (Table 3. 2).
Chapter 3 Reference genes of G. fossarum
108
Table 3. 2 Identification of Gammarus fossarum gene sequences
Genes
Hyalella azteca G. fossarum consensus
sequences NCBI Blastx
Accession
number Accession number Homology Identification Accession number
Actin XM_018157137.1 MF940257 97% actin, partial
[Hoplolaimus galeatus] AEM45650.1
Tubulin XM_018153872.1 MF940258
96%
PREDICTED: tubulin alpha-8
chain-like isoform X1
[Serinus canaria]
XP_009098159.2
Ubiquitin XM_018170409.1 MF940259 96%
ubiquitin conjugating
enzyme-3
[Eriocheir sinensis]
ADF45343.1
GAPDH XM_018154227.1 MF940254 95%
putative glyceraldehyde-3-
phosphate dehydrogenase
[Gammarus locusta]
CAQ60115.1
SDH XM_018156499.1 MF940255 96%
PREDICTED: succinate
dehydrogenase
[ubiquinone] iron-sulfur
subunit, mitochondrial-like
[Hyalella azteca]
XP_018011988.1
Clathrin XM_018171236.1 Sequence provided in
supplementary material 93%
PREDICTED: clathrin light
chain-like isoform X1
[Hyalella azteca]
XP_018025977.1
HSP90 XM_018155941.1 MF04256 98%
PREDICTED: heat shock
protein HSP 90-alpha-like
[Hyalella azteca]
XP_018022683.1
Chapter 3 Reference genes of G. fossarum
109
Table 3. 3 List of primers of the candidate reference genes and target gene HSP90
Name Sequence (5'→ 3') Amplicon Length (bp) Amplicon Tm (°C) PCR Efficiency Regression
Coeff. (R2)
Actin_F CTCATGCTATCCTTCGTCTTGA 103 78 2.02 0.999
Actin_R CGTTCAGCGGTGGTTACAA
Tubulin_F CGGCTGTTGTTGAACCTTAC 93 81 2.09 0.999
Tubulin_R AGATGGCCTCATTGTCAACC
GAPDH_F GTCCGTCTCGCTAAGGAGTG 94 85 1.91 0.999
GAPDH_R TGTATCCGAGGTAGCCCTTG
SDH_F GGAAGAAGCTGGATGGTCTG 87 84 1.98 0.998
SDH_R ACTTGTCTCCGTTCCACCAG
Ubiquitin_F CCCACGATACTCCCTTTGAA 82 79 2.01 0.991
Ubiquitin_R ACAATCGGTGGCTTGTTAGG
Clathrin_F ATCGCCAAGCTTTGTGACTT 107 85 1.99 0.999
Clathrin_R GCTTTGATAGGCGGACTCTG
HSP90_F CTGGTTTCTTCTCCCTGCTG 135 85 1.99 0,995
HSP90_R GATCTCGAGGTGCTTCTTGG
Chapter 3 Reference genes of G. fossarum
110
4.3.5. Primer design
All the primers were designed using Primer3Plus (http://primer3.ut.ee) with the following criteria: primer
size between 18 and 25 base pairs, GC content between 40% and 60 %, amplicon size from 80 to 150 base
pairs, primer annealing temperatures in the 58-61 ºC range. Primers were checked using NetPrimer
(http://www.premierbiosoft.com/netprimer/) for secondary unexpected structures. PCR efficiency was
evaluated using decreasing five-fold dilutions from cDNA pool (from 25 ng to 0.04 ng and no template
control). A melting curve was performed at the end of each run, in order to assess the specificity of the
amplified products. All tested genes displayed one clear peak and were therefore retained for analyses.
Primers sequences, amplicon size, and melting temperature are described in Table 3. 3.
4.4. RNA extraction, cDNA and RT-qPCR
G. fossarum tissues were ground on ice using a pellet pestle tissue grinder. Homogenates were centrifuged
at 250g for 5 min at 4°C to remove cuticle fragments as described in [18]. Total RNA was extracted using
RNeasy mini kit (Qiagen, Leusden, The Netherlands) according to the manufacturer’s instructions for cells
and animal tissues (including DNase treatment) [39,40]. RNA concentrations and purity were assessed
measuring the absorbance at 230, 260 and 280 nm using Nanodrop ND-1000 (ThermoScientific, Villebon-sur-
Yvette, France). Finally, RNA integrity was checked using the RNA Nano 6000 assay (Agilent Technologies,
Diegem, Belgium) and 2100 Bioanalyzer (Agilent, Santa Clara, CA, USA)[39]. All RNA samples displayed no
degradation patterns (sharp peaks and clean baseline).
The extracted RNA were reverse-transcribed into cDNA using the Protoscript II reverse transcriptase (New
England Biolabs, Leiden, The Netherlands) and random primers, according to the manufacturer’s instructions.
cDNAs were further diluted to 0.8 ng. µL-1 and used for RT-qPCR analyses using a 384-well plate design. An
automated liquid handling robot (epMotion 5073, Eppendorf, Hambourg, Germany) was used to properly
prepare the 384-well plates. qPCR were performed using the Takyon Low ROX SYBR MasterMix dTTP Blue Kit
(Eurogentec, Liège, Belgium) on a ViiA 7 Real-Time PCR System (Thermo-Fisher, Waltham, MA, USA) in a 10
µL final volume [41]. All reactions were performed in technical triplicates and repeated on four biological
replicates. The PCR conditions consisted on an initial denaturation step at 95°C for 3 min, followed by 40
cycles of denaturation at 95 °C for 15 sec and annealing/extension steps at 60 °C for 60 sec.
4.5. Stability of the candidates’ reference genes
The stability of the selected genes was analysed using five different methodologies. GeNormplus performs a
pairwise comparison and generate the M-value which consist on a comparison of the variation of a gene
compared to all the remaining candidates [22] while NormFinder calculates both a single best gene (best
Chapter 3 Reference genes of G. fossarum
111
gene) and an optimal gene pair (best pair) as the best pair may compensate the expression in the different
experimental groups [1,23]. BestKeeper on its side is based on assigning correlation factor of each gene with
the geometric means of all genes [7,42] while RefFinder is an online very easy to use platform which compiles
the three most popular algorithms for reference gene validation based on an input of Cq values only [24,42].
Additionally, a simple delta-Ct comparison approach was applied [25].
4.6. Statistical analyses
Normal distribution of data-set was checked using a Shapiro-wilk test. Homogeneity of variances was checked
using a Levene test. A one-way ANOVA followed by a Fisher-LSD post hoc test (P < 0.05) was performed on
the log2 transformed data (CNRQs) using Statistica 12 software (Statsoft Inc.).
5. CONCLUSIONS
This study provides of a set of reference genes suitable for normalization of RT-qPCR data obtained from G.
fossarum samples. Six candidate genes were tested and five different algorithms allowed the identification
of the most stable sub-set of genes. clathrin and SDH were identified as the most stable genes in our applied
experimental design, while widely used reference genes were unsuitable in G. fossarum in the present work.
Our results highlight how important and crucial is to experimentally define and validate of a set of reference
genes for each RT-qPCR experiment.
Chapter 3 Reference genes of G. fossarum
112
Acknowledgments: This work is supported by the Fonds National de la Recherche (NanoGAM AFR-PhD-
9229040). The authors are grateful to S. Contal and M. Fossépré for technical supports.
Author Contributions: K.M conceived, designed and performed the experiments, prepared tables and figures
and wrote the manuscript. K.M, S.L, S.C analysed the data; S.C. coordinated its revision; All authors read and
provided helpful discussions and approved the final discussion
Conflicts of Interest: The authors declare no conflict of interest.
Abbreviations
RT-qPCR Reverse transcription quantitative polymerase chain reaction
TUB Tubulin
UB Ubiquitin
SDH Succinate dehydrogenase
GAPDH Glyceraldehyde-3-phosphate dehydrogenase
HSP90 Heat shock protein 90
AgNPs Silver nanoparticles
AuNPs Gold nanoparticles
AgNO3 Silver nitrate
NF Normalization factor
Vn Pairwise variations
Chapter 3 Reference genes of G. fossarum
113
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COMPLEMENT 2
Identification of reference genes for RT-qPCR data normalization in Gammarus fossarum
(Crustacea Amphipoda)
Kahina Mehennaoui 1,2, Sylvain Legay 1, Tommaso Serchi 1, François Guérold 2, Laure Giambérini 2,
Arno C. Gutleb 1, Sébastien Cambier 1*
1 Environmental Research and Innovation (ERIN) Department, Luxembourg Institute of Science and
Technology (LIST), 5, avenue des Hauts-Fourneaux, Esch-sur-Alzette, Luxembourg
2 Laboratoire Interdisciplinaire des Environnements Continentaux (LIEC), CNRS UMR 7360,
Université de Lorraine – Metz, France
Article submitted to Scientific Reports
Chapter 3 Reference genes of G. fossarum
118
Table S1: Identification of Gammarus fossarum gene sequences
Chapter 3 Reference genes of G. fossarum
119
Table S2: Cq values (mean ± SD) of the different genes regarding the different exposure conditions.
a CIT: citrate stabilized NPs, PEG: polyethylene-glycol coated NPs b 0: this condition is the control condition for each compound
Compounda Concentrations(µg.L-1)b Actin Clathrin GAPDH SDH TUB UB HSP90
AgNO3
0 21.60± 0.57 21.40± 0.30 18.69± 0.52 19.99± 0.40 18.60± 0.28 24.16± 0.34 18.73± 0.48
0,5 20.93± 0.30 21.23± 0.04 18.93± 0.20 19.82± 0.13 18.01± 0.22 23.86± 0.11 18.48± 0.20
AgNPs40nm
0 21.35± 0.21 21.12± 0.57 18.65± 0.13 20.21± 0.08 18.39± 0.11 23.39± 0.60 19.22± 0.10
CIT 0,5 22.22± 0.24 21.32± 0.03 18.52± 0.38 20.23± 0.10 17.97± 0.21 23.85± 0.12 18.78± 0.09
PEG 0,5 21.67± 0.10 21.34± 0.05 18.60± 0.21 20.31± 0.03 18.02± 0.32 23.94± 0.12 19.03± 0.29
AuNPs40nm
0 21.55± 0.17 21.21± 0.13 18.78± 0.14 20.17± 0.06 18.07± 0.30 24.01± 0.05 18.57± 0.28
CIT 0,5 21.45± 0.15 21.25± 0.08 18.85± 0.18 20.07± 0.03 18.14± 0.29 24.00± 0.11 19.16± 0.09
PEG 0,5 21.56± 0.15 21.26± 0.13 18.71± 0.20 19.98± 0.08 18.69± 0.13 23.58± 0.60 19.04± 0.17
117
Chapter 4 AgNPs and AuNPs acute toxicity
118
CHAPTER 4
AgNPs and AuNPs acute toxicity
CHAPTER 4
AgNPs and AuNPs acute toxicity
Chapter 4 AgNPs and AuNPs acute toxicity
119
KEY FINDINGS
Aim of the study
This study aimed at investigating the influence of the size (20, 40 and 80 nm) and surface coating (CIT and
PEG) on the uptake, tissue distribution and molecular effects of AgNPs and AuNPs on G. fossarum exposed
for 72h to AgNPs and AuNPs, in the absence of food
Experimental design
Key findings
• Coating-dependent uptake of Ag and Au o CIT-AgNPs > PEG-AgNPs
o CIT-AuNPs > PEG-AuNPs
• No size-dependent effects on bioaccumulation
• Concentration-dependent uptake
• Influence of chemical composition on the internal distribution was observed as o AgNPs were found only in G. fossarum gills
o AuNPs were found only in G. fossarum gut
• A metal-dependent effect on gene expression was also observed with (e.g.) o AgNPs leading to an increase in MnSOD and CuZnSOD gene expression
o AuNPs leading to a decrease in MnSOD and CuZnSOD gene expression
Luxembourg
Gammarus fossarum
Organisms Materials
Schwaarzbaach
Sampling site
Ag20nm Ag40nm Ag80nm
Citrate stabilized AgNPs and PEG coated AgNPs
Au20nm Au40nm Au80nm
Citrate stabilized AuNPs and PEG coated AuNPs
Acclimation
Volvic water
D1
1/3
Exposure
T0h T24h T48h T72hD4 D10D2 D3
2/33/3 Temperature : 12 oC
Photoperiod : 16h/8h
Daily renewal of Volvic water
Temperature : 12 oC
Photoperiod : 16h/8h
Daily renewal of exposure medium
Stalling
Experimental design
Chapter 4 AgNPs and AuNPs acute toxicity
120
ARTICLE 3
Submitted to Science of the Total Environment
Influence of size and surface coating on uptake of silver and gold nanoparticles and the molecular effects
on Gammarus fossarum (Crustacea Amphipoda)
Kahina Mehennaoui a,b, Sébastien Cambier a, Tommaso Serchi a, François Guérold b, Johanna Ziebel a, Jean-
Sebastien Thomann c, Nathalie Valle c, Laure Giambérini b, *, Arno C. Gutleb a*
A Environmental Research and Innovation (ERIN) Department, Luxembourg Institute of Science and
Technology, 41 rue du Brill, Belvaux, Luxembourg
B Laboratoire Interdisciplinaire des Environnements Continentaux (LIEC), CNRS UMR 7360, Université de
Lorraine – Metz, France
C Material Research and Technology (MRT) Department, Luxembourg Institute of Science and Technology, 41
rue du Brill, Belvaux, Luxembourg
Chapter 4 AgNPs and AuNPs acute toxicity
121
Influence of size and surface coating on uptake of silver and gold nanoparticles and the
molecular effects on Gammarus fossarum (Crustacea Amphipoda)
Kahina Mehennaoui a,b, Sébastien Cambier a, Tommaso Serchi a, François Guérold b, Johanna Ziebel
a, Jean-Sebastien Thomann c, Nathalie Valle c, Laure Giambérini b, *, Arno C. Gutleb a*
A Environmental Research and Innovation (ERIN) Department, Luxembourg Institute of Science and
Technology, 41 rue du Brill, Belvaux, Luxembourg
B Laboratoire Interdisciplinaire des Environnements Continentaux (LIEC), CNRS UMR 7360,
Université de Lorraine – Metz, France
C Material Research and Technology (MRT) Department, Luxembourg Institute of Science and
Technology, 41 rue du Brill, Belvaux, Luxembourg
* Corresponding authors: Arno C. Gutleb
Postal Address: Environmental Research and Innovation Department
Luxembourg Institute of Science and Technology
5, avenue des Hauts-Fourneaux, L-4362 Esch-sur-Alzette, Luxembourg
E-mail : [email protected]
Laure Giamberini
Postal Address : Université de Lorraine, Laboratoire Interdisciplinaire des Environnements
Continentaux (LIEC, CNRS UMR 7360)
Campus Bridoux, rue du General Delestraint, F-57070 Metz
E-mail : [email protected]
Chapter 4 AgNPs and AuNPs acute toxicity
122
ABSTRACT
The specific and unique properties of silver nanoparticles (AgNPs) and gold nanoparticles (AuNPs), make
them of high interest for different scientific and industrial applications. Their increasing use will inevitably
lead to their release in the environment and aquatic ecosystems where they may represent a threat to
aquatic organisms. Being a widespread and important component of the aquatic macroinvertebrate
assemblage, amphipods and more specifically Gammarus fossarum will certainly be exposed to AgNPs and
AuNPs. For these reasons, G. fossarum was selected as model organism for this study.
The aim of the present work was the evaluation of the influence of both size (20, 40 and 80 nm) and surface
coating (citrate CIT, polyethylene glycol PEG) on the acute toxicity of AgNPs and AuNPs on G. fossarum. We
investigated the effects of AgNPs and AuNPs on the uptake by G. fossarum, NP tissue distribution and the
expression of stress related genes by the use of ICP-MS NanoSIMS 50 and Cytoviva, and RT-qPCR,
respectively.
Ag and Au bioaccumulation revealed a significant surface-coating dependence, with CIT-AgNPs and CIT-
AuNPs showing the higher bio-accumulation potential in G. fossarum as compared to PEG-NPs. Opposite to
that, no size-dependent effects on the bioaccumulation potential were observed. NanoSIMS 50 and CytoViva
revealed an influence of the type of metal on the tissue distribution after uptake, with AgNPs detected in the
cuticle and the gills of G. fossarum, while AuNPs were detected in the gut area. Furthermore, AgNPs were
found to up-regulate MnSOD and CuZnSOD gene expression while AuNPs led to their down-regulation.
Modulation of SODs may indicate generation of reactive species of oxygen and a possible activation of
antioxidant defence in order to prevent and defend the organism from oxidative stress. However, further
investigations are still needed to better define the mechanisms underlying the observed AgNPs and AuNPs
effects.
Keywords: Gammarus fossarum, Silver nanoparticles, Gold nanoparticles, bioaccumulation, gene expression,
NanoSIMS 50, Cytoviva®
Chapter 4 AgNPs and AuNPs acute toxicity
123
1 INTRODUCTION
Engineered nanoparticles (ENPs) are being more and more incorporated in consumers’ products such as
cosmetics, paints, food, computers, medicine, etc. (McGillicuddy et al., 2017; Vale et al., 2016). As a
consequence, their environmental release is raising concerns about their potential toxicity/effects on human
and other living organisms (Klaine et al., 2008; Moreno-Garrido et al., 2015). Among all ENPs, silver
nanoparticles (AgNPs) are the most commercialized and the most used ones in consumer products; such as
clothing, health-care products, food packaging, etc. (Fabrega et al., 2011; Vale et al., 2016) mostly due to
their well-documented antibacterial properties (Bone et al., 2012; Cleveland et al., 2012; McGillicuddy et al.,
2017; Zhang et al., 2016). Gold nanoparticles (AuNPs) represent another example of widely used
nanomaterials. They are mostly used for medical and molecular biology applications mainly for their
supposedly low toxicity and high bio-compatibility (García-Negrete et al., 2013; Lapresta-Fernández et al.,
2012).
The behaviour and fate of AgNPs and AuNPs in aquatic ecosystems and their toxicity towards aquatic
organisms are influenced by their physical-chemical properties such as size and surface chemistry (Angel et
al., 2013; Ellis et al., 2016; Makama et al., 2016). The surface of NPs is commonly modified by addition of
capping agents of surface functionalization in order to prevent aggregation and ion release (Gao et al., 2012;
Levard et al., 2012; Park et al., 2015a; Tolaymat et al., 2010). Common capping agents are citrate (CIT), which
is applied as stabilizer by charge repulsion (Tejamaya et al., 2012), and polyethylene glycol (PEG), which is
used as a sterically stabilizer and described to strongly bind to the core of NPs (Cumberland and Lead, 2009;
Römer et al., 2011; Tejamaya et al., 2012). Although surface-modified AgNPs and AuNPs are widely used, the
influence of these compounds on the behaviour, uptake and toxicity of AgNPs and AuNPs on aquatic
organisms is still not fully understood (Croteau et al., 2011; Kwok et al., 2012; Park et al., 2015).
Amphipods of the genus Gammarus such as Gammarus fossarum represent an important part of the aquatic
macroinvertebrates assemblage (Dedourge-Geffard et al., 2009; Ladewig et al., 2006). Recently, G. fossarum
have been used as a model organism due to their wide distribution in Europe (Janetzky, 1994) and their major
functional role in litter breakdown process and nutrient cycling (Forrow and Maltby, 2000; Kelly et al., 2002;
Lacaze et al., 2010; MacNeil et al., 1997). The ease of use, identification to the species level, differentiation
between gender, sampling and laboratory handling in addition to the well-documented sensitivity to different
kind of pollutants make Gammarus good model organisms for ecotoxicological studies (Arce Funck et al.,
2013; Mehennaoui et al., 2016). Gammarids have been recognized as a relevant tool for water quality and
risk assessment as they allow the assessment of their internal contamination level that is always dependent
and proportional to the amount of contaminants introduced in the environment (Kunz et al., 2010; Vellinger
et al., 2012a).
Chapter 4 AgNPs and AuNPs acute toxicity
124
The aim of the present study was to investigate both the influence of coatings (CIT and PEG) and size (20, 40
and 80 nm) on the uptake and molecular effects of AgNPs and AuNPs on the aquatic amphipod G. fossarum.
Effects of AgNO3 were also studied, as a positive control, in order to determine the contribution of the
released ions to the observed biological effects. The biological endpoints that were measured in this study
include mortality, bioaccumulation, localization of AgNPs and AuNPs in G. fossarum tissues and expression
of stress-related genes. Total and ionic Ag and Au were measured in the exposure media in order to
determine the behaviour and aging of AgNPs and AuNPs.
2 MATERIALS AND METHODS
2.1 Particles and Chemicals
Citrate stabilized Ag nanoparticles (CIT-AgNPs 20, 40 and 80 nm) and citrate stabilized Au nanoparticles (CIT-
AuNPs 20, 40 and 80 nm) were purchased from Nanocomposix (San Diego, USA). Coating of AgNPs and AuNPs
with PEG was performed at the LIST’s Material Research and Technology (MRT) Department, (Luxembourg
Institute of Science and Technology (LIST), Luxembourg). AgNO3 (CAS no.7761-88-8) was purchased from
VWR (Leuven, Belgium). Osmium tetroxide (OsO4, CAS no.20816-12-0), glutaraldehyde (CAS no.111-30-8) and
the Spurr low viscosity resin embedding kit were purchased from Sigma Aldrich (Bornem, Belgium).
Phosphate buffer saline (PBS) was obtained from Invitrogen (Merelbeke, Belgium).
2.2 Particle characterization
CIT-AgNPs and CIT-AuNPs were characterized by the manufacturer and the information provided include TEM
images (JEOL1010 Transmission Electron Microscope), mass concentration (Thermo Fisher X series 2 ICP-MS),
Hydrodynamic diameter and zeta potential (Malvern nanoZetasizer), UV-visible extinction spectrum and pH
were provided by Nanocomposix
(https://tools.nanocomposix.com:48/cdn/coa/Silver/Spheres/BioPure/AG20-SS-BP-CIT-
SCM0069.pdf?202%20903).
Functionalization of AgNPs and AuNPs surface was performed at the Material Science and Technology
department (MRT) in the Luxembourg Institute of Science and Technology (LIST). AgNPs and AuNPs were
coated with PEG (MW layer 3000) following a wet-chemistry (Thiol/metal chemistry) method that allowed
functionalization of each isolated NPs. AgNPs and AuNPs coatings were controlled using Scanning Electron
Microscopy (SEM, Figure S1 A) and size distribution in MilliQ water was checked using Nanoparticle Tracking
Analysis (NTA, measured on a NanoSIGHT instrument, Malvern Instrument Ltd, UK(Hole et al., 2013)), (Figure
S1 B).
Chapter 4 AgNPs and AuNPs acute toxicity
125
In addition to the information provided by the manufacturer, we performed complementary characterization
including size distribution and zeta potential measurement. Size distribution of CIT-AgNPs, PEG-AgNPs, CIT-
AuNPs and PEG-AuNPs in Volvic water (exposure medium), at T0h and T24h, were characterized using NTA.
Zeta potential in exposure medium was analysed using NanoZetasizer (Malvern Instrument Ltd, UK).
2.3 Organism sampling and acclimation
G. fossarum were collected in June 2015 (CIT-AgNPs and PEG-AgNPs exposure), September 2015 (CIT-AuNPs
exposure) and October 2015 (PEG-AuNPs exposure) at an unpolluted stream (49°48’24.9’’ N and 06°04’53.2’’
E, Schwaarzbaach, Colmar-berg, Luxembourg) (Dohet et al., 2008; Mehennaoui et al., 2016). Animals were
collected using a hand net and were sorted in the field to separate males, which were used for the
experiments, from females, which were placed back in the river. G. fossarum males were selected from pre-
copula pairs or based on sexual dimorphism like gnatopode size (Mehennaoui et al., 2016). Adult males were
kept for the experiments in order to avoid influence of gender on the studied parameters (Arce Funck et al.,
2013; Sornom et al., 2010). Collected male animals were immediately brought in river water to the laboratory
and were kept at 12 °C. Adult males were then acclimated to laboratory conditions (Andreï et al., 2016;
Mehennaoui et al., 2016). The acclimation was conducted in 2 steps. First, Gammarids were acclimated for
72h to mineral water (Volvic, France) by progressively changing field water to Volvic water (30% v/v, 50% v/v,
100% v/v). Then, a stalling period of 10 days was conducted in 100% Volvic water (Andreï et al., 2016;
Mehennaoui et al., 2016). The acclimation was performed under controlled conditions at 12 °C with a 16h
light 8h dark photoperiod. Volvic water was aerated and changed every 24h to avoid organic matter
accumulation and potential increase of ammonium, nitrite and nitrate. Gammarids were fed ad libitum with
alder leaves (Alnus glutinosa) up to 24h before experiment.
2.4 Acute toxicity test
In order to avoid as much as possible the adsorption of AgNO3, AgNPs and AuNPs during the exposure, the
tanks used for exposure were saturated for 96h before exposure (Andreï et al., 2016).
At the end of the acclimation period, groups consisting of 6 males each were transferred into plastic tanks
(250 ml polypropylene tanks) containing 70 ml of exposure medium (Volvic water with or without
contaminants) (Mehennaoui et al., 2016). A piece of mesh was added in each tank to provide a resting surface
for G. fossarum and to reduce as much as possible potential losses linked to cannibalism. Males were exposed
to different treatments: AgNO3 (0.5 – 1 – 2 – 4 – 8 µg.L-1); CIT-AgNPs and PEG-AgNPs 20 nm (1 – 2 -4 – 8 – 10
µg.L-1), CIT-AgNPs 40 nm, PEG-AgNPs 40 nm, CIT-AgNPs 80 nm, PEG-AgNPs 80 nm, CIT-AuNPs and PEG-AuNPs
20, 40 and 80 nm (1 – 5 – 10 – 25 – 50 µg.L-1), and a control for 72h at 12 °C with a photoperiod of 16h light
Chapter 4 AgNPs and AuNPs acute toxicity
126
and 8h darkness and in absence of feeding. The different ranges of concentrations tested were selected based
on previous results obtained by exposing D. magna to the same silver and gold nanoparticles (Table S1).
AgNO3, AgNPs and AuNPs stock solutions were diluted to the desired concentrations in mineral water (Volvic,
France). 100% of exposure medium were renewed daily. Dead animals were removed and survivors were
counted.
2.4.1 Total and dissolved silver and gold measurements
At the end of exposure period, 4 mL of exposure medium were analysed for ions release measurements.
Medium were centrifuged at 4000g for 40 min using a 3KDa cut off membrane (Millipore) followed by
mineralization step with 1% HNO3 for Ag ions measurements and 1% HCl for Au ions measurements.
Total Ag and Au concentrations were measured in the exposure medium sampled from the water column at
the end of the exposure period. 5 mL of exposure medium/replicate were mineralized using 1% (v/v) of HNO3
and HCl for Ag and Au measurements, respectively. All measurements of total and dissolved silver and gold
were performed with Inductively Coupled Plasma Mass spectrometer (ICP-MS) (ElanDRC-e, Perkin-
ElmerWaltham, MA, USA) as previously published (Mehennaoui et al., 2016). All measurements were
performed in the absence of animals. The quantification limit of ICP-MS for Ag was around 0.051 µg. L-1 and
0.255 µg. L-1 for Au. A calibration curve was performed following serial dilutions (from 0.01 to 50 µg. L-1 of Ag
and from to for Au) using a multi-element standard solution (Chemlab, certified ISO/IEC 17025). The validity
of the analytical method was checked every 10 measurements within the same series of measurements with
3 quality controls (0.5, 5 and 25 µg.L-1 of Ag and Au) of a second multi-element standard solution (Merck-
VWR, ISO/IEC 17025); Ag and Au values were consistent within the certified ranges (Cambier et al., 2018).
2.4.2 Silver and gold bioaccumulation
At the end of the exposure period, three pools of 4 gammarids each per conditions were rinsed with milliQ
water and dried on filter paper. Animals were frozen at -20 °C, freeze-dried and weighted. Samples were then
mineralized in HNO3 and H2O2 for Ag measurement and in acidic mixture of HNO3 and HCl for Au
measurement at a maximum pressure of 35 bars and maximum temperature of 200 °C using a microwave
(Anton Paar Multiwave Pro) as previously described (Mehennaoui et al., 2016). Results are expressed in µg.g-
1 of dry weight. All measurements were performed in triplicates. A calibration curve was performed following
serial dilutions (from 0.01 to 50 µg.L-1 of Ag and from 0.025 to 10 µg.L-1 for Au). Using a multi-element standard
solution (Chemlab, certified ISO/IEC 17025). The validity of the analytical method was checked every 10
measurements within the same series of measurements with 3 quality controls (0.5, 5 and 25 µg.L-1 of Ag and
Chapter 4 AgNPs and AuNPs acute toxicity
127
Au) of a second multi-element standard solution (Merck-VWR, ISO/IEC 17025); Ag and Au values were
consistent within the certified ranges (Cambier et al., 2018).
2.4.3 Nanoparticle uptake
2.4.3.1 Sample preparation
At the end of the exposure period, G. fossarum exposed to 10 µg.L-1 of AgNPs and AuNPs were fixed in 5%
(v/v) glutaraldehyde where they were kept at 4 °C until analyses. After a washing step with PBS, samples
were post-fixed with 1% osmium tetroxide for 2h at room temperature on a horizontal shaker (180 rpm).
After a second washing with PBS, samples were dehydrated with increasing concentrations of ethanol (10%,
20%, 30%, 50%, 70% 90% and 100% v/v) and a final wash in acetone (100% v/v) before being embedded in
Spurr resin for 16h at 60 °C.
2.4.3.2 NanoSIMS50 analyses
Semi-thin layers of 300 nm were cut using ultra-microtome and placed on silicon wafers (Siltronix, Archamps,
France) for Secondary Ion Mass Spectrometry (SIMS) analysis (Georgantzopoulou et al., 2013; Mehennaoui
et al., 2016). Samples were analysed with a NanoSIMS 50 (Cameca, Gennevilliers, France) using Cs+ as primary
ion source (8kev) sputtering the surface of the sample (-8 kev) with a raster of 40 x 40 μm2 to generate
secondary negative ions for silver analyses. Images were recorded in a pixel format of 256 x 256 image points
with a counting time of 30 ms per pixel. The instrument was tuned for a mass resolution (M/ΔM) up to 5000
for the elimination of atomic or molecular isobar interference. The mass calibration of the silver ion was
realized using a silver foil (Goodfelow, Huntingdon, UK). The isotopic ratio between 107Ag- (m= 106.9051,
51.8%) and 109Ag- (m=108.9048, 48.2%) was measured and verified (ratio = 1,074). The mass calibration of
gold ions 197Au- (m= 196.966543 amu) was realized using a gold leaf. The 12C14N- cluster was simultaneously
detected with the silver ions to allow the recognition of the essential anatomical features (Eybe et al., 2009;
Georgantzopoulou et al., 2013).
2.4.3.3 Cytoviva® dark field hyperspectral imaging
Semi-thin layers of 300 nm were cut using ultra-microtome and placed on microscope glass slide and let to
dray for 1 min on a drying workstation. Slides were covered with a glass cover slip, sealed, and kept at 4 °C
until analysis. Visualization was performed using Cytoviva® hyperspectral imaging system (Cytoviva Inc.,
Auburn, Alabama, USA) mounted on Olympus BX-43 optical microscope. Images of the different tissues of G.
fossarum were captured at 60x oil immersion magnification using hyperspectral camera controlled by
environment for visualization ENVI software (version 4.8 from Harris Corporation, Melbourne, FL, USA and
Chapter 4 AgNPs and AuNPs acute toxicity
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modified by CytoViva, Inc.). Spectral libraries of exposed animals were generated manually acquiring about
200 spectra per samples. Acquired libraries were filtered against non-exposed samples to filter out all spectra
non-related to AgNPs or AuNPs using a spectral angle mapper (SAM) algorithm with a 0.05 radians tolerance.
Filtered libraries were mapped onto images of exposed samples using SAM with a 0.05 radian tolerance which
allows highlighting similarities between the spectra in the image and in the spectral library.
2.4.4 Molecular responses
2.4.4.1 RNA extraction and cDNA synthesis
At the end of the exposure period, animals were gently dried and dived in RNA lysis Buffer (RLT buffer)
supplemented with 1% (v/v) β-mercaptoethanol, flash-frozen in liquid nitrogen and stored at -80 °C until RNA
extraction. For RNA extraction, G. fossarum tissues were grinded on ice using a pellet pestle tissue grinder
and the homogenates were centrifuged at 250g for 5 min at 4 °C to remove cuticle fragments as previously
described (Mehennaoui et al., 2016). Total RNA was extracted using RNeasy mini kit for cells and animal
tissues (Qiagen, Leusden, The Netherlands), according to the manufacturer’s instructions (including digestion
with DNase). RNA concentrations and purity were assessed measuring the absorbance at 230, 260 and 280
nm using Nanodrop ND-100 spectrophotometer (ThermoScientific, Villebon-sur-Yvette, France) (A260/280
and A260/230 ratios). RNA integrity was checked using the RNA Nano 6000 assay (Agilent Technologies,
Diegem, Belgium) with a 2100 Bioanalyzer (Agilent, Santa Clara, CA, USA) (Legay et al., 2015). For the
synthesis of cDNA, 1 µg of the extracted RNA were retro-transcribed into cDNA using the Protoscript II reverse
transcriptase (New England Biolabs, Leiden, The Netherlands) and random primers, according to the
manufacturer’s instructions.
2.4.4.2 Primer design and quantitative real-time PCR
All the primers for the RT-PCR were designed using Primer3Plus (http://primer3.ut.ee) with the following
criteria: primer size between 18 and 25 base pairs, GC content between 30% and 70 %, amplicon size from
80 to 150 base pairs, primer annealing temperatures in the 57-61 ºC range. Primers were analysed using
NetPrimer (http://www.premierbiosoft.com/netprimer/) for secondary unexpected structures. PCR
efficiency was checked using decreasing five-fold dilutions from cDNA pool (From 25 ng to 0.04 ng and no
template control). Melting curves were generated at the end of each experiment to assess the specificity of
the amplified products. All primers displayed one clear peak and were all retained for the analysis. All the
used primers are described in Table 4. 1
cDNAs were diluted to 0.8 ng. µL-1 and used for RT-qPCR analyses in 384-well plates. An automated liquid
handling robot (epMotion 5073, Eppendorf, Hamburg, Germany) was used for the preparation of the 384-
Chapter 4 AgNPs and AuNPs acute toxicity
129
well plate. cDNA were amplified using Takyon Low ROX SYBR MasterMix dTTP Blue Kit (Eurogentec, Liège,
Belgium) on a ViiA 7 Real-Time PCR System (Thermo Fisher, Waltham, MA, USA) in a final volume of 10 µL
(Behr et al., 2015). All reactions were performed in three technical replicates each one including four
biological replicates. The PCR conditions consisted on an initial denaturation step at 95°C for 3 min, followed
by 40 cycles of denaturation at 95°C for 15 sec and annealing/extension steps at 60°C for 60 sec. Three genes,
namely Ca2+ATPase, Clathrin and SDH, were selected among seven genes to be used as reference and they
were validated using the GeNorm module in the Biogazelle qBase PLUS software (Mehennaoui et al., subm.
Chapter 3). The relative gene expression was calculated including PCR efficiency using Biogazelle qbase Plus
software 2.5 with the ΔΔCT method.
Chapter 4 AgNPs and AuNPs acute toxicity
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Table 4. 1. Specific primer pairs used for RT-qPCR analyses on Gammarus fossarum exposed for 72h to AgNPs and AuNPs (coated with CIT or PEG, sizes: 20, 40 and 80
nm, F: forward sequence, R: reverse sequence)
Gene code Description Primer sequence (5’ – 3’) Amplicon size (bp) PCR efficiency
Actin Actin F: CTCATGCTATCCTTCGTCTTGA
103 2.02 R: CGTTCAGCGGTGGTTACAA
Tub Tubulin F: CGGCTGTTGTTGAACCTTAC
93 2.09 R: AGATGGCCTCATTGTCAACC
GAPDH glyceraldehyde 3-phosphate dehydrogenase F: GTCCGTCTCGCTAAGGAGTG
94 1.91 R: TGTATCCGAGGTAGCCCTTG
SDH succinate dehydrogenase F: GGAAGAAGCTGGATGGTCTG
87 1.98 R: ACTTGTCTCCGTTCCACCAG
UB Ubiquitin F: CCCACGATACTCCCTTTGAA
82 2.01 R: ACAATCGGTGGCTTGTTAGG
Ca2+ATPase Calcium ATPase F: CTACGCCAATCAGTTCGTCA
87 2.03 R: TCACAGCGAGAGGAAGACCT
Clathrin Clathrin F: ATCGCCAAGCTTTGTGACTT
107 1.99 R: GCTTTGATAGGCGGACTCTG
MnSOD manganese superoxide dismutase F: AAACGCGACTTTGGAAGTGT
150 1.99 R: TTCTAGCGGATCCTGATTGG
CuZnSOD copper zinc superoxide dismutase F: TTGCCTTGTTGTTGTTGGTG
95 1.92 R: GGCCGTGTTGATGTAGAGGT
P53 Tumour protein P53 F: ACGAGCTCTCCTTCAAGTGC
104 R: CTTGGAAGGTATCCAGTGCAG
HSP90 heat shock protein 90 F: CTGGTTTCTTCTCCCTGCTG
135 1.99 R: GATCTCGAGGTGCTTCTTGG
Gadd45 Growth arrest and DNA damage 45 F: CACACGACGACAAGTTCCTG
91 1.92 R: TGCTGCTTCAATTGCTTCTT
GPx7 glutathione peroxidase 7 F: AGCGTGCTAGTCCAGCTTTC
106 1.95 R: GCATTGATCACGACACCATC
Chapter 4 AgNPs and AuNPs acute toxicity
131
HEM hemocyanin F: GGAGAGGATATCGGCATGAA
124 1.94 R: GGTGATGAGCCCAGAAGAAA
CAT catalase F: CTCGTCAACAAGGACAATGC
130 1.97 R: GGTCTCTGATGGCGTAGTCG
CathL Cathepsin L F: GGTTCACTCGAAGGTCAACAC
83 1.95 R: TTGTCTGAACAATCGACCAAG
Cjun C-Jun F: GATCGGTCGTTTGGAAGAAA
94 1.95 R: TGCACACACTTGATCACGTAAC
Cox1 Cyclooxygenase 1 F: GGGAATCAATTTCATTACTACTTGTT
89 1.95 R: TTATTGCCCAAACGAAGAGG
GST Glutathione S transferase F: TGCCTCTAGGTCAACTTCCAG
80 1.98 R: AGCAACATAACGAGCAATCG
NaKATPase Sodium potassium ATPase F: ATTCGCTTGGCTGTGAGG
85 2.02 R: TCCTCATCGGTATCATCGTG
NfkB nuclear factor kappa B F: GCAGAGCATCGACTTGAATG
101 2.09 R: GCGAGACGACAGGAGTAAGG
Chitinase Chitinase F: CTCGTCAACAAGGACAATGC
130 1.96 R: GGTCTCTGATGGCGTAGTCG
Chapter 4 AgNPs and AuNPs acute toxicity
132
2.5 Statistical analyses
LC50 values were calculated on AgNO3 nominal concentrations based on a logistic curve-fitting procedure
applying the method described earlier (Isnard et al., 2001; Vindimian et al., 1999). The Excel macro REGTOX
allowed us to estimate both LC50 values using Hill model (Vindimian et al., 1983) and their 95% confidence
intervals (Vellinger et al., 2012b).
Bioaccumulation results are reported as mean ± SD (standard deviation) and relative gene expression results
are reported as mean ± SE (standard error). The observed Ag and Au bioaccumulation were compared using
two-way ANOVA followed by Tukey-HSD post-hoc test for multiple comparisons. The weighed means were
calculated to identify in one hand the “coating effects” when results of all sizes were combined and in the
other hand, the “size effects” were compared while the coatings were combined.
Effects on molecular responses were analysed on log2 transformed normalized relative quantities. Two-way
ANOVA followed by a Fisher LSD post hoc test was performed. All statistical calculations were performed
using Statistica software 12 (StatSoft Inc.).
3 RESULTS
3.1 Nanoparticle characterization
Analysis of the hydrodynamic diameter of NPs by NTA, showed that both PEG-AgNPs and PEG-AuNPs
were generally more stable in Volvic water, both at T0h and T24h, than CIT-AgNPs and CIT-AuNPs (Table 4. 2
and Figure S2). After 24h in exposure medium, CIT-AgNPs 20 nm appeared to be the most aggregating particle
with a size of 53 ± 3.7 nm, while CIT-AgNPs 40 and 80 nm appeared to be more stable, showing hydrodynamic
diameter of 45 ±1.5 nm and 72 ± 2.9 nm, respectively. Similarly, CIT-AuNPs 20 and 40 nm tented to aggregate
after 24h in exposure medium with CIT-AuNPs 20 nm being the one that aggregate the most with a size of
154 ± 28 nm. CIT-AuNPs 40 nm and CIT-AuNPs 80 nm had sizes of 67 ± 12.8 nm and 85 ± 0.3 nm, respectively.
Zeta potentials were ranging from -4.8 to -19.1 mV for all the studied AgNPs and AuNPs with CIT-NPs
presenting more negative charges at their surface than PEG-NPs (Table 4. 2).
Chapter 4 AgNPs and AuNPs acute toxicity
133
Table 4. 2. Size distribution of particles (mode ± SD, 3 replicates) and ζ potential of AgNPs and AuNPs in Volvic
water (exposure medium) at T0h and T24h.
3.2 Acute toxicity test
3.2.1 Survival
As expected, AgNO3 was the most toxic compound. A significant seasonal variability was observed with LC50
values ranging from 0.78 µg. L-1 (0.001 – 0.85) in autumn (November 2015) to 4.31 µg.L-1 (3.88 – 4.75) in
summer (June 2015) (Figure S3). None of the tested NPs, irrespective of the core material (Ag or Au), coating
or size, induced increase of mortality at any of the tested concentrations (from 1 to 50 µg. L-1).
3.2.2 AgNP exposure
3.2.2.1 Total and dissolved [Ag]
At the end of the experiment, exposure medium was analyzed by ICP-MS to quantify the total concentration
of total silver. Total Ag measured for control groups were under the limit of quantification (< 0.051 µg. L-1,
Table 4.3). Total Ag recovered from CIT-AgNPs 20 nm was higher than Ag recovered from PEG-AgNPs 20 nm
(Table 4.3) while a higher recovery of total Ag introduced as PEG-AgNPs 40 nm and PEG-AgNPs 80 nm was
observed compared to CIT-AgNPs 40 and 80 nm. The recovery rates of total Ag from AgNO3 was in the same
concentration range than PEG-AgNPs 40 and 80 nm (Table 4.3).
Concentrations of dissolved Ag ions from AgNPs were detectable (3.39 ± 0.4 µg. L-1 of Ag only at the highest
concentrations of PEG-AgNPs 40 nm (50 µg.L-1), while in all other treatments the amount of released Ag ions
were below the limit of detection (< 0.051 µg.L-1).
Particles Volvicwater(T0h) Volvicwater(T24h)
Coating Nominalsize(nm) Size(nm) ζpotential(mV) Size(nm) ζpotential(mV)
AgNPs
Citrate
20 61±2.1 -9.8 53±0.7 -19.1
40 41±0.3 -13.6 45±1.5 -7.2
80 78±2.9 -16.8 72±2.9 -18.6
PEG
20 22±0.7 -2.8 20±1.0 -5.3
40 43±1.2 -6.0 47±1.3 -7.2
80 77±1.2 -6.1 75±2.1 -8.6
AuNPs
Citrate
20 42±1.5 -6.0 154±28 -17.1
40 38±0.9 -14.1 67±12.8 -19
80 75±1.0 -14.3 85±0.3 -19
PEG
20 39±1.5 -0.1 40±2.5 -8.8
40 59±1.2 -6.7 38±19.2 -4.8
80 89±0.9 -11.8 86±1.5 -9.2
Chapter 4 AgNPs and AuNPs acute toxicity
134
Table 4. 3. Total ag concentrations (mean ± SD) and recovery rates (mean ± SD) of AgNO3, CIT-AgNPs and PEG-
AgNPs in Volvic water (exposure medium).
3.2.2.2 [Ag] in Body tissue
Small but detectable amounts of Ag could be detected in the negative control groups of G. fossarum, for
which the observed Ag body concentration was 0.27 ± 0.02 µg. g-1 dry weight (Figure 4. 1). Treatment with
AgNO3 (4 µg. L-1) lead to a significant increase of the Ag body concentrations that were of 7.56 ± 1.89 µg.g-1
d.w. Similarly, treatment with the AgNPs resulted in higher body concentrations of silver in treated animals.
A significant effect of “coating”, “size” and the interaction “coating x size” was observed on the Ag
concentrations in body tissues of G. fossarum (Figure 4. 1 A, Two-way ANOVA, P < 0.05) but the effects were
stronger for the coatings (Figure 4. 1, A, B; Two-way ANOVA, F = 237.3, P < 0.05) than for the size (Figure 4.
1A, C; Two-way ANOVA, F = 1721.4, P < 0.05).
0 <LOQ N.D
0.5 0.20±0.01 39.54±2.78
1 0.45±0.08 44.71±7.91
2 0.91±0.09 45.40±4.38
4 2.76±0.29 68.91±7.29
8 4.92±0.14 61.56±1.77
1 0.80±0.02 80.48±2.08
2 1.68±0.07 84.11±3.67
4 3.22±0.18 80.44±4.38
8 5.91±0.57 73.91±7.16
10 8.70±0.20 87.04±1.98
1 0.47±0.04 46.87±3.83
2 1.10±0.35 54.82±17.56
4 1.94±0.06 48.87±3.83
8 5.17±0.59 64.57±7.33
10 6.97±0.29 69.66±2.86
1 0.49±0.04 48.67±3.91
5 3.21±0.25 64.10±7.91
10 6.79±0.90 67.91±9.04
25 12.58±1.28 50.33±5.12
50 20.26±2.89 40.52±5.77
1 0.97±0.20 97.36±20.15
5 4.33±0.08 86.65±1.51
10 9.01±0.46 90.14±4.57
25 14.67±1.42 56.68±5.66
50 27.60±2.58 55.21±5.16
1 0.77±0.02 77.00±2.34
5 4.37±1.32 87.49±26.42
10 7.21±0.69 72.07±6.93
25 12.17±1.22 48.70±4.86
50 20.38±1.76 40.77±3.51
1 1.10±0.07 >100
5 5.46±0.14 >100
10 10.34±0.55 >100
25 23.68±1.11 94.71±4.45
50 24.43±6.36 48.86±12.72
AgNPs80nm
CIT
PEG
Compounds Nominalconcentrations(µg.L-1)
Control
AgNO3
AgNPs20nm
CIT
PEG
Totalsilverinwater(µg.L-1) Recovery(%)
AgNPs40nm
CIT
PEG
Chapter 4 AgNPs and AuNPs acute toxicity
135
Figure 4. 1. G. fossarum Ag bioconcentration (mean ± SD) after 72h exposure to CIT and PEG-AgNPs 20, 40 and
80 nm.
(A) Coating x size: mean Ag bioconcentration for each tested coating (CIT vs PEG) and sizes (20, 40 and 80 nm).
Vertical bars represent standard deviation
(B) Coating effect: Weighted mean Ag bioconcentration for each tested coating (results for all sizes tested were
combined). Vertical bars represent 0.95 confidence interval
(C) Size effect: Weighted mean Ag bioconcentration for each size tested (results for all coatings tested were
combined). Vertical bars represent 0.95 confidence interval. Different letters (from a to l) represent the significant
differences in mean Ag bioconcentration values (Two-way ANOVA + Tukey-HSD post hoc test at P < 0.05 level of
significance)
A significant concentration-dependent Ag bioconcentration was observed for CIT-AgNPs. The highest
concentrations (10 µg. L-1) of CIT-AgNPs 20 nm and (25 and 50 µg. L-1) of CIT-AgNPs 40 nm and CIT-AgNPs 80
nm led to a significant higher Ag bioconcentration (Figure 4.1, A, C). On the contrary, no significant
concentration-dependent Ag bioconcentration was observed for PEG-AgNPs 20 nm and PEG-AgNPs 80 nm.
Only the highest concentrations (25 and 50 µg. L-1) of PEG-AgNPs 40 nm led to a significant higher Ag
bioconcentration (Figure 4. 1A). At the same Ag concentrations (≈ 5 µg. L-1), G. fossarum exposed to AgNO3
accumulated more silver than those exposed to AgNPs (Figure 4. 1 A, C).
Control
AgN
O30.5
AgN
O31
AgN
O32
AgN
O34
AgN
O38
AgN
Ps201
AgN
Ps202
AgN
Ps204
AgN
Ps208
AgN
Ps2010
AgN
Ps401
AgN
Ps405
AgN
Ps4010
AgN
Ps4025
AgN
Ps4050
AgN
Ps801
AgN
Ps805
AgN
Ps8010
AgN
Ps8025
AgN
Ps8050
5
10
15
20
25
30
0
35
Size(nm)
1
2
3
4
5
6
0
Citrate PEG
AgBioconcentration(µg.gdw
-1)
Coatings
A
B C
Chapter 4 AgNPs and AuNPs acute toxicity
136
3.2.3 AuNP exposure
3.2.3.1 Total and dissolved [Au]
Total gold concentrations in the control tanks were under the limit of detection (< 0.255 µg. L-1). The
concentration of AuNPs in exposure medium were close to the limit of detection leading to missing values
for several conditions and limiting the possibility to exploit the data.
3.2.3.2 [Au] in body tissue
The Au bio-concentration was under the limit of quantification (< 0.075 µg. g-1 d.w) in the non-exposed
animals. Treatment with the AuNPs resulted in higher body concentrations of Au in treated animals. A
significant effect of “coating”, “size” and the interaction “coating x size” was observed (Figure 4.2, Two-way
ANOVA, P < 0.05). For CIT-AuNPs, a significant dose-dependent effect in bio-accumulation is observable, with
higher doses leading to higher body concentrations (Figure 4.2, A B, Two-way ANOVA, F = 669.28, P < 0.05)
while a trend to increasing Au internal concentrations in a dose-dependent manner is also observed for PEG-
AuNPs (Figure 4.2, A,B). On the contrary, no clear influence of NPs size on bio-accumulation is observable.
(Figure 4.2 , A, C).
Legend next page
Citrate PEG
1
2
3
4
5
6
0
7
-1
Coatings
Size(nm)
0
5
10
15
20
25
30
0 0.5 1 2 4 8 1 5 10 25 50 1 5 10 25 50 1 5 10 25 50 1 5 10 25 50 1 5 10 25 50 1 5 10 25 50
Citrate PEG Citrate PEG Citrate PEG
control AgNO3 AuNPs20nm AuNPs40nm AuNPs80nm
Exposureconditions (µg.L-1)
a ababababcabcabc abc abcabc abcd
abcdeabcde
abcde abcdebcdef
cdef def
fggh
gh
i
i
j
ef
h
abcde
AuBioconcentration(µg.gdw
-1)
Control
AgN
O30.5
AgN
O31
AgN
O32
AgN
O34
AgN
O38
AuNPs201
AuNPs205
AuNPs2010
AuNPs2025
AuNPs2050
AuNPs401
AuNPs405
AuNPs4010
AuNPs4025
AuNPs4050
AuNPs801
AuNPs805
AuNPs8010
AuNPs8025
AuNPs8050
5
10
15
20
25
30
0
A
B C
Chapter 4 AgNPs and AuNPs acute toxicity
137
Figure 4. 2. G. fossarum au bioconcentration (mean ± SD) after 72h exposure to CIT and PEG AuNPs 20, 40 and
80 nm.
(A) Coating x size: mean Au bioconcentration for each tested coating (CIT vs PEG) and sizes (20, 40 and 80 nm).
Vertical bars represent standard deviation
(B) Coating effect: Weighted mean Au bioconcentration for each tested coating (results for all sizes tested were
combined). Vertical bars represent 0.95 confidence interval
(C) Size effect: Weighted mean Au bioconcentration for each size tested (results for all coatings tested were
combined). Vertical bars represent 0.95 confidence interval. Different letters (from a to j) represent the significant
differences in mean Au bioconcentration values (Two-way ANOVA + Tukey-HSD post hoc test at P < 0.05 level of
significance)
3.2.4 Particle uptake
3.2.4.1 NanoSIMS50 analyses
Optical microscope images (OM) of the cross section of G. fossarum are presented in Figure 4. 3 A and Figure
4. 4 A. CN pictures are used to identify the anatomical features of G. fossarum (Figure 4. 3 B). Ag could be
detectable in gills of G. fossarum exposed to AgNPs 40 nm (Figure 4. 3). A higher signal of Ag was found in
gills of G. fossarum exposed to CIT-AgNPs 40 nm, more specifically in the cuticle (Figure 4. 3). A lower Ag
signal was found in gills of G. fossarum exposed to PEG-AgNPs 40 nm whereas control animals contained
considerably less Ag (Figure 4. 3).
Figure 4. 3. A) 300 nm cross section of G. fossarum observed under optical microscope. Red square represents
the area observed with NanoSIMS 50. Scale bar is 250 µm. B) Elemental distribution of 12C14n- clusters. C) 109Ag-
ions in 300 nm cross sections of G. fossarum gills. Animals were exposed to 10 µg. L-1 of CIT-AgNPs 40 nm and
PEG-AgNPs 40 nm in Volvic water. Scale bar is 2 µm.
Cit-AgNPs
40 nm
PEG-AgNPs
40 nm
107Ag
CN- 107Ag
Control
CN- 107Ag
CN-
A B C
Chapter 4 AgNPs and AuNPs acute toxicity
138
Au could only be found close to intestinal caeca of G. fossarum exposed to CIT-AuNPs 40 nm (Figure 4. 4 C)
more specifically between the external wall of two intestinal caeca. No Au signal could be found in G.
fossarum exposed to PEG-AuNPs or in control animals (Figure 4. 4).
Figure 4. 4. A) 300 nm cross section of G. fossarum observed under optical microscope. Red square represents
the area observed with NanoSIMS 50. Scale bar is 250 µm. B) Elemental distribution of 12C14N- clusters. C) 109Au-
ions in 300 nm cross sections of G. fossarum intestinal caeca. Animals were exposed for 72h to 10 µg. L-1 of CIT-
AuNPs 40 nm and PEG-AuNPs 40 nm in Volvic water. Scale bar is 2 µm
3.2.4.2 CytoViva® dark field hyperspectral imaging
In order to determine whether Ag and Au signals observed with NanoSIMS are internalized as nanoparticles,
Cytoviva dark field hyperspectral imagining microscope was used, since the hyperspectral imaging only allows
the detection of NPs and not of ions. Analysis were performed in different organs and tissues (gills, gut,
muscle and cuticle). Only results for gills and gut are presented in order to compare them with NanoSIMS
results (Figure 4. 5, Figure 4. 6).
The analysis of Ag in G. fossarum gills revealed the presence of a AgNPs in the gills of animals exposed to CIT-
AgNPs and PEG-AgNPs, highlighted with red spots (Figure 4. 5 C). As already observed with the NanoSIMS
analysis, the amount of AgNPs detected in animals exposed to CIT-coated AgNPs is higher than the amount
detected in animals exposed to PEG-coated AgNPs.
Control
Cit-AuNPs
40 nm
PEG-AuNPs
40 nm
CN-
197Au
197Au
CN-
A B C
Chapter 4 AgNPs and AuNPs acute toxicity
139
Figure 4. 5. A) 300 nm cross section of G. fossarum observed under optical microscope. Red square represents
the area observed with Cytoviva darkfield hyperspectral microscope. Scale bar is 250 µm. B) G. fossarum gills
viewed with Cytoviva (60x oil immersion magnification). C) AgNPs accumulation in G. fossarum gills (red spots
and white arrows). Animals were exposed for 72h to 10 µg. L-1 of CIT-AgNPs 40 nm and PEG-AgNPs 40 nm in Volvic
water. Scale bar is 6 µm
The analysis of the intestinal caeca of G. fossarum exposed to CIT-AuNPs 40 nm allowed the detection of
AuNPs within the epithelial tissue (Figure 4. 6 C). Identical to what was observed with NanoSIMS analysis,
PEG-AuNPs were not detectable in exposed animals.
Cit-AgNPs
40 nm
PEG-AgNPs
40 nm
Control
A B C
Chapter 4 AgNPs and AuNPs acute toxicity
140
Figure 4. 6. A) 300 nm cross section of G. fossarum observed under optical microscope. Red square represents
the area observed with Cytoviva darkfield hyperspectral microscope. Scale bar is 250 µm. B) G. fossarum intestinal
caeca viewed with Cytoviva (60x oil immersion magnification). C) AuNPs accumulation in G. fossarum intestinal
caeca (green spots and white arrows). Animals were exposed for 72h to 10 µg. L-1 of CIT-AgNPs 40nm and PEG-
AgNPs 40 nm in Volvic water. Scale bar is 6 µm
3.2.5 Molecular responses
In order to better understand the effects of sub-lethal concentrations AgNPs and AuNPs in G. fossarum and
elucidation of the acute toxic mechanisms, we assessed the expression of genes involved in oxidative stress,
general stress, apoptosis, DNA damage and repair, cytoskeleton remodelling and trafficking, osmoregulation
and respiration. The analysis of gene expression evidenced specific patterns in relation to the metal of the
particle (Table 4. 4 and Table 4. 5). Indeed, the effects observed on gene expression were dependent on the
chemical composition of the studied NPs (Ag or Au). AuNPs 80 nm led to a down-regulation of both MnSOD
and CuZnSOD in G. fossarum whereas AgNPs 80 nm induced both of them. MnSOD was induced in G.
fossarum exposed to 25 and 50 µg. L-1 of PEG-AgNPs 80 nm (1.5-fold increase for both concentrations)
whereas no significant effect of PEG-AuNPs 80 nm was observed. In contrast, MnSOD and CuZnSOD were
down-regulated (2.5- and 4-fold decrease) in G. fossarum exposed to CIT-AuNPs 80 nm (25 and 50 µg. L-1,
respectively), while no significant effect of CIT-AgNPs 80nm was observed. Chitinase was induced 2- and 1.7-
fold by exposure to 10 and 25 µg. L-1 of CIT-AgNPs 40 nm respectively and down regulated by 1.75-fold at 50
µg. L-1 of CIT-AgNPs 40 nm. A 1.4-, 2.3- and 2.8-fold down regulation of GST of was also observed in G.
A B C
Control
Cit-AuNPs
40 nm
PEG-AuNPs
40 nm
Chapter 4 AgNPs and AuNPs acute toxicity
141
fossarum exposed to 10, 25, 50 µg. L-1 of PEG-AgNPs 80 nm, respectively. Gpx7 was impacted only in G.
fossarum exposed to the highest concentrations of PEG-AuNPs 20 nm and CIT-AuNPs 40 nm with a 2-fold
increase. However, the observed responses were not coating or size-dependent.
Chapter 4 AgNPs and AuNPs acute toxicity
142
Table 4. 4. Relative gene expression of G. fossarum exposed for 72h to CIT-AgNPs and PEG-AgNPs 20, 40 and 80 nm
indicates significant (P < 0.05) up-regulated genes (fold change > 1.5) indicates significant (P < 0.05) down-regulated genes (fold change < 0.5)
indicates non-significant effect on relative gene expression.
exoskeleton/cuticle Lysososmes
Actin TUB UB Chitinase CAT MnSOD CuZnSOD GPx7 GST HSP90 GAPDH Gadd45 NfkB Cjun P53 CathepsinL NaKATPase Hemocyanin
0.5 + + - -
1 +
2
4 + + + + +
8
1 +
2
4 + +
8 +
10 -
1 +
2 +
4
8 +
10
1 +
5
10 + + + -
25 +
50 -
1 + + -
5 -
10 +
25 + + -
50 -
1 +
5 +
10 - + - -
25
50
1 + +
5
10 + + - -
25 + + - - -
50 + - -
AgNPs40nm
citrate
PEG
AgNPs80nm
citrate
PEG
Compound Concentrations(µg.L-1)
AgNO3
AgNPs20nm
Citrate
PEG
Generalstress DNAdamageandrepair OsmoregulationandRespiration
Cytoskeleton Antioxidantdefence
+ - n.s
Chapter 4 AgNPs and AuNPs acute toxicity
143
Table 4. 5. Relative gene expression of G. fossarum exposed for 72h to CIT-AuNPs and PEG-AuNPs 20, 40 and 80 nm
indicates significant (P < 0.05) up-regulated genes (fold change > 1.5) indicates significant (P < 0.05) down-regulated genes (fold change < 0.5)
indicates non-significant effect on relative gene expression
exoskeleton/cuticle Lysososmes
Actin TUB UB Chitinase CAT MnSOD CuZnSOD GPx7 GST HSP90 GAPDH Gadd45 NfkB Cjun P53 CathepsinL NaKATPase Hemocyanin
0.5 - - + -
1 +
2 + +
1 - +
5 +
10
25 - - +
50
1
5
10 -
25 - - +
50 - +
1
5 -
10 - -
25
50 - - + +
1
5 - -
10 -
25 - -
50 - -
1
5 -
10 - -
25 + - -
50 + - - +
1 - -
5 -
10 - -
25 -
50 - -
Compound
PEG
AuNPs80nm
citrate
PEG
AuNPs40nm
citrate
PEG
AuNPs20nm
Citrate
AgNO 3
OsmoregulationandRespirationDNAdamageandrepairGeneralstressAntioxidantdefenceCytoskeletonConcentrations(ug.L
-1)
+ - n.s
Chapter 4 AgNPs and AuNPs acute toxicity
144
4 DISCUSSION
The fate and behaviour of nanoparticles including AgNPs and AuNPs in the aquatic environment will
undoubtedly influence their toxicity towards aquatic organisms (Lapresta-Fernández et al., 2012;
McGillicuddy et al., 2017; Vale et al., 2016). In the present study, we investigated the influence of size and
surface coating on the uptake and acute effects of AgNPs and AuNPs on G. fossarum following a realistic
exposure scenario in terms of concentrations, which are expected to be in the low µg.L-1, respectively
(Gottschalk et al., 2013; Yang et al., 2016). Three different sizes (20, 40, 80 nm) and two different coatings
(CIT and PEG) were used both for AgNPs and AuNPs. The studied endpoints included uptake, tissue
distribution and short-term effects of the selected AgNPs and AuNPs on expression of stress-related genes of
G. fossarum.
The studied AgNPs and AuNPs appeared to be stable in the exposure medium (commercial mineral water,
Volvic®, France), with limited tendency to agglomeration and sedimentation in the experimental conditions.
Only CIT-AgNPs 20 nm and CIT-AuNPs 20 nm showed the tendency to aggregation with measured values for
hydrodynamic diameters of 53 and 154 nm and zeta potential of -19.1 and -17.1, respectively. It is usually
described that NPs with a zeta potential within -30mV and +30mV range have a tendency to agglomerate in
the tested experimental conditions (Lee and Ranville, 2012). However, in the present study, it seems that
zeta potential is not a good indicator of stability for CIT-AgNPs and CIT-AuNPs as for sterically repulsed NPs
(Sørensen et al., 2016). These results are consistent with previous studies that reported instability and an
increase in size of CIT-AgNPs (Baalousha et al., 2013; Park et al., 2015a; Sakka et al., 2016) and CIT-AuNPs
(Park et al., 2015b) compared to PEG-NPs (Park et al., 2015b; Tejamaya et al., 2012).
AgNPs used in commercial products are generally coated with organic compounds, such as PEG, in order to
increase their colloidal stability by reducing the surface energy (Ju-Nam and Lead, 2008). As a result, capping
agents prevent interaction of the NPs with the surrounding environment and between them, thus avoiding
or reducing their aggregation (Kvítek et al., 2008), which may explain the increased stability of PEG-AgNPs
and PEG-AuNPs in our exposure medium (Volvic water).
Differences in stability can affect behaviour, bioavailability and toxicity of AgNPs and AuNPs (Kwok et al.,
2012; Levard et al., 2012; Park et al., 2015a). Thus, Ag and Au total and ionic concentrations in exposure
medium were evaluated in order to determine the behaviour and fate of the AgNPs and AuNPs. Total Au
concentration could not be determined and further analyses are therefore needed. Ag concentrations in the
exposure medium were close to the nominal ones in the low µg. L-1 range while a plateau around 20 µg. L-1
was observable at the higher exposure concentrations (25 to 50 µg. L-1). This loss in Ag at the highest
concentrations may be linked to a rapid sedimentation (Bacchetta et al., 2017; Farmen et al., 2012; Scown et
Chapter 4 AgNPs and AuNPs acute toxicity
145
al., 2010; Zimmermann et al., 2017) or to the uptake by the animals. Furthermore, dissolution assessment
showed that only 6% of Ag ions are released at the highest AgNPs concentrations (50 µg. L-1) indicating that
the tested particles are stable in our experimental conditions as shown by our NTA results for CIT-AgNPs and
CIT-AuNPs 40 and 80 nm and PEG-AgNPs and PEG-AuNPs. This result is in accordance with previous studies
that showed low dissolution rate for CIT-AgNPs, PEG-AgNPs and PVP-AgNPs (Park et al., 2015a, 2015b).
Hence, some studies suggested that presence of organisms or organic matter decreased, through adsorption,
the amount of ionic Ag released from AgNPs (Bone et al., 2012; Cupi et al., 2015; Newton et al., 2013).
Exposure of G. fossarum to AgNPs and AuNPs for 72h led to a significant coating-dependent bioaccumulation
being of CIT-AgNPs and CIT-AuNPs, which were taken up in higher amounts than PEG-AgNPs and PEG-AuNPs,
which is in accordance with previous studies (Oliver et al., 2014; Park et al., 2015b; Sakka et al., 2016). A
coating dependent bioaccumulation of CIT-AgNPs and CIT-AuNPs was reported in Lymnea stagnalis (Oliver et
al., 2014), D. magna (Sakka et al., 2016) and Gammarus pulex (Park et al., 2015b). The observed difference in
the bioaccumulation rate between AgNPs and AuNPs could be due to the sedimentation of CIT-AgNPs and
CIT-AuNPs linked to the reduced colloidal stability of the particles (Croteau et al., 2011; Sakka et al., 2016) as
compared to PEG-coated NPs, which are more stable and thus have a lower tendency for agglomeration and
sedimentation. Since G. fossarum is a benthic invertebrate living at the water/sediment interface, this could
enhance the bioavailability of the CIT-NPs, which will be in the proximity of the animals (Kunz et al., 2010;
Park et al., 2015b). It has been suggested that PEG prevent the interaction of the NPs with the surrounding
environment leading, for instance, to the formation of a reduced and more labile protein corona than for CIT-
coated NPs, which is supposed to be important for the uptake of NPs by aquatic organisms (Lynch and
Dawson, 2008; Nasser and Lynch, 2016; Sasidharan et al., 2015).
In addition to influences of the surface coating, we also investigated the influence of size on the uptake of
AgNPs and AuNPs. While a decrease of uptake with increasing size of AgNPs and AuNPs could be expected
(Georgantzopoulou, 2015; Georgantzopoulou et al., 2013; Mehennaoui et al., 2016), no clear relationship
between size and bioaccumulation in the organisms was observed in the present work. Previous studies
reported that size did not influence the uptake of AgNPs by G. roeseli (Andreï et al., 2016). Same observations
were reported for the uptake of AuNPs by D. magna (Khan et al., 2014; Wray and Klaine, 2015) and G. pulex
(Park et al., 2015b). In these previous studies, rather than the size, a concentration-dependent uptake was
highlighted and it had been described that the particle number seem to be an important factor that influence
the uptake (Park et al., 2015b) and an increase of the particles number led to an increase of bioaccumulation.
When comparing Ag bioaccumulation after treatment with AgNPs or AgNO3, at similar exposure
concentrations (~5 µg. L-1), AgNO3 led to the higher uptake of Ag by G. fossarum. This result is in accordance
with previous studies in D. magna (Croteau et al., 2011; Zhao and Wang, 2010), snails (Chen et al., 2017; Khan
Chapter 4 AgNPs and AuNPs acute toxicity
146
et al., 2015) and rainbow trout (Bruneau et al., 2016). This was explained by the high affinity of Ag+ for Na+
channels in membrane transporters (Bianchini and Wood, 2003). Indeed, in our previous study, a decrease in
haemolymph osmolality of G. fossarum exposed to AgNO3 was observed and this was linked to the decrease
of Na+ concentration in haemolymph (Arce Funck et al., 2013; Mehennaoui et al., 2016; Ribeiro et al., 2017).
AgNO3 also showed to be more toxic to G. fossarum leading to a significant mortality rate (Andreï et al., 2016;
Arce Funck et al., 2013; Mehennaoui et al., 2016). The recorded LC50s during the present study showed a
significant seasonal variability in G. fossarum sensitivity as the LC50 recorded in autumn was lower than the
one recoded in summer. It is well established that the season greatly influence the physiological status and
metabolism of gammarids (Lebrun et al., 2015), which are normally field collected, leading to subsequent
changes in their responses to chemicals (Dedourge-Geffard et al., 2013; Gismondi et al., 2012; Lebrun et al.,
2015; Sroda and Cossu-Leguille, 2011). Despite the significant bioaccumulation of AgNPs and AuNPs used in
the present study, none of them were lethal to G. fossarum, nor did they show any sign of acute toxicity.
These results are also consistent with literature that reported an absence of mortality after exposing G.
roeseli to 0.5 and 5 µg.L-1 CIT-AgNPs 10 and 60 nm (Andreï et al., 2016) and G. pulex exposed to up to 1 mg.L-
1 of CIT-AuNPs and PEG-AuNPs (Park et al., 2015b).
To better understand the mechanisms underlying AgNPs and AuNPs bioaccumulation, it is important to know
how NPs behave upon internalization by the organism. NanoSIMS and Cytoviva are good and powerful
techniques to investigate the distribution of NPs within biological matrices. NanoSIMS allows the elemental
detection, with high sensitivity, of virtually the whole periodic table, but it cannot be used to distinguish ions
from intact particles. On the contrary, Cytoviva allows the detection of nanoparticles, through the
hyperspectral analysis, but it does not allow the detection of ions. The use of the two techniques allowed the
detection of AgNPs in gills of G. fossarum and AuNPs close to the intestine of G. fossarum. The obtained
signals, with both techniques (NanoSIMS and Cytoviva) were higher for CIT-AgNPs and CIT-AuNPs which is
consistent with ICP-MS results. The higher uptake of CIT-AgNPs and CIT-AuNPs could be explained by the
attachments on the cuticle of G. fossarum as reported in other studies that used D. Magna as model species
(Asghari et al., 2012; Baumann et al., 2014; Sakka et al., 2016; Zhao and Wang, 2012). The analysis performed
with the NanoSIMS and Cytoviva supports this hypothesis, as presence of CIT-AgNPs and CIT-AuNPs was
observed on the cuticle of G. fossarum with a higher signal obtained for AgNPs than AuNPs. Additionally, CIT-
AgNPs were observed in the gills, more specifically the cuticle and the epithelial tissue of G. fossarum gills.
This result is in accordance with many other studies that reported the presence of AgNPs in the gills of
mussels (Canesi et al., 2012), zebrafish and perch (Bilberg et al., 2010; Griffitt et al., 2009; Scown et al., 2010)
and around the nucleus of gill cells of rainbow trout (Farkas et al., 2011). Our results appear to be consistent
with previous reports since fish and crustaceans’ gills are in direct contact with the external environment and
Chapter 4 AgNPs and AuNPs acute toxicity
147
are described to play a protection role as a barrier. Crustaceans’ gills are also described as the major sites of
accumulation of heavy metals and toxic compounds as they can be trapped in the mucus layer or absorbed
by the epithelial cells (Ali, 2014; Georgantzopoulou et al., 2016; Henry et al., 2012). It is suggested that the
main uptake mechanism for AgNPs is the direct uptake from the water through the gills (Ribeiro et al., 2017).
Other studies suggested endocytosis process (pinocytosis and phagocytosis) to explain AgNPs uptake by D.
magna (Ribeiro et al., 2017).
AuNPs were observed in intestinal caeca of G. fossarum. Previous studies reported the presence of AuNPs in
the gut, carapace, antennae and appendages of D. magna (Botha et al., 2016; Jensen et al., 2017; Lee and
Ranville, 2012). AuNPs have been also detected close to the peritrophic membrane and in the microvilli of D.
magna gut, where NPs inhibited the intestinal clearance and thus the excretion of NPs. Ingestion is very likely
the main uptake route of AuNPs in freshwater crustaceans (Wray and Klaine, 2015). Thus, internalization of
nanoparticles by G. fossarum seems to be metal-dependent and can take multiple routes (Lee and Ranville,
2012; Lovern et al., 2008): AgNPs seem to be taken up through the gills while AuNPs may be ingested through
the mouth. It had been suggested that fish and crustaceans’ gills are the primary organ for AgNPs uptake
although Ag was present in the gut, liver, brain depending on the surface coating (Kwok et al., 2012) indicating
that Ag might be distributed to different organs after being taken up through the gills. The intestinal caeca
participated in the production of digestive fluids (Cristescu and Hebert, 2005) and were reported to be the
sites of Cd and Ca2+ uptake (Georgantzopoulou et al., 2016) and seem to be the main site for bioaccumulation
of AuNPs in G. fossarum.
There are many uncertainties about mechanisms of toxicity of AgNPs and AuNPs on aquatic organisms. A
transcriptomic approach, through targeted gene expression method, was used in order to investigate AgNPs
and AuNPs effects at sub-lethal concentrations. A set of genes was selected based on our previous work on
G. fossarum (Mehennaoui et al., 2016) in which a battery of biomarkers was used to asses AgNPs effects. The
set of genes included genes associated with oxidative stress (CAT, MnSOD, CuZnSOD, GPx7), general stress
and detoxification mechanisms (GST and HSP90), cytoskeleton remodelling and trafficking (Actin, TUB, UB),
genotoxicity (c-jun, Gadd45, NfkB, P53), lysosomes (Cathepsin L), moulting (Chitinase), osmoregulation and
respiration (Na+K+ATPase and HEM) were analysed. AgNPs led to an increase of both SODs expression while
AuNPs led to their down regulation. SODs are known to be antioxidant enzymes converting the superoxide
anion to H2O2 (Sroda and Cossu-Leguille, 2011; Vasseur and Cossu-Leguille, 2004) and oxidative stress is
described as one of the most frequent effects reported in aquatic organisms exposed to NPs (Vale et al.,
2016). The down-regulation of MnSOD in G. fossarum exposed to AuNPs was in accordance with a previous
study that reported a decrease of MnSOD expression in gills of C. fulminate (Renault et al., 2008) exposed to
contaminated algae with up to 1.6 x 105 amine-coated AuNPs/cell and in G. fossarum exposed through
Chapter 4 AgNPs and AuNPs acute toxicity
148
contaminated diet to up to 46 mg. L-1 of AuNPs 10 nm (Baudrimont et al., 2017). However, our results are in
contrast with other studies that reported the absence of significant effects of AgNPs on SOD expression in
the gills of M. galloprovincialls exposed for 15 days to 10 µg.L-1 of AgNPs (Bebianno et al., 2015) and in liver
of zebrafish exposed to up to 120 mg.L-1 of AgNPs (Choi et al., 2010). It has been described that gene
expression vary dependently of the studied tissue/organ (Guerriero et al., 2014). Therefore, these results
may be explained by the fact that authors looked to the effects of AgNPs and AuNPs in one specific organ
while in the present work gene expression were investigated in the whole animal. In the present study, the
increase in expression of CuZnSOD could indicate an induction of antioxidant defences. Ag is known to
interact with thiol group found in many antioxidant enzymes and it is also known to generate production of
intra and extracellular reactive oxygen species (ROS) (Bebianno et al., 2015; Connolly et al., 2015;
McGillicuddy et al., 2017). Previous studies suggested that generation of ROS was due to the presence of
AgNPs themselves in addition to Ag ions that could be dissolved within the organism (Ahn et al., 2014;
Bebianno et al., 2015). These compounds were described to lead to an overwhelming of antioxidant defence
and metallothionein, which plays an important detoxification role (Gomes et al., 2013; Gomes et al., 2014b).
Thus, it might be suggested that internalization of AgNPs or their adsorption on the cuticles of the gills in G.
fossarum induced to the production of ROS leading to the activation of antioxidant defences. Moreover, the
down-regulation of Cathepsin L and GST by AgNPs 80 nm could be linked to the interaction of AgNPs with
thiol groups of proteins and a potential disruption of glutathione metabolism leading to oxidative stress
(Gismondi et al., 2012; Vellinger et al., 2013). Additionally, the significant effect on Chitinase expression in G.
fossarum exposed to CIT-AgNPs 40 nm may indicate that CIT-AgNPs 40 nm are interacting/interfering with
the structure and the maintenance of exoskeleton. The presence of CIT-AgNPs 40 nm on the cuticle of the
gills observed with NanoSIMS 50 and Cytoviva could therefore explain this result. This effect may disturb ion
transport, gaseous exchange, growth process of G. fossarum that occur through successive moults (Sutcliffe,
1992) and lead to deleterious effects at the population level.
5 CONCLUSION
In the present study, we observed a coating-dependent effect that enhance the uptake of CIT-coated particles
as compared to PEG-coated particles. This uptake led to a metal-dependent distribution of AgNPs and AuNPs
within G. fossarum tissues. AgNPs were mainly found in the gills, while AuNPs were mainly observed in
intestinal caeca. AgNPs were found to up-regulate MnSOD and CuZnSOD gene expression while AuNPs led to
their down-regulation. The observed effects seemed to be directly linked to the NPs themselves as no ions
were detected with ICP-MS and Cytoviva allowed to confirm a significant presence of NPs within the tissues.
Modulation by AgNPs of SODs may indicate a generation of ROS and possible activation of SOD enzymes in
Chapter 4 AgNPs and AuNPs acute toxicity
149
order to defend and protect the organism from a potential oxidative stress. However, to affirm this
hypothesis, further investigations are needed in order to better define mechanisms of action at the molecular
level of AgNPs and AuNPs on G. fossarum.
Acknowledgments
This work was supported by National Research Fund, Luxembourg (AFR-PhD-9229040). The authors are
thankful to S. Contal and M. Fossépré for technical support, C. Guignard for chemical analyses, E. Lentzen and
A. Chauvière for NanoSIMS analyses.
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COMPLEMENT 3
Submitted to Science of the Total Environment
Influence of size and surface coating on uptake of silver and gold nanoparticles and the molecular effects
on Gammarus fossarum (Crustacea Amphipoda)
Kahina Mehennaoui a,b, Sébastien Cambier a, Tommaso Serchi a, François Guérold b, Johanna Ziebel a, Jean-
Sebastien Thomann c, Nathalie Valle c, Laure Giambérini b, *, Arno C. Gutleb a*
A Environmental Research and Innovation (ERIN) Department, Luxembourg Institute of Science and
Technology, 41 rue du Brill, Belvaux, Luxembourg
B Laboratoire Interdisciplinaire des Environnements Continentaux (LIEC), CNRS UMR 7360, Université de
Lorraine – Metz, France
C Material Research and Technology (MRT) Department, Luxembourg Institute of Science and Technology, 41
rue du Brill, Belvaux, Luxembourg
Chapter 4 AgNPs and AuNPs acute toxicity
158
Supplementary material
1- Coating of AgNPs and AuNPs by wet chemistry
Commercial citrate stabilized AgNPs and AuNPs (purchased from Nanocomposix) were coated with
polyethylene glycol (PEG layer MW 3000) via wet chemistry method that allowed the control of the correct
functionalization of isolated NPs. This method relied on the established thiol/metal chemistry.
The amount of added PEG was adjusted to ensure the surface saturation of the NPs (Figure S1 A). The
obtained PEG-AgNPs and AuNPs were characterized for size distribution using Nanoparticles Tracking analysis
(NTA, NanoSIGHT 500, Malvern, UK, Figure S1 B). PEG-AgNPs and PEG-AuNPs were stored at 4 °C until use.
Figure S 2 PEG coating on metal Nanoparticles (here silver NPs). SEM shows clearly the conformal silver coating
with the PEG layer. NTA analysis shows a small size increasing for the 40 and 80 nm silver nanoparticles as a
consequence of their coating with the PEG layer. The sensitivity of NTA was not enough to resolve the size increase
for 20 nm nanoparticles.
Chapter 4 AgNPs and AuNPs acute toxicity
159
2- Particle character ization
Figure S 3. Size distribution of CIT-AgNPs 20 nm (A), CIT-AgNPs 40 nm (B) and CIT-AgNPs 80 nm (C) in Volvic
Water (T0h) and CIT-AgNPs 20 nm (D), CIT-AgNPs 40 nm (E) and CIT-AgNPs 80 nm (F) in Volvic water after 24h
of incubation. Size distribution is expressed as particle concentration E6.mL-1. The red errors bars indicate ±SD of
the mean of triplicate measurements
A
B
C
D
E
F
Chapter 4 AgNPs and AuNPs acute toxicity
160
Figure S 4. Size distribution of PEG-AgNPs 20 nm (A), PEG-AgNPs 40 nm (B) and PEG-AgNPs 80 nm (C) in Volvic
Water (T0h) and PEG-AgNPs 20 nm (D), PEG-AgNPs 40 nm(E) and PEG-AgNPs 80 nm (F) in Volvic water after 24h
of incubation. Size distribution is expressed as particle concentration E6.mL-1. The red errors bars indicate ±SD of
the mean of triplicate measurements
A
B
C
D
E
F
Chapter 4 AgNPs and AuNPs acute toxicity
161
Figure S 5. Size distribution of CIT-AuNPs 20 nm (A), CIT-AuNPs 40 nm (B) and CIT-AuNPs 80 nm (C) in Volvic
Water (T0h) and CIT-AuNPs 20 nm (D), CIT-AuNPs 40 nm (E) and CIT-AuNPs 80 nm (F) in Volvic water after 24h
of incubation. Size distribution is expressed as particle concentration E6.mL-1. The red errors bars indicate ±SD of
the mean of triplicate measurements
A
B
C
D
E
F
Chapter 4 AgNPs and AuNPs acute toxicity
162
Figure S 6. Size distribution of PEG-AuNPs 20 nm (A), PEG-AuNPs 40 nm (B) and PEG-AuNPs 80 nm (C) in Volvic
Water (T0h) and PEG-AuNPs 20 nm (D), PEG-AuNPs 40 nm(E) and PEG-AuNPs 80 nm (F) in Volvic water after 24h
of incubation. Size distribution is expressed as particle concentration E6.mL-1. The red errors bars indicate ±SD of
the mean of triplicate measurements
A
B
C
D
E
F
Chapter 4 AgNPs and AuNPs acute toxicity
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3- Preliminary test: Daphnia magna Straus mobility inhibition test ( ISO
6341:1996)
In order to determine the concentrations range to be tested with G. fossarum, an acute Daphnia magna test
was performed. Daphnia magna acute test is the most widely used bioassay in ecotoxicology. It aims at
determining the effective concentration of a pollutant that inhibits the mobility of 50% of Daphnia magna
population (EC50).
3.1- Test Organisms
Experiments were conducted with a strain of Daphnia magna Straus 1820, which was maintained in the
laboratory for four years. Parthenogenetic cultures were carried out in 1L aquaria at 20 °C with LCV medium:
a mixture of (20/80) of Lefevre-Czarda (LC) medium and mineral water (Volvic, France). This mixture is
supplemented with:
• Ca and Mg in order to obtain a total hardness of 250 mg.L-1 and a Ca/Mg ration of 4/1
• A mixture of vitamins containing thiamine HCL, vitamin B12 and biotin.
Cultures were maintained under a 16h light 8h dark and at density of 40 animals per liter. The medium was
renewed three times a week and Daphnids were fed with a mixture of two algal species (Pseudokirchneriella
subcapitata and Scenedesmus subspicatus).
3.2- Biological assay
All experiments were performed according to the guideline ISO 6341 – 1996.
3.2.1- Reference test
Daphnia magna were exposed during 24-48h to an increasing range of concentration of potassium
dichromate (K2Cr2O7) (table 1). The aim is to provide data about the sensitivity and the validity of the acute
test.
Four replicates of five neonates (aged < 24h) were placed in glass test tubes containing 10mL of each test
concentration and control. The neonates were not fed during the test. The assessment endpoint examined
was immobilization. More precisely, juveniles that were able to swim were considered mobile and those
which still moved their antennae but did not swim within 15s after a gentle shaking were considered
immobile.
Table S 2 D. magna reference test
Concentration (mg/L) 0 0.58 0.67 0.78 0.9 1.04 1.2 1.39
K2Cr2O7 (µL) 0 290 335 390 450 520 600 695 ISO medium (mL) 10 9.71 9.66 9.61 9.55 9.48 9.5 9.3 Daphnia (n) 5 5 5 5 5 5 5 5
Chapter 4 AgNPs and AuNPs acute toxicity
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3.2.2- Final test
Acute toxicity of AgNPs and AuNPs was determined during 24 and 48h of exposure (serially diluted in ISO
medium, AgNPs 20nm: 1 – 10 µg. L-1, AgNPs 40 and 80nm: 1 – 50 µg.L-1, AuNPs: 10 – 100 µg.L-1). All tests were
conducted at 20 °C in full darkness. Four replicates of five neonates (aged < 24h) were placed in glass test
tubes containing 10mL of each test concentration and control. The neonates were not fed during the test.
The assessment endpoint examined was immobilization.
3.2.3- Results
The inhibition of mobility rate was determined relatively to the total number of Daphnia used. EC50 of each
NPs and their confidence intervals were calculated using Excel Macro REGtox®.
In all tests conditions, immobilisation in controls was lower than 10% and the EC50-48h after exposure to
K2Cr2O7 was 0.9 mg. L-1 which was in accordance with the validity criteria.
The EC50-48h values obtained, after exposure to AgNPs and AuNPs, are expressed in table 6. A size depend
effect was observed in Daphnids exposed to AgNPs (Table 6). AgNPs 20nm was the most toxic one followed
by AgNPs 40 nm. No full dose-responses curves could be obtained for AgNPs 80 nm and AuNPs. These results
were in accordance with previous studies performed in the Lab with the same particles and the same strain
of Daphnia magna.
Table S 3. EC50 values obtained after 48h exposure of D. magna to AgNPs and AuNPs
NPs AgNPs AuNPs
Size EC50 (µg. L-1) EC50 (µg. L-1)
20 nm 6.7 (5.1 – 8.5) > 100
40 nm 30.8 (25.9 – 36.07) > 100 80 nm > 50 > 100
Chapter 4 AgNPs and AuNPs acute toxicity
165
4- Acute toxicity of AgNO3 on Gammarus fossarum
Figure S 7. Effects AgNO3 on survival of Gammarus fossarum collected in A) June, B) September and C) November
after 72h of exposure.
0
20
40
60
80
100
120
0 2 4 6 8 10 12
Survivalrates(%)
Concentration (µg.L-1)
Effetcalculé Effetmoyen Réplicats
CE5 CE10 CE15
CE20 CE25
0
20
40
60
80
100
120
0 1 2 3 4 5 6
Survivalrates(%)
Concentration (µg.L-1)
Effetcalculé Effetmoyen Réplicats
CE5 CE10 CE15
CE20 CE25
0
20
40
60
80
100
120
0 2 4 6 8 10 12
Survivalrates(%)
Concentration (µg.L-1)
Effetcalculé Effetmoyen Réplicats
CE5 CE10 CE15
A B
C
166
Chapter 5 AgNPs and AuNPs chronic toxicity
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Chapter 5
Chronic toxicity of AgNPs and AuNPs
CHAPTER 5
AgNPs and AuNPs chronic toxicity
Chapter 5 AgNPs and AuNPs chronic toxicity
168
KEY FINDINGS
Aim of the study
The aim of the study was to investigate the sub-chronic toxicity of CIT and PEG-AgNPs 40 nm and CIT and
PEG-AuNPs 40 nm on G. fossarum. A multibiomarker approach including molecular effects (stress-related
gene expression), physiological (osmoregulation) and behavioural (locomotion and ventilation) responses
was used. The biological endpoints also included mortality, bioaccumulation and localization of AgNPs and
AuNPs in G. fossarum tissues.
Experimental design
Key findings
• Concentration-dependent uptake
• Coating-dependent uptake o CIT-AgNPs ≥ PEG-AgNPs
o CIT-AuNPs ≥ PEG-AuNPs
• Influence of chemical composition on the internal distribution o CIT-AgNPs40 nm were found only in G. fossarum gills
o CIT-AuNPs 40 nm were found only in the lumen of intestinal caeca
• A metal-dependent effect on gene expression was also observed with o AgNPs leading to an increase in Na+K+ATPase, CAT and Chitinase gene expression
o AuNPs leading to an increase in CuZnSOD gene expression
o Decrease in locomotion
o No effects on osmoregulation, ventilation and survival
Gammarus fossarum
Organisms Materials
Ag40nm
Au40nm
CIT and PEG-AgNPs 40 nm
CIT and PEG-AuNPs 40 nm
Luxembourg
Schwaarzbaach
Sampling site
Acclimation
D-10
Exposure
D-7 D-1
Temperature : 12 oC
Photoperiod : 16h/8h
Daily renewal of Volvic water
Stalling
D15D-30 D-20
Leave
conditioning
Temperature : 12 oC
Photoperiod : 16h/8h
Exposure medium renewal: every 3 days Addition of food: every 2 days
D0
Chapter 5 AgNPs and AuNPs chronic toxicity
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Sub-chronic effects of AgNPs and AuNPs on Gammarus fossarum (Crustacea Amphipoda): from
molecular to behavioural responses
Kahina Mehennaoui a,b, Sébastien Cambier a, Tommaso Serchi a, François Guérold b, Arno C. Gutleb a*, Laure
Giambérini b, *
a Environmental Research and Innovation (ERIN) Department, Luxembourg Institute of Science and
Technology, 41 rue du Brill, Belvaux, Luxembourg
b Laboratoire Interdisciplinaire des Environnements Continentaux (LIEC), CNRS UMR 7360, Université de
Lorraine – Metz, France
* Corresponding authors:
Arno C. Gutleb
Postal Address: Environmental Research and Innovation Department
Luxembourg Institute of Science and Technology
5, avenue des Hauts-Fourneaux, L-4362 Esch-sur-Alzette, Luxembourg
E-mail: [email protected]
Laure Giamberini
Postal Address: Laboratoire Interdisciplinaire des Environnements Continentaux, Universite de Lorraine,
Campus Bridoux 57000 Metz, France.
Email : [email protected]
Chapter 5 AgNPs and AuNPs chronic toxicity
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ABSTRACT
The aim of the present study was the assessment of the sub-chronic effects of silver nanoparticles (AgNPs)
and gold nanoparticles (AuNPs) of 40 nm primary size either stabilized with citrate (CIT-AgNPs and CIT-AuNPs)
or coated with polyethylene glycol (PEG-AgNPs and PEG-AuNPs). AgNO3 was used as a positive control in
order to study the contribution of ions potentially released from the AgNPs on the observed effects. A
multibiomarker approach was used to study the long-term effects of AgNPs and AuNPs 40 nm on molecular,
physiological and behavioural responses of the freshwater invertebrate Gammarus fossarum.
G. fossarum were exposed for 15 days to 0.5 and 5 µg. L-1 of CIT and PEG AgNPs and AuNPs 40 nm in the
presence of food. A significant uptake of Ag and Au was observed in exposed animals, with CIT-AgNPs
detected in gills and CIT-AuNPs 40 nm within the lumen of intestinal caeca of G. fossarum using hyperspectral
darkfield microscopy (Cytoviva). PEG-AgNPs 40 nm led to an up-regulation of CAT and Chitinase and CIT-
AgNPs and PEG-AgNPs 40 nm led to an up-regulation of Na+K+ATPase. CIT and PEG-AuNPs 40 nm led to an
increase of CuZnSOD gene expression. Both AgNPs and AuNPs 40 nm significantly impacted locomotor activity
of G. fossarum while no effects were observed on haemolymphatic ions and ventilation.
Keywords: Gammarus fossarum, silver nanoparticles; gold nanoparticles; gene expression; osmoregulation;
behaviour; Cytoviva; long term exposure.
Chapter 5 AgNPs and AuNPs chronic toxicity
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1. INTRODUCTION
The recent advances in nanotechnologies led to an increasing use of engineered nanoparticles (ENPs) in
various consumers’ products. Silver nanoparticles (AgNPs) are among the most promising and most widely
used ENPs due to their broad spectrum antimicrobial activities (Vance et al., 2015). They are currently found
in many products of daily life such as textiles, plastics, health care products, water filters or food packaging.
Gold nanoparticles (AuNPs) are used in a wide range of applications in the field of biomedicine, biology and
chemistry (Nel et al., 2006; Volland et al., 2015) mostly because of their low toxicity and high bio-compatibility
(García-Cambero et al., 2013; Lapresta-Fernández et al., 2012). The increased production and use of AgNPs
and AuNPs has raised concerns about their release in the environment (Farkas et al., 2011; Kaegi et al., 2010).
The predicted environmental concentrations in surface water for both AgNPs and AuNPs are expected to be
in the range from 0.01 µg.L-1 to 0.32 µg.L-1 and from 0.14 µg.L-1 to 1.43 µg.L-1, respectively for AgNPs and
AuNPs (Gottschalk et al., 2013; Tiede et al., 2009; Yang et al., 2016). The NPs’ unique physical-chemical
properties in terms of size, shape, surface coatings and charges will substantially influence their behaviour,
fate and their effects/toxicity towards living organisms (Ju-Nam and Lead, 2008; McGillicuddy et al., 2017).
These properties may influence AgNPs and AuNPs entry into the cells (Nel et al., 2006; Oberdörster et al.,
2006; Volland et al., 2015) potentially leading to the generation of reactive species of oxygen (ROS) and
resulting in oxidative stress that has been reported as one of the principal mechanisms of toxicity of NPs in
aquatic organisms (Klaine et al., 2008; Vale et al., 2016; Volland et al., 2015).
Many studies have investigated the impact of AgNPs and AuNPs on key freshwater. Most of the studies
focused on acute toxicity through direct exposure (Buffet et al., 2013). However, in order to better predict
the potential effects of ENPs at higher biological level, such as population and community levels, it is
important to include dietary exposure and assess, through a multi-biomarker approach, the chronic toxicity
of these ENPs (Blinova et al., 2012; Gaiser et al., 2011; Mackevica et al., 2015). It is recognized that one
important uptake route for metals and particles is through the diet (Croteau et al., 2011; Luoma et al., 2014;
Luoma and Rainbow, 2005). Ag was found in the form of nanoparticle in the gut of polychaete exposed
through the diet to AgNPs (García-Alonso et al., 2014). Some studies reported effects of AgNPs and AuNPs
after trophic exposure of aquatic organisms. For instance, trophic exposure of the clam Corbicula fluminae
to AuNPs led to their presence in the epithelia tract of digestive gland and gills of the animals (Renault et al.,
2008). Additionally, an activation of phase II antioxidant enzymes encoding genes and an increase in
metallothionein gene expression was observed (Renault et al., 2008). Trophic exposure (15 days) of zebrafish
to sub-lethal concentrations of AgNPs 20 nm resulted in a significant Ag accumulation in Liver blood vessels
and in the interstitial tissue between the intestine and the liver. Gene expression profiles revealed that AgNPs
Chapter 5 AgNPs and AuNPs chronic toxicity
172
20 nm impacted photo-transduction system, circadian clock regulation and photoreception (Cambier et al.,
2018).
Amphipods of the genus Gammarus such as Gammarus fossarum represent an important part of the aquatic
macroinvertebrate assemblage (Dedourge-Geffard et al., 2009; Ladewig et al., 2006). Recently, G. fossarum
have been used as a model species due to their wide distribution in Europe (Janetzky, 1994) and their major
functional role in litter breakdown process and nutrient cycling (Forrow and Maltby, 2000; Kelly et al., 2002;
Lacaze et al., 2010; MacNeil et al., 1997). The ease of use, identification to the species level, differentiation
between gender, sampling and laboratory handling in addition to the well-documented sensitivity to different
kind of pollutants make Gammarus a good model organisms for ecotoxicological studies (Alonso et al., 2010;
Arce Funck et al., 2013; Felten et al., 2008; Mehennaoui et al., 2016). In our previous study, influence of size
and surface coatings on the acute toxicity of AgNPs and AuNPs were assessed (Mehennaoui et al., subm.,
Chapter 4). Acute exposure of G. fossarum to AgNPs and AuNPs showed:
• A coating dependent bio-accumulation, which was higher for citrate stabilized NPs as compared to
polyethylene glycol-coated (PEG) NPs;
• A metal dependent tissues distribution, with AgNPs found in G. fossarum gills while AuNPs were
found in intestinal caeca;
• An induction of antioxidant genes by AgNPs while AuNPs led to their down regulation.
However, the previous study only investigated the acute effects of AgNPs and AuNPs, but there is still a lack
of information about the chronic toxicity of AgNPs and AuNPs in G. fossarum. Hence, the aim of the present
study was to investigate the chronic toxicity of differently coated AgNPs and AuNPs 40 nm (citrate (CIT) and
polyethylene glycol (PEG) as surface coatings) on G. fossarum. Effects of AgNO3 were also studied in order to
determine the contribution of the released ions to the observed biological effects. A multibiomarker
approach including molecular effects (stress-related gene expression), physiological (osmoregulation) and
behavioural (locomotion and ventilation) responses was used (Mehennaoui et al., 2016). Biological endpoints
also included mortality, bioaccumulation and localization of AgNPs and AuNPs in G. fossarum tissues.
2. MATERIALS AND METHODS
2.1. Particles and Chemicals
Citrate stabilized AgNPs (CIT-AgNPs 40 nm) and citrate stabilized AuNPs (CIT-AuNPs 40 nm) were purchased
from Nanocomposix (San Diego, USA). Coating of AgNPs and AuNPs with PEG was performed at the LIST’s
Material Research and Technology (MRT) Department, (Luxembourg Institute of Science and Technology
(LIST), Luxembourg, Mehennaoui et al., subm., see Chapter 4). AgNO3 (CAS no.7761-88-8) was purchased
Chapter 5 AgNPs and AuNPs chronic toxicity
173
from VWR (Leuven, Belgium). Osmium tetroxide (OsO4, CAS no.20816-12-0), glutaraldehyde (CAS no.111-30-
8) and the Spurr low viscosity resin embedding kit were purchased from Sigma Aldrich (Bornem, Belgium).
Phosphate buffer saline (PBS) was obtained from Invitrogen (Merelbeke, Belgium).
2.2. Particle characterization
Size distribution of CIT-AgNPs, PEG-AgNPs, CIT-AuNPs and PEG-AuNPs 40 nm in Volvic water (exposure
medium), at T0h and T72h, were characterized using Nanoparticle Tracking Analysis (NTA, measured on a
NanoSIGHT instrument, Malvern Instrument Ltd, UK using validated protocols (Hole et al., 2013). Zeta
potential analyses are still on progress and will be performed on a Zeta-sizer Nano (Malvern, Hoeilaart,
Belgium).
2.3. Alder leaves conditioning
Leaves of Alnus glutinosa (Alder) were collected, air-dried and kept at room temperature until use. Leaves
were rinsed in deionized water for few minutes and cut into 16 mm diameter disks. Sets of 50 leaf disks were
placed into fine-net bags (0.5 mm pore size) and immersed in an unpolluted stream (49°48’24.9’’ N and
06°04’53.2’’ E, Schwaarzbaach, Colmar-berg, Luxembourg) (Dohet et al., 2008; Mehennaoui et al., 2016) for
10 days to allow microbial colonization. At the end of the immersion period, leaves were returned to the
laboratory where they were kept at -20 °C until use (Andreï et al., 2016; Batista et al., 2017).
2.4. Organisms sampling and acclimation
G. fossarum were collected in November 2016 at the same unpolluted site in Schwaarzbaach studied in our
previous study (Mehennaoui et al., 2016). Animals were collected using a hand net and they were sorted in
the field to separate males, which were used for the experiments, from females, which were immediately
returned to the water stream. Collected male animals were transported in river water to the laboratory once
collection was finished and they were kept at 12 °C. The decision to keep only adult males for the experiment
is justified by the fact that we wished to avoid influence of gender on the studied parameters (Arce Funck et
al., 2013; Sornom et al., 2010). G. fossarum males were selected from pre-copula pairs or based on sexual
dimorphism like gnatopode size (Mehennaoui et al., 2016). Adult males were then acclimated to laboratory
conditions (Andreï et al., 2016; Mehennaoui et al., 2016). The acclimation was conducted in 2 steps. First,
Gammarids were acclimated for 72h to mineral water (Volvic, France) by progressively changing field water
to Volvic water (30% v/v, 50% v/v, 100% v/v). Then, a stalling period of 10 days was conducted in 100% Volvic
water (Andreï et al., 2016; Mehennaoui et al., 2016). Acclimation was performed under controlled conditions
at 12 °C with a 16h light 8h dark photoperiod. 100% of Volvic water was aerated and changed every 24h to
Chapter 5 AgNPs and AuNPs chronic toxicity
174
avoid organic matter accumulation and potential increase of ammonium, nitrite and nitrate. Gammarids were
fed ad libitum with alder leaves (Alnus glutinosa) up to 24h before experiment.
2.5. Trophic exposure
2.5.1. Experimental design
In order to avoid as much as possible the adsorption of AgNO3, AgNPs 40nm and AuNPs 40 nm, the tanks
used for exposure were saturated with AgNO3, AgNPs 40 nm and AuNPs 40 nm for 96h before exposure
(Andreï et al., 2016).
At the end of the acclimation period, groups consisting of 15 males (compartment: animals) were transferred
into the pre-saturated plastic tanks (500 ml polypropylene tanks) containing 400 ml of exposure medium
(Volvic water with or without contaminants) (Mehennaoui et al., 2016) and 6 leave disks (compart: food). A
piece of mesh was added in each tank to provide a resting surface for G. fossarum and to reduce as much as
possible potential losses linked to cannibalism. Males were exposed to different treatments: AgNO3 (0 – 0.5
µg.L-1); CIT-AgNPs 40 nm, PEG-AgNPs 40 nm, CIT-AuNPs 40nm and PEG-AuNPs 40 nm (0 – 0.5 –5 µg.L-1) for
15 days at 12 °C with a photoperiod of 16h light and 8h darkness. The different ranges of concentrations
tested were selected based on previous results obtained by exposing G. fossarum to the same Ag and Au
nanoparticles for 72h (Mehennaoui et al., subm., see Chapter 4). AgNO3, AgNPs and AuNPs stock solutions
were diluted to the desired concentrations in mineral water (Volvic, France). 100% of exposure medium
(compartment: water colomn) were renewed every 72h to maintain the exposure conditions stable, to ensure
sufficient amount of nutrients and to remove waste products. Dead animals were removed and survivors
were counted every 24h. 6 leave disks per replicates were added every 48h.
2.5.2. Total and dissolved silver and gold measurements
In order to better describe the fate of Ag and Au in the studied compartment (water column, food or animals),
water samples were filtered on a nitrate cellulose filter (47 mm diameter, 0.45 µm pore size) previously rinsed
with Milliq water, dried for 24h at 80 °C and weighed. After filtration, the filters were dried for 24h at 80 °C
and weighed in order to determine the quantity of fine particle suspended matter produced after 72h of
exposure (Andreï et al., 2016). 5 mL of filtered water were kept for total Ag and Au measurement and 4 mL
of filtered water were used for ion release measurements and were centrifuged at 4000 g for 40 min using a
3KDa cut off membrane (Millipore) followed by a mineralization step with 1% HNO3 for Ag ion measurements
and 1% of HCl for Au ion measurements (Georgantzopoulou et al., 2013). The analyses are currently ongoing.
All measurements will be performed on an Inductively Coupled Plasma Mass spectrometer (ICP-MS, ElanDRC-
e, Perkin-Elmer, Waltham, MA, USA) as previously described (Mehennaoui et al., 2016).
Chapter 5 AgNPs and AuNPs chronic toxicity
175
2.5.3. Silver and gold bioaccumulation
At the end of the exposure period, three pools of 4 gammarids per each condition were rinsed with milliQ
water and dried on filter paper. Animals were frozen at -20 °C, freeze-dried and weighed. Samples were then
mineralized in HNO3 and H2O2 for Ag measurement and in acidic mixture of HNO3 and HCl for Au
measurement at a maximum pressure of 35 bars and maximum temperature of 200 °C using a microwave
system (Anton Paar Multiwave Pro) as previously described (Mehennaoui et al., 2016). All measurements
were performed in triplicates and are expressed in µg.g-1 of dry weight. A calibration curve was performed
following serial dilutions (from 0.01 to 50 µg.L-1 of Ag and from to for Au). Using a multi-element standard
solution (Chemlab, certified ISO/IEC 17025). The validity of the analytical method was controlled every 10
measurements within the same series of measurements with 3 quality controls (0.5, 5 and 25 µg.L-1 of Ag and
Au) of a second multi-element standard solution (Merck-VWR, ISO/IEC 17025). Ag and Au values were
consistent within the certified ranges (Cambier et al., 2018).
2.5.4. Particles uptake: Cytoviva ® analyses
2.5.4.1. Sample preparation
At the end of the exposure period, G. fossarum exposed to 5 µg. L-1 of AgNPs and AuNPs were fixed in 5%
(v/v) glutaraldehyde where they were kept at 4 °C until analyses. After a washing step with PBS, samples were
post-fixed with 1% osmium tetroxide for 2h at room temperature on a horizontal shaker (180 rpm). After a
second washing with PBS, samples were dehydrated with increasing concentrations of ethanol (10%, 20%,
30%, 50%, 70%, 90% and 100% v/v) and a final wash in acetone (100% v/v) before being embedded in Spurr
resin for 16h at 60 °C.
2.5.4.2. Cytoviva® dark field hyperspectral imaging
Semi-thin layers of 300 nm were cut using ultra-microtome and placed on microscope glass slide and let dry
for 1 min on a drying workstation. Slides were covered with a glass cover slip, sealed, and kept at 4 °C until
analysis. Visualization was performed using a Cytoviva® darkfield hyperspectral imaging system (Cytoviva Inc.,
Auburn, Alabama, USA) mounted on an Olympus BX-43 optical microscope. Images of the different tissues of
G. fossarum were captured at 60x magnification with oil immersion using hyperspectral camera controlled
by ENVI software (version 4.8 from Harris Corporation, Melbourne, FL, USA and modified by CytoViva, Inc.).
Spectral libraries of exposed animals were generated manually acquiring about 200 spectra per sample.
Acquired libraries were filtered against non-exposed samples to filter out all spectra non-related to AgNPs or
AuNPs using a spectral angle mapper (SAM) algorithm with a 0.1 radians tolerance. Filtered libraries were
Chapter 5 AgNPs and AuNPs chronic toxicity
176
mapped onto images of exposed samples using SAM with a 0.1 radians tolerance which allows highlighting
similarities between the spectra in the image and in the spectral library.
2.5.5. Molecular effects
2.5.5.1. RNA extraction and cDNA synthesis
At the end of the exposure period, animals were gently dried and dived in RNA lysis Buffer (RLT buffer)
supplemented with 1% (v/v) β-mercaptoethanol, flash-frozen in liquid nitrogen and stored at -80 °C until RNA
extraction. For RNA extraction, G. fossarum tissues were grinded on ice using a pellet pestle tissue grinder
and the homogenates were centrifuged at 250g for 5 min at 4 °C to remove cuticle fragments as previously
described (Mehennaoui et al., 2016). Total RNA was extracted using RNeasy mini kit for cells and animal
tissues (Qiagen, Leusden, The Netherlands), according to the manufacturer’s instructions (including digestion
with DNase). RNA concentrations and purity were assessed measuring the absorbance at 230, 260 and 280
nm using Nanodrop ND-1000 spectrophotometer (ThermoScientific, Villebon-sur-Yvette, France) (A260/280
and A260/230 ratios). RNA integrity was controlled using the RNA Nano 6000 assay (Agilent Technologies,
Diegem, Belgium) with 2100 Bioanalyzer (Agilent, Santa Clara, CA, USA) (Legay et al., 2015). For the synthesis
of cDNA, 1 µg of the extracted RNA were retro-transcribed into cDNA using the Protoscript II reverse
transcriptase (New England Biolabs, Leiden, The Netherlands) and random primers, according to the
manufacturer’s instructions.
2.5.5.2. Primer design and quantitative real-time PCR
All the primers for the RT-qPCR were designed using Primer3Plus (http://primer3.ut.ee) with the previously
described criteria (Mehennaoui et al., subm., chapter 3; Mehennaoui et al., subm., chapter 4). Primers were
analysed using NetPrimer (http://www.premierbiosoft.com/netprimer/) for secondary unexpected
structures. PCR efficiency was checked using decreasing five-fold dilutions from cDNA pool (From 25 ng to
0.04 ng and no template control). Melting curves were generated at the end of each experiment to assess
the specificity of the amplified products. All primers displayed one clear peak and were all retained for the
analysis. All the used primers are described in our previous study (Mehennaoui et al., subm., chapter 3 and
4)
cDNAs were diluted to 0.8 ng. µL-1 and used for RT-qPCR analyses in 384-well plates. An automated liquid
handling robot (epMotion 5073, Eppendorf, Hamburg, Germany) was used for the preparation of the 384-
well plate. cDNA were amplified using Takyon Low ROX SYBR MasterMix dTTP Blue Kit (Eurogentec, Liège,
Belgium) on a ViiA 7 Real-Time PCR System (Thermo Fisher, Waltham, MA, USA) in a final volume of 10 µL
(Behr et al., 2015). All reactions were performed in three technical replicates each one including four
Chapter 5 AgNPs and AuNPs chronic toxicity
177
biological replicates. The PCR conditions consisted on an initial denaturation step at 95 °C for 3 min, followed
by 40 cycles of denaturation at 95 °C for 15 sec and annealing/extension steps at 60 °C for 60 sec. Two genes,
namely Clathrin and SDH, were selected among six genes to be used as reference and they were validated
using the GeNorm module in the Biogazelle qBase PLUS software (Mehennaoui et al., subm., chapter 3). The
relative gene expression was calculated including PCR efficiency using Biogazelle qbase Plus software 2.5 with
the ΔΔCT method.
2.5.6. Behavioural responses and osmoregulation
For each treatment, measurements of locomotor activity, ventilation and osmoregulation were performed
on the same pool of 10 organisms (Mehennaoui et al., 2016).
2.5.6.1. Locomotor activity
At the end of the exposure period, locomotor activity was firstly assessed by counting the number of animals
in movement in a 80 mL glass tank containing 10 organisms with a piece of net added to provide a resting
surface (Felten et al., 2008). Measurements were performed after 5 min of acclimation at the same time of
the day with similar light conditions and in a quiet environment. Moving G. fossarum were counted for a
period of 2 sec and this process was repeated 40 times (Mehennaoui et al., 2016).
2.5.6.2. Ventilation activity
Ventilation activity was recorded immediately after locomotion on the same animals, by measuring the
frequency of pleopod beats. Ten gammarids from each treatment group were placed individually in a glass
tube containing Volvic water and left for a 30 sec acclimation period. Then, pleopod beats were visually
counted three times for 10 seconds, using a manual cell counter, only when animals were at rest.
Measurements were performed at the same period of the day to avoid possible effects of a circadian rhythm
on respiration (Felten et al., 2008; Mehennaoui et al., 2016; Rosas et al., 1992).
2.5.6.3. Osmoregulation
2.5.6.3.1. Haemolymph sampling
Immediately after behavioural response measurements, animals were dried on filter paper. Samples of
haemolymph were taken from the telson of each individual (n = 10) using a modified microcapillary (Felten
et al., 2008). To prevent haemolymph coagulation, samples were Kept at 4 °C until Na+, Cl- and Ca2+
measurements (Arce Funck et al., 2013).
Chapter 5 AgNPs and AuNPs chronic toxicity
178
2.5.6.3.2. Haemolymph Na+, Cl- and Ca2+ concentrations measurements
Samples of haemolymph (0.8 – 1.2 µL) were transferred to a gauged 5 µL microcapillary tube and were
centrifuged for 10 min at 6600 g at 4 °C. Samples were then diluted in 2.5 mL of MilliQ water to determine
Na+ and Ca2+ concentrations using atomic absorption spectrophotometry (Perkin Elmer Analyst 100) and Cl-
concentrations using ionic chromatography (Dionec 4500i equipped with an Ion Pac AS4A column) (Arce
Funck et al., 2013; Felten et al., 2008).
2.6. Statistical analyses
Survival, bioaccumulation, osmoregulation and behavioural response results are reported as mean ± SD
(standard deviation) and relative gene expression results are reported as mean ± SE (standard error). Results
of locomotor activity were obtained from the same set of individuals. Therefore, a one-way ANOVA with
repeated measurements following a general linear model was used as statistical test. The observed Ag and
Au bioaccumulation, survival, ventilation and gene expression on log2 transformed normalized relative
quantities were compared using a one-way ANOVA followed by Tukey-HSD post-hoc test for multiple
comparisons when normality and homogeneity of variances were respected. When this was not the case, a
non-parametric Kruskal-Wallis ANOVA followed by a Mann-Whitney U test was used. All statistical
calculations were performed using Statistica software 12 (StatSoft Inc.).
3. RESULTS
3.1. Particle character ization
In our previous study CIT-AgNPs, PEG-AgNPs 40 nm were the most stable in Volvic water with sizes of 45 ±
1.5 and 47 ± 1.3 nm, respectively (Mehennaoui et al., subm., chapter 4). CIT-AuNPs 40 nm had a size of 67 ±
13 nm after 24h in Volvic water, while PEG-AuNPs was stable after 24 hours in Volvic water with a size of 38
± 19 nm. In the present study, size distribution, which is measured using the NanoSIGHT, could not be
characterized due to the presence of food preventing the distinction between AgNPs, AuNPs and leave
particles. Thus, further analyses are needed in order to define the behaviour of AgNPs and AuNPs in Volvic
water in the presence of food.
3.2. Survival
A general tendency for a decrease in survival rates following exposure time (15 days) was observed in all the
exposure conditions (Figure 5. 1). However, only 0.5 µg. L-1 of AgNO3 led to a significant decrease in survival
rates of G. fossarum at day 15 (65 ± 3.3 %, One-way ANOVA, F= 11.63, 8df, P < 0.001) (Figure 5. 1). Survival
Chapter 5 AgNPs and AuNPs chronic toxicity
179
rates of G. fossarum tended to decrease with increasing exposure time to AgNPs and AuNPs even though
effects of AgNPs and AuNPs on survival were not statistically significant.
Figure 5. 1. Survival rates (mean ± SD) of G. fossarum exposed for 15 days to 0.5 and 5 µg. L-1 of CIT and PEG-
AgNPs and PEG-AuNPs 40 nm
3.3. Silver and gold b ioaccumulation
At the end of the exposure period, small but detectable amounts of Ag could be detected in the negative
control groups of G. fossarum (0.24 ± 0.1 µg. g-1 dry weight), for which the observed Ag body concentration
was 0.24 ± 0.01 µg. g-1 dry weight (Table 5. 1). Treatment with AgNO3 (0.5 µg. L-1) lead to the highest increase
of silver body concentrations that were of 5.16 ± 1.20 µg. g-1 d.w. Similarly, treatment with the AgNPs resulted
in a significant concentration-dependent Ag bioconcentration for both CIT-AgNPs and PEG-AgNPs 40 nm
(One-way ANOVA, F=45.12, 17df, P < 0.01). The highest concentrations (5 µg. L-1) of CIT-AgNPs and PEG-
AgNPs 40 nm led to a significant higher Ag bioconcentration in exposed animals compared to non-exposed
ones (P < 0.0001, Table 5.1).
The Au bio-concentration was under the limit of quantification (< 0.075 µg. g-1 d.w) in the non-exposed
animals. G. fossarum exposed to the highest concentration (5 µg. L-1) of CIT and PEG-AuNPs 40 nm tended to
accumulate more Au than those exposed to 0.5 µg. L-1 of the same AuNPs (P > 0.05, Table 5. 1).
0
20
40
60
80
100
120
0 0.5 5 0 0.5 5 0 0.5 5 0 0.5 5 0 0.5
CIT PEG CIT PEG
AgNPs40nm AuNPs40nm AgNO3
Survival(%)
Exposureconditions
Chapter 5 AgNPs and AuNPs chronic toxicity
180
Table 5. 1. G. fossarum Ag and Au uptake (mean ± SD) after 15 days of exposure to CIT and PEG-AgNPs and
AuNPs 40 nm.
Different letters (a-d) indicate significant differences (one-way ANOVA + Tukey HSD post hoc test at P < 0.05). Bold
values indicate significant differences compared to control G. fossarum.
Nominalconcentrations
(µg.L-1)
Gammarus internal[Ag]and
[Au](µg.g-1gammaridsdry
weight)
0 0.24±0.01a
0.5 0.42±0.05ab
5 2.09±0.19c
0 0.24±0.01a
0.5 0.37±0.28ab
5 1.55±0.28bc
0 <LOQa
0.5 0.11±0.01a
5 1.03±0.18abc
0 <LOQa
0.5 0.10±0.03a
5 0.63±0.16ab
0 0.21±0.04a
0.5 5.16±1.21d
NPs
AgNO3
AgNPs40nm
CIT
PEG
AuNPs40nm
CIT
PEG
Chapter 5 AgNPs and AuNPs chronic toxicity
181
3.4. Cytoviva® dark f ie ld hyperspectral imaging
A Cytoviva dark field hyperspectral imaging microscope was used in order to localize the internalized AgNPs
and AuNPs 40 nm. Currently only results for G. fossarum exposed to CIT-AgNPs and CIT-AuNPs 40 nm are
available (Figure 5. 2, Figure 5. 3).
Figure 5. 2 A) 300 nm cross section of G. fossarum observed under optical microscope. Red square represents
the area observed with cytoviva darkfield hyperspectral microscope. Scale bar is 250 µm. B) G. fossarum gills or
intestinal caeca viewed with Cytoviva (60x oil immersion magnification). C) AgNPs accumulation in G. fossarum
gills (red spots and white arrows). Animals were exposed for 72h to 10 µg. L-1 of CIT-AgNPs 40nm and PEG-AgNPs
40 nm in Volvic water. Scale bar is 6µm.
Analyses of samples exposed to CIT-AgNPs and CIT-AuNPs 40 nm were performed in different organs and
tissues (gills, gut and muscle). Tissues were selected based on the results obtained in the previous study
where AgNPs were found in the gills while AuNPs were found in the gut area of G. fossarum (Mehennaoui et
al., subm., chapter 4). In the present study, in addition to the long-term effects, the influence of food on the
internal distribution of AgNPs and AuNPs was investigated.
The analysis of Ag revealed the presence of AgNPs in the gills of G. fossarum exposed to CIT-AgNPs 40 nm
(Figure 5. 2C) while the analyses of Au showed the presence of AuNPs within the lumen and epithelial tissues
of intestinal caeca on G. fossarum exposed to CIT-AuNPs (Figure 5. 3C).
Cit-AgNPs
40 nm
Gills
Control
A B C
Chapter 5 AgNPs and AuNPs chronic toxicity
182
Figure 5. 3. A) 300 nm cross section of G. fossarum observed under optical microscope. Red square represents
the area observed with CytoViva darkfield hyperspectral microscope. Scale bar is 250 µm. B) G. fossarum intestinal
caeca viewed with CytoViva (60x oil immersion magnification). C) AuNPs accumulation in G. fossarum caeca (red
spots and white arrows). Animals were exposed for 72h to 10 µg. L-1 of CIT-AgNPs 40nm and PEG-AgNPs 40 nm in
Volvic water. Scale bar is 6µm.
3.5. Molecular effects
In the present study, the same set of stress-related genes (Mehennaoui et al., subm., chapter 4) was tested
in order to define the influence of the presence of food and the longer exposure time (15 days) on the
molecular responses of G. fossarum to AgNPs and AuNPs 40 nm. Both concentrations (0.5 and 5 µg. L-1) of
PEG-AgNPs 40 nm led to an up-regulation of catalase and chitinase and CIT-AgNPs and PEG-AgNPs 40 nm led
to an up-regulation of Na+K+ATPase. The highest concentrations of CIT and PEG-AuNPs 40 nm (5 µg. L-1) led
to an increase of CuZnSOD gene expression (Table 5. 2).
3.6. Behavioural responses and osmoregulation
3.6.1. Locomotion and venti lation activity
G. fossarum exposed to CIT-AgNPs, PEG-AgNPs 40 nm, CIT-AuNPs and PEG-AuNPs 40 nm and AgNO3 showed
a significant reduced locomotor activity when compared to non-exposed animals (Figure 5. 4 A, GLM - one-
way ANOVA, P < 0.05). Animals exposed to 0.5 µg. L-1 of CIT-AgNPs 40 nm showed a significant reduced
locomotion activity when compared to G. fossarum exposed to 5 µg. L-1 CIT-AgNPs 40 nm and 0.5 µg. L-1 of
AgNO3 (GLM- Tukey-HSD post-hoc test, P < 0.05). At the highest concentration (5 µg. L-1), PEG-AgNPs 40 nm
led to a significant decrease in locomotor activity compared to CIT-AgNPs 40 nm (GLM- Tukey-HSD post-hoc
test, P < 0.05).
CIT-AuNPs
40 nm
Intestinal caeca
Control
A B C
Chapter 5 AgNPs and AuNPs chronic toxicity
183
Table 5. 2. Relative gene expression of G. fossarum exposed for 15 days to CIT- and PEG-AgNPs 40 nm and CIT and PEG-AuNPs 40 nm
Exoskeleton Lysosomes
Actin TUB UB Chitinase CAT CuZnSOD GPx7 GST HSP90 GAPDH Gadd45 Cjun CathL Ca2+ATPase NaKATPase HEM Cox1
0 1±0.18 1±0.13 1±0.06 1±0.14 1±0.12 1±0.32 1±0.08 1±0.17 1±0.12 1±0.1 1±0.05 1±0.12 1±0.19 1±0.15 1±0.14 1±0.12 1±0.26
0.5 1.44±0.29 1.33±0.28 1.16±0.13 0.71±0.17 0.73±0.17 0.66±0.16 0.83±0.08 1.5±0.21 1.07±0.16 0.77±0.08 1.04±0.05 1.37±0.11 1.54±0.35 0.89±0.44 1.13±0.08 0.95±0.04 1.07±0.14
Control 0 1±0.15 1±0.12 1±0.48 1±0.11 1±0.1 1±0.15 1±0.04 1±0.22 1±0.13 1±0.1 1±0.12 1±0.26 1±0.23 1±0.1 1±0.13 1±0.14 1±0.25
0.5 0.56±0.11 0.79±0.09 1.33±0.09 1.15±0.12 1.02±0.14 1.13±0.08 0.8±0.08 0.71±0.15 1.33±0.08 0.89±0.04 1.15±0.08 1.18±0.2 0.61±0.17 1.07±0.15 1.39±0.15 0.79±0.04 1.11±0.1
5 0.74±0.19 0.89±0.04 0.93±0.35 1.48±0.3 1.43±0.23 0.88±0.19 0.92±0.11 0.86±0.16 0.93±0.11 0.91±0.07 1.06±0.19 1.09±0.2 0.76±0.16 1.1±0.18 1.55±0.16 0.89±0.07 1.4±0.4
0.5 0.77±0.07 0.97±0.22 1.09±0.25 2.17±0.95 1.98±0.83 1.2±0.03 0.84±0.07 0.84±0.08 1.09±0.07 1.03±0.15 1.27±0.03 1.47±0.22 0.74±0.13 1.11±0.04 1.49±0.08 0.97±0.15 1.51±0.3
5 0.77±0.22 0.92±0.12 0.58±0.33 1.68±0.57 1.7±0.63 0.9±0.32 0.88±0.08 0.81±0.12 1.42±0.08 0.82±0.11 1.26±0.32 1.28±0.07 0.69±0.11 1.12±0.03 1.53±0.12 0.68±0.11 0.97±0.43
Control 0 1±0.19 1±0.15 1±0.22 1±0.11 1±0.06 1±0.24 1±0.11 1±0.13 1±0.11 1±0.09 1±0.24 1±0.07 1±0.15 1±0.05 1±0.13 1±0.09 1±0.39
0.5 1.02±0.16 0.93±0.22 0.96±0.06 1.2±0.2 1.31±0.24 1.71±0.35 1.04±0.16 1.02±0.09 0.66±0.16 0.91±0.11 0.89±0.35 1.05±0.13 1.09±0.23 0.97±0.06 0.95±0.09 0.83±0.11 0.68±0.15
5 1±0.21 0.77±0.13 0.85±0.06 1.46±0.13 1.43±0.09 1.21±0.28 1.05±0.12 0.83±0.14 0.86±0.12 1.11±0.04 1.15±0.28 1.43±0.27 0.88±0.16 0.89±0.06 0.81±0.14 1.18±0.04 0.97±0.36
0.5 0.97±0.04 0.61±0.24 1.23±0.05 1.29±0.23 1.24±0.27 1.25±0.18 0.99±0.08 0.87±0.04 0.69±0.08 1±0.14 0.81±0.18 0.83±0.15 0.82±0.09 0.79±0.07 0.66±0.04 1.23±0.14 1.12±0.15
5 1.26±0.2 0.94±0.1 1±0.14 1.42±0.27 1.44±0.23 2.14±0.54 1.01±0.08 1.13±0.14 0.96±0.08 0.68±0.17 1.06±0.54 1.27±0.17 1.31±0.27 0.88±0.07 0.8±0.14 0.92±0.17 1.15±0.56
Generalstress DNAdamageandrepair Osmoregulation Respiration
AuNPs40nmCitrate
PEG
Cytoskeleteon Antioxidantdefensecompound concentrations(µg.L
-1)
AgNO3
AgNPs40nmCitrate
PEG
Chapter 5 AgNPs and AuNPs chronic toxicity
184
A significant decrease in the number of moving Gammarids was observed when animals were exposed to 0.5
µg. L-1 of CIT-AuNPs 40 nm compared to those exposed to 5 µg. L-1 of CIT-AuNPs 40 nm (GLM- Tukey-HSD post-
hoc test, P < 0.05) while no significant effects were observed in Gammarids exposed to PEG-AuNPs 40 nm
(GLM- Tukey-HSD post-hoc test, P > 0.05).
Pleopod beats frequency of G. fossarum exposed to 5 µg. L-1 of CIT-AgNPs 40 nm and CIT-AuNPs 40 nm as not
significantly affected by AgNPs, AuNPs and AgNO3 (Figure 5. 4 B, Kruskal-Wallis ANOVA, P < 0.05).
Figure 5. 4. Behavioural responses of G. fossarum exposed for 15 days to CIT and PEG-AgNPs and AuNPs 40 nm.
A) Locomotor activity (mean percentage of moving G. fossarum ± SD) and b) Ventilation (mean pleopods beat
frequency ± SD). Different letters (a-f) indicates significant differences (one-way ANOVA + Tukey HSD post hoc test
at P < 0.05 level of significance, n = 10).
0
10
20
30
40
50
60
70
80
90
100
0 0.5 5 0 0.5 5 0 0.5 5 0 0.5 5 0 0.5
CIT PEG CIT PEG
AgNPs40nm AuNPs40nm AgNO3
%Gammarusinmouvement
Exposureconditions(µg.L-1)
a
a ab
f
c
c
c
cd
ee
ff
de
f
A
0
50
100
150
200
250
300
350
400
450
0 0.5 5 0 0.5 5 0 0.5 5 0 0.5 5 0 0.5
CIT PEG CIT PEG
AgNPs40nm AuNPs40nm AgNO3
Ventilation(Pleopodsbeatperminute)
Exposureconditions(µg.L-1)
abab
abab ab
b
aa
ab
ab
ab
abab b
B
Chapter 5 AgNPs and AuNPs chronic toxicity
185
3.6.2. Osmoregulation
The basal level of haemolymphatic Cl- in control gammarids was constant with values of 70.4, 101.9 and 95.5
mmol. L-1 in the three different control groups. The basal level of haemolymphatic Na+ was 108.8, 169.4 and
114.2 mmol. L-1 in the non-exposed animals. Basal level of haemolymphatic Ca2+ were 39.7, 72.7, and 28.7
mmol. L-1 in the non-exposed animals (Table 5. 3). No significant effects of AgNPs and AuNPs 40 nm were
observed on the concentration of the different haemolymphatic ions indicating that osmoregulation of G.
fossarum was not impacted by the applied treatment (Kruskal-Wallis ANOVA, P > 0.05).
Table 5. 3. Haemolymph [Cl-], [Na+] and [Ca2+] of G. fossarum exposed for 15 days to CIT and PEG-AgNPs 40 nm
and CIT and PEG-AuNPs 40 nm.
4. Discussion
The chronic toxicity of AgNPs and AuNPs on G. fossarum is still not fully understood. Hence, in the present
study we investigated the chronic effects of differently coated (CIT and PEG) AgNPs and AuNPs 40 nm on G.
fossarum. Individuals were exposed to realistic environmental concentrations (0.5 and 5 µg. L-1) of AgNPs and
AuNPs 40 nm for 15 days and in presence of food (leaves disk of A. glutinosa). AgNPs and AuNPs 40 nm were
selected based on the results of our previous study (Mehennaoui et al., subm., chapter 4). AgNPs and AuNPs
40 nm were selected as they were the most stable in exposure medium and the NPs with the highest uptake
rate by G. fossarum as described in our previous study (Mehennaoui et al., subm. Chapter 4). The studied
endpoints include mortality, bioaccumulation, tissue distribution, gene expression, osmoregulation and
behavioural responses (locomotion and ventilation).
Nominal
concentrations
(µg.L-1)
haemolymph[Cl-]
(mmol.L-1)
haemolymph[Na+]
(mmol.L-1)
haemolymph[Ca2+]
(mmol.L-1)
0 70.4±31 108.9±42 39.7±5
0.5 66.4±22 103.4±37 41.6±39
5 69.2±12 110.0±27 31.9±12
0 70.4±31 108.9±42 39.7±5
0.5 82.1±7 123.2±7 36.3±7
5 78.5±23 129.4±17 49.2±27
0 101.9±18 169.4±54 72.7±33
0.5 102.8±8 125.3±17 36.8±20
5 86.5±5 137.5±22 40.2±4
0 101.9±18 169.4±54 72.7±33
0.5 84.5±10 137.4±9 34.6±5
5 116.3±20 130.8±22 22.5±3
0 95.6±23 114.3±17 28.7±7
0.5 75.3±24 105.5±40 107.8±62
AuNPs40nm
CIT
PEG
AgNO3
NPs
AgNPs40nm
CIT
PEG
Chapter 5 AgNPs and AuNPs chronic toxicity
186
4.1. Particle character ization
AgNPs and AuNPs 40 nm were characterized in the exposure media (Volvic water) that contained food debris.
For the characterization of the size distribution in water we, first, used the nanoparticle tracking analysis
(NTA), which, however, did not allow a complete and relevant characterization of the studied NPs at the used
concentrations (0.5 and 5 µg. L-1) since the presence of the food debris interferes with the detection of
particles. Further analyses using NanoZetasizer are on progress. Thus, the necessity of feeding in the present
experiment can be considered as an influencing/confounding factor on the behaviour of AgNPs and AuNPs
as observed in previous studies (Mackevica et al., 2015; Sørensen et al., 2016). Potentially AgNPs and AuNPs
interact with the different compound present in the exposure medium and may be complexed or aggregated
(Andreï et al., 2016; Levard et al., 2012; Petersen et al., 2014) but further analyses are needed in order to
verify this hypothesis. This behaviour makes the commonly used techniques for characterization (DLS, NTA
and electron microscopy) unsuitable for environmental samples analysis due to their detection limits (Farré
et al., 2008; Gallego-Urrea et al., 2011; Yang et al., 2016). One possibility is to test new techniques such as
the combination of the ORION NanoFab helium ion microscope (HIM) with a secondary ion mass
spectrometer (SIMS) that may produce an elemental SIMS maps with a high resolution images (Dowsett et
al., 2016). This technique might allow distinction between metallic element and carbon element although it
might not allow the determination of the size distribution of AgNPs and AuNPs.
4.2. Biological endpoints
Trophic exposure of G. fossarum for 15 days to CIT and PEG-AgNPs and CIT-and PEG-AuNPs 40 nm led to a
concentration-dependent uptake of Ag and Au. These findings are in agreement with what was already
described in our previous study in which G. fossarum were exposed for 72h to an increasing gradient of
different sizes of differently coated AgNPs and AuNPs (Mehennaoui et al., subm., chapter4; Mehennaoui et
al., 2016). Dose-dependent accumulation was also reported in G. roeseli exposed for 72h to 0.5 and 5 µg.L-1
of CIT-AgNPs 10 and 60 nm (Andreï et al., 2016). G. fossarum exposed to the highest concentration of CIT-
AuNPs 40 nm (5 µg. L-1) tended to accumulate more Au than non-exposed animals and those exposed to PEG-
AgNPs and PEG-AuNPs. Similar observations were done in previous studies reporting a rapid uptake of Au by
the clams Ruditapes philippinarum exposed for 7 days to 0.75 µg.L-1 of CIT-AuNPs 20 nm and Scorbicularia
plana exposed for 16 days to 100 µg.L-1 of AuNPs 40 nm (Pan et al., 2012). The internal concentrations of Ag
tended to be higher than those of Au which could be linked to different metal speciation in the exposure
medium and/or their affinity to the binding sites (Sørensen et al., 2016). The uptake of Ag and Au after a
dietary exposure in G. fossarum may be relevant of a potential transfer of these NPs in the food chain.
Chapter 5 AgNPs and AuNPs chronic toxicity
187
AgNO3 led to the highest uptake compared to AgNPs and was associated with a significant mortality rate in
G. fossarum, which is consistent to the short-term exposure (Mehennaoui et al., 2016). Similarly, other
studies reported a significant uptake associated with a significant mortality of G. roeseli exposed to AgNO3
for 72h (Andreï et al., 2016) and G. fossarum exposed to up to 8 µg. L-1 of AgNO3 for 96h (Arce Funck et al.,
2013). However, despite an uptake of Ag and Au, no significant effects on survival were observed when
animals were exposed to CIT and PEG-AgNPs 40 nm and CIT and PEG-AuNPs 40 nm. The internal
concentration of Ag recorded in G. fossarum reached only 2 µg.g-1 gammarid d.w. which is less than the toxic
threshold of 8 µg.g-1 gammarid d.w. previously reported in G. fossarum (Arce Funck et al., 2013). Observed
differences in toxicities are most likely due to the ratio of NPs to free ions which has been shown to be an
important factor for toxicity. NPs used in the current study have low dissolution of below 6% (Mehennaoui
et a., subm., chapter 4) which may indicate the low toxicity observed for AgNPs.
Knowing the behaviour of AgNPs and AuNPs, once they are taken up by an organism, is crucial for better
definition of their fate and their potential toxic mechanisms. Thus, the internal distribution of AgNPs and
AuNPs was evaluated using enhanced darkfield hyperspectral microscopy. AgNPs were observed in the gills
of G. fossarum exposed to CIT-AgNPs 40 nm similar to short-term exposure (Mehennaoui et al., subm.,
chapter 4). However, in the presence of food during exposure of the organisms less AgNPs were observed
compared to our previous study with exposure via water. CIT-AuNPs 40 nm were observed within the lumen
of intestinal caeca of G. fossarum. This may suggest that AuNPs were ingested with the food and might not
cross the intestinal barrier and might be eliminated with faeces. This observation may also explain the
absence of significant effects of AuNPs on G. fossarum survival rates.
Acute waterborne exposure of G. fossarum to CIT-AgNPs and CIT-AuNPs 40 nm showed an uptake of AgNPs
in the gills and AuNPs in membrane of intestinal caeca of AuNPs (Mehennaoui et al., subm., chapter4). The
necessity of feeding in the present work had a significant effect on the pattern of internal distribution of
AgNPs and AuNPs. Fewer AgNPs were observed in G. fossarum gills when compared to gammarids exposed
to AgNPs in the absence of food (Mehennaoui et al., subm. Chapter 4), while AuNPs were observed only
within the lumen of intestinal caeca which may suggest a rapid elimination of AuNPs with faeces. AuNPs were
observed in the gut of D. magna after direct exposure and were not eliminated while the presence of food
during the experiment resulted in their elimination (Botha et al., 2016). Other study reported an adsorption
of AuNPs on algae surface which led to their ingestion by D. magna followed by a rapid elimination from the
gut (Khan et al., 2014; Volland et al., 2015; Wray and Klaine, 2015).
No significant effects of AgNPs or AuNPs on concentrations of haemolymphatic ions, namely Na+, Cl- and Ca2+,
were observed. As described above, only small amounts of AgNPs were observed in G. fossarum gills
indicating a potential lower Ag uptake through the gills. The dietary exposure may have decreased the
Chapter 5 AgNPs and AuNPs chronic toxicity
188
bioavailability of Ag, leading to the formation of different complex with biomolecules, thus in turn making Ag
less toxic or favouring a different uptake route (Mann et al., 2004; Webb and Wood, 1998). Thus, the low
presence of AgNPs in the gills may explain the absence of significant effects on concentrations of
haemolymphatic ions (Mann et al., 2004). Similarly, the presence of AuNPs within the lumen of intestinal
caeca of G. fossarum may indicate that these compounds are likely to be eliminated with faeces without
crossing intestinal barrier and reaching the circulatory system. Gammarids exposed to metallic contaminants
are usually described to exhibit an iono/osmotic disruption including changes in the Na+K+ATPase activity,
ions influx and gills surface permeability (Lignot et al., 2000). Osmoregulation impairment can therefore be
noticed by an impact on haemolymphatic Na+ and Cl- concentrations which represent 90% of the osmotic
pressure in gammarids (Mantel and Farmer, 1983). It is also known that an acute exposure to Ag generally
lead to the inhibition of Na+ uptake through the inhibition of Na+K+ATPase activity (Bianchini and Wood, 2002,
2003; Bury et al., 2002; Grosell et al., 2002; Webb and Wood, 1998). This mechanism was previously observed
in G. fossarum exposed to AgNO3 and AgNPs 23 nm for up to 96h (Arce Funck et al., 2013; Mehennaoui et al.,
2016). Even though, the major haemolymphatic ions were not impacted by AgNPs and AuNPs, an
osmoregulation impairment could occur as other elements present in haemolymph, such as amino-acids,
may be affected (Felten et al., 2008) and thus, further analyses are needed to better define the chronic effects
of AgNPs and AuNPs at the physiological level of G. fossarum.
A set of stress-related genes was used in order to investigate the effects of AgNPs and AuNPs 40 nm at the
molecular level (Mehennaoui et al., subm., chapter 3-4.). Both concentrations (0.5 and 5 µg. L-1) of PEG-AgNPs
40 nm led to an up-regulation of catalase and chitinase and CIT-AgNPs and PEG-AgNPs 40 nm led to an up-
regulation of Na+K+ATPase. The highest concentrations of CIT and PEG-AuNPs 40 nm (5 µg. L-1) led to an
increase of CuZnSOD gene expression. One previous study reported an over-expression of SOD and catalase
in G. fossarum exposed for 48h to 4.56 mg.L-1 of positively charged AuNPs 10 nm (Baudrimont et al., 2017).
Authors linked the up-regulation of antioxidant defense to the increase of the number of mitochondria,
observed via induction of 16S gene, as these mechanisms require energy (Baudrimont et al., 2017). Similarly,
catalase was induced in Chironomus riparus exposed to up to 1 mg.L-1 of AgNPs (Nair et al., 2013). At a higher
biological level, an activation of phase II antioxidant enzymes after exposing R. philippinarum to 0.75 µg.L-1 of
CIT-AuNPs 20 nm for 24h was reported (Volland et al., 2015). However, by increasing the exposure time,
authors observed significant alterations in the gene expression of GPx and genes related to inflammatory
responses (Gadd45) as well as MT (Volland et al., 2015). Additionally, digestive gland of bivalve appeared as
the target organ for AuNPs under short-term exposure. 1 day of exposure led to the activation of phase II
antioxidant enzymes (GST, CAT, GPx and GR) (Volland et al., 2015) while 7 days of exposure led to a persistent
oxidative stress linked to an increase in GPx and SOD expression and a decrease of Gadd 45 (Volland et al.,
Chapter 5 AgNPs and AuNPs chronic toxicity
189
2015). However, the absence of effects on CAT, GPx and GR in addition to an activation of MT may suggest
an adaptative handling and/or detoxification strategies due to the low doses of AuNPs administered for a
prolonged exposure time (Pan et al., 2012; Volland et al., 2015). This hypothesis may also explain the increase
in Na+K+ATPase gene expression in G. fossarum exposed to AgNPs. As previously described, the acute toxicity
of Ag is exerted through the inhibition of Na+K+ATPase activity. A prolonged exposure may have led to an
adaptative strategy leading to the up-regulation of Na+K+ATPase in order to offset the inhibition of the
enzymatic activity. These results highlight the differences or evolution of gene expression responses observed
after an acute waterborne exposure of G. fossarum to AgNPs and AuNPs (Mehennaoui et al., subm., chapter
4) and chronic exposure in the presence of food to the same particles.
Exposing G. fossarum for 15 days to CIT and PEG-AgNPs 40 nm led to a significant decrease in the number of
G. fossarum in movement. This result is in accordance with our previous finding as a decrease in locomotor
activity was observed in G. fossarum exposed for 72h to AgNPs 23 nm (Mehennaoui et al., 2016). Similarly, a
decrease in locomotor activity of G. roeseli exposed for 72h to CIT-AgNPs 10 and 60 nm was observed (Andreï
et al., 2016). Authors reported a size-dependent effect on locomotion as CIT-AgNPs 10 nm was more effective
than CIT-AgNPs 60 nm (Andreï et al., 2016). The effects of CIT-AgNPs 10 nm was as strong as the effects
observed for AgNO3 (Andreï et al., 2016; Arce Funck et al., 2013). Therefore, it may be suggested that the
readily taken up AgNPs dissolve internally and release Ag+ ions that are more potent in inhibiting G. fossarum
locomotor activity (Arce Funck et al., 2013; Mehennaoui et al., 2016). The effects on locomotion could also
be linked to sensorial disruption linked to the adsorption of AgNPs and AuNPs on the carapace of G. fossarum
as observed in our previous study (Mehennaoui et al., subm., chapter 4) and which is consistent with the
significant effect of AgNPs on the expression of chitinase. The decrease in locomotor activity could also be a
result of an energy reallocation in favour of defence mechanisms. Indeed, an increase in the internal
concentrations of metals or NPs usually lead individuals to invest their energy in highly energy consuming
mechanisms such as detoxification, homeostasis maintenance (iono/osmoregulation) or cellular repairing
process (Arce Funck et al., 2013; Vellinger et al., 2013).
This study highlights once more the great sensitivity of behavioural responses that are fast, simple to perform,
cheap and non-invasive biological endpoints that can highlight sub-lethal effects of contaminants.
Behavioural responses allow linking responses observed at the physiological level to potential effects at the
population and community levels regarding the central position of G. fossarum in aquatic ecosystems.
Indeed, an impairment of locomotion may have consequent effects on the fitness of an organism and lead to
“ecological death” (Scott and Sloman, 2004).
Chapter 5 AgNPs and AuNPs chronic toxicity
190
5. CONCLUSIONS
The present study allowed obtaining first insights on key mechanisms of sub-chronic effects of AgNPs and
AuNPs. Gene expression and locomotion appeared good biological endpoins as they allow the detection of
sub-lethal effects of AgNPs and AuNPs at very low concentration (0.5 µg. L-1). Similarly to the acute exposure
(Chapter 4), chemical composition of AgNPs and AuNPs appear as the main factor influencing the fate and
effects of these compounds on G. fossarum. Indeed, different patterns of internal distribution were observed
as AgNPs were observed in the gills whereas AuNPs were localized within the lumen of intestinal caeca
indicating potential elimination with faeces. Gene expression results may allow an early prediction of effects
of AgNPs and AuNPs on G. fossarum. Locomotion may predict effects at the population and community level
regarding the central position of G. fossarum in aquatic ecosystems.
Acknowledgments
This work was supported by National Research Fund, Luxembourg (AFR-PhD-9229040). The authors are
thankful to S. Contal for technical support, P. Rouselle, C. Guignard and J. Ziebel for chemical analyses and A.
Chauvière for ultramicrotome sections.
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CHAPTER 6
General discussion
CHAPTER 6
General discussion
Chapter 6 General discussion
199
Amphipods are one of the most important components of freshwater ecosystems. Among them, gammarids
are the most widespread group in Europe and are often used as bioindicators of the quality of aquatic
ecosystems and model organisms in many ecotoxicological studies due to their well-known ecology and
sensitivity to different kind of pollutants (Farkas et al., 2003; Kunz et al., 2010). However, their use, especially
of Gammarus fossarum for the study of the environmental impact of nanoparticles has been rather limited
so far (Mehennaoui et al., 2016).
Engineered nanoparticles (ENPs) are being increasingly incorporated in consumer products and as a
consequence, their environmental release is raising concerns about their potential toxicity/effects on living
organisms (Klaine et al., 2008; Moreno-Garrido et al., 2015). Silver nanoparticles (AgNPs) are among the most
promising and used ENPs for application in numerous consumer products due to their well-documented
antibacterial properties (Bone et al., 2012; Cleveland et al., 2012; Vance et al., 2015). Gold nanoparticles
(AuNPs) are used in a wide range of applications in the field of biomedicine, biology and chemistry (Nel et al.,
2006; Volland et al., 2015) mostly for their low toxicity and high bio-compatibility (García-Cambero et al.,
2013; Lapresta-Fernández et al., 2012).
The research conducted within this thesis aimed at the evaluation of potential threats of AgNPs and AuNPs
and identification of their key mechanisms of toxicity on G. fossarum. The influence of size, surface coating
and chemical composition on the toxicity of AgNPs and AuNPs was assessed. An integrated approach
including molecular, enzymatic, physiological and behavioural responses, was applied in order to define
and potentially predict effects of AgNPs and AuNPs at higher biological organisation (population,
community, ecosystems) regarding the central and important position of G. fossarum in the aquatic
ecosystems.
Regarding the lack of information on the toxicity mechanisms of ENPs on G. fossarum and its scare use as a
model organism in nanotoxicology studies, the first study of this thesis (Chapter 2) aimed at evaluating the
effects of a well-characterised and well-studied set of AgNPs (Georgantzopoulou, 2015). Synthetically
produced AgNPs of different sizes (20 and 200 nm) as well as AgNPs synthetized by a biological method (using
plant leaf extracts of Ocinum sanctum and Azadirachta indica, AgNPs 23 and 27 nm, respectively;
(Balachandran et al., 2012, 2013) were used in order to elucidate the relation between size, synthesis
method, NPs surface properties, ions dissolution and their toxicity (Georgantzopoulou et al., 2013;
Mehennaoui et al., 2016). A multi-biomarker approach was applied in order to determine the effects at the
physiological and behavioural level. Biological endpoints including survival, bioaccumulation, tissue
distribution, osmoregulation, antioxidant responses, defence mechanisms, cellular damage, energy reserves
and behaviour were investigated and allowed obtaining first information about the effects of AgNPs on G.
fossarum (Figure 6. 1).
Chapter 6 General discussion
200
Figure 6. 1. Physiological and behavioural effects of synthetic and biological AgNPs on G. fossarum exposed for
72h: summary of results presented in Chapter 2. G. fossarum were exposed for 72h to synthetic AgNPs 20 and
200 nm and biologically synthetized from plant leaf extract AgNPs 23 and 27 nm. Effects on antioxidant responses
(GPx, TAC, CAT), defense mechanism (GST, ACP), cellular damage (LDH, CASP3, LOOH), energy reserves (Prot, Chol,
Trig, ETS), osmoregulation and behaviour (Locomotion and ventilation) were assessed.
Despite the rapid advances in the “omics” field marked by the development of next generation sequencing
(NGS) methods, RT-qPCR remains one of the most accurate and reliable techniques for target gene expression
and NGS data validation. This method is recognized for its capacity to highlight sensitive changes in gene
transcription levels (Guerriero et al., 2014; Plusquin et al., 2012) which represent an early warning of
pollution-induced stress and may be linked to important physiological changes (Lacroix et al., 2014).
Therefore, a set of stress related genes, selected based on the battery of biomarkers developed earlier by
our research group (Garaud et al., 2015a; 2015b, 2016), was used to investigate the toxicity/effects of
different sized AgNPs and AuNPs (20, 40 and 80 nm) with two different coatings (CIT and PEG) on G. fossarum
(Chapter 4). This study aimed, firstly, at investigating the influence of the size, chemical composition and
surface coating on the uptake, tissue distribution and molecular effects of AgNPs and AuNPs on G. fossarum
exposed for 72h to AgNPs and AuNPs, in the absence of food (Figure 6. 2). Additionally, in order to obtain
accurate results at the molecular level, a set of reference genes for RT-qPCR data normalization was
identified. To our best knowledge, this study was the first attempt of identifying a set of reference genes for
Sub-individual
Physiologicaleffects
• ↓Osmoregulation
• Antioxidantresponses
• Defensemechanisms
• Cellulardamage
• Energyreserves
Individual
Behavioraleffects
• ↓ Locomotion
• Ventilation
• ↓Survival
SyntheticAgNPs Biological AgNPs
20 200 23 27
Chapter 6 General discussion
201
the freshwater crustacean G. fossarum. Six candidate genes (actin, TUB, UB, SDH, clathrin and GAPDH) were
tested in order to determine the most stable ones in different stress conditions and to increase the
robustness of obtained RT-qPCR data (Chapter 3).
Figure 6. 2. Influence of size (20, 40 and 80 nm) and surface coating (CIT and PEG) of AgNPs and AuNPs on their
uptake, tissue distribution and molecular effects of G. fossarum exposed for 72h: summary of the results
presented in Chapter 4. G. fossarum were exposed for 72h to up to 50 µg. L-1 of AgNPs and AuNPs in absence of
food. Influence of size and surface coating on bioaccumulation was assessed using ICP-MS, internal distribution of
AgNPs and AuNPs were evaluated using NanoSIMS50 and Cytoviva, and molecular effects were assessed using RT-
qPCR. A set of stress-related genes expression including genes implied in cytoskeleton trafficking (Actin, TUB, UB),
exoskeleton cuticle (Chitinase), antioxidant defence (CAT, MnSOD, CuZnSOD, GPx7), general stress (GST, HSP90,
GAPDH), DNA damage and repair (Gadd45, NfkB, c-jun, P53), lysosomes (Cathepsin L), osmoregulation and
respiration (Na+K+ATPase, HEM) was used.
The second aim of this approach was the identification of sub-chronic effects of AgNPs and AuNPs on G.
fossarum. Therefore, a more realistic exposure scenario was applied by exposing G. fossarum for 15 days
to CIT and PEG-AgNPs 40 nm and CIT and PEG-AuNPs 40 nm in presence of food. Effects on the molecular
(gene expression), physiological (osmoregulation) and behavioural responses (locomotion and ventilation)
were investigated (Chapter 5, Figure 6. 3).
AgNPs
CIT PEG
AuNPs
CIT PEG
20
40
80
Bioaccumulation
• CIT-AuNPs>PEG-AuNPs
• Concentration
dependentuptake
• Nosizeeffect
TissueDistribution
• ↑↑Intestine
• ↑Cuticle
• Gills
Moleculareffects
• ↑GPx7
• ↓MnSOD
• ↓CuZnSOD
• ↓HSP90
• ↓Hemocyanin
Bioaccumulation
• CIT-AgNPs>PEG-AgNPs
• Concentration
dependentuptake
• Nosizeeffect
TissueDistribution
• ↑↑Gills
• ↑Cuticle
• Intestine
Moleculareffects
• ↑MnSOD
• ↑CuZnSOD
• ↑Chitinase
• ↓GST
• ↓Hemocyanin
• ↓CathepsinL
20
40
80
20
40
80
20
40
80
Chapter 6 General discussion
202
Figure 6. 3. Sub-chronic toxicity of CIT and PEG-AgNPs and CIT and PEG-AuNPs 40 nm on molecular, physiological
and behavioural responses of G. fossarum: summary of the results presented in Chapter 5. G. fossarum were
exposed for 15 days to 0.5 and 5 µg. L-1 of CIT and PEG-AgNPs and AuNPs 40 nm in presence of food.
Bioaccumulation was assessed using ICP-MS, internal distribution of AgNPs and AuNPs were evaluated using
Cytoviva and molecular effects were assessed using a set of stress-related genes expression including genes
implied in cytoskeleton trafficking (Actin, TUB, UB), exoskeleton cuticle (Chitinase), antioxidant defence (CAT,
MnSOD, CuZnSOD, GPx7), general stress (GST, HSP90, GAPDH), DNA damage and repair (Gadd45, NfkB, c-jun, P53),
lysosomes (Cathepsin L), osmoregulation and respiration (Na+K+ATPase, HEM).
AgNPs
CIT PEG
AuNPs
CIT PEG
Molecularresponses
• ↑CuZnSOD
Physiologicalresponses
• [Na+]
• [Cl-]
• [Ca2+]
Behavioralresponses
• ↓ Locomotion
• Ventilation
Molecularresponses
• ↑CAT
• ↑Chitinase
• ↑Na+K+ATPase
Physiologicalresponses
• [Na+]
• [Cl-]
• [Ca2+]
Behavioralresponses
• ↓ Locomotion
• Ventilation40 40
40 40
Chapter 6 General discussion
203
1. NPS SIZE, SURFACE COATING AND CHEMICAL COMPOSITION INFLUENCE THE
UPTAKE
As reported in many different studies, NPs size, surface coating and chemical composition appeared as critical
factors controlling the uptake and effects of AgNPs and AuNPs (Baumann et al., 2014; Blinova et al., 2013;
Ivask et al., 2014; Silva et al., 2014; Chapter 2 and 4). A size-dependent toxicity was observed on G. fossarum
exposed for 72h to AgNPs 20, 23, 27 and 200 nm with AgNPs 23 nm, as the one that aggregated the less
(smallest hydrodynamic diameter of 53 nm), being the most toxic one with an LC50 of at least 2.3 µg. L-1
(Figure 6. 1). The order of toxicity was the same as reported in a previous study using V. fisheri, D. subspicatus
and D. magna as model organisms (Georgantzopoulou et al., 2013). Size-dependent effects for the bare
AgNPs 20 and 200 nm with AgNPs 20 nm being more potent to D. magna than AgNPs 200 nm were reported.
Moreover, biologically synthetized AgNPs 23 and 27 nm were always more toxic to D. magna with AgNPs 23
nm being the most toxic one (Georgantzopoulou et al., 2013). Similarly, in addition to dose-dependent
effects of AgNPs on a multi-trophic test battery including algae and D. magna, size-dependent effects were
observed with the smallest AgNPs (10 nm) being more toxic to D. magna than the bigger ones (20, 40, 60 and
80 nm) (Ivask et al., 2014).
Interestingly, using commercial AgNPs, stabilized with citrate or coated with PEG, did not induce size-
dependent effects. Indeed, short-term exposure of G. fossarum to three different sizes of CIT and PEG-AgNPs,
namely 20, 40 and 80 nm, led to a concentration-dependent rather than size-dependent Ag uptake (Chapter
4, Figure 6. 2). The same results were obtained for AuNPs with a concentration-dependent rather than a size-
dependent Au uptake. A surface coating-dependent bioaccumulation of Ag and Au was observed with CIT-
AgNPs and CIT-AuNPs being more taken up by G. fossarum than PEG-AgNPs and PEG-AuNPs. Our results were
consistent with previous studies that reported a coating dependent bioaccumulation of CIT-AgNPs and CIT-
AuNPs in aquatic organisms (Oliver et al., 2014; Park et al., 2015; Sakka et al., 2016). These differences may
be linked to the behaviour of CIT-NPs in the water column. As they appeared to aggregate more than PEG-
NPs, it could be expected that CIT-NPs settle down and reach the water/sediment interface where G.
fossarum tend to be found. Therefore, it could be expected that CIT-NPs may be more bioavailable for benthic
macro-invertebrates than PEG-NPs. It has been reported that G. roeseli accumulated more cerium (Ce) than
Dreissena polymorpha exposed to CeO-NPs (Garaud et al., 2015). The higher bioaccumulation could be linked
to the ingestion of aggregate that may settle in the aquarium or via the ingestion of faeces and pseudo-faeces
of contaminated bivalves by gammarids (Garaud et al., 2015). However, in chapter 2, we showed that the
bare AgNPs 20 and 200 nm were less absorbed by G. fossarum than biological AgNPs 23 and 27 nm despite a
higher aggregation (Mehennaoui et al., 2016). This result enhances the key role played by surface coating of
AgNPs and AuNPs bioaccumulation by G. fossarum. Exposing G. fossarum to CIT and PEG-AgNPs 40 nm and
Chapter 6 General discussion
204
CIT and PEG-AuNPs 40 nm for 15 days and in the presence of food confirmed the role of surface coating on
the bioaccumulation of Ag and Au (Chapter 5, Figure 6. 3). Indeed, G. fossarum exposed to the highest
concentration of CIT-AgNPs and CIT-AuNPs 40 nm (5 µg. L-1) tended to accumulate more Ag and Au than
animals exposed to PEG-AgNPs and PEG-AuNPs.
One important step to understand NPs’ fate, behaviour and adverse effects is by investigating their internal
distribution in tissues. Usually, transmission electron microscopy (TEM) is used for internal localization of NPs
(Heinlaan et al., 2008; Lovern et al., 2008). However, this method does not allow a direct elemental
identification of NPs leading to ambiguous results (Georgantzopoulou et al., 2013). In the beginning of the
present thesis, we have used one of the latest generation of Secondary Ion Mass Spectrometry instruments,
the NanoSIMS 50. This probe provides information about the elemental distribution of NPs within biological
materials. However, it does not allow the distinction between ions and NPs. Later during this work, a dark-
field hyperspectral microscope, Cytoviva, was acquired, which allows detection of only the NPs in biological
tissues but not of the ions. In Chapter 2, NanoSIMS allowed localization of Ag from AgNPs 23 and 27 nm in
the gills of G. fossarum. The other organs could not be analysed due to sample preparation limits. Thus, the
protocol of sample preparation for NanoSIMS 50 analyses was optimized and different tissues of G. fossarum
were analysed, namely gills, intestines, muscle and external cuticle (Chapter 4, Figure 6. 2), when Gammarus
were exposed to CIT and PEG AgNPs and AuNPs. NanoSIMS50 and Cytoviva allowed the confirmation of
bioaccumulation results from ICP-MS as the signal of Ag and Au obtained in G. fossarum exposed to CIT-
AgNPs and CIT-AuNPs was higher than PEG-AgNPs and PEG-AuNPs.
Interestingly, the chemical composition of the studied NPs appeared also as a critical factor determining their
internal distribution. Indeed, different uptake routes were observed (Chapter 2, 4 and 5). AgNPs were found
in the gills (cuticle and epithelial tissue) while AuNPs were observed in the epithelial tissue of intestinal caeca
(Chapter 4, Figure 6. 2). The sub-chronic exposure of G. fossarum to CIT and PEG-AgNPs and CIT and PEG-
AgNPs 40 nm in the presence of food confirmed the different uptake routes as AgNPs were observed in the
gills while AuNPs were observed in the gut area (Figure 6. 3). However, the presence of the food led to the
localization of AuNPs within the lumen on intestinal caeca indicating a potential elimination with faeces while
in the absence of the food, it seems that AuNPs were able to cross the gut barrier.
Chapter 6 General discussion
205
2. SYNTHESIS METHODS AND CHEMICAL COMPOSITION INFLUENCE AGNPS AND
AUNPS EFFECTS
The NPs synthesis method proved to be an important parameter in determining their toxicity. The biological
AgNPs 23 and 27 nm synthesised using two different plant leaf extracts, A. indica (AgNPs 27 nm) and O.
sanctum (AgNPs 23 nm) were the most potent ones (Chapter 2, Figure 6. 1). Using two different plant leaf
extracts led to the presence of different biomolecules on the surface of AgNPs 23 and 27 nm as one additional
molecule was observed on the surface of AgNPs 27 which may confer it more stability compared to AgNPs 23
nm (Georgantzopoulou et al., 2013; Mehennaoui et al., 2016). These observations confirm once more the
importance of surface coating on bioaccumulation potential of AgNPs. Even though both AgNPs 23 and 27
nm led to a significant mortality of G. fossarum, the sub-lethal effects were different. Only AgNPs 23 nm led
to a higher mortality rate, osmoregulation impairment and a decrease in locomotion (Mehennaoui et al.,
2016). In the same way, the chemical composition played an important role on the sub-individual effects of
AgNPs and AuNPs on G. fossarum (Chapter 4 and 5). The acute exposure of G. fossarum to AgNPs led to an
up-regulation of MnSOD, CuZnSOD and Chitinase, and a down-regulation of GST, HEM and Cathepsin L gene
expression whereas AuNPs led to an up-regulation of GPx 7, and a decrease in MnSOD, CuZnSOD, HSP90 and
HEM gene expression (Chapter 4, Figure 6. 2). It is important to note that these effects were independent
from AgNPs and AuNPs size and surface coatings.
Gene expression appear as a very sensitive biomarker level. Although the effects at the protein and enzyme
activity level were not performed in the experiments described in Chapter 4 and 5, the observed effects at
the gene level allow to highlight that AgNPs and AuNPs induced different responses. The analysis of the gene
expression in G. fossarum also revealed the importance of the exposure route. For instance, acute exposure
of G. fossarum to AgNPs led to the induction of SODs while AuNPs led to their down-regulation. The chronic
exposure led to the up-regulation of Na+K+ATPase by AgNPs whereas no effects were observed on this gene
when G. fossarum were exposed to AuNPs. However, it can be suggested that both AgNPs and AuNPs might
activate antioxidant defense, which is coherent with literature as oxidative stress via ROS generation was
reported as the main mechanism of action of NPs (Vale et al., 2016).
Chapter 6 General discussion
206
3. CONTRIBUTION OF SOLUBLE IONS TO THE EFFECTS
In the present work, AgNO3 was used as a positive control in order to compare AgNPs effects to Ag ions effects
while ion release from AuNPs was not expected. Ions release from AgNPs is size, surface coating and exposure
media dependent (Georgantzopoulou, 2015). In the present thesis, significant ions release was observed only
with biologically synthetized AgNPs 23 nm (Chapter 2) with a dissolution rate of up to 29% in Volvic water
(Mehennaoui et al., 2016). Bare AgNPs 20 and 200 nm, AgNPs 27 nm, CIT and PEG-AgNPs and CIT and PEG-
AuNPs showed a dissolution rate of maximum 6% (24h) with very low concentration of Ag and Au ions
(chapter 2 and 4). According to the results obtained in Chapter 2 and 4, it seems that surface coating is the
main factor influencing ions release as bare, CIT and PEG-AgNPs in addition to AgNPs 27 nm appeared to be
more stable in the exposure medium (Volvic water) than AgNPs 23 nm.
Interestingly, AgNPs 23 nm were more toxic to G. fossarum than AgNPs 20 and 27 nm. Indeed, the EC50s
causing effects were similar for AgNPs 23 nm and AgNO3. Both of them led to a decrease in survival rates and
locomotion in addition to an osmoregulation impairment at similar concentrations (Mehennaoui et al., 2016).
These results highlight the implication of Ag ions in the potency of AgNPs to G. fossarum. However, ion
release seems to be not the only mechanism of effects of AgNPs on G. fossarum. Indeed, gene expression
results showed that AgNPs affect G. fossarum in a different way compared to AgNO3 (chapter 4 and 5)
although common mechanisms were reported. For instance, both AgNPs and AgNO3 led to a down-regulation
of hemocyanin and an up-regulation of CuZnSOD and Chitinase whereas only AgNPs increased MnSOD gene
expression and decreased GST, Cathepsin L and AgNO3 led to an up-regulation of TUB, GST and Cathepsin L
and to a down-regulation of NfkB. These results are in accordance with results observed on the MXR
responses of in vitro gastrointestinal co-culture model exposed to AgNPs (Georgantzopoulou et al., 2016).
AgNPs 23 nm were more potent than AgNO3 and AgNPs 27 nm while the level of dissolved Ag in the media
were comparable for AgNO3 and AgNPs 23 nm (Georgantzopoulou et al., 2016). This suggest that the
differences in the observed effects are not only linked to Ag ions release as AgNPs appeared to be sometime
more toxic than AgNO3 at the same nominal exposure concentrations (García-Alonso et al., 2014; Ivask et al.,
2014). Moreover, different gene expression profiles were observed in D. magna exposed to AgNO3 and AgNPs
(Poynton et al., 2012) which is in support of our findings for the gene expression study (Chapter 4).
Chapter 6 General discussion
207
4. LIMITS OF THE MULTI-BIOMARKERS APPROACH
The application of a multi-biomarkers approach allowed the assessment of effects of AgNPs and AuNPs on
different biological levels (from molecular to behavioural responses). Biomarkers need to be sensitive
indicators and are defined as early warning of pollution-induced stress on ecosystems (Lagadic, 2002; Vasseur
and Leguille, 2004). The results obtained in the present work allowed to confirm the sensitivity of sub-
individual and individual responses (behaviour) as effects were observed at very low and environmentally
realistic concentrations of AgNPs and AuNPs (from 0.5 µg. L-1). Behavioural responses allow obtaining rapid
and non-invasive responses even though no effects on ventilation could be detected in the present work
(Chapter 2 and 5). However, locomotion was one of the most sensitive responses as effects were detected at
very low concentrations of AgNPs and AuNPs (0.5 µg. L-1). The effects on locomotion may suggest an energy
reallocation strategy in favour of defense mechanisms in G. fossarum. Our results were in accordance with
previous studies performed within the ANR MESONNET project by exposing G. roeseli to different sizes of
CIT-AgNPs (Andreï et al., 2016). Effects on locomotion may also allow predicting the effects of AgNPs and
AuNPs at the population and community level. Indeed, locomotion is a highly ecologically relevant biomarker
as it is necessary for foraging, mate finding and escaping from predators. By interfering with locomotion
AgNPs and AuNPs may reduce the fitness of the organism and have a potential effect at the population level.
The use of numerous stress-related genes allowed to highlight the influence of chemical composition on the
effects of AgNPs and AuNPs on G. fossarum. However, the obtained results should be considered with
precaution as one of the difficulties of the target genes approach is the important number of potential
isoforms present encoding for each protein family. As no functional studies (proteomic analyses, enzymatic
analyses; Chapter 4 and 5) were performed, no clear conclusions can be formulated as an induction of a given
gene does not systematically result in the expression/activation at the protein level. Hence, further analyses
are still needed in order to better define and predict the effects of AgNPs and AuNPs on G. fossarum and to
better link responses observed at the tissue level to those observed at the whole individual level.
Furthermore, the responses studied in chapter 2 seem not to be the most appropriate endpoints for studying
the effects of NPs on G. fossarum. The absence of significant effects may be due to exposure time and the
chemical composition of the studied materials. Indeed, it has been reported that antioxidant responses were
significantly impacted in goldfish after 6h of exposure to a mixture of pesticides while no significant effects
on antioxidant responses were observed after 96h or 16 days of exposure (Gandar et al., 2017). This may
suggest the setting up of different sampling times in order to unsure detection of early effects on G. fossarum.
Chapter 6 General discussion
208
5. CONCLUDING REMARKS AND FUTURE PERSPECTIVES
The present work provides a substantial amount of data which open new windows for further exploratory
work. The results of this thesis highlight the importance of considering on one hand the NPs characteristics
such as size, methods of synthesis, surface coating and chemical composition and on the other hand, the
exposure mode of the studied AgNPs and AuNPs regarding their significant influence on the behaviour of NPs
in the exposure medium and uptake routes.
It can be concluded that the chemical composition is the main factor influencing the effects of NPs on aquatic
organisms. It is important to keep in mind that for a specific NP, physico-chemical characteristics should be
considered as these parameters play a crucial role in modulating the behaviour of NPs in the environment
and their interactions with organisms. The results of the present thesis show that both the NPs and the
released ions play a role for the toxicity to G. fossarum. It seems that surface coating and chemical
composition, and to a lower extent size, are the main factors favouring uptake, strong interaction with
biological sites and thus relative toxicity of AgNPs and AuNPs.
The knowledge acquired in the present work through the exposure of G. fossarum to realistic concentrations
of NPs contributed to a better understanding of mechanisms of action of AgNPs and AuNPs. However,
establishment of more realistic exposure scenarios including river water and organic matter should provide
complementary information with a biological continuum linking molecular to functional responses as these
latter appeared as sensitive responses (Andreï et al., 2016). Linking molecular to individual and functional
responses will allow a better prediction of the effects of NPs at the population and community level. With
the acquirement of new sequences of G. fossarum, it might be valuable to apply NGS methods like RNAseq
(Illumina) in order to identify new biomarkers in nano-ecotoxicology. This highly sensitive method might
allow a better coverage and determination of mechanisms of action of NPs. The use of SIMS in combination
with Cytoviva allows localisation of AgNPs and AuNPs within G. fossarum. However, these techniques do not
allow a determination of effects of the studied NPs on the studied tissues. Hence, the use of histopathological
methods might also bring complementary data to better link responses at the molecular level to those
observed at the individual level.
Chapter 6 General discussion
209
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Aquatic Toxicology 170 (2016) 162–174
Contents lists available at ScienceDirect
Aquatic Toxicology
j o ur na l ho me pag e: www.elsev ier .com/ locate /aquatox
Manufactured nanoparticles in the aquatic environment-biochemicalresponses on freshwater organisms: A critical overview
Gonc alo Vale a,b,∗, Kahina Mehennaoui c, Sebastien Cambier c, Giovanni Libralatod,Stéphane Jomini e, Rute F. Domingos a,f
a Centro de Química Estrutural, Instituto Superior Técnico, Universidade de Lisboa, Torre Sul Av. Rovisco Pais, 1049-001 Lisboa, Portugalb Department of Molecular Genetics, University of Texas Southwestern Medical Center, Harry Dallas, TX 75390, USAc Luxembourg Institute of Science and Technology, Environmental Research and Innovation (ERIN) Department, Belvaux, Luxembourgd Department of Environmental Sciences, Informatics and Statistics, University Ca’ Foscari Venice, Via Torino 155, 30172, Mestre, Venice, Italye Laboratoire Interdisciplinaire des Environements Continentaux (LIEC), Université de Lorraine, UMR 7360, Campus Bridoux rue du Général Delestraint,
57070 Metz, Francef Institut de Physique du Globe de Paris, Sorbonne Paris Cité, UMR CNRS 7154, Université Paris Diderot, 75205 Paris Cedex 05, France
a r t i c l e i n f o
Article history:
Received 26 August 2015
Received in revised form
16 November 2015
Accepted 19 November 2015
Available online 23 November 2015
Keywords:
Nanoparticles
Transformations
Freshwater organisms
Nanotoxicity
Reactive oxygen species (ROS)
Anti-oxidant enzymes
a b s t r a c t
The enormous investments in nanotechnology have led to an exponential increase of new manufactured
nano-enabled materials whose impact in the aquatic systems is still largely unknown. Ecotoxicity and
nanosafety studies mostly resulted in contradictory results and generally failed to clearly identify biologi-
cal patterns that could be related specifically to nanotoxicity. Generation of reactive oxygen species (ROS)
is one of the most discussed nanotoxicity mechanism in literature. ROS can induce oxidative stress (OS),
resulting in cyto- and genotoxicity. The ROS overproduction can trigger the induction of anti-oxidant
enzymes such as catalase (CAT), superoxide dismutase (SOD) and glutathione peroxidases (GPx), which
are used as biomarkers of response. A critical overview of the biochemical responses induced by the pres-
ence of NPs on freshwater organisms is performed with a strong interest on indicators of ROS and general
stress. A special focus will be given to the NPs transformations, including aggregation, and dissolution,
in the exposure media and the produced biochemical endpoints.
© 2015 Elsevier B.V. All rights reserved.
Contents
1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 163
2. NPs transformations in aquatic systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 163
3. Nanotoxicity toward aquatic organisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 164
3.1. Generation of ROS. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .164
3.2. Omics endpoints . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 165
4. NPs toxicity on freshwater organisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 165
4.1. Silver NPs (nAg) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 165
4.2. Titanium dioxide NPs (nTiO2) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 166
4.3. Zinc oxide NPs (nZnO) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 167
4.4. Copper oxide NPs (nCuO) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 170
5. General remarks and conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 172
Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 172
References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 172
∗ Corresponding author.
E-mail addresses: [email protected] (G. Vale), [email protected] (K. Mehennaoui), [email protected] (S. Cambier), [email protected]
(G. Libralato), [email protected] (S. Jomini), [email protected] (R.F. Domingos).
http://dx.doi.org/10.1016/j.aquatox.2015.11.019
0166-445X/© 2015 Elsevier B.V. All rights reserved.
G. Vale et al. / Aquatic Toxicology 170 (2016) 162–174 163
1. Introduction
Nanotechnology has emerged as a fast growing sector impactingkey economical fields and providing new engineered nano-enabledproducts, constituted by nanoparticles (NPs), with novel andunique functions that reach the market every day (Bour et al.,2015). NPs are defined as materials with a size between 1 and100 nm on at least one dimension, having unique physicochemi-cal properties differing from their bulk forms due to their greatersurface area to volume ratio. This size related-properties results inlarger reactivity and higher mobility (Rauscher et al., 2014), lead-ing to numerous applications in medical diagnostics, electronics,computers, cosmetics and environmental remediation. The world-wide consumption of NPs is expected to grow from 225,060 metrictons in 2014 to nearly 584,984 metric tons in 2019 represent-ing an annual growth rate of 21.1% (Research, 2015). Althoughimpressive, these numbers are in fact “expected” values obtainedby estimation or modeling. The lack of legislation for nanotech-nologies gives the manufacturers no onus to reveal the real figures,thus, indeed, these predicted values are most probably signifi-cantly higher. The absence of real numbers hinders the predictionof the NPs amount that are actually being released into the envi-ronment (Piccinno et al., 2012). Even though several studies havebeen performed with the goal of modeling NPs environmentalconcentrations (Gottschalk et al., 2013), they should only be con-sidered as guidelines, since they derive from uncertain data aboutthe NPs production (often obtained by surveys to the producers)and extrapolations used to scale up regional to worldwide amounts(Piccinno et al., 2012; Gottschalk et al., 2011; Keller et al., 2013).
When released in natural media NPs will be subjected to adynamic physical and chemical environment that consequentlyresults in different and unknown endpoints far from their pris-tine or as released state. Therefore, environments and humans arenot facing pristine manufactured NPs but rather transformed nano-enabled products, which is factually accepted but so far neglected.In fact, the large majority of the physicochemical and toxicity dataobtained so far was focused on simple nanoscale particles andnot on relevant nano-enabled products. This includes not onlythe NP embedded in the manufactured matrix but also the mate-rials resulting from the interaction with biotic and abiotic (bio)molecules composing the natural systems. To further complicatethe interpretation of the NPs studies, there are two distinct mecha-nisms that should be considered but are not easily differentiated: (i)chemical toxicity by the release of possible ions and/or formationof reactive oxygen species (ROS) (Fu et al., 2014), and (ii) physicalstress or stimuli caused by NPs size, shape and surface properties(Vale et al., 2014; Libralato et al., 2013). These materials are gener-ally associated with cellular perturbations such as ROS generation,and gene expression and proteome profiles alterations. For thesereasons, the NPs escalating production and applications has raisedconcerns about their environmental and human safety, which haveled to large investments in nanosafety-related projects resultingin a considerable amount of data assessing their potential hazard(Savolainen et al., 2013). However, the establishment of relation-ships between bioavailable NP-containing species and the specificbioadverse or biocompatible endpoints is still lacking, mainly since,the effects are NP-dependent and also specie-dependent (Buricet al., 2015).
This work provides an overview of the latest studies on theimpact of NPs onto freshwater ecosystems, considered by many asthe ultimate sink of these particles, with a special focus on (i) NPstransformations and characterization in the different test media,and (ii) toxicological effects such as generation of ROS, genotoxic-ity, metallomic and proteomic changes. This survey is focused onmetallic NPs including nAg, nTiO2, nZnO and nCuO, mostly due tothe great number of studies dedicated to these particles.
2. NPs transformations in aquatic systems
NPs can enter in an aquatic compartment from (i) wastewa-ter treatment plants effluents, (ii) direct use (e.g., application ofNPs-containing paintings on boats), and (iii) deposition from theair compartment. When entering aquatic compartment, NPs willbe exposed to a highly dynamic physical and chemical environ-ment that leads to several transformations that will change theirpristine or as released physicochemical properties (Fig. 1). Thesetransformations, including dissolution, aggregation and sedimen-tation, are dependent on both physicochemical properties of the NP(and nanoforms thereof) and those of the environment into whichthey were released.
Colloidal particles, including organic and inorganic matter,are ubiquitous in the aquatic environment and can be origi-nated from both natural and anthropic sources. These colloidscan strongly interact with NPs, thereby determining their formsover space and time (dynamic speciation), and greatly affectingtheir bioavailability. Thus, the NP will have a specific speciationin each environmental compartment, and this speciation is alwaysdynamic with reaction rates that depend upon the chemical natureand physical sizes of the engineered and natural colloids. Althoughit is clear that dynamic speciation must be considered in order tomake relevant predictions of NPs fate, toxicity and risk, until nowthis critical issue, was mostly neglected (see detailed explanationon Pinheiro and Domingos, 2015).
Dissolution, which is one of the main transformations of metallicNPs such as nZnO, nCuO, and nAg, is mainly due to (i) the formationof partially soluble metal-oxide (Heinlaan et al., 2008; Domingoset al., 2013a; Wang et al., 2015), (ii) the oxidation of the parti-cle constituents (Ma et al., 2014; Wang et al., 2013a; Dale et al.,2013; Lok et al., 2007; Derfus et al., 2004), and (iii) the complexa-tion of the particle constituent metal by complexants present inthe environmental compartment or even in the NPs embeddedmatrix (including the manufactured stabilizers) (Domingos et al.,2013b; Domingos et al., 2014). The sulfidation of the metallic NPscan retard their oxidation and, thus, their dissolution (Ma et al.,2014; Wang et al., 2013a; Dale et al., 2013; Thalmann et al., 2014).This dissolution mechanism results in the release of toxic cations,such that their persistence is reduced but the toxicity is increased.Evidently, complete dissolution of the NPs allows the prediction oftheir impact using already existing models for metal speciation andtoxicity.
Photoreactions can also be important transformations affectingthe NPs coatings, oxidation state, generation of ROS, and persis-tence, which is the case of the innately photoactive TiO2 and ZnOparticles (Hund-Rinke and Simon, 2006; Zhang et al., 2007a).
Aggregation is other critical transformation, which mainly byinteraction with naturally occurring bio- or geomacromoleculesaffect NPs size and surface chemistry. For example, organic mat-ter (OM) provides both charge and steric stabilization (Mohammedet al., 2008; Domingos et al., 2009a) of the NPs, although they mayalso result in bridging flocculation when in presence of multiplecharged cations and anions (Domingos et al., 2010). OM effects arecomplex and difficult to predict, however, is of extremely impor-tance to explore these interactions since, OM concentrations aretypically orders of magnitude higher in concentration than engi-neered NPs, and, thus, likely to substantially modify their propertiesand behaviors. Despite the significance of these interactions, withboth organic and inorganic matter, and to the best of our knowl-edge, no relevant toxicity studies are available.
Dissolution and aggregation are dynamic processes that candecrease the NPs available surface area, thereby decreasing theirreactivity. However, this decrease is dependent on the surface prop-erties, particle number, size distribution, and the fractal dimensionsof the aggregate (Hotze et al., 2010). The NP size will affect its
164 G. Vale et al. / Aquatic Toxicology 170 (2016) 162–174
Fig. 1. Representative chemical and physical transformations of NPs when entering in natural aquatic systems: dissolution, phosphatization, sulfidation, homo- and heteroag-
gregation, and sedimentation. Important constituents with which NPs can interact governing their fate and transport includes hardness cations (e.g., Ca2+ , Mg2+), alkalinity,
phosphate and sulfide anions, pH, dissolved organic carbon (DOC), organic matter (OM) and mineral surfaces (such as iron and manganese oxides, and clays). Legend: blue
circles: engineered NPs; yellow circles: humic substances (HM); brown circles: natural inorganic colloids; blue lines: rigid biopolymers; gray surroundings: representing
sulfidation; Mz+: free metal ion. Adapted from Domingos et al. (2015). For interpretation of the references to color in this figure legend, the reader is referred to the web
version of this article.
bioavailability to the organisms; when aggregates become too largefor direct transport across the cell wall and/or membrane, uptakemay be prevented, whereas partial dissolution, which will lead tosmaller sizes, would facilitate this cellular transport. Since thesetransformations are most often not in equilibrium, they requirereal-time kinetic measurements, limiting the methodology to beused: (i) the storage of whole unfractionated samples for ion anal-ysis may not be possible since the dissolution rate may be fastor not attain the equilibrium during the experimental time, and(ii) the aggregation rate can be fast or the aggregates size distri-bution may not reach equilibrium within the experimental timewindow.
Despite the large number of studies focused on nanotoxicol-ogy, most of them disregard the particles kinetic physicochemicalcharacterization under the exposure conditions, hindering theestablishment of crucial predictive structure-activity relationshipsthat can be used afterwards in the categorization and function forrisk assessment studies. In absence of these realistic studies, twoless constrain conditions were used to select the literature studiesthat will be discussed in this critical overview:
i) studies where the NPs characterization was performed in thesame media as the bioassays.
ii) quantification of the NPs dissolution for studies using metallicNPs with propensity for dissolution such as nAg and nZnO.
3. Nanotoxicity toward aquatic organisms
3.1. Generation of ROS
Despite the large number of studies on NPs toxicity both in cellline systems and organisms, a complete understanding about themechanisms behind is still lacking (see reviews (Manke et al., 2013;Fu et al., 2014; Bour et al., 2015; Schultz et al., 2014)). ROS genera-tion, whose overproduction can lead to oxidative stress (OS) in theorganism tissues, is unquestionably the most studied nanotoxicitymechanism.
Molecular oxygen is used as an oxidizing agent for the pro-duction of adenosine triphosphate (ATP) in the organism cells,being afterwards reduced to water. The non reduced oxygen resultsin the formation of superoxides
(
O−
2•)
that can be further con-verted to hydroxyl radicals (HO•), which has the highest reductionpotential of all the physiological relevant ROS. When under con-trol, these species are easily scavenged by (i) antioxidant agentssuch as polyphenols (Fu et al., 2014; Lipinski, 2011) (e.g., elim-ination of HO•), and (ii) enzymes such as superoxide dismutase(SOD), catalase (CAT) and glutathione peroxidase (GPx). The SODenzymes catalyze the dismutation of O−
2• into oxygen or hydro-
gen peroxide (H2O2), which is decomposed by CAT into water andoxygen. Even though H2O2 is less reactive than the radical species,is still a strong oxidant that needs further elimination. The GPx,also plays a role in the detoxification of H2O2 by using glutathione
G. Vale et al. / Aquatic Toxicology 170 (2016) 162–174 165
(GSH) as a reductant. During the process, GSH is oxidized and con-verted to glutathione disulfide (GSSG) being latter reduced back toGSH by glutathione s-transferase enzymes (GST), thus completingthe cycle (Brigelius-Flohe and Maiorino, 2013; Deponte, 2013). Theantioxidant-enzymes activity is considered a reflection of the redoxstate of the cells and is frequently studied as a biomarker of OS.
When ROS is overproduced beyond the organism antioxidantresponse capacity, it leads to several deleterious effects on the cellscomponents such as lipids, proteins and DNA, possibly resultingin lipid peroxidation, apoptosis and/or cancer initiation processes,respectively. The production of ROS can be enhanced by the pres-ence of NPs, depending mainly on their size, aggregation, solubilityand coating. It is commonly accepted that smaller particles can eas-ily penetrate cell membranes, and thus induce cytotoxicity (Sakaiet al., 2011; Buzea et al., 2007). However, this correlation betweensize and toxicity is still controversial. Shi et al. (2013) reportedthat 5–10 nm nAg had higher toxicity to Tetrahymena pyriformis
when compared to slightly larger nAg (15–25 nm), while no size-dependent response on Danio rerio was obtained when exposedto 20, 50 and 110 nm (Bowman et al., 2012). Most studies showthat toxicity increases with decreasing particle size. However, oth-ers reported that either the size has no role on toxicity or thatsmaller NPs are less toxic (Ivask et al., 2014). A consensus aboutthe size effect is still lacking, and, most probably, will be unlikelyto be attained since the effects seem to be NP- and even specie-dependent (Buric et al., 2015).
The dissolution of metallic NPs such as nCuO and nAg resultsin the release of Cu and Ag ions, which are known to catalyzeFenton, Fenton-like and Haber–Weiss reactions, leading to the for-mation of ROS (Fu et al., 2014; Lipinski, 2011; He et al., 2012; Wanget al., 2013b,c). Moreover, the highly reactive surface of NPs andthe presence of manufactured and/or natural coatings can lead tothe adsorption/complexation of trace metals present in the envi-ronmental compartment also acting as a catalyzer platform to theabove-mentioned reactions, and thus increasing the concentrationof ROS in the system. Photoactive NPs such as nTiO2 and nZnO,can also induce the formation of ROS. When exposed to visible orUV light, these NPs can be photo excited resulting in the forma-tion of electron–holes, which are powerful oxidants that can reactwith surface bounded molecules forming radicals (Clemente et al.,2014). All these processes are schematized in Fig. 2.
3.2. Omics endpoints
The omics tools, such as toxicogenomic, metallomic and pro-teomic, are very useful on the establishment of toxic endpoints.A toxicogenomics approach allows the identification of gene andprotein activities in the organisms cells induced when in the pres-ence of a certain xenobiotic. A central assumption is that chemicalsgenerating toxicity by the same mechanism will produce similargene expression responses under a given set of conditions, bring-ing new insights about their mode of action that can be linked totheir specific physicochemical properties. A metallomic and pro-teomic approach will allow a complete analysis on the metal andmetalloid species composition within a cell or tissue and the estab-lishment of metalloproteins profiles leading to the identification ofnew biomarkers (e.g., proteins expressed by the NP itself (Shepardet al., 2000)). This will allow a better understanding and profil-ing of NPs toxic mechanisms discriminating them from their bulkcomponents.
Despite the large number of studies about nanotoxicology, onlya few number have reported the use of omics tools to evaluatethe NPs toxic effects at the molecular level on freshwater organ-isms (Rainville et al., 2014; Kuznetsova et al., 2014). Indeed, the“omics” approach generates a huge amount of data whose inter-pretation is not always straightforward, being, most probably, the
main reason moving nanotoxicologists away from using these tools.This large amount of data together with the fact that usually anappropriate physicochemical characterization prior to the biolog-ical assays is not performed, results in an escalating number ofunknown variables impeding a comprehensible understanding ofthe biochemical responses.
4. NPs toxicity on freshwater organisms
4.1. Silver NPs (nAg)
nAg are known for their antifungal and antimicrobial proper-ties, being extensively used in several products such as clothing,cosmetics, medical devices, paints, humidifiers and refrigerators(Reidy et al., 2013; Fabrega et al., 2011), leading to predicted con-centrations of 0.088–2.16 ng L−1 in European and North Americansurface waters (Gottschalk et al., 2009).
Despite nAg is one of the most studied NPs, the toxicity mech-anisms are still not fully clear; some assign the nAg toxicity to therelease of Ag+ ions to the media, while others assume that intactnAg particles are responsible for the induction of toxic responsesin the organisms (Schultz et al., 2014). It is well known that Ag+
ions has a great propensity to bioaccumulate in the tissues leadingto ROS generation, genotoxicity and inhibition of Na+/K+-ATPaseactivity by blocking the Na+ uptake by the cells (Luoma, 2008;Morgan et al., 1997). Similar toxic effects were also observed injuvenile Oncorhynchus mykiss fishes exposed to citrate coated nAg,with claimed low propensity for dissolution (Schultz et al., 2012).However, it was not possible to confidently establish if these inhi-bitions were caused by the nAg per se or by dissolved Ag+. In fact,it is very difficult to distinguish between the toxic effects inducedby particulate or ionic Ag, and, thus, is crucial not only to evaluatethe size distribution but also to quantify the nAg dissolution in thebioassay (Table 1).
It was possible to identify two distinct routes that could inducebiochemical responses in freshwater organisms (Table 1) (1) pres-ence of the NPs per se, and (2) presence of both nAg and dissolvedAg+. For the first route the identification of responses that areuniquely due to the presence of nAg and do not occur in matchedAg+ exposures is crucial. This is possible by (i) using nAg with lowdissolution rates, so, that the leached Ag+ in the media is insufficientto induce toxicity to the organisms, or (ii) identifying endpointsspecific to nAg, such as internalization of NPs, cytotoxicity andgenotoxicity. The work performed by Kumar et al. (2014) is anexample on the application of these strategies. They reported a sig-nificant increase of ROS and SOD and morphological alterations onfreshwater bacteria exposed to nAg coated with PVP. Since the NPsdissolution was very low (leached Ag+ < 1 �g L−1), it was concludedthat the results were related to the silver particulate form. Mor-phological changes and alteration in genes profiles, related to thepresence of nAg itself, were also observed in Cyprinus carpio (Leeet al., 2012), zebrafish (Griffitt et al., 2009; Choi et al., 2010), Daph-
nia magna (Poynton et al., 2012), medaka fish (Pham et al., 2012)and rainbow trout (Gagne et al., 2012). The exposure of 7 day oldD. magna to citrate-coated nAg (size 10 nm, 30 �g L−1) and AgNO3
(2.5 �g L−1) during 24 h showed that both Ag forms act throughdifferent pathways; both forms increased proteins thiol content,but only particles increased proteins carbonyl levels (Gundel et al.,2007).
For the second route, both forms, particulate and ionic Ag, caninduce OS and genotoxicity being the distinction between theseeffects a truly challenge. For example, in Chlorella vulgaris a positivecorrelation between ROS production and LPO on the tissues after24 h of exposure to uncoated nAg (1 and 10 mg L−1) (Oukarroumet al., 2012) was found, but with no possibility to establish a positive
166 G. Vale et al. / Aquatic Toxicology 170 (2016) 162–174
Fig. 2. Potential routes for the generation of ROS due to the presence of NPs. (1) Internalization of NPs–ROS generation could occur due to the NPs dissolution inside the
cells and/or due to the NPs photocatalytic activity. (2) Dissolution of the NPs leads to an increase concentration of metal ions in the media; some of these metals can also be
uptake by the organisms. (3) NPs and/or their surrounding coatings can adsorb/complex other metals present in the media, being taken up by the cells. (4) Photocatalytic
activity of the NPs in the presence of UV and/or natural light.
correlation with solely the particulate or ionic Ag. Similar difficul-ties were also observed in the snail Lymnaea luteola (Ali et al., 2014),where the exposure to nAg resulted in DNA damage and inductionof OS but without the possibility to identify which Ag forms has thedominant role on the observed effects. In these situations, a com-plete physicochemical characterization of the NP in the exposuremedia along with the use of controls containing ionic Ag are cru-cial to understand which Ag form (or both) is responsible for theobserved biochemical responses.
Although in the majority of the studies is not possible to estab-lish if the nAg potential toxicity is due to its particulate or ionicform, it is clear that the presence of nAg in freshwater systemspresents a high risk to aquatic life.
4.2. Titanium dioxide NPs (nTiO2)
nTiO2 is one of the most produced NPs in the world, withan expected production of 201,000 tons during 2015 (Epa, 2011;Markets, 2015). Some studies have reported evidences that thisNP have a low toxicity toward aquatic organisms even at concen-trations higher than the ones expected to occur in the freshwatersystems (3 ng L−1 to only 1.6 �g L−1) (Gottschalk et al., 2013).Federici et al. (2007) have showed that even at 0.1–1 mg L−1 ofnTiO2 the rainbow trout, defense system can naturally scavengeROS species avoiding OS. For this reason, most of the nanotoxicol-ogy studies (if not all) uses much higher concentrations with theobjective of obtain more straightforward acute toxicity responses.Evidently, as used NPs concentrations are largely higher than theabove predicted values, the environmental impact of these stud-ies decreases significantly. Moreover, higher nTiO2 concentrationsimpacts on their own undergone transformations; the presence ofa higher number of particles can lead to homoaggregation (seeSection 2), most probably, resulting in a misinterpretation of theobtained results. Dalai et al. (2013) showed that for nTiO2 concen-trations larger than 16 mg L−1, ROS levels in the Ceriodaphnia dubia
decreased significantly due to the agglomeration of the NPs becom-ing less bioavailable to the organisms. The exposure of O. mykiss
(rainbow trout) to fairly low concentrations of nTiO2 (0.1–1 mg L−1)resulted in biochemical disturbances, respiratory distress and sev-eral organ pathologies (Federici et al., 2007; Boyle et al., 2013).However, the same did not occur, when other organisms wereexposed to higher concentrations (Dalai et al., 2013; Hao et al.,2009). Table 2 resumes the biochemical responses of freshwaterorganisms, mostly daphnids and fishes exposed to nTiO2. In mostof these studies, the nTiO2 characterization was solely based on theevaluation of the NPs size distribution since aggregation is the mostimportant transformation for this particle.
Three main routes leading to nTiO2 toxicity were identified: (i)physical stress, associated with the NPs size and surface properties(cytotoxicity), (ii) photocatalytic activity (phototoxicity), and (iii)NPs capacity do adsorb xenobiotics in the media.
The internalization of nTiO2 by an organism may result in itsaccumulation in different organs leading to physical stress and tis-sues damage. For instance, accumulation of nTiO2 in daphnid’s guts(Dalai et al., 2013; Tan and Wang, 2014; Zhu et al., 2010), bivalve’sdigestive gland (Vale et al., 2014) and fish’s gills (Federici et al.,2007; Boyle et al., 2013; Hao et al., 2009) constricts their alimentarycanal affecting their breathing capacity. The possible mechanismsinvolved in the NPs internalization in the organisms were alreadydescribed (Reidy et al., 2013; Chen and Bothun, 2014; Shang et al.,2014), and will not be further discussed here.
Due to their photocatalytic activity, nTiO2 can generate radi-cal species when exposed to natural and/or UV light radiation. Alinear relationship between illumination and the hydroxyl radi-cal generation in the media was established in presence of nTiO2,with the extracellular radical species generated inducing oxida-tive damage on the gill tissue cells of D. rerio (Xiong et al., 2011).Similar effects were also observed in other freshwater organismssuch as zebrafish (Bar-Ilan et al., 2013), rainbow trout (Boyle et al.,2013) and daphnids (Dalai et al., 2013). Moreover, nTiO2 can leadto toxicity assumed to be promoted by OS under dark conditions inlaboratory and in microcosm study (Jomini et al., 2012).
The nTiO2 can adsorb or interact with different elements, (e.g.,Cd, Zn, Pb, Cu, Ni and As (Vale et al., 2014; Engates and Shipley,
G. Vale et al. / Aquatic Toxicology 170 (2016) 162–174 167
Table 1
Impact of nAg in freshwater organisms, including physicochemical characterization of the NPs in the media and biochemical responses. The biomarkers of response expressed
as ↑ and ↓, indicates an increase or decrease on the biomarkers activity or concentration, respectively, when compared with the control groups (organism not exposed to
nAg). The biomarkers here listed are: superoxide dismutase (SOD), catalase (CAT), glutathione (GSH), glutathione-s-transferase (GST), total glutathione peroxidase (GPx),
metallothioneins (MTs), lipid peroxidation (LPO) and reactive oxygen species (ROS).
Supplier information
(supplier; coating)
nAg characterization
(Size (nm);
concentration
(mg L−1); medium)
Bioassay
Organism Medium Biomarkers of response
(nAg concentration
(mg L−1))
Ref.
Sigma–Aldrich
uncoated
260 ± 26 (DLS)a;
0.001–0.08; TWc
32 (TEM)b; 0.001–0.08;
TWc
Lymnaea luteola L. TWc After 24 h exposure:
↓GSH (0.012; 0.024;
0.036); ↑CAT (0.012;
0.024; 0.036); ↓GPx
(0.024; 0.036); ↓GST
(0.036); ↑LPO (0.012;
0.024; 0.036); DNA
damage
After 96 h exposure:
↓GSH (0.012; 0.024;
0.036); ↑CAT (0.012;
0.024; 0.036); ↓GPx
(0.024; 0.036); ↓GST
(0.024; 0.036); ↑LPO
(0.004;0.012; 0.024;
0.036); DNA damage
Ali et al. (2014)
MTI Corporation
uncoated
307 (DLS)a; 1; BG-11d
50 (TEM)b BG-11d
Chlorella vulgaris BG-11d ↑ROS (1 and 10); ↑LPO
(1 and 10)
Oukarroum et al.
(2012)
ABC Nanotech
citrate-capped
70 (DLS)a; 5; FW e
90 (DLS)a; 10; FW e
12 (TEM)b; 5 and 10;
Freshwater e
Cyprinus carpio FWe Brain: ↓GST (0.1 and
0.2)
Liver: ↑GST (0.05; 0.1
and 0.2); ↓CAT (0.025;
0.05 and 0.1)
Lee et al. (2012)
Sigma– Aldrich
PVP coated
443 ± 15 (DLS)a; 1;
FSLWf
20-100 (TEM)b: 1;
FSLWf
B. thuringiensis and
B. aquimaris
FSLWf ↑ROS (1)
↑SOD (1)
Kumar et al. (2014)
Nanopoly
uncoated
5–20 (TEM)b D. rerio BWg ↑LPO (60,120)
↑GSH (120)
↓CAT (60,120)
↓GPx (120)
Choi et al. (2010)
a Dynamic light scattering.b Transmission electron microscopy.c Tap water.d Culture medium for Cyanobacteria (Rippka et al., 1979).e Freshwater.f Filtered and sterile lake water.g Bottled water.
2011; Gao et al., 2004; Sun et al., 2006)), changing their specia-tion in the media and affecting their bioavailability and toxicityto organisms. For instances, it was shown a correlation betweenthe decrease on Cd toxicity to the green algae Pseudokirchneriella
subcapita and the presence of nTiO2 not internalized by the algae.But on the other hand, if nTiO2 is already present in the organ-ism tissues, an increased bioaccumulation of xenobiotics due toadsorption onto the internalized NPs can occur (Tan and Wang,2014). Dispersed nTiO2 can also act as a carrier of xenobiotics inthe media increasing the metal uptake rate (Zhang et al., 2007b).
Independently of the route taken, nTiO2 toxicity is mostly asso-ciated with ROS generation, and the most common biochemicalendpoints are related with the anti-oxidant enzymes activity (seeTable 2). In D. magna, an increase of CAT, GST and GPx activity wasobserved when animals were exposed to 5–10 mg L−1 of nTiO2 (Kimet al., 2010), whereas in C. dubia an increase in SOD activity was alsoobserved after exposure to 1–64 mg L−1 nTiO2 (Dalai et al., 2013).Similar responses were also found in fishes, D. rerio (Clemente et al.,2014; Xiong et al., 2011), C. carpio (Hao et al., 2009) and O. mykiss
(Federici et al., 2007).
4.3. Zinc oxide NPs (nZnO)
As for nTiO2 also, nZnO are among the most used NPs withan estimated production of 30,000 metric tons per year (2015),originating predicted environmental concentrations in surfacewaters ranging from 0.008 to 0.055 �g L−1 in Europe and 0.001 to0.003 �g L−1 in US (Gottschalk et al., 2013, 2009). However, and asfor the other NPs, the concentrations used in ecotoxicology studiesare in general far higher than the ones predicted (Table 3).
Besides being an essential microelement, when at higher con-centrations Zn2+ is known to be toxic to aquatic organisms (Brunet al., 2014; Mortimer et al., 2010a). Being dissolution one of themain transformations of this particle is necessary to consider thattoxicity effects may be provoked by the NP per se and/or by the ionicfraction (Bondarenko et al., 2013; Franklin et al., 2007). Whereasdissolution of nZnO is hindered by the increase of the particle sizeor agglomerates size (Brun et al., 2014), the presence of proteinscan enhance it due to the proteins binding ability toward Zn (Reedet al., 2012). The presence of manufactured stabilizers can alsogreatly affect the NPs dissolution, with largest dissolved Zn usuallyobserved for uncoated nZnO followed by polymer-stabilized par-
168 G. Vale et al. / Aquatic Toxicology 170 (2016) 162–174
Table 2
Impact of nTiO2 in freshwater organisms, including physicochemical characterization of the NPs in the media and biochemical responses. The biomarkers of response expressed
as ↑ and ↓, indicates an increase or decrease on the biomarkers activity or concentration, respectively, when compared with the control groups (organism not exposed to
nTiO2). The biomarkers here listed are: superoxide dismutase (SOD), catalase (CAT), glutathione (GSH), glutathione-s-transferase (GST), total glutathione peroxidase (GPx),
metallothioneins (MTs), lipid peroxidation (LPO) and reactive oxygen species (ROS).
Supplier information
(supplier; chrystal
phase)
nTiO2 characterization
(Size (nm);
concentration
(mg L−1); medium)
Bioassay
Organism Medium Biomarkers of response
(nTiO2 concentration
(mg L−1))
Ref.
Sigma–Aldrich
30% rutile + 70%
anatase
400–800 (DLS)a; 5;
MHWb
D. magna MHWb ↑CAT (10); ↑GST (5 and
10); ↑GPx (5 and 10)
Kim et al. (2010)
Sigma–Aldrich
100% anatase
400–700 (DLS)a; 100;
EEMc
D. rerio EEMc Visible light: ↓CAT (1
and 10); ↑GST (10)
UV + visible light: ↓GST
(10)
Clemente et al. (2014)
Sigma–Aldrich
100% anatase
20.9 ± 2.86 (DLS)a; 0.1;
SM7d
218 ± 47.3 (DLS)a; 1;
SM7d
21 (TEM)d;; 0.1 and 1;
SM7d
D. magna SM7d No significant effects
on ROS and MTs
Tan and Wang (2014)
Sigma–Aldrich
100% anatase
248–293(1); FSLWe
517–925 (16); FSLWe
697–1090 (64); FSLWe
C. dubia FSLWe Photoperiod (16 h
light: 8 h dark): ↑SOD
(1–64); ↑ROS (1–32); ↓
ROS (>32)
Dark period:
↑SOD(1–64); ↑ROS
(1–64)
Dalai et al. (2013)
Degussa Evonik
80% rutile + 20%
anatase
>1000 (DLS)a; 100;
EEMc
D. rerio EEMc UV + visible light: ↓CAT
(1 and 10); ↓GST (1)
Clemente et al. (2014)
Degussa Evonik
75% rutile + 25%
anatase
24 (TEM)f; 0.1;0.5 TWg O. mykiss TWg Liver tissues: ↓GSH (0.5
and 1)
Gill tissues: ↑ GSH (1);
↑LPO (0.1 and 5)
Brain tissues: ↑LPO
(0.1; 0.5 and 1)
Intestine tissues: ↑LPO
(0.1; 0.5 and 1)
Federici et al. (2007)
Nanjing University
100% anatase
249–488 (DLS)a; 1;
DWh
270–535 (DLS)a; 10;
DWh
251–630 (DLS)a; 50;
DWh
402–633 (DLS)a;100;
DWh
245–617 (DLS)a; 300;
DWh
20–70 (TEM)f; DWh
D. rerio DWh Liver tissues: ↓CAT
(50), ↓SOD (50), and
↓GSH (50)
Gut tissues: ↑SOD (50);
↑GSH (50); ↑LPO (50)
Gill tissues: ↑LPO (50)
Xiong et al. (2011)
a Dynamic light scattering.b Moderately hard water.c Embryo exposure medium.d Elendt simplified M7 medium.e Filtered and sterile lake water.f Transmission electron microscopy.g Tap water.h Distillated water.
ticles (Merdzan et al., 2014). These NPs can also dissolve after itsinternalization in the extra- or intracellular compartments (Fig. 2),dramatically changing the organism’s metallome profile.
The exposure of the algae P. subcapitata to nZnO, ZnSO4, ZnCl2and bulk ZnO resulted in similar toxic effects indicating that ionic Znwas the responsible for the observed toxicity (Franklin et al., 2007;Aruoja et al., 2009a; Neale et al., 2015). Dose–dependent adverseeffects of nZnO on D. rerio embryos and eleuthero-embryos wereobserved (Brun et al., 2014), resulting in hatching and inflammationreactions (Yu et al., 2011; Xia et al., 2011). Zn may act as a competi-tor with Ca influx in skin chloride cells by blocking its transportthrough the pore canals and affecting a large surface of the skin. The
observed toxic effects were related to the ionic Zn, since the NPssize distribution was too large to pass through the chorion pores.In contrast, zebrafish embryos showed a greater acute toxicity tonZnO than when in presence of ionic Zn (Zhu et al., 2008).
nZnO can establish hydrogen bonds and ligand exchanges withthe bacterial cell wall, modifying the protein structure and conse-quently changing its function (Jiang et al., 2010). These particlescan also bind to the cytoplasmic bacterial membrane possibly dis-rupting its integrity, and interrupt the fundamental role of electrontransport phosphorylation and energy transduction process (Lyonet al., 2007). nZnO can generate ROS due to its photocatalyticalproperties (Suresh et al., 2015), but can also suffer photocorrosion
G. Vale et al. / Aquatic Toxicology 170 (2016) 162–174 169
Table 3
Impact of nZnO in freshwater organisms including physicochemical characterization of the NPs in the media and biochemical responses. The biomarkers of response expressed
as ↑ and ↓, indicates an increase or decrease on the biomarkers activity or concentration, respectively, when compared with the control groups (organism not exposed to
nZnO). The biomarkers here listed are: superoxide dismutase (SOD), catalase (CAT), glutathione (GSH), glutathione-s-transferase (GST), total glutathione peroxidase (GPx),
lipid peroxidation (LPO), nitric oxide (NO), reactive oxygen species (ROS), tumor necrosis factor � (TNF�), myxovirus-resistance protein A (MxA), interleukin-1 beta (IL-1�),
mitochondrial uncoupling protein 2 (Ucp-2), heat shock proteins (HSP) and macrophage scavenger receptor (MSR).
Supplier information
(supplier; chrystal
phase)
nZnO characterization
(Size (nm);
concentration
(mg L−1); medium)
Bioassay
Organism Medium Biomarkers of response
(nTiO2 concentration
(mg L−1))
Ref.
Genes’Ink 218–316 (NTA)a; 5;
HM b
196–211 (NTA)a; 0.2;
HMAc
214–236 (NTA)a; 1;
HMA c
223–242 (NTA)a; 5;
HMAc
D. rerio
(embryos)
HMAc ↑ CAT (0.2, 1 and 5)
48 hpf; ↓ CAT (5) 96 hpf
↑ Cu/Zn-SOD (5)
48 hpf; ↓ Cu/Zn-SOD
(5) at 96 hpf
↑ mt2 gene (0.2, 1 and
5) 48 and 96 hpf
↓ TNF� (5) 96 hpf
↑ c-jun protein (0.2, 1
and 5) 96 hpf
↓ MxA (1 and 5) at
48 hpf; ↓ MxA (5) 96
hpf
↓ IL-1� (5) 96 hpf
Brun et al. (2014)
Genes’Ink 218–316 (NTA)a; 5;
HMWAb
196–211 (NTA)a; 0.2;
HMAc
214–236 (NTA)a; 1;
HMAc
223–242 (NTA)a; 5;
HMAc
D. rerio
(eleuthero-embryos)
HMA c ↑ CAT (0.2 and 1)
120 hpf
↑ mt2 gene (5) 120 and
168 hpf
↑ TNF� (1 and 5)
120 hpf
↓ Cu/Zn-SOD (5)
120 hpf
↓ c-jun protein (1 and
5) 120 and 168 hpf,
↓ MxA (5) 120 and
168 hpf,
↓ IL-1� (1 and 5) 120
and 168 hpf
Brun et al. (2014)
Nanjing High
Technology
2196–3144 (DLS)d; 10;
ZCMe
D. rerio (embryos) ZCMe ↑ ROS (1 and 10) at
96 hpf
↓ GSTp2 (10) at 96 hpf
↓ Nqo1 gene (10) at 24
and 96 hpf
↑ GSTp2 and Nqo1
gene (sediment + nZnO
10) at 96 hpf
Zhu et al. (2009)
Sigma–Aldrich 50–100 (DLS)d; 50;
ZCMe
D. rerio (embryos)
at 144 hpf
ZCMe ↑ SOD (1–100,
dose–dependent from
1 to 50)
↑ MDA (20–100); ↑ GPx
(20–100); ↑ DNA
damages (100)
↑ ROS (1 and 10; ↑
Ucp-2 (1–100);
↓ CAT (100); ↓ GSTp2
(50–100); ↓ Nqo1 gene
(100);
↓ Bcl-2 genes (50–100)
Zhao et al. (2013)
BASF UK
Z-COTE©
150 ± 60 (TEM)f; 100;
NHg
E. coli NHg ↑ SoxS (100) McQuillan and Shaw
(2014)
Nanostructured
& Amorphous
Material, Inc.
24–72 (FCS)h; 1; EPi
10–13 (AUC)j; 1; EPi
C. reinhardtii EPi Moderate Visible light:
↑ GSTs1 (1); ↑ HSP22C
(1);
↑ HSP70A (1); ↑
MSR1(1); ↑ MSR2(1);
and ↑ HSP90(1)
Simon et al. (2013)
nCuO; Nanostructured
& Amorphous
Materials, Inc.
∼130 (DLS)a; 10; MESb
∼140 (DLS)a:10;
MOPSc
∼140 (DLS)a;10; TAPd
∼140 (DLS)a;10; OECDe
∼140 (DLS)a; 10; LWf
Chlamydomonas
reinhardtii
MESb; MOPSc; TAPd;
OECDe; LWf
↑ROS showing
OECDe > MOPSc > MESb > TAPd > LWf
after 24 h
von Moos et al. (2015)
MTI Corporation 148(DLS)a; 40; HSMg C. reinhardtii HSMg ↑ROS (0.004 g L−1) Perreault et al. (2012)
Nanostructured
& Amorphous
Material, Inc.
27.2 ± 6.7 (TEM)f; 22;
MHRWk
D. magna MHRWk ↓ Ferritin 3 (2.2 and
9.0); ↓ C1q proteins
(2.2 and 9.0)
Poynton et al. (2011)
170 G. Vale et al. / Aquatic Toxicology 170 (2016) 162–174
Table 3 (Continued)
Supplier information
(supplier; chrystal
phase)
nZnO characterization
(Size (nm);
concentration
(mg L−1); medium)
Bioassay
Organism Medium Biomarkers of response
(nTiO2 concentration
(mg L−1))
Ref.
Sigma–Aldrich 264.8 (DLS)d; 1000;
DWl
L. luteola L. DWl Digestive gland cells
↑ DNA Damage (10–32)
24 h and 96 h
↑ LPO (21–32) 24 h; ↑
LPO (10–32) 96 h
↑ CAT (21–32) 24 h and
96 h
↓ GPx (21–32) 24 h; ↓
GPx (10–32) 96 h
↓ GST (21–32) 24 h; ↓
GST (10-32) 96 h
↓ GSH (21–32) 24 h and
96 h
Ali et al. (2012)
Sigma–Aldrich 71.11 (DLS)d; 1000;
DWl
B. alexandrina DWl Hemolymph: ↓ CAT
(7–35); ↓ SOD (7) and ↑
SOD (35);
↓ GSH (7–35); ↓ GST
(7–35); ↑ NO (7–35)
↑ LPO (7–35)
Tissues: ↓ CAT (7) and ↑
CAT (35); ↓ SOD (7) and
↑ SOD (35); ↓ GSH
(7–35); ↓ GST (7–35); ↑
NO (7–35)
↑ LPO (7–35)
Fahmy et al. (2014)
a Nanoparticle tracking analysis.b Holtfreter’s medium.c Holtfreter’s medium with alginate.d Dynamic light scattering.e Zebrafish culture medium.f Transmission electron microscopy.g Neidhardt’s medium.h Fluorescence correlation spectroscopy.i Experimental media.j Analytical ultra centrifugation (Saison et al., 2010).k Moderately hard reconstituted water.l Destilated water.
when exposed to UV light, decreasing their photocatalytic activ-ity in aqueous suspensions (Hariharan, 2006). OS induced by bothnZnO and/or released Zn2+, has already been observed in severalfreshwater organisms, such as bacteria (Lyon et al., 2007; Cabiscolet al., 2000; Adams et al., 2006; Zhang et al., 2014; Gunawanet al., 2013), algae (Simon et al., 2013), crustacean (Mwaanga et al.,2014), mussel (Gagné et al., 2013), snail (Fahmy et al., 2014), frog(Bacchetta et al., 2012, 2013; Nations et al., 2011) and fish (Brunet al., 2014; Xia et al., 2011; Hao and Chen, 2012; Zhu et al., 2009;Bai et al., 2010; Zhao et al., 2013; Nel et al., 2006). In bacteria,the OS effect of nZnO may occur under dark conditions as wellas under natural or artificial light and affect both gram-positiveand negative bacteria (Barnes et al., 2013). After exposure to UVlight radiation, a loss of viability in Escherichia coli and ROS gener-ation results in the expression of genes encoding SoxS and CAT,which are part of the OS response mechanism (Gunawan et al.,2013; Kumar et al., 2011). In the algae Chlamydomonas reinhardtii,the presence of nZnO led to elevated levels of transcripts genes,GSTS1, HSP22C, and HSP70A, and the transcripts encoding sub unitsMSR1, MSR2, and HSP90 which are involved in the organism defenseresponse against ROS (Simon et al., 2013). In bivalves, the presenceof nZnO led to an increase of LPO in the digestive gland (Gagnéet al., 2013). Similar findings were also observed in snails (Fahmyet al., 2014), with a decrease in GSH content, a significant inhi-bition of SOD and CAT activity and a decrease in total protein
and albumin contents. A significant increase of LPO levels and adecrease of GSH activity in the gills, liver and brain of carps werealso reported after a 14 days exposure to 50 mg nZnO L−1 (Haoand Chen, 2012). A concentration-time exposure effect on SOD,CAT, GPx, and Ucp-2 expression with an increase in MDA contentwas shown for zebrafish embryos (Brun et al., 2014; Zhao et al.,2013).
As found for nTiO2, nZnO can also adsorb several elementssuch as As, Al, Mo, Hg, Pb, Cu, Ni and Cd (Gagné et al., 2013;Ghiloufi, 2013; Hua et al., 2012; Mahdavi et al., 2012; Sheelaet al., 2012), changing their speciation in the media and therebytheir bioavailability. When internalized, these NP-metal complexesmay undergo dissociation followed by NPs dissolution due to theacidic conditions found internally resulting in the release of theadsorbed xenobiotics, and drastically increasing their concentra-tion in organism tissues.
Overall, these studies highlight the fact that, despite the vastliterature, more investigation is needed to fully elucidates themechanisms of nZnO toxicity from the cellular level to the organ-isms or population level.
4.4. Copper oxide NPs (nCuO)
nCuO is a widely used metal oxide presenting biological activ-ity (Moschini et al., 2013; Blinova et al., 2010; Chen et al., 2008;
G. Vale et al. / Aquatic Toxicology 170 (2016) 162–174 171
Table 4
Impact of nCuO in freshwater organisms including physicochemical characterization of the NPs in the media and biochemical responses. The biomarkers of response expressed
as ↑ and ↓, indicates an increase or decrease on the biomarkers activity or concentration, respectively, when compared with the control groups (organism not exposed to
nCuO). Lipoprotein (LP) and reactive oxygen species (ROS).
Supplier information
(supplier)
nCuO characterization
(Size (nm);
concentration
(mg L−1); medium)
Bioassay
Organism Medium Biomarkers of response
(nCuO concentration
(mg L−1))
Ref.
Nanostructured &
Amorphous
Materials, Inc.
∼130 (DLS)a; 10; MESb
∼140 (DLS)a:10;
MOPSc
∼140 (DLS)a;10; TAPd
∼140 (DLS)a;10; OECDe
∼140 (DLS)a; 10; LWf
Chlamydomonas
reinhardtii
MESb; MOPSc; TAPd;
OECDe; LWf
↑ROS showing
OECDe > MOPSc
>MESb > TAPd > LWf
after 24 h
von Moos et al. (2015)
MTI Corporation 148(DLS)a; 40; HSMg C. reinhardtii HSMg ↑ROS (0.004 g L−1) Perreault et al. (2012)
Sigma–Aldrich 209 ± 10 (DLS)a; 100;
MQWh
1230 ± 200 (DLS)a;
100; OMMi
25.5 ± 0.8 m2 g−1 (BET)j
T. thermophila OMMi ↑ROS (80); ↑LP (60) Mortimer et al. (2011)
Sigma–Aldrich 30 (SEM)k T. thermophila
(strain BIII)
OMMi Cytotoxicity:
fluorescence (4 h: 127
(124–144); 24 h: 97.9
(80.4–138) mg Cu L−1);
ATP (4 h: 129
(111–149); 24 h: 101
(91.1–190a mg Cu L−1)
Mortimer et al. (2010b)
MTI Corporation 523-800 (DLS)a; NRm ,
SFWn
−39.7 ± 3.8 mV (EPM)o;
NRm; SFWn
Lemna gibba SFW 150 ↑ROS (0.7 g L−1 of total
Cu) > esterase activity
Perreault et al. (2014)
a Dynamic light scattering.b 2-(N-morpholino) ethanesulfonic acid.c 3-(N-morpholino) propanesulfonic acid.d TAP x4 algae growth media.e OECD algae growth media.f Lake water (filtered and sterile).g High salt medium (Saison et al., 2010).h MilliQ water.i Osterhout’s mineral medium.j Brunauer–Emmet–Teller analysis.k Scanning electron microscopy.
m Not reported.n Synthetic freshwater (Frankart et al., 2002).o Electrophoretic mobility.
Midander et al., 2009), and producing DNA damages and cell deathcompared to micro-sized particles, and nTiO2 and nZnO (Midanderet al., 2009; Ahamed et al., 2010; Nel et al., 2009). However, itsenvironmental hazard has been poorly investigated (Kahru andDubourguier, 2010) as can be observed in Table 4.
The presence of nCuO (0.1 mg Cu L−1) induced formation ofsuperoxide anions, hydrogen peroxide and single-stranded DNAin different recombinant luminescent E. coli (Bondarenko et al.,2012). However, the dissolution of these particles was the key fac-tor triggering ROS and DNA damage. In other study is was alsoobserved that the presence of nCuO and released Cu2+ induced anapproximately 5-fold increase in ROS in E. coli compared to thebacteria-only control (Gunawan et al., 2011). The levels of non-viable cells exposed to nCuO and released Cu2+ were very similarsuggesting that adverse effects were originated by the ionic form asfound by (Ivask et al., 2010). Also Hu et al. (2009) showed that par-ticles significantly inhibited viable count of bacteria (36.8 − 81.9%)when using 25–200 mg nCuO L−1.
A low mutagenic potential to Salmonella typhimurium TA 97aand TA100 (marginal effects between 100 and 1600 �g/plate) wasdisplayed by nCuO (Pan et al., 2010). The OS provoked by nCuOon diverse Saccharomyces cerevisiae strains was investigated andcom pared with CuSO4 and bulk CuO (Kasemets et al., 2013). The
cup2� (Cu stress response-deficient strain) was the most sensi-tive strain (approximately 16-fold than the wild type), suggestingthat nCuO effect proceeds also via dissolved Cu-ions. EC50s of nCuO(16–19 mg Cu L−1) and CuSO4 (10–12 mg Cu L−1) differed from thebulk CuO (918–1082 mg Cu L−1). Effects on yeast growth were alsoreported (Kasemets et al., 2009a).
The paradigm of nCuO OS (10 mg L−1) was also investigatedon C. reinhardtii (von Moos et al., 2015), with results (24 h expo-sure) showing an immediate cell size increase, OS, and chlorophyllbleaching, while membrane permeabilization was observed after5 h. Agglomerated nCuO was toxic and the exposure media wasdecisive in whether or not particles or ionic Cu act as the main tox-icity mediators. C. reinhardtii was also used to show the effect ofthe presence of a manufactured coating on nCuO (polystyrene-co-butyl acrylate) comparing the outcomes with bare nCuO (Perreaultet al., 2012). Higher toxicity was obtained in presence of coatednCuO mainly due to its internalization in the cytosolic membranestructures. ROS formation was observed at 4 mg L−1 of coated nCuOreaching (392 ± 12)% at 40 mg L−1, whereas 4 mg L−1 of bare nCuOformed ROS that increased only up to (160 ± 15)% compared to thecontrol (Perreault et al., 2012). Coated nCuO was also 10-fold moretoxic for Lemna gibba than bare nCuO (Perreault et al., 2014); the48 h exposure of 0.4 g L−1 of coated-nCuO led to a 50% growth inhi-
172 G. Vale et al. / Aquatic Toxicology 170 (2016) 162–174
bition, while 4.5 g L−1 of bare nCuO were required to reproduce thesame effect. But particles dissolution played the major role in thetoxicity.
Tetrahymena thermophila was used to assess nCuO, bulk CuOand Cu salt effects on fatty acid profiling, ROS and LPO (Mortimeret al., 2011). All Cu forms induced ROS, but a larger induction wasobtained in presence of nCuO, which could not be explained solelyby the released Cu. Protozoa exposure (24 h) to 80 mg L−1 of nCuOsignificantly decreased the proportion of the two major unsatu-rated fatty acids, increasing the relative amount of two saturatedfatty acids. Analogous effects were not observed with other Cuforms. Several studies confirmed that ionic Cu contributed onlypartially to nCuO toxicity (Shi et al., 2011; Kasemets et al., 2009b;Aruoja et al., 2009b; Heinlaan et al., 2011). However, Mortimeret al. (2010b) showed that nCuO is 10–20-fold more toxic for T.
thermophila than bulk CuO mainly due to the released Cu.The D. magna exposed (96 h) to nCuO (<50 nm; 1.97 mg L−1)
resulted in significant alterations in (i) encoding protein acting inmetal ion binding, (ii) OS, and (iii) exoskeleton (Adam et al., 2015).In all cases, the alteration up-regulated the transcription level ofthe transcripts genes. Glycogen, lipid and protein concentrationin exposed daphnids was lower than unexposed organisms, butwhen comparing nCuO and metal salt exposures, the microarrayevidenced no significant differences in transcribed gene fragments.Thus, the toxicity of nCuO to D. magna was caused by Cu ions(Adam et al., 2015). nCuO (26.7 nm; 100 �g L−1) exposure (48 h)led to significant changes in gills transcriptome of adults D. rerio,with induction of genes involved in apoptosis, mitogenesis and cellproliferation (Griffitt et al., 2009).
Despite the very limited available data about nCuO ROS gen-eration, cytotoxicity and genotoxicity in for freshwater organisms,it can be concluded that nCuO provoke ROS toxicity mediated byionic Cu.
5. General remarks and conclusions
Most of the biochemical responses reported are related to theorganism’s ROS defense mechanisms, mainly through gene expres-sion or changes on anti-oxidant enzymes activities. As mentionedbefore (Section 3.2), data related with changes in freshwater organ-ism genome and proteome due to the presence of NPs is very scarce.This is indeed surprisingly since omics techniques have alreadyproved a great potential on the recognition of signatures related tospecific stress, eventually leading to the discover of new biomark-ers. The few data available suggest that the interaction with theorganisms is NP-specific. For example, it seems that both nTiO2
and nAg have similar targets in the organism physiology affectingboth protein synthesis and circadian regulation, while nZnO andnCuO effects are generally related with OS.
Photoactive NPs such as nTiO2 and nZnO, when under naturalor UV light, can generate ROS inducing OS. However, NPs composi-tion can also play an important role on ROS generation, since somemetals constituting the NP can instigate Fenton and Weiss-typereactions releasing ROS in the intra or extracellular media. Evi-dently, and as usually performed, the bioassays should contain acontrol group exposed to the salt form of the metal constitutingthe NP, allowing the distinction between the effects provoked bythe NP per se and/or by the NP dissolution products.
Despite the large number of studies dealing with ecotoxicologyof NPs, it is evident that is still not possible to establish cru-cial predictive structure-activity relationships. This is mainly dueto the fact that the majority of the available studies have criti-cal deficiencies on their experimental designs; a comprehensivephysicochemical characterization of the particles under the expo-sure conditions is mostly miscarried or restricted to a secondary
task. Despite the scientific community is already aware about theimportance of the NPs physicochemical characterization prior andduring the bioassay, this is still frequently neglected in the mostrecent studies giving rise to more confusing and contradictory data.Clearly, this pushes back the possibility to establish a proper envi-ronmental risk assessment plan for these current early generationson ENPs (1st and 2nd generation passive and active nanostructures,respectively), while advanced generations of ENPs (3rd and 4thgeneration nanosystems and molecular nanosystems) may not befar away, bringing additional challenges that require further novelapproaches.
The dynamic speciation of the NPs should be assessed in thesame exposure media of the bioassay by following key NPs transfor-mations: (i) dissolution, (ii) homo- and heteroaggregation, and (iii)sedimentation. Several analytical tools are nowadays available forthe quantification of these physicochemical transformations, eachof them having their specific advantages and limitations being ableto provide different information on ENPs properties (see reviewsPinheiro and Domingos, 2015; Tiede et al., 2008; Domingos et al.,2009b). This physicochemical characterization approach allows(i) to assess the bioavailable NP-containing species to which theorganisms will be exposed, and (ii) to relate the biocompatibleor bioadverse effects with the NP-containing species permitting aNP categorization and function. Nanotoxicology is indeed a mul-tidisciplinary field where the study of the NPs physic, chemistryand biological impact is crucial for a complete toxicological assess-ment. Unfortunately, there is a lack of legislation controlling theproduction, use and release of these materials to the environ-ment, and new NPs are commercialized every day without anappropriate assessment about their impact in environment andhuman health. The establishment of national and international lawsregulating the production of these materials is mandatory. Further-more, it is also urgent to increase the number of comprehensivenano(eco) toxicology studies under natural more environmentally-realistic conditions implying the co-presence of ENPs (at low andenvironmentally-realistic doses) and environmental constituentssuch as natural organic and inorganic dissolved and colloidal mat-ter. Only with these approaches a comprehensive risk assessmentwill be possible with production of environmentally safe-by-designENPs.
Acknowledgments
All the authors acknowledge to the European COST actionES1205, Engineered Nanomaterials—From Wastewater Treatment& Stormwater to Rivers (ENTER). G. Vale and R.F. Domingosacknowledges to Fundac ão para a Ciência e Tecnologia (FCT,Portugal) the post—doctoral grant SFRH/BPD/73117/2010 and theproject PTDC/AAC-AMB/111998/2009, respectively. The contribu-tion of K. Mehennaoui and S. Cambier were in part possible withinNanoGAM (AFP-PhD-9229040) and NANION (FNR/12/SR/4009651)respectively, both founded by Fonds National de la Recherche Lux-embourg.
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Annexe 2: Identification of Gammarus sp. in Luxembourg streams
1. Sampling sites:
Gammarus sp. were collected from different unpolluted sites in Luxembourg (Figure 1). Animals were
collected using a hand net. They were quickly transported to the laboratory where they were kept at 12 °C.
Only adults were selected. Ten individuals from each site were kept and frozen in liquid nitrogen and
conserved at – 80 °C until analysis.
Figure 1: Unpolluted rivers in Luxembourg. Stars indicates sites where Gammarus sp. were collected for
molecular identification
Schwaarzbaach
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2. Gammarus sp. identification
2.1. Genomic DNA extraction and PCR
Gammarids were digested overnight at 56°C in lysis buffer supplemented with proteinase K (10% v/v).
Genomic DNA (gDNA) was extracted using QIAmp DNA mini kit® (QIAgen), following manufacturer’s
instructions. The extracted DNA was quantified using Nanodrop. Primers for Gammarus sp. were designed
based on a set of six pairs of universal primers: COI, 12S, 16S, 18S and 28S (Berschick, 1997; Hou et al., 2007).
Primers used for amplification are listed in Table. 1.
Table 1: primers used for amplification and sequencing
Gene Primer Sequence (5’ – 3’) Reference
COI * HCO2198_F TAAACTTCAGGGTGACCAAAAAATCA
Folmer et al., 1994 LCO1490_R GGTCAACAAATCATAAAGATATTGG
12S 12S-2-SSU _ F GTGGATCCATTAGATACCC Berschick, 1997 12S-2-SSU_ R ACTGGTACCTTGTTACGACTT Hou et al., 2007
16S 16ST_F GGTAWHYTRACYGTGCTAAG MacDonald et al., 2005 16SB_R CCGGTTTGAACTCAGATCATGT Palumbi, 1996
18S 18S_ F CCTAYCTGGTTGATCCTGCCAGT
Englisch et al., 2003) 18S_R TAATGATCCTTCCGCAGGTT
28S 28S _F TTAGTAGGGGCGACCGAACAGGGAT
Hou et al., 2007 28S_R GTCTTTCGCCCCTATGCCCAACTGA
* COI : cytochrome c oxidase subunit 1
Polymerase chain reactions (PCR) were performed in a reaction volume of 25 µL containing 12.5 µL of 2X Q5®
master mix, 1.25 µL of each primer (0.5 µM) and 10µL of gDNA (1ng). PCR settings were:
2.2. Agarose gel electrophoresis and DNA profile analysis
Electrophoresis of PCR products was performed on a 3% (w/v) agarose gel containing ethidium bromide and
they were visualised under UV light. The PCR product sizes were determined by using Mass ruler TM DNA
ladder Mix ready to use #SM0403.
Denaturation 98°C 5min
Denaturation 98°C 30s
Hybridization 55°C 60s
Elongation 72°C 60s
Final extension 72°C 5 min
Cooling 4°C ∞
35 cycles
253
2.3. PCR purification and sequencing
PCR products were purified by classical ethanol precipitation method (Cambier et al., 2010). The purified DNA
fragments were quantified using Nanodrop, then amplified and purified using BigDye Xterminator® kit
following manufacturer instructions. The obtained fragments were sequenced using ABI genetic Analyser
3500.
3. Results
3.1. DNA profile analysis
One Gammarus was used to test the primers. Among the five-targeted genes, 18S was not amplified, and
many PCR products were obtained for 28S (Figure 2). This could be linked to the low specificity of the primers
allowing the amplification of various gDNA regions. COI, 12S and 16S genes were successfully amplified. Low
intensity fragments were obtained for COI (800bp) and 12S (800bp). An amplified fragment of 400bp was
obtained for 16S gene (Figure 2). Therefore, these 3 genes were selected for further steps.
Figure 2: Agarose gel electrophoresis of amplification products generated with universal primers designed for
Gammarus sp. White arrows indicate the selected genes for sequencing analyses. Black arrows indicate the size
of the amplified fragments
Low intensity PCR products of about 800bp of COI (Figure 3) and 12S (Figure 4) were obtained for 10
Gammarus. This confirms the low efficiency of these two primers for the amplification of COI and 12S genes
of Gammarus sp.
L L 28S COI 12S 18S 16S
800 bp
400 bp
254
Figure 3: Agarose gel electrophoresis of amplification products generated with COI primers for Gammarus sp.
Arrows indicate the amplified fragments.
Figure 4: Agarose gel electrophoresis of amplification products generated with 12S primers for Gammarus sp.
Arrows indicate the amplified fragments
Fragment of about 400 bp with a high intensity were obtained for the amplification of 16S gene for all
individuals (Figure 5). This gene was kept for the sequencing analysis and identification step.
Figure 5: Agarose gel electrophoresis of amplification products generated with 16S primer pair for Gammarus
sp.
L L L G1 G2 G3 G4 G5 G6 G7 G8 G9 G10
L L L G1 G2 G3 G4 G5 G6 G7 G8 G9 G10
L L L G1 G2 G3 G4 G5 G6 G7 G8 G9 G10
255
3.2. Sequencing results
Sequencing analysis (Figure 6) showed that the obtained PCR products, using 16S primers, corresponded to
~400bp mitochondrial partial rRNA16S gene.
Figure 6: Gammarus fossarum 16S chromatogram obtained with ABI genetic analyser 3500.
The obtained sequences were analysed using the Blastn program available on the NCBI website
(http://blast.st-va.ncbi.nlm.nih.gov/Blast.cgi) using the the GenBank database. Results are showed in table
2.
Homology of 97% to 100% with very low e-value were obtained for all 10 Blastn analysis. These Blastn results
on these 10 DNA sequences allowed to identify the Gammarus sp. as Gammarus fossarum. Furthermore,
among these 10 animals 9 were identified as Gammarus fossarum type B, and only one (G2) was
characterised as a Gammarus fossarum type C. These results show that the Gammarus species present in The
Attert River at Colmar berg (Luxembourg) is Gammarus fossarum with a majority of the cryptic species B.
256
Table 2: Species identification and GenBanK Accession number
Individual Gene Description Query
cover E- value Identification
Accession
number
G1 16S Gammarus fossarum type B
mitochondrial partial rRNA 16S gene 84% 2e-155 100% AJ269607.1
G2 16S Gammarus fossarum type C
mitochondrial partial rRNA 16S gene 82% 1e-124 100% AJ269622.1
G3 16S Gammarus fossarum type B
mitochondrial partial rRNA 16S gene 85% 1e-150 99% AJ269612.1
G4 16S Gammarus fossarum type B
mitochondrial partial rRNA 16S gene 82% 1e-118 98% AJ269610.1
G5 16S Gammarus fossarum type B
mitochondrial partial rRNA 16S gene 81% 1e-112 99% AJ269607.1
G6 16S Gammarus fossarum type B
mitochondrial partial rRNA 16S gene 99% 1e-150 99% AJ269607.1
G7 16S Gammarus fossarum type B
mitochondrial partial rRNA 16S gene 98% 2e-161 97% AJ269607.1
G8 16S Gammarus fossarum type B
mitochondrial partial rRNA 16S gene 84% 2e-142 99% AJ269610.1
G9 16S Gammarus fossarum type B
mitochondrial partial rRNA 16S gene 84% 1e-144 99% AJ269607.1
G10 16S Gammarus fossarum type B
mitochondrial partial rRNA 16S gene 82% 1e-124 99% AJ269610.1
257
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