Dissertation

Dissertation
Dissertation
submitted to the Combined Faculties for the Natural Sciences
and for Mathematics of the Ruperto-Carola University of Heidelberg, Germany,
for the degree of Doctor of Natural Sciences
Presented by:
Diploma biologist Jan Wölz
Born in:
Heidelberg
Oral-examination: ……………………………
Impact of contaminants on aquatic systems and inundated sites
with respect to flood events
In vitro biotests, chemical target analysis
and fractionation methods
Referees:
Prof. Dr. Thomas Braunbeck
Heidelberg Institute of Zoology, University of Heidelberg
Prof. Dr. Heinz Karrasch
Geographical Institute, University of Heidelberg
Acknowledgment
This PhD thesis was very pleasant, since it was not only a product of laboratory work and
computing, but also the result of an inspiring and friendly atmosphere with colleagues and
friends. Therefore, I would like to thank the workgroup Aquatic Ecology & Toxicology
Section at the Heidelberg Institute of Zoology, especially Prof. Dr. Thomas Braunbeck and
the workgroup Department of Ecosystem Analysis at the RWTH Aachen University, in
particular Prof. Dr. Henner Hollert. Both owe my great thanks for their support, help and
supervision of my thesis.
Beyond this, I would like to thank Prof. Dr. Henner Hollert and Dipl. biol. Thomas-Benjamin
Seiler for the perfect collaboration in our in 2007 established Department of Ecosystem
Analysis at the RWTH Aachen University.
Furthermore, I would like to thank Prof. Dr. Heinz Karrasch for his deep interest in the
research topic of this PhD and for being my second supervisor.
Good scientific work is supported by friendly and helpful people. My special thanks go to my
dear colleagues at both departments Lisa Bragenheim, Markus Brinkmann, Andrea Gerstner,
Katja Großhans, Sebastian Hudjetz, Dr. Steffen Keiter, Dr. Thomas Kosmehl, Eva Lammer,
Sibylle Maletz, Paula-Suares Rocha, Hanno Zielke and in particular Marit Ernst and ThomasBenjamin Seiler.
I furthermore like to express my sincere thanks to Ulrike Diehl, Susanne Miller and Kerstin
Winkens for their support in the lab.
I also like to thank the project partners of the RIMAX-Hot cooperation partners and the
BMBF for funding this project and, thus, the majority of my PhD thesis.
Beyond, I want to thank, in particular, Dr. Werner Brack, Tobias Schulze and Dr. Urte
Lübcke-von Varel (UFZ Helmholtz Centre for Environmental Research, Leipzig) for their
collaboration, the possibility for effect-directed analysis and their assistance with chemical
analysis. Furthermore, PD Dr. Lothar Erdinger and Dr. Andrew Rastall (Institute of Hygiene
and Medical Microbiology, University of Heidelberg) for support with the YES assay.
Accordingly, my thanks go to Evelyn Claus and Dr. Georg Reifferscheid (Federal
Hydrological Institute, Koblenz) for chemical analysis and provision of the Ames Fluctuation
assay.
Finally, I am very grateful to my parents for all their encouragement and support of my
doctoral thesis.
Contents
I
Contents
Abstract………………………………………………………………..……………………….1
Zusammenfassung....…………………………………………………..………………………3
1
Introduction ......................................................................................................................... 5
1.1
Chemicals in the (aquatic) environment and legal handling ................................................... 7
1.1.1
Excursus 1: REACH guideline of the European Union for regulation of chemical ........ 8
1.2
Chemicals related to suspended particulate matter.................................................................. 9
1.2.1
Excursus 2: EU Water Framework Directive (WFD) − Integrated river basin
management for Europe ................................................................................................................ 13
1.3
Flood events − Impact on flood plains .................................................................................. 14
1.4
Contaminants in the groundwater and aquifers ..................................................................... 16
1.5
Objectives of the study .......................................................................................................... 17
1.6
References ............................................................................................................................. 20
2
Influence of hydrodynamics on sediment ecotoxicity ...................................................... 29
2.1
Role of sediments in freshwater quality ................................................................................ 31
2.2
Factors affecting mobilization of sediments and (bio-)availability of contaminants ............ 32
2.3
Ecotoxicological methods to assess sediment contamination ............................................... 33
2.4
Combined approaches to investigate the influence of hydrodynamics on sediment
ecotoxicity ......................................................................................................................................... 34
2.5
Case Study River Neckar (Germany) .................................................................................... 35
2.5.1
Methods ......................................................................................................................... 36
2.6
Results and Discussion .......................................................................................................... 37
2.7
Case Study Morava Catchment Area (Czech Republic) ....................................................... 39
2.8
Conclusions ........................................................................................................................... 42
2.9
References ............................................................................................................................. 43
3 Changes in toxicity and Ah receptor agonist activity of suspended particulate matter
during flood events at the rivers Neckar and Rhine ................................................................. 49
3.1
Abstract ................................................................................................................................. 51
3.2
Introduction ........................................................................................................................... 53
3.3
Materials and methods........................................................................................................... 55
3.3.1
Suspended particulate matter sampling ......................................................................... 55
3.3.2
Sample extraction .......................................................................................................... 56
3.3.3
Water samples ............................................................................................................... 56
3.3.4
Multilayer fractionation ................................................................................................. 56
3.3.5
PCBs and PCDD/Fs − HRGC-HRMS analysis ............................................................. 57
3.3.6
Neutral Red Retention assay ......................................................................................... 58
3.3.7
DR-CALUX assay ......................................................................................................... 58
3.3.8
GPC.2D assay................................................................................................................ 59
3.3.9
7-ethoxyresorufin-o-deethylase assay ........................................................................... 59
3.3.10 Bio-TEQ values ............................................................................................................. 60
3.3.11 Chem-TEQ values ......................................................................................................... 60
3.4
Results ................................................................................................................................... 61
3.4.1
Neutral Red retention assay ........................................................................................... 61
Contents
II
3.4.2
AhR-mediated activity .................................................................................................. 62
3.4.3
DR-CALUX and GPC.2D assay with SPM .................................................................. 63
3.4.4
Multilayer and carbon on celite fractionation................................................................ 64
3.5
Discussion ............................................................................................................................. 66
3.5.1
Cytotoxic effects of complex samples ........................................................................... 66
3.5.2
Ah receptor agonist activity of water samples............................................................... 67
3.5.3
AhR-mediated activity of SPM ..................................................................................... 68
3.5.4
Modification of pollutant composition .......................................................................... 69
3.5.5
Comparison of AhR-mediated activity .......................................................................... 69
3.5.6
Sources of the remobilized PCDD/PCDF ..................................................................... 70
3.5.7
Relevance of persistent compounds analyzed ............................................................... 72
3.6
Conclusions ........................................................................................................................... 72
3.7
Recommendations and perspectives ...................................................................................... 73
3.8
References ............................................................................................................................. 74
4 Effect-directed analysis of Ah receptor-mediated activities caused by PAHs in suspended
particulate matter sampled in flood events ............................................................................... 79
4.1
Abstract ................................................................................................................................. 81
4.2
Introduction ........................................................................................................................... 81
4.3
Material and Methods............................................................................................................ 82
4.3.1
Chemicals used .............................................................................................................. 82
4.3.2
Sampling and preparation .............................................................................................. 82
4.3.3
Fractionation.................................................................................................................. 82
4.3.4
Primary Fractionation .................................................................................................... 83
4.3.5
Secondary fractionation ................................................................................................. 84
4.3.6
PAH analysis ................................................................................................................. 84
4.3.7
EROD induction assay .................................................................................................. 84
4.3.8
Bio-TEQ values ............................................................................................................. 85
4.3.9
Chem-TEQ values ......................................................................................................... 85
4.4
Results ................................................................................................................................... 85
4.4.1
AhR-agonist activities in primary fractions................................................................... 85
4.4.2
Distribution of activities among secondary PAH fractions ........................................... 86
4.4.3
Quantification of EPA-PAHs ........................................................................................ 86
4.4.4
Contributions of EPA-PAHs to determined AhR-agonist activity ................................ 87
4.5
Discussion ............................................................................................................................. 88
4.5.1
Active fractions ............................................................................................................. 88
4.5.2
Evaluation of prioritized compounds ............................................................................ 89
4.6
Conclusions ........................................................................................................................... 90
4.7
References ............................................................................................................................. 92
5
Flood Retention and Drinking Water Supply − Preventing Conflict of Interests ............. 97
5.1
5.2
5.3
5.4
5.5
Background ........................................................................................................................... 99
Aim of the joint research project ........................................................................................... 99
Framework of investigation................................................................................................. 100
Structure of the joint research project ................................................................................. 101
References ........................................................................................................................... 102
Contents
III
6 Impact of suspended particulate matter sampled at the river Rhine with respect to
operation of retention basins and drinking water safety ........................................................ 105
6.1
Abstract ............................................................................................................................... 107
6.2
Introduction ......................................................................................................................... 108
6.3
Material and methods .......................................................................................................... 109
6.3.1
Chemicals used ............................................................................................................ 109
6.3.2
SPM sampling ............................................................................................................. 109
6.3.3
Preparation of crude extracts ....................................................................................... 111
6.3.4
Clean-up of extracts and automated fractionation ....................................................... 111
6.3.5
Chemical analysis of HCB and PCBs ......................................................................... 112
6.3.6
GC-MS analysis for PAHs .......................................................................................... 112
6.3.7
EROD-induction assay ................................................................................................ 113
6.3.8
Bio-TEQ values ........................................................................................................... 113
6.3.9
Ames Fluctuation assay ............................................................................................... 113
6.4
Results ................................................................................................................................. 115
6.4.1
SPM sampled in 2006.................................................................................................. 115
6.4.2
SPM sampled in the context of the flood event in August 2007 ................................. 115
6.4.3
Identification of effective fractions ............................................................................. 116
6.4.4
Mutagenic potentials of fractions ................................................................................ 117
6.5
Discussion ........................................................................................................................... 118
6.5.1
Chemical loads of crude extracts................................................................................. 118
6.5.2
Biological hazard potential in crude extracts .............................................................. 119
6.5.3
AhR-agonists and mutagenic potential in fractions..................................................... 119
6.6
Conclusions ......................................................................................................................... 120
6.7
References ........................................................................................................................... 122
7
Pollution of riparian areas in consequence of inundation by extreme flooding ............. 125
7.1
Abstract ............................................................................................................................... 127
7.2
Introduction ......................................................................................................................... 128
7.3
Materials and methods......................................................................................................... 129
7.3.1
Chemicals used ............................................................................................................ 129
7.3.2
Soil sampling ............................................................................................................... 129
7.3.3
Soil extraction for assessment of total samples ........................................................... 130
7.3.4
Soil extraction and clean-up for fractionation ............................................................. 130
7.3.5
Automated fractionation procedure ............................................................................. 130
7.3.6
GC-MS analysis of fractions ....................................................................................... 131
7.3.7
EROD induction assay ................................................................................................ 131
7.3.8
Bio-TEQ values ........................................................................................................... 131
7.3.9
Ames Fluctuation assay ............................................................................................... 131
7.4
Results ................................................................................................................................. 131
7.4.1
AhR-mediated activities and identified compounds .................................................... 131
7.4.2
EROD inducing potential by soil fractions ................................................................. 133
7.4.3
Mutagenic potential of individual fractions ................................................................ 133
7.5
Discussion ........................................................................................................................... 134
7.5.1
Chemical contamination of crude extracts .................................................................. 134
7.5.2
Biological hazard potentials by crude extracts ............................................................ 136
7.5.3
Identification of active fractions .................................................................................. 136
Contents
7.6
7.7
IV
Conclusions ......................................................................................................................... 138
References ........................................................................................................................... 140
8 Contaminant entry into and transport in the saturated groundwater zone subsequent to
extreme flood events .............................................................................................................. 143
8.1
Abstract ............................................................................................................................... 145
8.2
Introduction ......................................................................................................................... 145
8.3
Materials and methods......................................................................................................... 147
8.3.1
Chemicals used ............................................................................................................ 147
8.3.2
Sampling and Preparation............................................................................................ 147
8.3.3
Water extraction .......................................................................................................... 148
8.3.4
Automated fractionation of SPM and soil samples ..................................................... 149
8.3.5
Chemical analysis – Carbamazepine (CBZ)................................................................ 149
8.3.6
Method for the instrumental analysis of estrogenic compounds ................................. 149
8.3.7
Yeast Estogen Screen (YES) assay ............................................................................. 149
8.4
Results ................................................................................................................................. 150
8.4.1
Investigation of groundwater samples ......................................................................... 150
8.4.2
Estrogenic activity in individual fractions and target analysis .................................... 151
8.4.3
Target analysis in fractions.......................................................................................... 152
8.5
Discussion ........................................................................................................................... 153
8.5.1
Carbamazepine as a tracer for riverine contamination ................................................ 153
8.5.2
Estrogenic activities in the groundwater ..................................................................... 154
8.5.3
Active fractions and target analysis ............................................................................. 155
8.6
Conclusions ......................................................................................................................... 156
8.7
References ........................................................................................................................... 157
9
Contaminant transport to public water supply wells via flood water retention areas ..... 161
9.1
Abstract ............................................................................................................................... 163
9.2
Introduction ......................................................................................................................... 163
9.2.1
Contaminant transport ................................................................................................. 163
9.2.2
Kastenwoert-Rappenwoert study area ......................................................................... 164
9.3
Phase 1: Entry of contaminants into the retention area ....................................................... 166
9.3.1
Characterization........................................................................................................... 166
9.3.2
Chemical analysis ........................................................................................................ 168
9.3.3
Ecotoxicological analysis ............................................................................................ 168
9.3.4
Modelling .................................................................................................................... 169
9.4
Phase 2: Passage through the soil zone ............................................................................... 170
9.4.1
Characterization........................................................................................................... 170
9.4.2
Chemical analysis ........................................................................................................ 171
9.4.3
Ecotoxicological analysis ............................................................................................ 171
9.4.4
Modeling ..................................................................................................................... 172
9.5
Phase 3: Groundwater flow ................................................................................................. 174
9.5.1
Characterization........................................................................................................... 174
9.5.2
Chemical analysis ........................................................................................................ 175
9.5.3
Ecotoxicological analysis ............................................................................................ 176
9.5.4
Modeling ..................................................................................................................... 177
9.6
Discussion and conclusion .................................................................................................. 178
9.7
References ........................................................................................................................... 180
Contents
10
11
V
Conclusions of the study ............................................................................................. 183
List of references ......................................................................................................... 187
Contents
VI
Abstract
1
Abstract
Scope of the present study is the development and application of aquatic in vitro bioassays and
methods of effect-directed analysis (EDA). It aims at investigating contamination of suspended
particulate matter (SPM) and pollution of inundated sites and riparian aquifer, respectively. In the first
part of this study, SPM was sampled during flood events and toxicological activities were determined.
The second part of the study dealt with possible conflict of interests between flood management
(operation of retention basins) and drinking water supply (sustainment of water protection areas).
Cytotoxic potencies were determined with the Neutral Red retention assay and dioxin-like and aryl
hydrocarbon receptor mediated activities with the 7-ethoxyresorufin-o-deethylase (EROD) assay, both
using RTL-W1 cells derived from rainbow trout (Oncorhynchus mykiss). Both bioassays indicated
elevated potencies associated with SPM sampled during flood events. Highly active samples were
fractionated in order to determine effective compounds. Strongly persistent compounds had an only
minor contribution to total biological effects, whereas less persistent substances caused the bulk of
biological activity. Chemical analysis showed that compounds analyzed with priority are not capable
of adequately explaining the biological effects measured. Non-priority and a priori unknown
compounds were mainly effective.
The second part of the study aimed to investigate impacts of river contaminants to inundated sites and
aquifer in flood events. For this end, the biotest battery was extended with the Ames Fluctuation assay
and the bacterial tester strains TA98 and TA100 (Salmonella typhimurium) to detect mutagenic
activity, as well as the Yeast Estrogen Screen (YES) assay with bakery yeast (Saccharomyces
cerevisiae) to determine endocrine activity. Further, a recently developed method of effect-directed
analysis (EDA) was used to separate more polar compounds in SPM and soil. Less persistent
compounds were shown to be highly active. However, more polar compounds caused the highest
effects. In accordance to findings of the first part of the study, chemical analysis showed that priority
compounds only made a minor contribution to biological effects.
River contaminant infiltration in the aquifer was assessed following a flood event with a recurrence
interval of ten years by measurement of a tracer compound and hormonal activity. Both parameters
indicated contamination of the aquifer following the flood. Water that was sampled in the hinterland
showed delayed effects and, thus, indicated mass transport in groundwater layers over elevated
distances.
The findings of this study document high contamination of flood SPM that may be deposited at
inundated sites. In particular, increased biological effects and chemical loads of more polar
compounds indicate an increased impact of contaminant transfer through soil and aquifer
contamination. Furthermore, infiltration and increased toxicological effects indicate a general risk of
groundwater contamination in consequence of flood events.
The results of the present study directly contribute to a manual assisting stakeholders and operators of
retention basins and waterworks to a priori avoid potential conflict of interests and, thus, could
directly be implemented in practical work.
Abstract
2
Zusammenfassung
3
Zusammenfassung
Die vorliegende Arbeit befasst sich mit der Entwicklung und Anwendung von In vitro-Biotests und
Methoden der Effekt-dirigierten Analyse (EDA). Ziel ist die Untersuchung der Belastung von
Schwebstoff (SPM) sowie des Schadstoffeintrags auf Retentionsflächen und in den flussnahen
Aquifer. Dazu wurden im ersten Teil der Studie SPM aus Hochwasser entnommen und hinsichtlich
toxischer Wirksamkeiten untersucht. Im zweiten Teil wurde an einem Modellstandort untersucht, ob
es zu einem Interessenkonflikt zwischen Hochwassermanagement (Betrieb von Retentionsräumen)
und Trinkwasserversorgung (Erhaltung von Wasserschutzgebieten) kommen kann.
Zytotoxische Schädigungspotentiale wurden im Neutralrot-Test und Dioxin-ähnliche und Arylhydrocarbonrezeptor-vermittelte Wirksamkeit im 7-ethoxyresorufin-o-deethylase (EROD)-Assay mit
RTL-W1-Zellen der Regenbogenforelle (Oncorhynchus mykiss) untersucht. In beiden Biotests wurden
erhöhte Wirksamkeiten durch SPM aus Hochwasser ermittelt. Zur Bestimmung der Effektverursachenden Substanzklassen wurden hoch wirksame Proben fraktioniert. Während Fraktionen mit
sehr persistenten Schadstoffen nur geringe Effekte bewirkten, wurde der Großteil der biologischen
Wirksamkeit durch mäßig persistente Verbindungen verursacht. Mittels chemischer Analytik konnte
gezeigt werden, dass prioritär untersuchte Kontaminanten die ermittelten Effekte nicht hinreichend
erklären konnten. Nicht-prioritäre und a priori unbekannte Substanzen wiesen folglich die größten
Schädigungspotentiale auf.
Der zweite Teil der Studie befasste sich mit der Fragestellung, ob Kontaminanten aus dem Fluss im
Hochwasserfall zu einer Belastung von Überflutungsflächen sowie des Aquifers führen können. Zur
Bearbeitung dieser komplexen Fragestellung wurde die eingesetzte Biotestbatterie erweitert und
zusätzlich mutagene Potentiale im Ames-Fluktuationstest mit den Bakterienstämmen TA98 und
TA100 (Salmonella typhimurium) sowie hormonelle Aktivität im Yeast Estrogen Screen (YES)-Assay
mit Bäckerhefe (Saccharomyces cerevisiae) gemessen. Weiterhin wurde eine neuartige, kürzlich
entwickelte Methode der Effekt-dirigierten Analyse (EDA) zur Anwendung gebracht, mittels derer
erstmals auch polarere Substanzen in SPM und Bodenproben für die Biotestung aufgetrennt wurden.
Hohe Wirksamkeiten wurden für mäßig persistente Schadstoffe ermittelt. Die größten Wirksamkeiten
wiesen jedoch polare Verbindungen auf. Mittels erweiterter chemischer Analytik konnte gezeigt
werden, dass, in Übereinstimmung mit den Ergebnissen aus dem ersten Teil der Studie, die prioritären
Schadstoffe nur zu einem sehr geringen Anteil zur biologischen Gesamtwirksamkeit beitrugen.
Weiterhin wurden die Auswirkungen der Infiltration von Schadstoffen aus Flüssen in den Aquifer
anhand eines Hochwassers mit einem Wiederkehrintervall von zehn Jahren untersucht. Zu diesem
Zweck wurde das Vordringen von Flusswasser durch den Nachweis einer Leitsubstanz sowie die
Veränderung hormoneller Wirksamkeiten infolge des Hochwassers erfasst. Beide Parameter wiesen
auf eine direkte Beeinflussung des Aquifers durch das Hochwasser hin. Effekte von Proben die mit
größerem Abstand zum Fluss entnommen worden waren wiesen eine zeitliche Verzögerung auf, die
auf Stoffinfiltration in den Grundwasserleiter hinweist.
Die Ergebnisse dieser Dissertation belegen eine hohe Schadstoffbelastung von Schwebstoffen aus
Hochwasser, die auf Überflutungsflächen abgelagert werden können. Insbesondere die hohen
Belastungen mit eher polaren Verbindungen machen einen Stoffeintrag in den Aquifer über längere
Zeiträume wahrscheinlich. Weiterhin weisen die Befunde der Grundwasseruntersuchungen nach dem
Hochwasser auf das generelle Risiko von Schadstoffinfiltration auch ohne den Betrieb von
Retentionsbecken hin.
Die Resultate dieser Dissertation fließen unmittelbar in ein Handbuch ein, das Entscheidungsträger
und Betreiber von Rückhaltebecken und Wasserwerken dabei unterstützen soll mögliche
Interessenkonflikte von vornherein zu vermeiden. Diese Arbeit konnte somit direkt in die praktische
Anwendung eingebunden werden.
Zusammenfassung
4
5
Chapter 1
1 Introduction
6
1.1 Chemicals in the (aqutic) environment and legal handling
7
1.1 Chemicals in the (aquatic) environment and legal handling
Water is the most fundamental substance for life, and fresh water is, in particular, most
important for organisms and subject to multiple use by humans as drinking water, agriculture
and industrial processes (Baron et al. 2002, Gleick et al. 2008). Although pure water is
absolutely essential for human beings, anthropogenic activities have resulted in significant
impairment of the aquatic environment. Due to dynamics of riverine systems, rivers are, e.g.,
used as 'solvent' and means of transportation for human waste (Oki & Kanae 2006,
Schwarzenbach et al. 2006).
Chemical pollution has profound impacts on aquatic ecosystems, and hazardous compounds
are released intendedly (application products) or accidentally into the aquatic environment
(Bendz et al. 2005, Vassiliadou et al. 2009). Chemicals are used to make virtually every manmade product and play an important role in everyday life of people around the world. On the
one hand, chemical products provide protection for crops and increase yields, prevent and
cure diseases, provide insulation to reduce energy use and so forth. On the other hand,
chemicals can also negatively impact human and environmental health, leaving a considerable
footprint, when their production and use are not managed responsibly (Helland et al. 2007,
Weber et al. 2008b).
Over the entire life cycle of a chemical product – from “cradle to grave” – there is a potential
for adverse effects on man and environment. Chemical risk reduction requires a management
that involves continuous review of each compound from conceptual design in the laboratory
over development up to distribution, marketing and a handling guide for degradation and/or
waste disposal (Norgate et al. 2007, Schiefer et al. 1997). Today, about 100,000 compounds
are on the market in the EU and worldwide, 30,000 to 70,000 out of which are in daily use
(Schwarzenbach et al. 2006). For thousands of chemicals that are sold or used in products
today, incomplete information exists on the volumes released to the environment, the targets
of exposure and the toxic properties. This means that the risk of many chemicals has neither
been thoroughly evaluated nor have they been adequately managed, because the necessary
information to do so is not available (Hofer et al. 2004, Petry et al. 2006).
Almost every country has chemicals industries; yet, almost 80% of the world’s total output is
currently being produced by only 16 countries: USA, Japan, Germany, China, France, UK,
Italy, Korea, Brazil, Belgium, Luxembourg, Spain, The Netherlands, Taiwan, Switzerland and
Russia. Usage and consumption of chemicals is far higher in OECD countries than in nonOECD countries. It is assumed that OECD countries will remain both the largest chemical
producers and consumers until at least 2020, while production and consumption will grow
much faster in non-OECD countries. Therefore, it was necessary to develop chemical
management at a world-wide level (OECD 2001).
Overall, the chemicals industry in OECD countries has made significant progress in reducing
releases of pollutants to the environment from manufacturing processes. Although there are
no consolidated data on emissions of known hazardous substances across OECD countries, it
is probable that, overall, such releases from chemical industries in these countries are
1.1 Chemicals in the (aqutic) environment and legal handling
8
declining. Nevertheless, releases of hazardous substances per unit of output still rank high
compared to other industries (OECD 2001). Further, direct discharges via municipal treatment
plants represent an important source of contaminants to the aquatic environment. This holds
even more true, since specifically acting substances such as drugs are usually released by
households via waste water and are not (completely) degraded (Heidler & Halden 2007,
Zuccato et al. 2006).
Over the years, policies have been designed to protect man and the environment from both the
hazardous emissions released during the production of chemicals and the risks posed by
chemicals which are contained in consumer products (Kocha & Ashford 2006). Industries are
also subject to regulations aimed at managing risks posed by the chemicals themselves, e.g.
collection and assessment of data on hazard and exposure, material safety data sheets,
labeling, marketing and use restrictions (Foth & Hayes 2008).
Historically, most of the management approaches used for controlling emissions during
production have dealt with “end-of-pipe” solutions (Lee & Rheeb 2005, Sarkis & Cordeiro
2001). Recently, governments and industries in the European Union (EU) implemented more
holistic approaches to minimize impacts on health and the environment throughout the
lifecycle of a product – from raw material use to final disposal – by designing more
environmentally benign chemicals and adopting integrated product policies, including
extended producer responsibility (Clift & France 2008, EU 2006).
Given increasing trade volumes of chemical products and the growing awareness of pollutant
transportation across national borders, the last three decades have seen a significant increase
in international efforts by governments to co-ordinate the management of chemicals. Overall
direction for this work was provided by the 1992 United Nations Conference on Environment
and Development (UNCED) held in Rio de Janeiro when it adopted Chapter 19 of Agenda 21
(UN 2001). Among other things, this chapter calls for accelerating international work on the
assessment of chemical risks, harmonization of classification and labeling of chemicals,
establishing risk reduction programs and strengthening national capacities for managing
chemicals. The REACH guideline of the European Union (EU), enacted in June 2007, aims at
harmonizing such regulations and at minimizing risks of chemicals at least within the
European Union markets (European Union 2006).
1.1.1 Excursus 1: REACH guideline of the European Union for regulation of chemical
Industrial chemicals have been used for many decades, and new products are regularly
introduced to the market. However, there was a confusing patchwork of current legislations in
the EU used for regulation. Thus, the old chemical regulation is currently replaced by the new
regulations on industrial chemicals control, the EU guideline REACH (EC 1907/2006). This
guideline deals with Regulation, Evaluation, Authorization and Restriction of Chemical
substances and eventually came into force on June 1, 2007. It aims at overcoming limitations
in testing requirements of former regulations on industrial chemicals in order to enhance
competitiveness and innovation with regard to the manufacture of safer substances and,
furthermore, at promoting the development of alternative testing methods (Ahlers et al. 2008,
Foth & Hayes 2008).
1.2 Chemicals related to suspended particulate matter
9
Without registration, commercialization of chemicals is no longer possible, while
authorization is limited to 'hazardous' substances. Information on properties and possible
impacts of chemicals are registered in a central database which is run by the European
Chemicals Agency of the EU (ECHA) in Helsinki, Finland (ECHA 2008, Kemmlein et al.
2009).
The purpose of REACH is to guarantee that substances produced, put onto the market and
used are not hazardous to human and environmental health. Therefore, this regulation is based
on the principle that confers responsibilities of product safety to producing industries and
downstream users. Furthermore, REACH intends to increase transparency and to extend
information publications about applied chemicals (Hansen et al. 2007, Homa et al. 2009).
Notably, hazardous compounds with carcinogenic, and mutagenic properties as well as
reprotoxic (CMR), along with those substances with persistent, bioaccumulative and toxic
(PBT) properties, as well as very persistent and very bioaccumulative compounds (vPvB)
have to be identified and authorized by ECHA, regardless of the production volume (Hansson
& Ruden 2006, Pouillot et al. 2009).
Producers and importers of chemicals with production volumes of > 1 ton per annum have to
accomplish registration including a technical document. Existing products will have to be
registered until 2018 (Black 2008). Compounds of volumes > 10 tons per annum need an
additional 'Chemical Safety Report'. Compounds with volumes > 100 tons have to be
registered until 2013; such with > 1000 tons until 2010. Substances with production volumes
< 1 tons are excluded from registration (Edser 2008, Gubbels-van Hal 2007).
REACH has been estimated to affect about 30,000 compounds (> 1 tons/a) out of a total of
100,000 old chemicals on the market as listed in the European EINECS index (Loewenberg
2006, Wolf & Delgado 2003). To manage these substances, a base set of data is required
containing information on identity, classification and labeling, as well as exposure
assessments. Further on, chemical safety reports have to be provided containing more
specified information as well as management strategies to minimize risks (Laamanen et al.
2008, Petry et al. 2006,).
Since estimations of potential exposure and effects of each chemical have to be provided, an
effect-assessment is carried out using biotests with at least a subset of species and exposure
scenarios to determine either biological effects or non-effect levels. In general, depending on
the production volume, all trophic levels need to be considered, and effects are determined
with at least destruents (bacteria), producers (algae) and consumers (daphnids, fish) providing
a base set of information (Wei et al. 2006).
1.2 Chemicals related to suspended particulate matter
In this study, suspended particulate matter (SPM) is defined as the particles that, given at any
time, are maintained in suspension by turbulent currents in a river. With decreasing flow,
SPM tend to settle down on the ground and, thus, becomes part of sediment. However, with
increasing discharge in flood events, sediments are eroded and re-contribute to SPM. This
1.2 Chemicals related to suspended particulate matter
10
definition is close to that of the Environmental Specimen Bank of the Federal Republic of
Germany as given by Schulze et al. (2007).
SPM is a heterogeneous mixture of compounds of various origin, size, shape, density and
surface structure. In detail these components may be topsoil, sand, carbonates, clay minerals
and organic matter in various stages of composition (Cornelissen et al. 2005, Doxarana et al.
2002). These solid matters originate from biogenic or geogenic erosion processes in
catchment areas and cover sizes of about 0.02 - 2 mm. Translocation and downstream
deposition in direction to the coast is mainly given for particles with grain sizes < 2 mm.
Transportation distances of greater particles are reduced, and deposition takes place close to
the sources, except in the case of mountain catchments or during flood events with increased
sheer-stress and flow intensities (SedNet 2004, Slattery & Burt 1997,).
SPM components provide more or less numerous binding sites for substances and, thus, also
for contaminants. Whereas sand provides less sportive surfaces, clay and organic matter are
minor sized and provide a higher surface-to-volume ratio and manifold binding sites for
compounds such as organic contaminants. In particular, humic substances which, e.g., bear
polymerized organic complexes of nitrogen and oxygen represent attractive binding sites for
organic pollutants (Gagne et al. 1999, Cornelissen et al. 2005).
SPM are an integral part of aquatic ecosystems and the raw material of habitats for numerous
benthic species (Heise & Apitz 2007). They are fundamental in the cycle of inorganic and
organic matter in aquatic systems (Netzband 2007), and, thus, adsorbed pollutants can have a
negative influence on ecosystem functions. Since SPM play a major role in aquatic
ecosystems, they have to be integral parts of characterization and environmental evaluation
(Hakanson 2006).
Furthermore, SPM are most relevant with respect to safety of waterway transportation.
Introduction and transportation of SPM causes sedimentation filling waterways and in
particular harbors. Ports have many still-water zones where fine-grained material accumulates
(Aria et al. 2009, Koethe 2003). Regular dredging is necessary to ensure that ports remain
fully functional. However, materials mainly consist of sediments and contaminants from the
whole course of the river. At the Hamburg Port, every year 3 to 5 Mio. m3 of dredged
materials are removed. For that portion of the materials that is too highly contaminated for
relocation in the North Sea, storage at land is necessary which causes expenditures of
35 Mio. € per year (Netzband 2007).
Certain pollutants have always been of concern in aquatic systems, including the
16 EPA-PAHs as defined to be of superior relevance by the United States Environmental
Protection Agency (US-EPA). Further on, Persistent Organic Pollutants (POPs) including
toxic, persistent and accumulating compounds as defined and restricted by the Stockholm
Convention in 1995, are prioritized. These are further characterized by the so-called
'grasshopper effect' which means that they can travel long distances in the environment by the
repeated processes of evaporation and deposition (Mackay & Wania 1995, Gouin et al. 2004).
Thereby, POPs are found anywhere in the environment as well as in human and animal tissues
all over the world (Weber et al. 2008a, b). At present, dioxins, furans, polycyclic aromatic
hydrocarbons (PAHs), polychlorinated biphenyls (PCBs) and some other hazardous
11
1.2 Chemicals related to suspended particulate matter
compounds, pose a threat to the environment. Worldwide, the number of (highly) polluted
sites continuously increases, causing a considerable problem with a need for regulation and
management (Apitz & White 2003, Bridges et al. 2006, Weber et al. 2008b).
Pollutants can be introduced into rivers through various pathways which include the
atmosphere (emissions, aeolian transportation), effluents (wastewater treatment plants) or
surface runoff (precipitation, irrigation; Boxall & Maltby 1995, Brown & Peake 2006, Kay et
al. 2006). Furthermore, a distinction can be made between rural areas with soil and channel
bank erosion as well as atmospheric deposition on soils, urban areas with leaching (sewer
drainage) and direct inputs from industries and shipping (Vink 2001).
Sources of contaminants differ between point and diffuse sources (see also fig. 1): Point
sources are identified sources of steady inflow over the scale of years, and the magnitude of
pollution is not influenced by meteorological conditions. Thus, point sources include
municipal and industrial wastewater effluents (Förstner et al. 2004, Stronkhorst & van Hattum
2003,).
Aerial
sources
Flood event
Soil
Point
sources
Non-point
sources
Floodplain
Retention basin
Groundwater
Dissolved
Particles
Colloids
Pelagic/
benthic
fauna
Sedimentation
Bioturbation
Diffusion
Resuspension
Pore water, sediments,
colloids, diffusion
Sedimentation
Remobilization/Erosion
Contaminant sink
Secondary contaminant source
Fig. 1 Pathways of contaminant introduction into the aquatic environment and fate in the river system,
with emphasize on the impacts through flood events (modified; Power & Chapman 1992).
Diffuse sources are highly dynamic and widely spread pollution sources with a close link to
meteorological factors such as precipitation. These sources include loads of surface runoff as
of cultivated fields, erosion and paved urban areas with traffic and atmospheric deposition
(Ferreira et al. 2003).
Whereas dissolved compounds remain easily available in the free water column for aquatic
organisms, less solvable hydrophilic compounds as, e.g., PAHs tend to adsorb to non-polar
surfaces of inorganic and organic particles (Knezovich et al. 2004, Tusseau-Vuillemin et al.
1.2 Chemicals related to suspended particulate matter
12
2007). Thus, they become less available for organisms. With currents and turbulences below
certain thresholds, SPM particles deposit on the river bed, thus acting as a sink of pollutants
and reducing their availability in the water column (Schneider & Reincke 2006, von der
Heyden & New 2004). Therefore, only bottom feeders and benthic fauna may still be exposed
to these particle-bound substances, whereas particles remaining in the water column (SPM)
have an impact on pelagic organisms (Caldas et al. 1999). Deposited particle-bound pollutants
may further consolidate with sediment and become part of new sediment layers. SPM act as a
buffer with respect to, e.g., nutrients, but also to persistent organic pollutants and act as sinks
and important secondary sources of contaminants due to absorption of settling particulate
matter (Ahlf et al. 2002b, Westrich & Foerstner 2005). As a consequence, in densely
populated and industrialized regions, a great variety of man-made hazardous compounds are
typically associated with aquatic SPM and sediments. Thus, contamination is a serious
problem in areas of rivers with intensive sedimentation, as is given in naturally (floodplain) or
artificial (retention areas) inundated areas along the rivers and, in particular, in deltas and
estuaries that act as large 'sediment traps' (Santschi et al. 2001, Vigano et al. 2003, Yang et al.
2008).
Previously deposited and consolidated SPM can be remobilized following sediment dredging
(Koethe 2003), with bioturbation being a major post-sedimentation process with normal
discharge (Butcher & Garvey 2004) and flood events (Hollert et al. 2003b, Wölz et al. 2008).
In fluvial systems, cycling of pollutants is dominated by processes of resuspension, settling
and burial of particulate matter (Heise et al. 2004).
Particle-bound contaminants are most relevant when assessing the pollution of aquatic
systems, in particular since they act as sinks and sources for contaminants. However, they
were not considered in the EU Water Framework Directive (EU-WFD 2000/60/EC), with
dissolved compounds being prioritized (EU 2000). Although, in a daughter directive to the
EU-WFD, the evaluation of water pollution is proposed including the assessment or priority
compound concentrations in sediments, the major relevance of sediments as secondary source
of pollution is not accounted for (Bergmann & Maass 2007, Hollert et al. 2007b). Further, the
implementation of sediments in the WFD underlines the particular need to focus on particlebound contaminants. Since sediments are often highly polluted, certain contaminants may
cause a failure to reach the quality goals in the WFD and may require additional measures for
its control (Heise & Foerstner 2006, Hollert 2007).
At the river Rhine, hexachlorobenzene (HCB) is such a candidate, since it has a significant
effect on the quality of sediments downstream and, e.g., on the dredged materials from
Rotterdam harbor. Therefore, HCB is assumed to be categorized as Category 1 contaminant
according to Article 16 Source/Pathway p. 11.1 in the WFD (Heise & Foerstner 2006). HCB
accumulates with depositing particulate matters in particular at the barrages along the Upper
Rhine; they can be remobilized during floods. Impacted downstream areas are highly polluted
and may become areas of important challenges for sediment management. Hamburg Port
Authorities are anxious that HCB may cause exceeded contamination levels that would
prevent relocation of sediments at sea. The relocation at sea is the least expensive option
(Netzband 2007). HCB loads in flood SPM were determined with up to 50 µg/kg in floods
1.2 Contaminants in the groundwater
13
with recurrence intervals of 1 year and increasing concentrations up to 350 µg/kg in floods
with recurrence intervals of 20 years. Thus, most HCB concentrations would exceed
thresholds for quality goals by Ahlf et al. (2002a) that would be reached with < 5 µg/kg.
Using the Chemical Toxicity Test (CTT) approach as well as the Uniform Content Test
(UCT) action levels (20 µg/kg) would also clearly be exceeded (Stronkhorst & van Hattum
2003).
1.2.1 Excursus 2: EU Water Framework Directive (WFD) − Integrated river basin management for Europe
Since the end of the year 2000, a novel legislative approach has guided European water
protection policy: The Water Framework Directive (WFD; 2000/60/EC). The new concept
includes replacing, merging and renewing of the previous protection policies from the 1970s
and provides a more consistent, transparent and comprehensive conception. The WFD aims at
a holistic and integrated water protection and sets ambitious high-quality goals to achieve a
'good status' for European lakes and rivers until 2015 (Bald et al. 2005, Wilby et al. 2006).
Thereby, ecological terms are of superior relevance, essential processes are detailed, as well
as instruments to reach the set aims, and, finally, there is a strict time schedule to be followed
(Anderson et al. 2006).
The WFD European Union legislation is double-tracked. On the one hand, community-wide,
substances of concern were selected (European Commission 2000), and, on the other hand,
each member state has to take measures at river basin level to manage prioritized pollutants
(Westrich & Foerstner 2005). Policies concerning hazardous substances in European waters
were introduced in a previous 'old' Framework Directive (76/464/EEC). In so-called daughter
directives of the 1980s, certain substances were regulated defining emission limits and quality
objectives in the surface and coastal waters valid community-wide. Until 2013, the 'old'
directive will be in force and subsequently replaced by the 'new' Water Framework Directive
(2000/60/EC) that integrates the ' old' directive (as codified under 2006/11/EC).
Identification of substances of concern for surface waters (significant risk for or via the
aquatic environment) and development of control measures are set out in Article 16 of WFD,
and priority compounds are listed and were adopted in Decision No. 2455/2001/EC (EU
2001).
To reach the aim set, the European Commission adopted another directive, setting
environmental quality standards for the priority substances in July 2006, which each member
state will have to achieve by 2015, to ensure a 'good chemical surface water status'.
Furthermore, within 20 years, emissions as well as losses and discharge of priority substances
will have to be reduced (Crane 2003).
The WFD implements the assessment of sediment contamination, e.g., in Article 16: “The
Commission shall submit proposals for quality standards applicable to the concentrations of
the priority substances in surface water, sediments or biota” (EU 2001). If quality criteria
were to be defined for sediment, then monitoring would be required to establish compliance
with such criteria. Certain compounds were selected and categorized as priority, priority
substances subject to review to priority hazardous substances and priority hazardous (see
1.3 Flood events – Impacts on flood plains
14
tab. 2). According to the WFD, good chemical status for a water body is obtained when the
concentrations of the priority substances in water, sediment or biota are below the
Environmental Quality Standards (EQSs): this is expressed as ‘‘compliance checking’’
(Coquery et al. 2005, Lepom et al. 2009).
The member states shall also set quality standards for river basin-specific pollutants and take
action to meet the WFD standards at latest by 2015 as part of the ecological status (Article 4,
11 and Annex V, WFD). Measures shall be in place by 2009 and become operational by 2012,
replacing Directive 76/464/EEC (Mostert 2003).
Table 2 Priority Substances List (according to EU 2008)
Priority substances
Alachlor
Benzene
Chlorfenvinphos
1,2-Dichloroethane
Dichloromethane
Fluoroanthene
Nickel and its
compounds
Trichloromethane
Priority substances subject to review
to priority hazardous substances
Anthracene
Atrazine
Chlorpyrifos
Di(2-ethylhexyl)phthalate
Diuron
Endosulfan
Pentabromobiphenylether
Cadmium and its compounds
C10 - 13-chloroalkanes
Hexachlorobenzene
Hexachlorobutadiene
Hexachlorocyclohexane
Isoproturon
Mercury and its compounds
Lead and its compounds
Naphthalene
Nonylphenols
Pentachlorobenzene
Polyaromatic hydrocarbons
Tributyltin compounds
Octylphenols
Priority hazardous substances
Pentachlorophenol
Simazine
Trichlorobenzenes
Trifluralin
1.3 Flood events − Impact on flood plains
Next to the more general introduction of the previous chapter, here the focus is on processes
in times of flood events. First of all, floods are characterized by modified hydrological
conditions and may occur when snow melts in spring or in consequence of (extreme)
meteorological events such as intense rain or storm (Barnett et al. 2005, Dankers & Feyen
2008). Precipitation volumes may no longer be retained by the soil and later be evaporated by
vegetation, respectively. Water level rises and river water inundates the surrounding
floodplains, often causing considerable socio-economic damage. At least when flooded areas
are agriculturally used or in particular when populated areas, industries and cities are affected.
In the past, many parts of European river catchments have severely been flooded (Becker &
Grunewald 2003, Frei et al. 2000, Hilscherova et al. 2007). Thus, rivers have been
straightened, dikes were erected and meanwhile housing at river sites was hindered or
prohibited, at least as long as dikes often failed to protect man and buildings (Faeh 2007,
Merz & Didszun 2005).
1.3 Flood events – Impacts on flood plains
15
Next to socio-economic aspects, flood events have to be evaluated in the context of sediment
erosion, sediment contamination and sediment deposition on inundated areas with different,
but nevertheless significant consequences. Flood events caused by intense rain come along
with extensive surface run-off at first (streets, farmland) introducing especially PAHs and
pesticides to the river and causing considerable toxicological impact (Boxall & Maltby 1997,
Brown & Peake 2006, Donald et al. 2005, Maltby et al. 1995a, b, Rocha et al. 2007). With
increasing discharge, critical sheer stress thresholds are exceeded and sediment erosion
becomes the dominant source of contaminant remobilization (Gerbersdorf et al. 2007a, Haag
et al. 2001, Hollert et al. 2003b, Lick & McNeil 2001, Ulrich et al. 2002,). Whereas less
intensive and more frequent flood events cause minor sediment erosion, more extreme floods
can cause deeper incisions remobilizing older and possibly more highly contaminated
sediments. These are transported downstream and deposited later on the surface sediments.
Thus, even older sediments can easily be (re-)eroded during subsequent floods and may
constitute significant sources of contaminants (Alekseevskiy et al. 2008, Hollert et al. 2003a).
Contaminant introduction and deposition on catchments may require regulation and
management of land use. At the river Rhine, serious flooding can leave up to 10 cm of
sediments along its banks. A few millimeters per flood cycle is deposited on the remaining
flood plain. Thus, over the centuries, a layer of sedimentary clay more than a meter thick has
been built up at various sites (Knepper 2006). Downstream transportation and deposition
through flood events may cause the removal of contaminated sediments and, thus, an
improvement of the water quality (Müller et al. 2002).
SPM begins to deposit with reduction of the flow following the flood peak (Jacoub &
Westrich 2006). With respect to inundated sites, the bulk of sedimentation takes place above
flat areas on floodplains. Furthermore, introduced sediments and adsorbed contaminants do
not necessarily remain on inundated sites. Subsequent flood events and after reflow of water,
dry sediments can be drifted to other areas, causing considerable pollutions elsewhere
(Asselman & Middelkoop 1995, Coulthard & Macklin 2003).
On inundated floodplains, introduced SPM preferentially deposit and accumulate in
depression zones from which aeolian export is minor. Thus, these geomorphologic structures
act as SPM sinks and tend to be highly polluted with particle-bound contaminants. Further on,
they are closer to the aquifer layer, are loaded with elevated contaminant concentrations and
increase the probability of passage through the unsaturated soil zone into the aquifer
(Hallfrisch 2008).
Past flood events take an impact on adjacent areas and cause tremendous damage in times
with extreme flooding (hundred-year-floods). As a result of extreme precipitation in August
2002, e.g., major flooding occurred in the catchment area of the rivers Elbe, Vltava (Moldau)
and Mulde (Navratil et al. 2008, Stachel et al. 2004). Pollutants from industrial sites and
municipal sewage treatment works entered the Elbe and led to serious pollution problems in
the river. At the Mulde river, the Spittelwasser-Schachtgraben and communities north of
Bitterfeld, a region with numerous and highly contaminated soils and sediments, was
inundated. Residential areas and farmland was contaminated through deposition of polluted
sediments (Brack et al. 2002a, Götz et al. 1998, Grote et al. 2005). Contaminated sediments
1.4 Contaminants in the groundwater
16
were remobilized and flood plains were considerably impacted. This resulted in
contamination of milk from cows grazing on the flood plains, and from two farms the milk
had to be destroyed, since the toxicity equivalency concentration (TEQ) values were above
the thresholds given by European regulation (Stachel et al. 2006, Umlauf et al. 2005).
Alluvially deposited matter of the Mulde River in the region of Bitterfeld was assessed by
Brack et al. (2002a, 2003b). The load of heavy metals and organic contaminants was shown to
exceed thresholds for sewage sludge. Ecotoxicological assessments by Heise et al. (2003)
with sediments sampled close to Brunsbuettel just before and up to 1.5 months after the Elbe
flood indicated a significantly increased toxicity after the flood event using standard luminous
bacteria (Vibrio fischeri), algae (Pseudokirchnella subcapitata), bacterial (Bacillus cereus)
and nematode assays (Caenorhabditis elegans).
1.4 Contaminants in the groundwater and aquifers
Organic contaminants that were introduced into the aquatic environment are of high diversity
with respect to their physico-chemical properties and molecular structures, as well as to
environmental compartments, e.g., for transformation and transportation processes. Thus, fate
and distribution of compounds are complex and require differentiated evaluation
(Reichenberg & Mayer 2006).
In this context, groundwater pollution has been regulated in the 1976 Dangerous Substances
Directive, which was later replaced by the Groundwater Directive 80/86/EEC (Lanz &
Scheurer 2001). With this new legislation, contaminants were divided into two categories:
(a) substances that must be prevented from entering groundwater ('black list') and (b)
substances the introduction into the groundwater of which must be limited ('grey list'). The
Groundwater Directive is of limited value and will expire and be repealed by the WFD by the
end of 2013 (Lanz & Scheurer 2001, Mostert 2003).
Contaminants may enter the riverine groundwater through pressure of river flow towards the
bank, at least when landside groundwater flow and ground gradient in direction to the river is
plain. Besides, contaminants can pass the unsaturated zone and enter the aquifer through the
soil, in particular at frequently inundated sites, e.g. floodplains and retention areas (Boulding
& Ginn 2004). Compounds are either hydrophilic or hydrophobic and, thus, accumulate in the
water phase or on the surface of suspended matters. Vertical flux in the water-unsaturated
zone as well as horizontal flux in the saturated zone has to be expected; an associated
transport of dissolved and particle-bound contaminants can be observed (Schwarzbauer
2006). In particular, the heterogeneous soil composition caused by many flood dependent
shifting is of major importance for the evaluation of the risk of contaminant transportation
through the soil (McKee et al. 1967).
Many studies worked on groundwater contamination and focused on inorganic contaminants
such as arsenic, lead and copper pollution (Black & Williams 2001, Fernandez-Galvez et al.
2007, Pasternack & Brown 2006). In contrast, groundwater and aquifer water has usually
been found contaminated with a variety of personal care products, pharmaceuticals, herbicides
and others more (Scheytt et al. 2007, Zhang et al. 2008a). In contrast, to date, organic
1.5 Objectives of the study
17
contaminant introduction from inundated floodplains and retention areas into the groundwater
has hardly been assessed.
One of the rare studies (Rudis et al. 2009) analyzed the deposition of zinc-polluted sediments
which were transferred by a catastrophic flood; this was related to changes in groundwater
quality. Modeling the fate of pollution leached from settled sediments to groundwater could
be show that zinc can leach to the groundwater aquifer both from bottom sediments of a pond
into the saturated zone and from the flood-pool sediments into the unsaturated zone and
thereafter into the saturated zone.
A more holistic approach was applied in the AquaTerra project that worked on a better
understanding of river-sediment-soil-groundwater systems (Gerzabek et al. 2007). New field
and laboratory observations as well as historical data are assembled and addressed for the
catchments of the Ebro, Meuse, Elbe and Danube Rivers and the Brevilles Spring. In the same
project, Barth et al. (2007) showed that for sediment transportation highest deposition rates
were given for β-hexachlorocyclohexane (β-HCH) in river sediments at hot-spot areas of the
Mulde River in the Bitterfeld region (Elbe Basin, Germany). However, no clear answer was
provided with respect to contaminant entry of floodplain soils in the groundwater.
Meanwhile, conflicts of interests emerge at different sites when, e.g., drinking water
protection areas and flood retention areas overlap. On the one hand, retention areas have to be
provided to manage in particular extreme flood events (Frerichs et al. 2003); on the other
hand, these areas frequently overlap with other land usages. Such a conflict of interests is
given next to Karlsruhe, were a waterworks (Kastenwoert) shall be constructed close to a
retention area (Bellenkopf-Rappenwoert), which has been designed to be used in the case of
an extreme flood event (events with a return period of more than a 100 years and a highly
destructive potential). This potential for conflict was subject of a joint BMBF research
project: “Flood retention and drinking water supply − Preventing conflict of interest” (Maier
et al. 2006).
1.5 Objectives of the study
This study aimed to investigate on the hazard potential of contaminants that are bound to
suspended particulate matter (SPM), in particular with respect to impact of flood events.
Therefore, in a first part flood SPM was sampled during floods with recurrence intervals of
one and eight years at the rivers Rhine and Neckar, Germany, and assessed using a broad
battery of in vitro biotests to detect both, acute and mechanism-specific adverse effects.
Exposure of RTL-W1 cells of the rainbow trout (Oncorhynchus mykiss) were used to
determine cytotoxicity in the Neutral Red retention assay. Dioxin-like and aryl hydrocarbon
receptor (AhR)-mediated activity was assessed with the EROD assay. Samples causing
elevated effects were chosen for effect-directed analysis (EDA) and fractionated in order to
reduce the complexity of the environmental samples. Subsequent, fractions were investigated
with biotests to identify biologically active fractions and possibly the effective contaminants.
1.5 Objectives of the study
18
Thus, hypotheses to be verified were:
• SPM sampled in flood events indicate elevated toxicities
• Elevated toxicities can be determined using in vitro biotests and different
(eco-)toxicological endpoints
• Effect-directed analysis is a powerful tool that can be used to identify active fractions
and effective compounds in complex environmental mixtures
• Chemical analysis allows to determine active compounds within the fractions
• The concept of toxicity equivalency concentrations (TEQs) can be used to determine the
quota of the analyzed compounds to the total biological activity of the extracts and
fractions
In a second part of this thesis, the angle of view was extended to the impacts of river
contaminants on inundated sites, in particular areas that are planned to be used as retention
basins in the case of extreme flood events. In a pilot study, a potential hazard of river
contaminants to the groundwater and, thus, drinking water resources, was indicated at one of
these projected retention basins that, therefore, was used as model site. Therefore, a long-term
study was carried out and samples were taken throughout two years. Further, soil was
sampled at inundated and non-inundated sites at the basin and groundwater at wells within the
retention basin.
Samples were assessed using an extended set of in vitro biotests. Next to the tests named
above, the Ames Fluctuation assay using the bacterial tester strains TA 98 and TA 100
(Salmonella typhimurium) was used to determined mutagenic potentials. Further, the Yeast
Endocrine Screen (YES) assay with bakery yeast (Saccharomyces cerevisiae) was used to
investigate endocrine activities. These investigations aimed to show hazard potentials of
contaminants that are transported in the fluent wave in floods to inundated sites and are
deposited there. Further, contaminant introduction in the groundwater and drinking water
resources was in the focus of research. Thus, hypotheses to be verified were:
• Biological hazards and chemical loads increase with more extreme flood events
• Long-term assessment of SPM allows to determine biological hazard potentials in flood
events that can be ranked with hazards in times without floods
• Deposition of flood SPM can be shown with soil sampled at inundated sites and
compared to soil of non-inundated sites
• Long-term investigation of groundwater shows modifications in contaminant
composition and biological activity in consequence of flood events
• The application of a recently developed fractionation method allows usage of EDA in
order to determine active fractions and to compare contamination pattern among
samples
• Further, this fractionation method provides new findings with respect to more polar and
polar compounds
1.5 Objectives of the study
19
• Finally, these investigations assist to evaluate the risk of contaminant transfer from the
fluent wave to retention basins and the hazard of groundwater contamination caused by
flood events
1.6 References
20
1.6 References
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challenges under REACH. Environ Sci Pollut Res Int 15: 565-572
Ahlf W, Braunbeck T, Heise S, Hollert H (2002a): Sediment and soil quality criteria. In: Burton F,
McKelvie I, Förstner U, Guenther A (Eds.), Environmental Monitoring Handbook. McGrawHill, New York, pp. 17-18
Ahlf W, Hollert H, Neumann-Hensel H, Ricking M (2002b): A guidance for the assessment and
evaluation of sediment quality - A german approach based on ecotoxicological and chemical
measurements. J Soils Sediments 2: 37-42
Alekseevskiy NI, Berkovich KM, Chalov RS (2008): Erosion, sediment transportation and
accumulation in rivers. Int J Sed Res 23: 93-105
Anderson BS, Phillips BM, Hunt JW, Worcester K, Adams M, Kapellas N, Tjeerdema RS (2006):
Evidence of pesticide impacts in the Santa Maria River watershed, California, USA. Environ
Toxicol Chem 25: 1160-1170
Apitz S, White S (2003): A conceptual framework for river-basin-scale sediment management. J Soils
Sediments 3: 132-138
Aria HA, Yuksel Y, Cevik EÖ, Guler I, Yalciner AC, Bayram B (2009): Determination and control of
longshore sediment transport: A case study. Ocean Engineering 34: 219-233
Asselman NME, Middelkoop H (1995): Floodplain sedimentation: Quantities, patterns and processes.
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27
Section A
Impact of sediment dynamics and hazard potentials of
suspended particulate matter
28
29
Chapter 2
2 Influence of hydrodynamics on sediment
ecotoxicity
H. Hollert1, M. Duerr2, I. Haag3, J. Wölz1, K. Hilscherova4, L. Blaha4, S. Gerbersdorf5
1
Department of Zoology, Institute of Zoology, University of Heidelberg, Im Neuenheimer Feld 230,
69120 Heidelberg, Germany
2
Hygiene Institute Halle, Johann-Andreas-Segner Street 12, 06097 Halle (Saale), Germany
3
Consultant Engineer Dr.-Ing. Karl Ludwig Water Economy - Hydraulic Engineering, Herrenstr. 14,
76133 Karlsruhe, Germany
4
Masaryk University, RECETOX, Kamenice 3, 62500 Brno, Czech Republik
5
University of St. Andrews, School Biology, Gatty Marine Laboratory, East Stands, Fife, St.
Andrews, Scotland, KY16 8LB, United Kingdom
Hollert H, Duerr M, Haag I, Wölz J, Hilscherova K, Blaha L, Gerbersdorf S (2007): Influence of hydrodynamics
on sediment ecotoxicity. In: Westrich B , Foerstner U (Editors), Sediment hydraulics and pollutant mobility in
rivers. Springer-Verlag, Heidelberg, pp. 401-416; ISSN 1863-5520, ISBN 978-3-540-34782-8
30
2.1 Role of sediments in freshwater quality
31
2.1 Role of sediments in freshwater quality
There is general agreement that sediment-bound substances are of major importance for the
fate and effects of trace contaminants as well as water quality in aquatic systems. Sediments
can act as sinks for various pollutants but could also become a contamination source under
certain circumstances such as dredging or flood events (Ahlf et al. 2002a,b, Foerstner and
Müller 1974, Hollert et al. 2000a, 2003a). Contaminated sediments are known to cause
various adverse effects on organisms even when contaminant levels in the overlying water are
low (Chapman 1989). Thus, monitoring and assessment of sediment quality is of prime
significance for national legislation in general and for the implementation of the European
Water Framework Directive in particular (SedNet 2004). Especially through the activities of
SETAC North America (Wenning et al. 2005, Wenning and Ingersoll 2002) and the European
SedNet network (Salomons and Brils 2004, SedNet 2004) sediments related issues were given
increasing attention in both science and the public. While water quality has notably improved
over the past three decades, the sediments in many European river basins still retain the toxic
heritage from the past era of uncontrolled industrial production, and which will continue to
influence the quality of waters significantly for many years to come (Salomons and Brils
2004, SedNet 2004).
Since the 1970s several chemical analytical studies revealed elevated concentrations of
dominant environmental contaminants such as heavy metals and organic pollutants in marine
and riverine sediments using chemical analyses (Foerstner et al. 2004, Foerstner and Müller
1974, Giger et al. 1974, Haag et al. 2001, Stoffers et al. 1977). These hazardous contaminants
are often accumulated in deeper layers covered by relatively unpolluted sediments, and, thus,
are sequestered from the bioavailable oxic sediment surface zone (Haag et al. 2001, Ziegler
2002). However, these chemicals are mostly persistent in the natural environment, and can
enter the oxic water column after an erosion events such as bioturbation (Chapman et al.
1992), flood events (Hollert et al. 2000a, 2003b) or dredging and relocation of sediments
(Koethe 2003). Consequently, toxicants can become bioavailable (Calmano et al. 1993,
Simpson et al. 1998, Ziegler 2002) and may result in detrimental effects on aquatic organisms
at various trophic levels. Furthermore, downstream transport and deposition of contaminated
particles in inundated areas may also result in negative effects on biota in these regions
(Japenga and Salomons 1993).
With a delay of more than one decade to the first geochemical studies, the assessment of
biological consequences of particle-bound pollutants has become a major topic in
international water research (Burton 1991, Giesy and Hoke 1989, Power and Chapman 1992).
To date, most studies focused on the development of suitable bioanalytical methods and the
assessment of their potential to investigate sediment-bound contaminants. However, the role
of sediment remobilization and possible ecotoxicological effects of contaminants bound to
suspended material has been scarcely investigated.
2.2 Factors affecting sediment mobilization and bioavailability
32
2.2 Factors affecting mobilization of sediments and (bio-)availability of
contaminants
In many river systems, hazardous contaminants are predominantly transported in association
with suspended particulate matter. The suspended particles and the sediment-bound pollutants
accumulate in regions of low turbulence, such as groyne fields, harbors, and river reservoirs
forming thus sites with high levels of contamination.
An important issue related to the role of sediments in water quality is their potential to be
subject to remobilization, transport and redistribution during certain environmental events
such as floods. Although these processes increase accessibility and bioavailability of
contaminants, the conditions under which these processes occur, their amplitude and possible
role in contaminant accessibility and effects are still poorly understood. The complexity of
cohesive sediments, which are biologically active and chemically reactive, precludes the
definition of a general analytical theory for their resuspension behavior. Moreover, the
sediment properties of cohesive sediments vary on a number of spatial, temporal and vertical
scales (Gerbersdorf et al. 2005, 2007) and empirically based field and laboratory experiments
are needed to elucidate the mechanisms which govern the erosion resistance of cohesive
sediments. As well, interdisciplinary studies are needed, to obtain better and realistic
conceptual understanding of natural sediments and their inherent physical and biological
complexity (Black et al. 2002). However, either physico-chemical or biological sediment
properties have been in the focus of research on their impact on sediment stability, and only
recently, the first comprehensive investigative approach to derive master-variables affecting
sediment stability was published (Gerbersdorf et al. 2005, 2007).
Over the past decades a numerous studies have been conducted that primarily addressed
isolated aspects of sediment pollution issues. Recently the fate of particle-bound pollutants
and hydrodynamic transport processes has been addressed increasingly in interdisciplinary
joint projects. These studies documented that particle-bound priority pollutants (e.g., EPAPAHs) are major contributors to both the overall contamination and transport of lipophilic
pollutants in rivers. Work that significantly contributed to these findings were, among others,
the DFG-Research Group 371 or the interdisciplinary BMBF-funded joint project, SEDYMO
(Foerstner et al. 2004, Foerstner and Westrich 2005). However, the questions regarding
physico-chemical surface properties of suspended particles, chemical mobilization and
biological degradation of pollutants as well as regarding the related bioavailability of
contaminants and their toxicity have not been satisfactorily addressed to date. Especially, the
important link between the erosion potential and hazard potential of sediments/distinctive
sediment horizons originating from contaminated riverine sites, need to be addressed in future
studies if a realistic risk assessment is to be derived.
The fate of the contaminants associated with sediments is strongly influenced by the amount
and type of the sedimentary organic matter, which reflects the environmental evolution in the
drainage area and fluviative or lake depositional systems (Martínek et al. 2006, Stout et al.
2002). The geochemical parameters of organic matter are controlled by the interplay of
2.3 Ecotoxicological sediment assessment
33
biomass productivity, weathering during transport, and microbial reworking during and
shortly after deposition (Peters et al. 2005).
Valuable monitoring data have been collected on the contaminants in sediments. However,
only limited data exist on the associated organic matter and the role of different organic
matrices for the fate of pollutants is insufficiently documented and not well understood (Stout
et al. 2004). Fresh sedimentary particles behave in a different way when compared to the redeposited older sediments, even if the content of pollutants is similar. It is, therefore, highly
desirable to integrate the role of natural organic matter of different biological origin, mainly
terrestrial plants, woody material, bacteria and algae (Gonzáles-Vila et al. 2003, Meyers
2003) into the ecotoxicological assessment of complex sedimentary systems of rivers and
their relevance for potential contaminant bioavailability. The extracellular polymeric
substances (EPS) excreted by microorganisms such as microalgae or bacteria, can be a
significant part of the total organic pool. These polymeric substances have received more and
more attention over the last years due to their role in biostabilization of sediments (e.g., De
Brouwer et al. 2000, Paterson et al. 2000). Only recently, the importance of EPS for the
erosion resistance could be shown for several contaminated freshwater sites (Gerbersdorf et
al. 2005, 2007). Concerning the fate of the contaminants, these polymeric substances
influence as well the nature of the eroded material, but this work is at an early stage (Perkins
et al. 2004). Depending on floc characteristics such as floc size and floc strength, the
adsorption/degradation processes of the associated contaminants will change as well as the
lateral particle transport until deposition (Droppo 2004). Thus, the binding capacity of the
polymeric substances, as well as their influence on the nature of the erodible material should
be addressed by investigating the quantity and the quality of the EPS in order to contribute to
the questions on the bioavailability of contaminants.
2.3 Ecotoxicological methods to assess sediment contamination
As discussed in the previous paragraphs, cohesive stability of sediments and their
mobilization leads to increased bioavailability of hazardous contaminants. Sediment
mobilization is affected by numerous physico-chemical, geochemical and biological
parameters that are poorly understood and that have been scarcely investigated by complex
interdisciplinary research projects. In spite of intensive research and development of
numerous model testing systems, little is still known about possible ecotoxicological
consequences of mobilized sediment contaminants. To evaluate adverse effects on
ecosystems, neither biotests nor chemical-analytic techniques alone are sufficient. In contrast,
a combination of biotests and chemical methods allows comprehensive insights into the
hazard caused by sediment contamination.
To monitor the sediment quality, ecotoxicological bioassays are first applied to screen if
contamination had significant effects on biological functions of the model organisms/
systems. A broad spectrum of test batteries of standardized bioassays has been used to assess
the possible hazardous effect of particulate matter and elutriate. The bioassays included in
vivo tests at different levels of the aquatic food chain and in vitro tests. Various
2.4 Influence of hydrodynamics on sediment ecotoxicity
34
microbiological toxicity tests have been developed and validated for use in sediment risk
assessment during past 20 years (Ahlf et al. 1989, van Beelen 2003). It was shown that
contamination correlates with the shift in microorganism communities toward toxicantresistant species and that persistent toxic effects on the micro flora caused for example by
zinc, cadmium and copper often occur at concentrations lower than European Community
limits (van Beelen 2003). Other assays for ecotoxicological studies include the algae growth
inhibition assay, the bacterial bioluminescence bioassay, and the Daphnia assay (den Besten
et al. 2003, Koethe 2003). Since fish are representing vertebrates, and can be linked via
bioaccumulation to humans, large efforts have been undertaken to develop fish-based test
systems for the assessment of sediment bound substances (Chen and White 2004, Davoren et
al. 2005, Hilscherova et al. 2000, Hollert et al. 2000a, 2005, Kammann et al. 2005a, Kosmehl
et al. 2004, US-EPA 2002). In addition to in vivo sediment exposure tests with fish, a number
of suborganismal assays are in use such as cell-based in vitro systems (Davoren et al. 2005,
Hollert et al. 2000a, Kosmehl et al. 2004, Segner 1998), the fish egg assay with Danio rerio
(Hallare et al. 2005, Hollert et al. 2003b). While acute toxicity was of major concern in the
last decades, recently for many river basins a change in focus to more subtle specific chronic
non-lethal effects occurred (Brack et al. 2005a). While these effects are difficult to assess
using in vivo tests, they can be relatively easily determined by in vitro techniques that allow to
predict toxic potentials of complex environmental mixtures (Janošek et al. 2006). The in vitro
bioassay approach serves as efficient, fast and cost effective screening for evaluation of the
receptor-mediated activities of the complex mixtures (Hilscherova et al. 2002). We have
successfully used this approach to prioritize contaminated sediment sites (Hilscherova et al.
2003, Hollert et al. 2002a) and to study novel endocrine disruptive effects observed in situ
(Blaha et al. 2006). A further advantage of a bioassay approach is, that the combination of
different bioanalytical methods allows to investigate multiple endpoints such as genotoxic or
mutagenic (Chen and White 2004, Kosmehl et al. 2006), dioxin-like (Hilscherova et al. 2002,
Hilscherova et al. 2001, Hilscherova et al. 2000), or various endocrine effects (Ankley et al.
1998, Sumpter and Johnson 2005) in parallel in the same sample.
2.4 Combined approaches to investigate the influence of hydrodynamics on sediment
ecotoxicity
Recently, in several studies toxicity has been evaluated at various sediment depths (Burton Jr.
et al. 2001, Hollert et al. 2003a, Kosmehl et al. 2004), showing for at least some of the
locations a dramatically increase of chemical contamination and toxicity with the sediment
depth. For several European river basis, including Neckar, Rhine and Elbe, highly
contaminated old sediments can be described as 'potential chemical time bombs' (Cappuyns et
al. 2006, Japenga and Salomons 1993). An important process which may remobilize such
sediments and which is of still increasing importance in relationship to the global climate
change is more often occurrence of stronger floods in Europe as well as in other parts of the
world. To understand and predict possible toxicological and ecotoxicological consequences of
contaminants mobilized from sediments by flood events it is necessary to develop scientific
2.5 Case Study River Neckar (Germany)
35
approaches for the assessment of regularly flooded rivers. The combination of hydrodynamics
and ecotoxicological investigations is devolving to an emerging field of research. Recently, it
was shown that hydrodynamic aspects can be involved as additional Line-of-Evidence in
Weight-of-Evidence studies assessing the impact of sediments (Chapman and Hollert 2006).
In the last five years several studies were published addressing the ecotoxicological impact of
flood events (Brack et al. 2002, Grote et al. 2005, Hollert et al. 2000a, Oetken et al. 2005,
Matthaei et al. 2006, Sect. 10.2) or using combined approaches for evaluating flood events
and the risk of erosion (Babut et al. 2006, Haag et al. 2001, Hollert et al. 2000b, 2003a).
In this context, studies on the Elbe flood in 2002 indicated elevated effects in bioassays
(Heise et al. in prep). Moreover, cellular changes could be found in livers from flounder
(Platichthys flesus) and digestive glands of blue mussels (Mytilus edulis), 5 month after the
flood disaster in the Elbe Estuary and the Wadden Sea (Einsporn et al. 2005). In comparison
to earlier data from long-term studies at the same stations, a significant impairment in the
function of cell organelles (lysosomes), involved in the detoxification and elimination of
pollutants in fish liver, was found. In addition, in a long time study, EROD activity was
measured in livers of dab (Limanda limanda) from the German Bight (North Sea) from 1995
to 2003 (Kammann et al. 2005b). In autumn 2002, significantly elevated EROD activities
were detected, possibly related to effects of the river Elbe flood event.
These findings support the hypothesis that extreme flood events can affect not only freshwater
ecosystems but also marine systems and have deleterious effects on animal health.
Furthermore, flood events can influence floodplains and wetlands negatively (Schwartz et al.
2006, Ulrich et al. 2002). Consequently, the risk of extreme flood events for drinking water
supply will be an emerging topic in the future (Maier et al. 2006).
In conclusion, research should consider the potential of sediments to serve as sources of
contamination for the aquatic ecosystem, for drinking water supply but also for the floodplain
soils and other flooded areas. In the following case studies, two examples for such integrated
approaches addressing the risk of erosion are presented briefly.
2.5 Case Study River Neckar (Germany)
During the seventies, the river Neckar in Southern Germany ranged among the most strongly
contaminated rivers in Germany with high loads of both organic pollutants and heavy metals
(Förstner and Müller 1974). For instance, cadmium loads were increased by a factor of up
to 300, when compared to pre-industrial clay stone sediments. As a consequence of sewage
treatment, the quality of water and sediments improved significantly, and today the Neckar
can be classified among Germany’s moderately contaminated rivers, however, with heavily
loaded old sediments at some sites (Hollert et al. 2000a). Hence, earlier studies within the
Neckar catchment area or the river Neckar itself, revealed moderate to strong ecotoxicological
effectiveness in several bioassays for mutagenic, genotoxic, endocrine, teratogenic, and
dioxin-like responses as well as correlations between biological effects and concentrations of
organic pollutants (Hollert et al. 2005, 2002a,b, 2003b).
2.5 Case Study River Neckar (Germany)
36
The objective of the presented study was to develop a combined ecotoxicological and
hydraulic approach by the cooperation between the Universities Stuttgart and Heidelberg to
elucidate the ecotoxicological implications associated with the risk of erosion of contaminated
sediments (Hollert et al. 2000b, 2003a). This integrated strategy was applied to the lockregulated river Neckar in Southern Germany (Haag et al. 2002, 2001, Hollert et al. 2000b,
2003a, Knauert et al. 2004). For this purpose, sediment cores of the heavily contaminated
Lauffen reservoir/river Neckar were investigated (A) as well as suspended particulate matter
during a flood event in the river Neckar (B) in order to give the potential and effective
pollution risk under different hydraulic scenarios (Fig. 1).
Assessment of the risk of erosion and
damage potential of highly loaded old sediments
Sediment cores
from reservoirs
Flood events
Non-intrusively measured density
profile of sediment cores
Cumulative parameters (AOX, TOC,
SAC254, KMnO4 consumption)
Chemical analyses
Chemical analyses of water and SPM
In vitro bioassays: toxicity,
genotoxicity & dioxin-like potential
Determination of the risk of erosion
(SETEG & hydraulic calculations)
In vitro- bioassays: toxicity, genotoxicity
& dioxin-like potential
Determination of SPM concentration
A
B
Separate evaluation
Separate evaluation
Integrated assessment
Fig. 1 Test strategies for examination and evaluation of the remobilization risk of old sediments in
lock-regulated river systems (redrawn from Hollert et al. 2000).
2.5.1 Methods
a) Two undisturbed sediment cores (13.5 cm in diameter and 150 cm in length) were taken
from one location in the backwater region of the Lauffen reservoir/river Neckar in southwest of Germany (in total 7 locations and 16 sediment cores). In both cores, vertical
profiles of bulk densities were measured in 1 cm steps non-intrusively by using a γ-raydensitometer. Thus, similar sediment layering within the parallel cores was ensured as well
as subsequent sampling of the appropriate sediment layers (Haag et al. 2001). If, on the
basis of the density profiles, parallel cores were considered to be similar, one of them
served to experimentally determine the critical shear stress of mass erosion (τc,e) as a
function of sediment depth. Erosion experiments were carried out in a rectangular water
flume, the so called SETEG-system (Kern et al. 1999). The second one of the parallel cores
2.6 Results and discussion
37
was sectioned into layers of almost uniformly texture, thus, the core was cut at depths of
significant bulk density changes. From this material, concentrations of heavy metals and
PCBs were identified by chemical analyses while the cytotoxicity, dioxin-like activity and
mutagenicity were investigated by bioanalytical methods (Kosmehl et al. 2004, Seiler et al.
2006). By comparison of the critical shear stress/sediment stability of the investigated
sediment cores with the natural occurring bottom shear stresses, calculated by the 1-D flow
and transport model COSMOS (Kern and Westrich 1997), the possible resuspension risk of
contaminated sediment layers could be predicted.
b) In order to gain insight into the ecological effects of a possible remobilization of heavily
contaminated old sediments, suspended particulate matter (SPM) was collected in SPM
traps from two sites of the lock-regulated section of the river Neckar: downstream the
Lauffen reservoir with its high cadmium contaminations and downstream the less polluted
Heidelberg reservoir (reference site). Parameters investigated are presented in fig. 1.
2.6 Results and Discussion
The combined hydraulic and ecotoxicological approach revealed the high risk of erosion
down to depth of 70 cm as well as an ecotoxicological hazard potential of the associated
contaminants (Haag et al. 2002, Hollert et al. 2000, 2003).
Clear cut changes in bulk densities, the percentage of particles size d < 20 µm and 137Cs
content support the hypothesis of an erosional unconformity (Fig. 2).
An erosional unconformity is the result of a flood event, where fine grained sediments are
resuspended and non-cohesive particles are re-deposited (Haag et al. 2000). In the vertical
sediment profiles, layers with coarse particles, low TOC and consequently increased bulk
densities could be detected mostly below 25 cm depth. Often these layers were also
characterized by sudden decreases of τc,e in the corresponding parallel core, indicating the
predominance of non-cohesive particles (Fig. 2).
2.6 Results and Discussion
38
Fig. 2 Non-intrusively measured density profile, d < 20 µm, critical shear stress, cytotoxicity,
mutagenicity, heavy metals and PCBs of core LN4K2 from the Lauffen Reservoir on the Neckar River
depending on the depth (according to Hollert et al. 2003).
Bioanalytical and chemical investigations (Fig. 2) were showing clear-cut changes of the
ecotoxicological hazard potential below the depth of the erosional unconformity. The younger
sediments within the top 25 cm depth revealed neither strong cytotoxicity nor mutagenicity.
In contrast, for the older sediments below that zone, a strong cytotoxicity, dioxin-like and
mutagenic potential could be determined. PCBs and anthropogenic influenced heavy metals
such as Cd and Pb showed up to 100 times higher concentrations in the sediment layers below
the erosional unconformity. Concentrations above 10.8 mg/kg of cadmium and 193 mg/ kg
copper, respectively, allowed the classification of these sediment layers to the older, highly
contaminated sediments (HCS). In contrast, the upper layers represented low contaminated
sediment layers (LCS, Haag et al. 2001, Hollert et al. 2003a). Since this unconformity
happens in a transition zone between younger, less contaminated and older, heavily
39
2.7 Case study Morava catchment ares
contaminated sediment layers, the last flood must have exposed not only deeper sediment
layers but also their contamination load.
The suspended matter of the high discharge - return periods of 15 to 20 years (Hollert et al.
2003a) - exerted significantly higher cytotoxicity and mutagenic activity (Fig. 3) than a
moderate flood with a 1-year return period (Hollert et al. 2000a). These findings supported the
conclusion that the observed ecotoxicological effects during major floods may be due to the
in-stream erosion of highly contaminated bottom sediments.
Recently, SPM of a flood event at the Neckar in 2004 with a recurrence interval of five year
was sampled using a sediment trap. Highest EROD activities of the extracts could be found
for the peak of the flood, with a ten time higher dioxin-like activity when compared to other
SPM samples (Wölz et al. 2008). The two samples with the highest effects have been used for
effect directed analyses. Using the shown strategy it is possible to investigate the risk of
erosion. However, the identity of the pollutants causing effects in the bioassay is still
unknown.
Effect-directed analyses is a strategy to gain insight into the character of the noxious
substances (Brack 2003, Brack et al. 2005b). Organic extracts from SPM sampled during the
2004 flood events was fractionated for polarity and aromatic properties according to a
previously developed methodology (Brack et al. 2005b). Only the fractions revealing high
toxicity on bioassays are used for chemical analyses in order to identify the toxic substance
class or substance. Using this approach, it was possible to elucidate PCBs and dioxins/furans
to contribute only for less than 1 % of the biologically derived EROD activities. The EROD
activities of the fractions with PAHs explained the major part of the Dioxin-like potential of
the crude extracts. However, the measured US-EPA priority PAHs contributed less than 20 %
to the total EROD activities (Wölz et al. 2008).
3,5
2000
Induction
m³/s
Discharge
3,0
TA98 w/o S9
1500
2,5
1000
2,0
500
1,5
TA98 w S9
1,0
0
20.Feb
21.Feb
22.Feb
23.Feb
24.Feb
Fig. 3: Time-course of the mutagenicity during the flood event of Oct/Nov 1998. Since for a moderate
flood event (HQ = 1) no mutagenicity could be found, several SPM extracts revealed genotoxic effects
in the Ames-Test without S9 mix.
2.7 Case Study Morava Catchment Area (Czech Republic)
Major flooding events also occur regularly in the catchment area of the river Morava (Czech
Republic). Water and sediment quality in this area has been impacted by historical industrial
40
2.7 Case study Morava catchment ares/2.8 Conclusion
activities within the watershed. In July 1997 the region was affected by disastrous floods
caused by two periods of exceptionally heavy rainfalls that resulted in great material and
ecological damages. Extensive rainfall plagued the north part of the Morava River basin and
the situation was even more complicated by the second flood wave within 10 days period. In
historical context it was very rare event but due to human landscape interventions it is
possible to expect similar events still more frequently. During the flooding period lasting for
several days, older sediments were washed away and new silt materials were deposited up to
several centimeters layer. Because of our earlier monitoring of this area, the situation brought
unique opportunity to evaluate the changes in contaminant levels and the toxic effects in
relation to flood events. Initial evaluations of the target contaminant profiles in sediment and
water samples from several sites revealed that there was a gradient of concentrations along the
Morava River from upstream to downstream, and suggested that the tributary of the little
stream Drevnice serves as a source of pollution to the Morava River (Hilscherova et al. 2001,
Holoubek et al. 1998). There are no limit values for sediment contaminants in the Czech
Republic but the concentrations of polycyclic aromatic hydrocarbons (PAHs) as well as other
organic compounds were above the maximal permissible limits that apply for instance in the
Netherlands, as were the concentrations of Cd and Zn for soils. Most studies have been
performed with the freshly contaminated top sediment layers, but still there is only little
information on the deeper layers that might be mobilized during frequent floods.
25000
ng.g
-1
20000
15000
before flood
after flood
10000
5000
12
8
6
2
1
0
Sediment No.
25000
ng.g
-1
20000
15000
10000
5000
26
24
22
20
18
16
13
8
7
1
0
Soil No.
Fig. 4: Effect of floods in 1997 on concentrations of PAHs (sum of 16 US-EPA PAHs) in sediments
and floodplain soils of the river Morava Catchment area.
Previous investigations have also shown the impact of floods on the periodically flooded soils
with significantly elevated contaminated levels namely with persistent organic compounds
(Hilscherova et al. 2001, Holoubek et al. 1998), and some heavy metals (recent unpublished
2.7 Case study Morava catchment ares/2.8 Conclusion
41
data). The most obvious changes related to major floods in 1997 were observed for PAHs, the
dominant contaminants in the area. The results clearly showed that in some regions there was
significant decrease in PAHs concentrations in riverine sediments after the floods while the
concentrations in the surrounding soils at most sites within the flood affected area
significantly increased (Fig. 4). Application of in vitro biotests has shown significant toxic,
genotoxic, dioxin-like, and estrogenic potentials in sediments collected from numerous sites
(Hilscherova et al. 2002, 2001) and the bioassay results confirmed significant effects of
floods. Both dioxin-like and estrogenic activities in sediments were generally either
unaffected or significantly decreased after floods (Fig. 5 and 6) showing removal of upper
contaminated layers and their transport downstream by the flood water.
The greatest added value of in vitro assays is that they provide an integrative measure of the
potential of the complex mixture of compounds within the sample that may cause a negative
effect through the specific mechanism of action. They serve as rapid, sensitive and relatively
simple screening systems evaluating the presence of chemicals and their mutual interactions
with specific mode of action. Fractionation of extracts enables separation of compounds
present in the complex mixture and allows determination of the most active classes of
compounds. In the study in part of the Morava catchment area, the simple fractionation
procedure also revealed the important role of mediate polar PAHs and pesticides for both the
estrogenic and dioxin-like effects (Hilscherova et al. 2001, 2002) which was confirmed by the
mass balance calculations (Hilscherova et al. 2002, 2001). Further, mechanism-specific
bioassays were confirmed to be an effective tool in initial screening of river sediments
compared to the more time- and cost-demanding instrumental analyses.
25000
TCDD-EQ
[pg TCDD/g]
20000
TEQ
15000
10000
5000
10A
9A
8A
5A
4A
3A
9B
8B
7B
6B
5B
4B
3B
0
Sample No.
Fig. 5: Dioxin-like equivalents determined in bioassay with H4IIE.luc cells (TCDD-TEQs) and by
chemical analysis (chem-TEQs) of organic extracts from sediments sampled in Morava Catchment
area before ('B' samples) and after ('A' samples) the major floods in 1997 (Hilscherova et al. 2001).
In vitro biotests have shown significant toxic, genotoxic, dioxin-like, and estrogenic potentials
in sediments collected from numerous waters (Hilscherova et al. 2002, 2001). However, most
studies have been performed on the freshly contaminated top sediment layers, and there are
42
2.8 Conclusions
no information regarding the deeper layers of sediments that might be mobilized during
frequent floods.
E2-EQ-chem
1400
E2-EQ-bio
E 2-EQ [pg E 2/g]
1200
1000
800
600
400
200
10A
9A
8A
5A
4A
3A
9B
8B
7B
6B
5B
4B
3B
0
Sample No.
Fig. 5: Estrogenic equivalents determined in bioassay with MVLN cells (E2-EQ-bio) and by chemical
analysis (E2-EQ) of organic extracts from sediments sampled in Morava Catchment area before (B)
and after (A) the major floods in 1997 (according to Hilscherova et al. 2002)
2.8 Conclusions
The present article features the urgent need to cross disciplinary boundaries in order to derive
a realistic assessment regarding the erosion risk of old deposited sediment layers as well as
the bioavailablitiy and hazard potential of their associated contaminants at different aquatic
sites. Especially the combination of hydrodynamic and ecotoxicological methods will give (i)
comprehensive insights into the effects of flood events on biota and ecosystems and (ii) allow
evaluation of sediment and, thus, water quality with regard to the global change and the
expectations of more severe floods in the near future.
2.9 References
43
2.9 References
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49
Chapter 3
3 Changes in toxicity and Ah receptor agonist
activity of suspended particulate matter during
flood events at the rivers Neckar and Rhine
Jan Wölz1,2 & Magnus Engwall2 & Sibylle Maletz1,2 & Helena Olsman Takner2
& Bert van Bavel2 & Ulrike Kammann3 & Martin Klempt4 & Roland Weber2,6
& Thomas Braunbeck2 & Henner Hollert1
1
Department of Ecosystem Analysis, Institute for Environmental Research (Biology V), RWTH
Aachen University, Worringerweg 1, 52074 Aachen, Germany
2
Man-Technology-Environment Research Centre (MTM), Department of Natural Science, University
of Oerebro, 70182 Oerebro, Sweden
3
Institute for Fishery Ecology, Federal Research Institute for Rural Areas, Forests, and Fisheries,
Johann Heinrich von Thuenen-Institute (vTI), Palmaille 9, 22767 Hamburg, Germany
4
Department of Safety and Quality of Milk and Fish Products, Federal Research Institute of Nutrition
and Food, Max-Rubner-Institute (MRI), Hermann-Weigmann-Straße 1, 22767 Kiel, Germany
5
Department of Zoology, Aquatic Toxicology and Ecology Section, University of Heidelberg, Im
Neuenheimer Feld 230, 69120 Heidelberg, Germany
6
POPs Environmental Consulting, Ulmenstraße 3, 73035 Goeppingen, Germany
Environ Sci Pollut Res (2008) 15:536-553; DOI 10.1007/s11356-008-0056-6
50
3.1 Abstract
51
3.1 Abstract
Background, aim, and scope
As a consequence of flood events, runoff and remobilized sediments may cause an increase of
ecotoxicologically relevant effects from contaminant reservoirs. Aquatic and terrestrial
organisms as well as cattle and areas of settlement are exposed to dislocated contaminants
during and after flood events. In this study, the impact of two flood events triggered by
intense rain at the rivers Neckar and Rhine (Southern Germany) were studied. Effects in
correlation to flood flow were assessed at the river Neckar using samples collected at frequent
intervals. River Rhine suspended particulate matter (SPM) was sampled over a longer period
at normal flow and during a flood event. Three cell lines (H4L1.1c4, GPC.2D.Luc, RTL-W1)
were used to compare Ah receptor agonist activity in different biotest systems. Multilayer
fractionation was performed to identify causative compounds, focusing on persistent organic
contaminants.
Materials and methods
Native water and SPM of flood events were collected at the river Neckar and at the
monitoring station (Rheinguetestation, Worms, Germany) of the river Rhine. Water samples
were XAD-extracted. SPM were freeze-dried and Soxhlet-extracted using acetone and finally
dissolved in dimethyl sulfoxide. Resulting crude extracts were analyzed for cytotoxicity with
the neutral red assay. Aryl hydrocarbon receptor (AhR)- agonist activity was measured in a
set of biological test systems (DR-CALUX, GPC.2D, and ethoxyresorufin-o-deethylase
(EROD) assay and different cell lines. In addition, crude extracts were fractionated using a
combined method of multilayer (sequence of acidified silica layers) and carbon fractionation.
Fractions from the multilayer fractionation contained persistent organic compounds
(polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs), polychlorinated biphenyls
(PCBs), and some polycyclic aromatic hydrocarbon (PAHs), fractions from the carbon
fractionation were separated into a PCDD/F and a PCB fraction. Dioxin-like activity of
multilayer and carbon fractions was determined in the EROD assay and expressed as
biological toxicity equivalency concentrations of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD;
bio-TEQs). The calculation of chemical equivalency concentrations (chem-TEQs) and
comparison to bio-TEQ values allowed the determination of the contribution of the analyzed
persistent compounds to the total biological effects measured.
Results
Soluble compounds in native and extracted water samples resulted in no or minor activity in
the toxicity tests, respectively. Filter residues of native water caused increased AhR-mediated
activity at the peak of the flood. Activities of SPM of the river Neckar correlated well with the
flow rate indicating a flood-dependent increase of toxicity culminating at the peak of flow.
River Rhine SPM showed a decrease of activity regarding an SPM sample of the flood event
compared to a long-term sample. Excellent correlations with AhR-agonistic activity were
3.1 Abstract
52
determined for DR-CALUX and EROD assay, while the GPC.2D assay did not correlate with
both other biotests. The activity of persistent dioxin-like acting compounds in multilayer and
carbon fractionated PCDD/F and PCB fractions was low if compared to corresponding crude
extracts. The congener pattern of PCDD/Fs revealed that the contaminations mainly
originated from products and productions of the chlorine and organochlorine industries.
Discussion
Native and extracted water samples could be shown to contain little or no cytotoxic or AhRagonistic compounds. In contrast, particle-bound compounds were shown to be the relevant
effective fraction, as indicated by the activities of filter residues of native water and SPM.
Compounds other than fractionated persistent PCBs and PCDD/Fs were more relevant to
explain AhR-mediated activities of crude flood SPM at both rivers assessed. Biologically
detected activities could at least in part be traced back to chemically analyzed and quantified
compounds.
Conclusions
The calculation of the portion of persistent PCBs and PCDD/Fs in multilayer fractions
causing the high inductions in the EROD assay in combination with chemical analysis
provides a suitable tool to assess dioxin-like activity of persistent compounds in SPM
sampled over the course of flood events. Depending on the catchment area and annual course
of flood events, end points may either indicate an increase or a decrease of activity. In order to
determine the ecological hazard potential of mobilized contaminants during flood events, the
focus should be set on particle-bound pollutants. Furthermore, PCDD/Fs and PCBs,
commonly expected to be the most relevant pollutants in river systems, could be shown to
contribute only to a minor portion of the overall AhR-mediated activity. However, they might
be most relevant for human exposure when considering persistence and bioaccumulation/biomagnification in the food chain.
Recommendations and perspectives
As a consequence of climate change, flood events will increase in frequency and intensity at
least in some regions such as Central Europe. Thus, it is crucial to identify the potential
hazard of (re-)mobilized contaminants from reservoirs dislocated via floods and threatening
especially aquatic organisms and cattle grazing in flood plains. Since other less persistent
compounds seem to be more relevant to explain AhR-mediated activities in flood SPM,
nonconventional PAHs and more polar compounds also need to be considered for risk
assessment. Effect-directed analysis using broad-range fractionation methods taking into
account compounds from polar to non-polar should be applied for identification of pollutants
causing biological effects, thus integrating biological and chemical parameters.
3.2 Introduction
53
3.2 Introduction
Anthropogenic changes of the atmosphere are expected to cause profound climate changes
and the first consequences are assumed to occur at present (Cracknell and Varotsos 2007,
Wittig et al. 2007). In particular, an intensification of global water cycling associated with an
increased risk of floods is anticipated (Hulme 2003, Wilby et al. 2006). Climate change will
result in a further increase of both
extent and frequency of floods in many regions across the globe (Ikeda et al. 2005, Kay et al.
2006a, b, Senior et al. 2002, Wilby et al. 2006). In Central Europe, increased flooding is
expected at least for most major rivers such as the rivers Rhine and Elbe. Severe weather
conditions with rainstorms in 2002 in the Alps, the Erz Mountains, and the Grand Mountains
in Eastern Germany resulted in an extreme flood event along the Elbe River. The Elbe River,
however, has been known to be severely contaminated by, e.g., magnesium chloride
electrolysis, and organochlorine production in the Bitterfeld area since the 1940s via the
tributary Mulde (Goetz et al. 1998, 2007, Wilken et al. 1994). Over the last 70 years, these
contaminants have repeatedly been translocated in-stream and deposited on inundated land
and flood plains (Goetz et al. 2007). During the 2002 flooding event, contaminated sediments
were remobilized from the riverbed of the Elbe and were deposited on flood plains (Umlauf et
al. 2005). As one of the many consequences, the milk produced by two farms had to be
destroyed since the thresholds of admitted dioxin levels set by European regulation have been
exceeded. Currently, a research project has been initiated to evaluate if some of the flood
plains can again be used for grazing cows or for sheep farming (Wojahn 2007).
Furthermore, during the 2002 flood, highly contaminated Mulde River sediments were
remobilized and deposited near Bitterfeld, Germany. Brack et al. (2002) were able to
document an increased ecotoxicological impact for the deposited sediments, with metal and
organic compound burdens exceeding threshold values for sewage sludge. As a consequence,
the sediments had to be categorized as potentially hazardous.
One strategy of governmental authorities to cope with the increased risk of floods is to
construct retention areas along major river systems to temporally retain and, thus, defuse the
peak flow of extreme flood events (Disse and Engel 2001, Maier et al. 2006). In 2000, the
European Union passed the Water Framework Directive (WFD, 2000/60/EG) as a basis for
the reconstitution of an ecologically good status for European freshwater systems by the year
2015. Meanwhile, the WFD has been complemented by a daughter directive, listing 33 socalled priority pollutants which are important contaminants of river sediments. Thus, the
contamination of sediments and suspended particulate matter (SPM) were enhanced in
political perception (Foerstner 2007, Hollert et al. 2007b). At least in this context, compounds
adsorbed to sediments and SPM have to receive special attention. Sediments of major rivers
in Germany (e.g., Elbe, Rhine, Main, Neckar) are known to contain elevated loads of
contaminants released mainly from the chlorine and organochlorine industries from the 1940s
to 1970s or via their products (Einsporn et al. 2005, Haag et al. 2001, Heinisch et al. 2004,
2007, Kosmehl et al. 2007, Weber et al. 2008a). These historic reservoirs are relevant for
3.2 Introduction
54
contemporary polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs)
contamination (Weber et al. 2008b). Today, fewer contaminated sediment surface layers are
known to cover older contaminated sediments, deposited at sites with low flow in the rivers,
such as flood plains, river reservoirs, and groyne fields, e.g., at the river Rhine.
Nevertheless, the risk of resuspension of old contaminated sediment layers as well as the
transport of particle-bound downstream translocation in the river system via in-stream erosion
increases with water discharge and the frequency of flood events. Furthermore, remobilized
sediments may intoxicate organisms and facilitate contaminant translocation in the water
column via SPM (cf. Gerbersdorf et al. 2007, Gerhardt 2007, Haag et al. 2001, Heise et al.
2004, Heise and Foerstner 2006, Hofmann and Wendelborn 2007, Hollert et al. 2007a, Lick
and McNeil 2001, Schwartz et al. 2006, Witt and Westrich 2003).
Contaminants in sediments and SPM of the rivers Neckar and Rhine are usually expected to
include chlorinated hydrocarbons comprising a large group of ubiquitously persistent organic
pollutants including polychlorinated biphenyls (PCBs), PCDDs, and PCDFs (Habe et al.
2001, Safe 1990, 1994). Due to their hydrophobic character and persistence, these chemicals
tend to adsorb to mineral and organic surfaces, e.g., SPM, which aquatic organisms are then
exposed to and which accumulate along the food chain (Hollert et al. 2000). Furthermore,
Heise et al. (2004) reported that hexachlorobenzene (HCB) contributes as a major toxicant to
the overall contamination of the river Rhine. HCB concentrations in SPM were shown to be
increased already at normal discharges. Concentrations of these contaminants have been
reported to significantly increase as a consequence of flood events and increasing sheer stress.
These compounds then are often translocated downstream and deposited in the riverbed or on
inundated floodplains (Hollert et al. 2007a, Gerbersdorf et al. 2007).
In this study, extreme meteorological conditions, causing intense rainfalls in the southwest of
Germany, were assessed in January 2004. As a consequence of the intensive precipitation,
river flow rates increased and flood events resulted along the rivers Neckar and Rhine.
Subsequent precipitation along the Neckar river lead to an intensive flood event with a hazard
quotient (HQ) of 1:10 (HQ = ratio of a flood with a certain discharge at a certain site) and, at
the rive Rhine, with an HQ of 1:2. Two flood events at the rivers Neckar and Rhine, two
major German river systems, were investigated to assess their potential influence on the
toxicity of water samples and SPM. For this end, a set of selected ecotoxicological end points
was recorded in a variety of cell-based monitoring systems. In detail, basic toxicity was
determined using the Neutral Red retention assay (NR assay) as detailed by Babich and
Borenfreud (1992). Acute cytotoxicity is quantified via uptake and retaining of the neutral red
stain in lysosomes of cells. In damaged cells, the stain is no longer retained in cytoplasmic
vacuolar membranes and the plasma membrane does not act as a barrier to retain the dye
within the cell (Babich and Borenfreud 1992). Furthermore, aryl hydrocarbon receptor (AhR)mediated toxicity was measured using different mechanism-specific cell-based test systems.
The 7-ethoxyresorufin-o-deethylse (EROD) assay was used to determine the overall and, in
particular, the dioxin-like enzyme activity, caused by native and extracted water of the river
Neckar and Soxhlet-extracted SPM sampled at the rivers Neckar and Rhine. The test is based
on the increased de novo synthesis and activity of cytochrome P450 enzymes (CYP) by cells
3.3 Material and methods
55
being exposed to contaminants, which is determined using the fluorescamine method to
quantify protein and the photometric detection of the enzyme reaction product resorufin
(Lorenzen and Kennedy 1993, Kennedy and Jones 1994). In addition, the GPC.2D assay and
the DR-CALUX assay were used to determine AhR-mediated activity of SPM samples,
quantifying the CYP activity via luminescence measurement.
Thus, the purpose of the present study was:
(1) to determine the portion of toxic effects of dissolved and particle-bound contaminants
with samples of two flood events at different catchments
(2) to compare the hazard potential of the more active samples using a set of cell-based
biotest systems and
(3) to identify and compare the portion of persistent compounds among the rivers Neckar and
Rhine
3.3 Materials and methods
3.3.1 Suspended particulate matter sampling
Surface water of the river Neckar were sampled in the period from January 14, 2004, to
February 3, 2004, (Tab. 1) with 20 L bottles or a water sampler according to Hollert et al.
(2000). SPM were collected with an SPM trap installed to the floating bridge of a power
station at Heidelberg. SPM was transferred to glass bottles, transported at 4 °C, and protected
from light. Samples were frozen at -20 °C, freeze-dried immediately (beta 1 - 8 K; Christ,
Osterode, Germany), and stored at 4 °C in darkness until analyzed.
Table 1 Sampling times and encoding of native and extracted water as well as SPM from the Neckar
River and of SPM from the Rhine River
Sample code
Neckar 0
Neckar 1
Neckar 2
Neckar 3
Neckar 4
Neckar 5
Neckar 6
Neckar 7
Neckar 8
Neckar 9
Neckar 10
Rhine 1
Rhine 2
Sampling period
Native water Extracted water
14.01.2003
10:00
10:00
14.01.2004
14:00
n.a.
14.01.2004
17:00
n.a.
14.01.2004
18:00
n.a.
15.01.2004
22:00
n.a.
15.01.2004
02:00
02:00
15.01.2004
08:00
08:00
15.-16.01.2004
12:00
n.a.
16.-20.01.2004
13:30
n.a.
20.01.-03.02.2004 12:00
n.a.
03.02.2004
14:00
n.a.
November 2003 to February 2004
15.-19.01.2004
Sampling date
n.a.—not assessed
SPM
n.a.
10:00-14:00
14:00-17:00
17:00-18:00
18:00-22:00
22:00-02:00
02:00-08:00
08:00-12:00
12:00-13:30
13:30-12:00
12:00-14:00
3.3 Material and methods
56
The sampling site at the river Rhine is located within the Rhine monitoring station at Worms,
a governmental institution founded for the continuous monitoring of water quality of the river
Rhine. Along the pillars of the bridge, four pumps were installed to provide a continuous
water supply from four lanes inside the river (for details, cf. Pawlowski et al. 2003). SPM was
allowed to settle down in a tank. Two mixed samples were collected, one of which was
sampled over an extended period from November 2003 to February 2004, including the flood
event (subsequently termed 'Rhine 1'). The other sample contains SPM over the complete
flood event (January 15-19, 2004; subsequently termed 'Rhine 2').
3.3.2
Sample extraction
Twenty grams of each freeze-dried SPM samples were extracted with 250 ml acetone (Merck,
Darmstadt, Germany) for 14 h using standard reflux (Soxhlet) extractors at approximately
eight to ten cycles per hour. The solvent was reduced in volume, and residues were
evaporated under a gentle N2-stream close to dryness. Finally, extracts were reconstituted in 1
ml dimethyl sulfoxide (DMSO, Sigma- Aldrich, Deisenhofen, Germany) and stored at -20°C
until testing. Empty extraction thimbles were extracted and processed in parallel to serve as
process controls. Subsequent provider and location of chemicals and medium used in this
study will only be named when differing from Sigma-Aldrich.
3.3.3 Water samples
Sample volumes of 20 L each were used for the preparation of XAD water extracts and
extracts of the filter residues according to the methods by Keiter et al. (2006) and Hollert et
al. (2005). All samples were cooled in a refrigerator to 4 °C immediately after return to the
laboratory and filtered using a 0.4 µm fiber glass filter (Sartorius, Goettingen, Germany) at a
pressure of approximately 1 bar. The effluent particle phase was extracted as detailed above
for SPM samples. In each case, the sample filtrate was adjusted to pH 2 using 1 M HCl.
1 L methanol (Fluka, Buchs, CH) was added to each sample before organic compounds were
extracted by using a 1:1 (v:v) mixture of Amberlite™-XAD 4 and XAD 7 resins (Serva,
Heidelberg, FRG). After extraction, the solvent volume was reduced close to dryness by
evaporation with N2, the extract was reconstituted in DMSO and stored at -20 °C until
analysis. Due to the reduction of the volume (20 L to 2 ml), a concentration factor (CF) of
10,000 was achieved for each sample. Additionally, three 20 ml subsamples of each water
sample were stored at -20 °C for the investigation of native water samples.
3.3.4 Multilayer fractionation
In order to provide information on the identities of unknown substances inducing AhRmediated activities in whole extracts, a multilayer fractionation was performed according to
the methods shown by Keiter et al. (2008) to remove acid-degradable non persistent
compounds (e.g., polycyclic aromatic hydrocarbon (PAHs)). The complete fractionation
method is shown in Fig. 1. Samples were cleaned on an open 'sandwich' silica column (mesh
15 mm, Merck, Darmstadt, Germany) packed with KOH silica (30 mm), neutral silica (5
3.3 Material and methods
57
mm), 40 % H2SO4 silica (30 mm), 20 % H2SO4 silica (15 mm), neutral silica (10 mm), and
Na2SO4 (10 mm) and were eluted with n-hexane (Riedel-de Haen, Seelze, Germany).
Soxhlet-extract
Silica gel multilayer
f ractionation
Acidity limited
degredation
F1: Acid-resistent
compounds
Carbon on celite
f ractionation
F2-1: PCB
Separation according
to planarity
F2-2: PCDD/F
Fig. 1 Fractionation procedure to separate Ah receptor-activating and persistent compounds in
complex flood SPM samples using multilayer and carbon on celite fractionation. F1 contained all
compounds being resistant to oxidation by sulfuric acid. Finally, fractions solely contained PCBs
(F2-1) and PCDD/Fs (F2-2). Dashed arrows describe the methodology, drawn through arrows describe
those related to fractionation
The remaining fraction contained persistent dioxin-like active compounds and included, e.g.,
PCDD/Fs and PCBs. Some PAHs, however, are known not to be retained in the multilayer
fraction (Windal et al. 2005). The solvents were divided into two equal portions, and the first
portions were evaporated under a nitrogen stream, and the sample was transferred to DMSO
(Fluka) for subsequent bioassays. To separate planar and non planar compounds, the second
portions of the samples were fractionated using an open carbon column (Carbopack C, Fluka).
Adsorbent (1.5 g), carbon on Celite™ (1:9, Celite 545, 20 - 45 µm, Fluka), and a layer of
Na2SO4 were packed in glass columns. The sample was eluted into two subsequent fractions
with 10 ml of hexane (non-planar compounds) and 80 ml of toluene (coplanar compounds).
After the addition of recovery standards (13C-labeled 1,2,3,4-TCDD, 1,2,3,7,8,9-HxCDD), the
samples were evaporated and transferred to amber glass autosampler vials in 25 µl of
tetradecane. The extracts and standards were stored at -18 °C until high-resolution gas
chromatography (HRGC)-high-resolution mass spectrometry (HRMS) analysis.
3.3.5 PCBs and PCDD/Fs − HRGC-HRMS analysis
HRGC-HRMS analysis was performed with a Micromass Autospec Ultima instrument
(Autospec Ultima, Waters Micromass, Manchester, UK) operating at greater than 10,00012,000 resolution using electron impact ionization at 35 eV. All measurements were
performed in the selective ion recording mode, monitoring the two most abundant ions in the
chlorine cluster. Splitless injection of 1 µl of the final extract was used on a 60 m Rtx dioxin 2
3.3 Material and methods
58
column (0.25 mm inner diameter, 25 µm). In addition to a blank sample with each set of
samples (five to ten), a procedure control that was treated like the samples was analyzed.
3.3.6 Neutral Red Retention assay
The acute cytotoxicity of the sediment extracts was determined with the neutral red retention
assay as detailed by Babich and Borenfreund (1992), with modifications described by Klee et
al. (2004) and Seiler et al. (2006). The permanent cell lines RTL-W1 (Lee et al. 1993) and
RTG 2 (Wolf and Quimby 1962), both from the rainbow trout (Oncorhynchus mykiss), were
used for biotesting. Cell culture was carried out as described by Klee et al. (2004). Sediment
extracts were serially diluted in L15 medium along seven wells in six replicates of a 96-well
microtitre plate (TPP, Trasadingen, Switzerland) to give a final concentration range of
0.78-50 mg/ml 3,5-dichlorophenol (Riedel-de-Haën) was used as a positive control at a
maximum concentration of 80 mg/L medium. Confluent cultures of RTL-W1 cells were
trypsinized and the resulting cell suspension was added to each well of the microtitre plate.
After incubation at 20 °C for 48 h, cells were incubated with neutral red (2-methyl-3-amino7-dimethylaminophenanzine) for 3 h, and neutral red retention was measured at 540 nm with
a reference wavelength of 690 nm using a Spectra™ III multiwell plate reader (Tecan,
Crailsheim, Germany). Second-order polynomial dose-response curves expressing the
viability of the cells, compared to controls, were plotted using Prism 4.0 (GraphPad, San
Diego, USA), and the cytotoxic potential of individual extracts was subsequently calculated
as NR50 values (= effective concentration for 50 % cell death in the neutral red test compared
to the negative controls with non-exposed cells). Subsequent NR50 values, with concentration
units of milligram SPM equivalents (SPM-EQ) per milliliter test medium in the well, will be
given as milligram per milliliter.
3.3.7 DR-CALUX assay
The DR-CALUX assay utilizes a rat hepatoma cell line (H4L1.1c4), with a luciferase reporter
gene controlled by the AhR (Biodetection Systems, Amsterdam, the Netherlands, Murk et al.
1996), and was applied with the protocol shown by Gustavsson et al. (2004, 2007). Cell
culture was carried out as described by Aarts et al. (1995). DR-CALUX cells were seeded in
96-well plates (TPP), 24 h prior to exposure under standard conditions of 37 °C and 5 % CO2,
and allowed to attain 100 % confluence. Thereafter, sample dilutions were prepared in culture
medium supplemented with 10 % fetal calf serum (FCS) and added to the cells in triplicate
wells. The extracts were tested in ten concentrations in 3-fold dilution series, with a maximum
concentration of 2 mg/ml. In each assay, a calibration curve of TCDD (0 - 300 pM) was
tested. The final concentration of DMSO (Fluka) was 0.8 %. After 24 h of exposure, the
medium was removed and the cells were washed twice with phosphate buffered saline (PBS)
and lysed in PBS at -20 °C overnight. After the substrate luciferin had been added, followed
by incubation for approximately 1 h in darkness at 20 °C, the activity of luciferase was
measured using the Luclite™ assay kit (PerkinElmer, Upplands Vaesby, Sweden). Cell
lysates were transferred to a white 96-well plate (TPP), and luminescence was determined in a
3.3 Material and methods
59
multiwell plate reader (Wallace 1420, Victor2, PerkinElmer, USA). Dose-response curves
were analyzed with GraphPad Prism 4, using the Hill equation (Olsman et al. 2007). The
bottom value was set to the response of the solvent controls. The luciferase-inducing potency
of the samples was converted to biological toxic equivalents (bio-TEQs) as described below.
3.3.8 GPC.2D assay
The guinea pig GPC.2D.Luc liver cells were provided by Dr. T.A. Gasiewicz (University of
Rochester, NY, USA) and cultivated in Dulbecco’s modified Eagle’s medium containing
10 % FCS and 2 mM glutamine in 75 cm2 culture flasks (Sarstedt AG and Co, Nuembrecht,
Germany) at 37 °C and 5 % CO2 (Gasiewicz et al. 1996). Cells were trypsinized with trypsinethylenediaminetetraacetic acid (EDTA) and transferred to 96-well plates (25,000 cells per
well, TPP). The cells were allowed to grow overnight. Subsequently, cells were exposed to a
series (0.0014 - 80 nM, 0.0042 - 25 pg/ml) of TCDD (Ehrenstorfer, Augsburg, Germany) or
sediment extracts (DMSO concentration 0.6 %) for 24 h. The maximum sediment
concentration in the assay was 0.12 mg SEQ per milliliter test medium. All samples were
tested in triplicates of five different concentrations according to the protocol by Olsman et al.
(2007). Every experiment was repeated independently at least three times. After incubation,
cells were washed with PBS and lysed in 50 µl lysis buffer (25 mmol glycoglycerine, 1 %
Triton X-100, 10 % glycerine). Luminescence measurements were performed using a
FLUOstar Optima plate reader (BMG Labtech, Offenburg, Germany) before and after
injection of 30 µl of substrate (Tris-(hydroxymethyl) aminomethane (20 mM, pH 7.8), MgCl2
(5 mM), EDTA (0.1 mM), dithiothreitol (33 nM), coenzyme A (270 µM), luciferin (470 µM)
and ATP (530 µM)). Data were calculated by subtracting the arithmetic mean for 5 s before
injection of the substrate from the mean for 13 s after substrate injection. The data were
analyzed using the GraphPad Prism 4 program. EC25 values were calculated from logtransformed data using a sigmoid curve fit with variable slopes. The results were expressed
relative to TCDD as bio-TEQ values as described below.
3.3.9 7-ethoxyresorufin-o-deethylase assay
Induction of EROD was measured in the CYP1A-expressing cell line RTL-W1, with slight
modifications to the method described by Behrens et al. (1998). RTL-W1 cells were kindly
provided by Drs. Niels C. Bols and Lucy Lee (University of Waterloo, Canada) and cultured
at 20 °C in 75 cm2 plastic culture flasks (TPP) in L15 medium supplemented with 8 % FCS,
1 % penicillin-streptomycin, and 1 % neomycin sulfate (Keiter et al. 2006). Before exposure
to the standards, cells were seeded in 96-well plates (TPP) and allowed to grow to 100 %
confluence for 72 h. Subsequently, the medium was removed and the cells were exposed for
72 h to the SPM extracts diluted in medium using eight dilutions with six replicates each. The
DMSO content in the wells was less than 0.1 %. Cytotoxicity of DMSO with RTL-W1 cells
can be determined with concentrations above 2-3 % in the well (analyzed in this study).
TCDD (Promochem, Wesel, Germany) was serially diluted to give a final concentration range
of 3.13 - 100 pM on two separate rows of each plate as a series of positive controls. Induction
3.3 Material and methods
60
was terminated by removing the growth medium and freezing at -70 °C to lyse the cells. The
deethylation of exogenous 7-ethoxyresorufin was initiated by adding the substance to each
well and incubating in the dark at room temperature for 10 min before addition of NADPH.
The plates were incubated for a further 10 min, and the reaction was stopped by adding
fluorescamine dissolved in acetonitrile. EROD activity was measured fluorometrically after
another 15 min using a GENios plate reader (Tecan, Crailsheim, Germany, excitation 544 nm,
emission 590 nm). Protein was determined fluorometrically using the fluorescamine method
(excitation 355 nm, emission 590 nm, Lorenzen and Kennedy 1993, Kennedy and Jones
1994) with the protocol detailed in Hollert et al. (2002). The concentration-response curves
for EROD induction in the RTL-W1 bioassay were computed by nonlinear regression
(GraphPad Prism 4) using the classic sigmoid curve or Boltzmann curve as model equations.
The luminescence-inducing potency of the samples was converted to bio-TEQs as described
below.
3.3.10 Bio-TEQ values
Bioassay-derived TCDD equivalents (bio-TEQs) were calculated by relating biological
activities (luminescence in the DR-CALUX assay and GPC.2D assay, fluorescence in the
EROD assay) caused by samples to the positive control 2,3,7,8-TCDD (cf. Keiter et al. 2008).
Bio-TEQs for concentration-response curves were calculated following the fixed effect level
quantification method, using the EC25 of the maximum response in the TCDD standard curves
as the fixed level (Brack et al. 2000, Engwall et al. 1996). The bio-TEQs given in this study
are means of n = 3 (EROD assay) and n ≥ 3 (GPC.2D assay) individual assays. Bio-TEQs of
the DR-CALUX assay were calculated using EC values of n = 1, as remaining extract
volumes were insufficient for more individual assays. Data from the bioassays were evaluated
with the GraphPad Prism 4.0 program.
Mean TCDD-EC25 values were determined as well as standard deviations (SD) between
individual EROD assays and used to calculate bio-TEQs (Eq. 1). In contrast, TCDD-EC25
values of each individual assay were used to calculate equivalent concentrations with the DRCALUX assay and the GPC.2D assay. Subsequent, bio-TEQs with concentrations in
picogram TCDD per gram of SPM-EQ will be given as picogram per gram.
Eq. 1 Bio-TEQ [pg TCDD /g SPM-EQ] = TCDD25[pg/ml] / EC25[g SEQ/ml]
3.3.11 Chem-TEQ values
Safe (1990) described the AhR-pathway as structure dependent, as the most potent congeners
were 2,3,7,8-substituted tetra- and penta-chlorinated PCDD/Fs as well as meta- and parasubstituted coplanar PCBs. Based on in vivo and in vitro data to each compound included,
relative toxic potencies (REP values) were assigned (Eadon et al. 1986, NATO/CCMS 1998).
REPs are given as values which are related to the toxic potency of 2,3,7,8-TCDD (REP = 1).
Aiming to explain the determined bio-TEQ levels, chemically derived TEQs (chem-TEQs)
were calculated by using the relative potency (REP) factors shown in tab. 1. Chem-TEQs are
3.4 Results
61
calculated by multiplying compound concentrations and REP values that were applied as
determined by Clemons et al. (1997), specific for RTL-W1 cells and each compound assessed
(Eq. 2). When cell-specific REPs were not available, World Health Organization
(WHO)-REPs were used as given by van den Berg et al. (1998). Subsequently, chem-TEQs
with concentrations in picogram TCDD per gram SPM-EQ will be given as picogram per
gram.
Eq. 2 Chem-TEQ [pg TCDD / g SPM-EQ] = Conc. [pg/g SPM-EQ] x REP
3.4 Results
3.4.1 Neutral Red retention assay
Concentration-response curves of native water samples, XAD-extracted water samples, and
XAD filter residue extracts of the water samples Neckar 0 and Neckar 5 are illustrated in
fig. 2. Viability of the cells is given in relation to the viability of cells that were not exposed
to samples. Significantly increased cytotoxic effects could not be observed for native Neckar
water samples, both in RTL-W1 and RTG 2 cells. In contrast, XAD-extracted water caused a
50 % diminished cell viability with RTL-W1 cells (Neckar 5) when exposed to water
concentrated 84.6 times (concentration factor: CF50 = 84.6) and a CF50 = 120.6 with RTG 2
cells (details not shown). Further on, extracted water filter residues caused CF50 = 50.3
(Neckar 0) and CF50 = 48.3 (Neckar 5) using RTL-W1 cells. Exposure of Neckar 5 extracts to
RTG 2 cells caused CF50 = 65.8. No effect was observed with extracts of water sample Neckar
6. Neckar 0 and 5 indicated higher toxicities by extracts of particle-bound contaminants and
filter residue extracts. Dissolved compounds in native and extracted water were shown to be
significantly less toxic.
125
Neckar 0
Cell viability [% of control]
Cell viability [% of control]
125
100
75
50
Filter CF50 = 50.3
XAD CF50 = n.d.
Water CF50 = n.d.
25
1
10
Concentration factor
100
Neckar 5
100
75
50
Filter CF50 = 48.3
XAD CF50 = 84.6
Water CF50 = n.d.
25
1
10
100
Concentration factor
Fig. 2 Cytotoxicity of native water samples (filled inverted triangles), XAD-extracted water (empty
circles) and the XAD filter residue extract (filled circles) of the samples Neckar 0 and Neckar 5,
illustrated as concentration-response curves with common logarithmic scale. A 50 % decrease of cell
viability is given as CF50 (CF = concentration factor) for native water, XAD-extracted water, and filter
residues. Native water samples were diluted 1:2 in the test, and, therefore, highest concentrations
3.4 Results
62
correspond to 50 % of native waters. The highest concentration of XAD extracts is corresponding to
100-fold concentrated water and water filter residue extract. n.d. − not detectable
Cytotoxic effects of river Neckar SPM collected during the flood event as well as
corresponding discharge is shown in fig. 3, using RTL-W1 and RTG-2 cells. SPM extracts
damaged RTL-W1 (average NR50 = 22 mg/ml) cells significantly (≥ twofold) more as
compared with RTG-2 cells (mean NR50 = 50 mg/ml). After the flood event, cytotoxicity is
decreased by a factor of 3 in both RTL-W1 (NR50 = 97 mg/ml) and RTG-2 cells
(NR50 = 144 mg/ml).
Comparing both cell lines used, RTL-W1 cells were shown to be more sensitive than RTG-2
cells by a factor 1.5 to 2. Cytotoxicity of the Rhine 1 extract was not significantly different in
RTL-W1 cells (mean NR50 = 50.6 mg/ml) and RTG-2 cells (NR50 = 74.1 mg/ml). In contrast,
cell toxicity significantly differed after exposure of RTL-W1 and RTG-2 cells to the extract of
Rhine 2 SPM to (NR50 = 40.6 and 97.8 mg/ml, respectively).
HQ 10
200
Discharge
2000
1600
160
HQ 2
120
1200
10
800
RTG-2
80
9
3
1 2 4 5 6
7
40
Discharge [m3/s]
NR50 [mg SEQ/ml test medium]
Increasing cytotoxicity
240
400
8
RTL-W1
1
100
1000
Flood lapse and sampling times
[h after 14.01.04, 12 a.m.]
Fig. 3 AhR-agonistic activity of SPM samples of the river Neckar derived from the EROD assay and
RTL-W1 cells in the context of the flood discharge.
3.4.2 AhR-mediated activity
An increase of EROD activity could not be observed with native water samples and XADextracted water samples from the Neckar. Filter residue extracts of the water sample Neckar 0
(CF25 = 10 ± 1.9) and Neckar 5 (CF25 = 3.7 ± 1.3) were shown to contain high quantities of
compounds interacting with the AhR-activity. No activity was observed with water sample
Neckar 6 (details not shown). AhR-mediated activity in RTL-W1 cells upon exposure to SPM
total extracts from the river Neckar is illustrated in fig. 4. Enzyme activities corresponding to
25 % of the maximum induction of the positive control (effective concentration, EC25) are
shown as normalized bio-TEQs and are given in picogram TCDD per gram sediment. SPM
3.4 Results
63
extracts caused a concentration-dependent increase of AhR-mediated activity following the
flow at the Neckar. Bio-TEQs indicated two peaks of AhR-mediated activity with
5,457 ± 2,213 pg/g (Neckar 2) and the maximum flow 8,341 ± 2,817 pg/g (Neckar 6),
respectively. At the end of the flood event, reaching mean flow levels, bio-TEQs decreased
significantly but still reached 3,313 ± 135 pg/g (mean value of Neckar SPM 9 and 10).
Finally, Neckar 11, sampled another 14 days later, displayed an EROD activity of 930 ± 119
pg/g. Highest and lowest measured enzyme inductions differed by a factor of 6.7, indicating a
flood-dependent increase of AhR-mediated activity. The SPM bio-TEQ of the Rhine SPM
sampled over a longer period (Rhine 1 3,693 ± 520 pg/g) was elevated compared to the flood
sample (Rhine 2 2,331 ± 328 pg/g), indicating a decrease of activity through the investigated
flood event.
Discharge
1600
Bio-TEQ
10000
5
1400
6
1200
2
7
3
1000
9
8
800
4
600
1
400
1000
10
1
Discharge [m3/s]
Bio-TEQ [pg TCDD/g SPM-EQ]
1800
100
200
1000
Time of SPM sampling
[h after 14.01.2004, 12 a.m.]
Fig. 4 Acute cytotoxicity of flood SPM from the river Neckar in the neutral red retention assays
determined with the acute Neutral Red retention test after 48-h exposure, using the RTL-W1 and
RTG-2 cell lines. Time of sampling is shown in hours on a common logarithmic scale, beginning 14
January 2004, 12 a.m. (= 1 on x-scale). Furthermore, discharge is given and highlighted by dots at the
times of SPM sampling. HQ values, illustrated by dashed lines, are provided to permit a classification
of the recent flood event discharge. Cell viability is given as percent of controls. Numbers indicate
Neckar SPM sample encodings (cf. Table 1).
3.4.3 DR-CALUX and GPC.2D assay with SPM
Bio-TEQ values of the SPM extracts determined in the DR-CALUX and the GPC.2D assays
in comparison with results of the EROD assay are given in fig. 5. All SPM extracts caused a
concentration-dependent increase of AhR-mediated activity. The three cell lines displayed
differential sensitivity to the SPM extracts. Highest activities were detected using the EROD
assay. The AhR-mediated activity of SPM extracts in the H4L1.1c4 cells correlated well with
RTL-W1 cells, with the Neckar 6 sample (bio-TEQ = 3,833 pg/g) displaying the highest
activity. In contrast, for H4L1.1c4 cells, no significant increase could be observed for the
3.4 Results
64
Neckar 2 sample. The lowest induction rates for all samples were obtained in the GPC.2D
assay. Furthermore, the DR-CALUX and the GPC.2D assays were used to assess Rhine SPM.
Using H4L1.1c4 cells, the bio-TEQ of Rhine 1 equaled a concentration of 2,007 pg/g, while
Rhine 2 was significantly less toxic (908 pg/g). Rhine 1 and GPC.2D cells demonstrated a
bio-TEQ of 145 ± 73 pg/g. Rhine 2 showed a significantly decreased bio-TEQ of
29 ± 30 pg/g. Thus, all applied assays indicated higher AhR-agonist activities for Rhine 1.
RTL-W1
Bio-TEQ [pg TCDD / g SPM-EQ]
10000
1000
H4L1.1C4
10000
1000
10000
GPC.2D.LUC.2DLuc
1000
100
* *
1
2
3
4
5
6
7
8
9
10
Neckar SPM
Fig. 5 AhR-mediated activity of SPM extracts of the Neckar River as determined in the DR-CALUX®
(H4L1.1C4 cells, n = 1), GPC.2D (GPC.2D.Luc cells, n = 3-8) and 7-ethoxyresorufin-o-deethylase
(EROD, RTL-W1 cells, n=3) assays. Bio-TEQ values were calculated as the concentration resulting in
25 % of the maximum induction of 2,3,7,8-TCDD. Data are presented as means ± SD in the case of
the DR-CALUX® and EROD assays. * − activity not detectable
3.4.4 Multilayer and carbon on celite fractionation
Fractions from silica multilayer fractionation contain persistent organic and dioxin-like active
compounds being resistant to oxidization by sulfuric acid. Neckar 1 and Neckar 6 as well as
Rhine 1 and Rhine 2 were selected for multilayer open column chromatography. Primary
multilayer fractions (F1), containing persistent sulfuric acid oxidation-resistant compounds
(PCBs and PCDD/Fs), and secondary fractions (F2-1 containing PCBs, F2-2 containing
PCDD/Fs) were analyzed by means of biological and chemical methods. Data about
chemically analyzed and quantified persistent chemical compounds are provided in tab. 2.
Concentrations of PCDDs increased with the degree of chlorination. Octachlorodibenzo-pdioxin (OCDD) was concentrated the highest, reaching a maximum concentration of
5,013 pg/g (Rhine 1). Regarding PCDFs, concentrations were highest for octachlorodibenzo-
3.4 Results
65
p-furans (OCDFs) in sample Rhine 1 (999 pg/g). Among all PCBs detected, #77 had the
highest concentrations of all samples, 31,998 pg/g (Rhine 2).
Table 2 Concentrations of priority PCDD/Fs and PCBs given for selected samples from the rivers
Neckar and Rhine
Compounds
analyzed
PCDD
2,3,7,8-TCDD
1,2,3,7,8-PeCDD
1,2,3,4,7,8-HxCDD
1,2,3,6,7,8-HxCDD
1,2,3,7,8,9-HxCDD
1,2,3,4,6,7,8HpCDD
OCDD
Sum PCDD
PCDF
2,3,7,8-TCDF
1,2,3,7,8-PeCDF
2,3,4,7,8-PeCDF
1,2,3,4,7,8-HxCDF
1,2,3,6,7,8-HxCDF
2,3,4,6,7,8-HxCDF
1,2,3,7,8,9-HxCDF
1,2,3,4,6,7,8HpCDF
1,2,3,4,7,8,9HpCDF
OCDF
Sum PCDF
PCB
PCB 77
PCB 126
PCB 169
Sum PCB
Total sum
Specific
REP1
WHO
REP2
1.0
2.6
1.1
0.2
-
0.1
-
0.2
Neckar 1
[pg/g SPMEQ]
Neckar 6
[pg/g SPMEQ]
Rhine 2
[pg/g SPMEQ]
Rhine 1
[pg/g SPMEQ]
9
5
9
23
16
8
4
7
29
20
18
3
5
15
9
10
2
1
5
4
343
497
230
67
2719
3123
4597
5162
5013
5292
3278
3367
-
0.0001
0.2
0.2
1.9
1.1
-
0.1
0.1
0.1
48
28
34
55
21
15
5
58
24
33
57
18
18
5
109
94
59
61
27
12
7
23
16
14
25
8
4
3
0.01
164
181
104
34
0.01
17
14
9
3
432
819
511
918
999
1479
746
874
890
39
7
936
14327
121
5
14454
610
14
2
627
31998
0
0
31998
4878
20535
7398
36239
-
0.0001
0.0034
0.023
0.00016
-
In order to elucidate whether the analyzed compounds were responsible for the dioxin-like
activity of each fraction tested in the bioassay applied, measured concentrations were
expressed as chem-TEQs. Bio- and chem-TEQs of fractionated SPM extracts of the rivers
Neckar and Rhine are illustrated in fig. 6. Regarding all multilayer fractions, both TEQ values
showed good accordance.
Furthermore, the multilayer fractions of both Neckar samples (including PCDD/Fs and PCBs)
showed similar inductions of dioxin-like activity. However, alterations in compound
concentrations within these fractions were obvious using chemical analysis. Bio- and
3.5 Discussion
66
10000
10000
Neckar 1
Chem-TEQ
1000
1000
1
*
1
1000
100
100
CE
ML
PCB
PCDF
1000
10
PCDD/F
10
1
PCDF
Rhine 2
10000
Neckar 5
PCDD
10000
PCDD
10
CE
ML
PCB
PCDF
PCDF
PCDD
10
Multilayer
100
100
1
Rhine 1
Bio-TEq
Bio-TEQ
PCDD
TEQ values
[pg TCDD/g SPM-EQ]
TEQ values
[pg TCDD/g SPM-EQ]
chem-TEQ values of PCBs increased by a factor of 13 during the flood event in both Neckar
samples, whereas bio-TEQ values of PCDD/F remained constant (approximately 200 pg/g).
Similarly, Rhine samples showed an increase of PCB bio- and chem-TEQs in Rhine 1 with a
bio-TEQ increasing by not detectable amounts to 268 pg/g in Rhine 2. Total PCDD/F
bio-TEQs were similar for Rhine 1 (289 pg/g) and Rhine 2 (278 pg/g).
PCDD/F
Fig. 6 Bio-TEQ values determined using EC25 values from the EROD assay (black bars) and chemTEQs calculated by multiplying compound concentrations and REPs (gray bars) for SPM samples
from the Neckar and Rhine rivers are shown. Neckar samples were selected as indicating highest
dioxin-like activities at around the discharge peak of the flood. Comparison of TEQs for each fraction
allowed for an explanation of the determined bio-TEQ levels related to the contribution of the
chemically analyzed dioxin-like activity for the overall biological activities. Bio-TEQs of total extracts
allow grading of each fraction activity. Bio-TEQs are given as a mean of three independent
experiments ± SD on a common logarithmic scale. CE − crude extract, ML − multilayer fraction,
* − activity not detectable
3.5 Discussion
3.5.1 Cytotoxic effects of complex samples
Cytotoxic effects with RTL-W1 cells were not detected for native water samples from the
river Neckar, indicating that concentrations of water-borne toxic compounds were low during
the flood event. Cytotoxic effects were only observed at the peak of flow with XAD-extracted
water samples. The impact of extracted water-soluble compounds seemed negligible for
cytotoxicity. In contrast, filter residue extracts of the samples Neckar 0 and Neckar 5
produced toxic effects, indicating particle-bound contaminants to be more relevant than
3.5 Discussion
67
dissolved compounds. Since the particle-bound contaminants were more toxic than the water
samples, further analyses had been focused on SPM. RTL-W1 and RTG-2 cells revealed
constantly dose-dependent decreased cell viabilities with SPM sampled during flood
discharge. SPM extracts were also toxic when sampled at mean annual discharge directly after
the flood event but became less toxic thereafter. Hollert et al. (2003b, 2007a) showed that
SPM from the 1998 flood event at the Neckar (HQ = 1:20) caused NR50 values for RTG-2
cells between 20 and 60 mg/ml test medium. These results were comparable to the toxicity
caused by the 2004 flood SPM assessed in this study. Compared to a moderate Neckar flood
event (HQ = 1:1) with NR50 values for RTG-2 cells between 40 and 150 mg/ml, cytotoxicity
was increased. As cytotoxic effects could not be correlated with the discharge in all three
studies, toxic effects seem to increase in the beginning of flood events, but to remain on a
certain level with increasing discharge.
Cytotoxic effects in this study indicated a good correlation between the RTL-W1 and RTG-2
cell line (rPearson = 0.93, c. f. fig. 7). The NR50 values of all but three samples were within the
95% confidence interval. Furthermore, RTL-W1 cells were shown to be more sensitive to
SPM extracts, possibly indicating effects of metabolic processing. Cytotoxicity by Rhine 1
was similar in both cell lines, whereas Rhine 2 proved to be 2.6-times more toxic to RTL-W1
cells.
3.5.2 Ah receptor agonist activity of water samples
As for the neutral red assay, an increase of activity could not be observed with native and
XAD-extracted water samples of the Neckar flood event, indicating a minor relevance of
water-dissolved compounds for dioxin-like activity. This finding is in accordance with earlier
studies investigating basic toxicity (Hollert et al. 2000, 2003a). In contrast to the solved
Ah-agonists, EROD activities of fig. 6 Bio-TEQ values determined using EC25 values from
the EROD assay (black bars) and chem-TEQs calculated by multiplying compound
concentrations and REPs (gray bars) for SPM samples from the Neckar and Rhine rivers are
shown. Neckar samples were selected as indicating highest dioxin-like activities at around the
discharge peak of the flood.
Comparison of TEQs for each fraction allowed for an explanation of the XAD filter residue
extracts and particle-bound compounds were significantly increased. An increased AhRmediated activity was observed for the extracted and SPM sample Neckar 0, whereas the
SPM sample Neckar 5 was even threefold more toxic than minimally concentrated water.
Thus, EROD activity was increased, indicating that flood events resulted in an elevated
exposure of aquatic organisms to dioxin-like compounds. Uptake of toxic compounds into
organisms can be intensified by exposure via external surfaces and by ingestion in the
digestive tract. Uptake conditions of contaminants via gills have been addressed in other
studies (Erickson et al. 2006a, b), also regarding influences of natural organic matter (Klinck
et al. 2005). So far, little research has been performed to bridge the gap between in vitro tests
in the laboratory and effects on organisms and populations in the field involving the context
of flood events. However, there is an urgent need to investigate the relevance of in vitro
results for the situation in the field and to answer the question if a short flood event period
3.5 Discussion
68
[mg SPM-EQ / ml test medium]
NR
50 of RTG 2
with an elevated hazard by particle-bound pollutants may be able to cause adverse effects in
fish in the field.
150
120
90
60
rPearson = 0,93
30
R2= 0,87
30
60
90
120
NR of RTL-W1
50
[mg SPM-EQ / ml test medium]
Fig. 7 Regression and correlation analysis for cytotoxicity data obtained with the rainbow trout cell
lines RTG 2 and RTL-W1 in the Neutral Red cytotoxicity assay, using ten acetonic Neckar SPM
extracts, are shown with the 95 % confidence interval (dashed lines). Correlation coefficients were
calculated and are given as an r value (Pearson) with 95 % confidence interval and R2 values.
3.5.3 AhR-mediated activity of SPM
Compounds causing AhR-mediated activity and more specific dioxin-like activity were the
focus of this study since they are ubiquitous and highly concentrated in aquatic ecosystems
(Weber et al. 2008b). This end point was mainly measured by means of the EROD assay,
which is also accepted to verify AhR-mediated activity by the WHO In contrast to the neutral
red assay, AhR-mediated activity of SPM was strongly increased in correlation with the flood
flow and indicated a maximum activity during the peak of the flood event (bio-TEQ =
6,300 pg/g). In the present study, decreasing activities were observed upon average flow after
the flood (Neckar 8 and 9), remaining on a significantly elevated level of 930 pg/g
(Neckar 10) another 14 days later. However, it should be mentioned that elevated activities
and high loads of SPM could be documented for several days following the maximum flood
flow, indicating the high relevance of flood events for the mass transport of particle-bound
contaminants causing cytotoxic and dioxin-like activity within the investigated Neckar flood
event. Floods with intensified flow were shown to increase the toxicity associated with SPM
assessing different biological end points (Hilscherova et al. 2007, Hollert et al. 2007a, Keiter
et al. 2006, Rao et al. 1990). Whereas Neckar SPM induced an increase of activity, Rhine
samples indicated a twofold decrease of AhR-mediated activity by flood flow comparing the
time integrated SPM sample (including SPM of the flood event) and the SPM sample of the
flood event alone. Most likely, the EROD activity of the Rhine flood sample was decreased as
a consequence of dilution, alternatively, it might be suppressed by the presence of
antagonistic compounds as was assumed in other studies (Chen and Bunce 2004, Peters et al.
2006). Apart from this, Rhine 2 SPM mainly consisted of significantly larger grain sizes,
mostly sand, providing less adsorptive surfaces for contaminants. Thus, decreasing AhRmediated activity might also have been caused by SPM composition. Sand load could be
3.5 Discussion
69
traced back to sediment erosion. Erosive processes seemed to be realistic regarding findings
of Witt and Westrich (2003), who developed a method to determine erosion rates directly
from experiments conducted with undisturbed sediment cores of the Rhine in a laboratory
flume.
3.5.4 Modification of pollutant composition
PCB-TEQs of Neckar SPM increased 12-fold at the peak of flood flow, and Rhine PCB-TEQs
increased even 40-fold, indicating a significant influence of flow on the amount and
composition of the compound mixture. PCDD/F-TEQs of Neckar SPM were constant with
flood flow, whereas TEQs of Rhine PCDD/F decreased 2.5-fold. As discussed above,
compound mixtures might mainly be influenced by sediment erosion and remobilization of
contaminants. Keiter et al. (2008) assessed near-surface sediment samples at the river Danube,
showing that concentrations of persistent PCBs #77, #126, and #169 varied with a minimum
of 23 pg/g and a maximum of 243 pg/g of dry sediment. All Neckar SPM indicated that
concentrations of PCB #77, #126, and #169 increased by 4 to 52 pg/g at the peak of the
discharge and were between 2 and 109 pg/g in samples collected from the Rhine. Thus,
concentrations in Neckar and Rhine flood SPM were in a similar range to those in Danube
near-surface sediments. Furthermore, Keiter et al. (2008) measured mean PCDD/F
concentrations of 1,620 pg/g and maximum concentrations of 5,419 pg/g.
Both in the Neckar 6 and the Rhine 1 sample, PCDD/F concentrations were four to fivefold
higher (6,080 and 6,771 pg/g, respectively) than mean concentrations at the peak of the flood
than in the Danube sediments. While Danube sediment contamination levels highlight the
high possible burden of sediments with organic chemicals, Neckar SPM indicate increased
loads in consequence of the flood event (surface runoff, sediment erosion). In contrast, Rhine
SPM showed decreased AhR-mediated activities, which might be traced back to dilution
effects caused by the larger catchment area.
3.5.5 Comparison of AhR-mediated activity
Luminescence-based measurement of activities with the two mammalian cell lines GPC.2D
and H4L1.1c4 confirmed the high toxicities of flood-borne SPM samples as determined by
the EROD assay with RTL-W1 cells. Similarly to RTL-W1 cells, the rat hepatoma cell line
H4L1.1c4 revealed a significant peak of activity at maximum flood flow (Neckar 6). The
GPC.2D cell line indicated one peak of activity with SPM sampled before the peak of flood
water flow (Neckar 3), thus differing from both other biotest systems used. Regression
analysis with all bio-TEQ values of Neckar SPM indicated a good correlation between RTLW1 and H4L1.1C4 cell line (Fig. 8, Pearson rank correlation coefficient r = 0.81). In contrast,
there was no correlation between RTL-W1 and GPC.2D (r = 0.096) and H4L1.1c4 and
GPC.2D (r = 0.004). Keiter et al. (2008) assessed Danube river sediment extracts and showed
comparable regression coefficients using the same cell lines: RTL-W1 and H4L1.1C4
(r = 0.84), RTL-W1 and GPC.2D (r = 0.04), and H4L1.1C4 and GPC.2D (r = 0.21). Both
studies indicated good correlations between RTL-W1 and H4L1.1C4 cell lines when testing
3.5 Discussion
70
SPM and sediment extracts with respect to AhR-mediated activity, whereas no correlation
was found for results of GPC.2D cell lines and the two other cell lines. These differences
might be explained by species-specific induction mechanisms and
different exposure times as well as substrate inhibition in RTL-W1 cells at higher
concentrations (Keiter et al. 2008). Furthermore, the GPC.2D assay indicated a 100-fold less
induction than the other two bioassays. Reasons for this low induction might be differences in
the transfected responsive elements as well as cell-specific and, therefore, chemicaldependent trans-acting factors and receptors (Garrison et al. 1996, Zhou et al. 2003).
In the DR-CALUX assay, the AhR-mediated activity of SPM samples ranged between
100 and ≥ 1,000 pg/g. Each year, authorities of The Netherlands have to decide whether
25 × 106 m3 sediments of the port of Rotterdam get permission for disposal at sea. It has been
suggested that 100 pg/kg dry weight, as measured in the DR-CALUX assay subsequent
removal of acid-degradable non-persistent compounds, could serve as a threshold for
prohibiting the disposal (Stronkhorst et al. 2002). In the present study, the bio-TEQs for SPM
contributed by persistent compounds in the ML fractions are significantly increased compared
to these values and, thus, the contamination levels of the river sediments would be too high
for dumping at sea.
1200
GPC.2D
900
r Pearson = 0,096
600
300
0
-300
0
3000
6000
9000
RTL-W1
Fig. 8 Regression and correlation analysis for AhR-mediated activity, as determined using the EROD
assay (RTL-W1) cells and the DR-CALUX assay H4L1.1c4. Correlation coefficients were calculated
and are given as r values (Pearson) with a 95 % confidence interval and R2 values
3.5.6 Sources of the remobilized PCDD/PCDF
The high contribution of the two fully chlorinated OCDD and OCDF (90 %, 85 %, 84 %, and
45 %, respectively, of the total 2,3,7,8-substituted PCDD/Fs) indicate that processes and
products from the chlorine and organochlorine industry were responsible for a large extent of
the PCDD/F contamination. Since only the 2,3,7,8-substituted congeners were analyzed in
this study, a more detailed identification of sources, e.g., by principal component analysis,
was not possible. However, the dominating 1,2,3,4,7,8-HexaCDF in all of the samples
measured (Fig. 9a) also shows, for the lower chlorinated homologues, that elemental chlorine
containing processes were also the source for most of the toxic HexaCDFs. Specifically,
processes containing elemental chlorine favor the formation of the 1,2,3,4,7,8-HexaCDF via
3.5 Discussion
71
chlorination of dibenzofurans, whereas the formation of the 1,2,3,4,7,8-HexaCDF is not
favored in incineration processes or processes from the metal industry (Fig. 9b).
1,0
1,2,3,7,8,9-HexaCDF
2,3,4,6,7,8-HexaCDF
Distribution within
2,3,7,8-HexaCDF
60
30
0,8
1,2,3,4,7,8-HexaCDF
1,2,3,6,7,8-HexaCDF
1,2,3,7,8,9-HexaCDF
2,3,4,6,7,8-HexaCDF
0,6
0,4
0,2
0
0,0
ar 1 eckar 5 Rhine 1 Rhine 2
Neck
N
PC
P
PC 1
P
C
hl C 2
o
C ral NP
hl ka
o
VC ral li 1
M ka
In s li 2
ci lu
n d
In era ge
ci t o
ne r
ra 1
t
Si or 2
nt
er
El Sin 1
t
ec er
tri 2
c
Ar
c
2,3,7,8-substituted HexaCDF
[pg TCDD/g SPM-EQ]
90
Fig. 9 a Concentrations and pattern of 2,3,7,8-HexaCDFs in SPM samples from the rivers Neckar and
Rhine during the flood events of the present study. 9 b Pattern of 2,3,7,8-HexaCDFs congeners in
some organochlorines (pentachlorophenols (PCP), chloronitrophene (CNP)) and processes (chloroalkali process, vinyl chloride monomer (VCM) production) from the chlorine-organochlorine industry
in comparison to patterns from incineration and metal industry
This indicates, for most of the PCDD/Fs in the monitored flooding events of the measured
sectors of the Rhine and Neckar river, that the historical releases of the chlorine and
organochlorine industry are still likely to be responsible for the major portion of the PCDD/F
load (which, of course, also holds true for the PCB contamination). During the 1970s, the
Neckar River in southern Germany ranked among the most strongly contaminated rivers in
Germany, with high loads of both organic pollutants and heavy metals. These particle-bound
contaminants accumulate in river-bottom sediments and, thereby, decrease the bioavailability
of toxicants for a broad range of aquatic organisms. Sediments which act as contaminant sinks
are more or less immobile and cohesive under normal hydrological conditions. The release of
adsorbed compounds to the free water column is usually of minor relevance. Admittedly,
increasing discharges as related to flood events may remobilize sediments via in-stream
erosion and remobilize highly contaminated sediments.
Hollert et al. (2003a) worked on bottom-sediment cores elucidating ecotoxicological
implications associated with the risk of erosion of contaminated sediments. It could be shown
that samples below an erosional unconformity revealed a clear increase of chemical and
biological indices. Furthermore, flood events with an HQ 1:5 and higher were in principle
said to erode even older, well-consolidated, and highly contaminated sediments. Thus,
deposited contaminants can pose a serious threat in the future, with increasing incidences of
extreme flood events caused by climate change (Hollert et al. 2007a). Hence, a
comprehensive erosion risk assessment of contaminated sites is of crucial importance
(Gerbersdorf et al. 2007).
3.6 Conclusions
72
3.5.7 Relevance of persistent compounds analyzed
The major aim of the chemical analysis was to obtain an overview of concentrations of
persistent organic pollutants in SPM from the Neckar and Rhine River, which are commonly
expected to be most relevant in sediments and SPM. To this end, PCDD/Fs and PCBs were
investigated which are known to induce dioxin-like activity. Bio- and chem-TEQs of
multilayer PCB and PCDD/F fractions were compared with one another. In fig. 6, chemTEQs of the multilayer PCB and PCDD/F fractions in the EROD assay were shown to equal
bio-TEQ values very well, thus confirming the concept of possible additive effects in complex
compound mixtures. The persistent compounds analyzed could, thus, be made responsible for
the AhR-mediated activity caused by compounds in the multilayer fractions and PCB, as well
as by PCDD/F fractions and analysis comprised relevant persistent pollutants in the river
SPM assessed. The bio-TEQs of the multilayer fractions were low compared to crude extracts
of bio-TEQs of each SPM: 5.6 % (Neckar 1), 11.8 % (Neckar 2), 12.8 % (Rhine 1), and 7.4 %
(Rhine 2). Hence, less persistent AhR agonists have to be mostly effective. Other studies
showed comparable or even lower contributions of these compounds quantified analytically.
Keiter et al. (2008) determined bio-TEQs of PCBs and PCDD/Fs fractions clearly below 8 %
assessing sediment crude extracts. Nevertheless, high proportions of the effects by crude
extracts cannot be explained by persistent organic pollutants in the latter and the present
study.
Furthermore, Brack et al. (2005) worked on sediments of the river Neckar and used the
approach of effect-directed analysis (EDA) to prove that less persistent non-priority
compounds caused the majority of AhR-mediated toxicity, underlining the recent results.
Hence, in a subsequent study of Woelz et al. (2008a) that is currently underway, a combined
EDA approach was used to identify the portion of less persistent compounds to the overall
effect of crude extracts.
3.6 Conclusions
The combination of multilayer fractionation and biological and chemical analysis is a suitable
tool to assess AhR-mediated activity by persistent PCDD/Fs and PCBs in flood event SPM
samples. Flood events translocate considerable amounts of SPM and, thus, particle-bound
contaminants. Depending on differences in the catchment area and the intensity of floods, the
load of erodible old sediments, as well as biological effects, can be significantly increased, as
shown in assessing a river Neckar flood event with high resolution sampling. In contrast, a
total sample of a river Rhine flood indicated a decrease. Thus, the assessment of flood events
needs a highly elaborated sampling design and intensive sampling, regarding contaminant
translocation and remobilizing behavior in the context of surface runoff and discharge. The
use of total samples to evaluate complete flood events is inadequate and may result in
undervaluation of pollutant release, translocation, and impact on aquatic organisms.
Furthermore, PCDD/Fs and PCBs were shown to contribute only to a minor portion to the
AhR-mediated activity of the raw extract without an oxidative cleanup step. As a
consequence, there is a need for the identification of other compounds (classes) with AhR
3.7 Recommendations and perspectives
73
agonistic activities and, in our study, to investigate the contribution of these unknown
compounds to the overall AhR-mediated toxicity of the SPM extracts investigated in this
study, and, more generally, to complex environmental samples in aquatic systems (Brack et
al. 2005, 2007).
3.7 Recommendations and perspectives
As it cannot be excluded, an increase of flood events in the context of climate change has to
be regarded in future flood risk management, at least in some regions such as Central Europe.
Thus, it is crucial to determine the potential hazard of (re-)mobilized contaminants dislocated
via floods and posing a threat to organisms and man. Furthermore, the majority annual load of
particle-bound contaminants has been determined to be translocated within a few days of
flood events in many catchment areas. Thus, floods remobilizing older highly contaminated
sediments, may pose a risk at these rivers and lead to the fact that the good surface water
status, as demanded by the EU-Water Framework Directive until the year 2015, will not be
achieved (Hollert et al 2007b, Heise and Foerstner 2006, Netzband et al. 2007). Consequently,
sediment mobility and particle-bound contaminants adsorbed to sediments ought to be more
in the focus of recent monitoring (Babut et al. 2007, Chapman and Hollert 2006, Hilscherova
et al. 2007, Hollert et al. 2007a, Westrich and Foerstner 2005). Since other less persistent
compounds seem to be more relevant to explain AhR-mediated activities in flood SPM, the
focus should be on PAHs and more polar compounds. Thus, effect-directed analysis using
broader fractionation methods, integrating biological and chemical analysis methods, and
covering compounds from polar to non-polar will be applied for the identification of causative
pollutants in a subsequent study.
In order to determine the possible ecological relevance of (re-)mobilized contaminants during
flood events, more focus should be put on particle-bound contaminants, at least with respect
to the fate and effects of more lipophilic compounds such as those inducing AhR-mediated
activity. The use of organic extracts and in vitro assays allows one to evaluate the potential
worst-case scenario with contaminants being detached from adsorptive surfaces easily
available for cellular systems. Nevertheless, the link of in vitro to in vivo approaches has been
poorly subjected in the context of flood events so far. There is a need for studies closing this
gap and assessing the possible effect to aquatic organisms being exposed to contaminants
originating from surface runoff and sediment erosion. Recently, a feasibility study was started
that connects the assessment of sediment erosion and effects on fish using a circular flume
(Pathfinder Project FLOODSEARCH, Woelz et al. 2009).
Acknowledgements
The authors would like to express their thanks to Drs. Niels C. Bols and Lucy Lee (University
of Waterloo, Canada) for providing RTL-W1 cells and the biodetection systems for the
DR-CALUX cells. Furthermore, we want to thank, in particular, Reviewer 1 for valuable
comments that helped to improve the article.
3.8 References
74
3.8 References
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vitro methods of toxicology. CRC, Boca Raton, pp. 237-251
Babut M, Oen A, Hollert H, Apitz S, Heise S (2007): Prioritization at river basin scale, risk assessment
at site-specific scale: suggested approaches, Chapter 4. In: Heise S (Ed.) Sustainable
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79
Chapter 4
4 Effect-directed analysis of Ah receptor-mediated
activities caused by PAHs in suspended
particulate matter sampled in flood events
J. Wölz1,4, W. Brack2, C. Moehlenkamp3, E. Claus3, Th. Braunbeck4, H. Hollert1,4
1
RWTH Aachen University, Institute for Environmental Research (Biology V), Department of
Ecosystem Analysis, Worringerweg 1, 52074 Aachen, Germany
2
UFZ Helmholtz Centre for Environmental Research, Permoserstraße 15, D-04318 Leipzig, Germany
3
German Federal Institute for Hydrology, Am Mainzer Tor 1, D-56068 Koblenz, Germany,
4
Department of Zoology, Aquatic Toxicology and Ecology Section, University of Heidelberg, Im
Neuenheimer Feld 230, D-69120 Heidelberg, Germany
To be submitted to The Science of the total Environment
80
4.1 Abstract / 4.2 Introduction
81
4.1 Abstract
Suspended particulate matter (SPM) sampled during a flood event in the year 2004 at the
rivers Neckar and Rhine (Southwest Germany) was assessed for aryl hydrocarbon receptor
(AhR)-mediated activities using EROD-induction in the rainbow trout liver cell line RTL-W1.
All EROD-inductions were normalized to the positive control TCDD and given as bio-TEQ
values. Since all samples indicated elevated AhR-mediated toxicities, an effect-directed
analysis (EDA) was applied to identify compounds causing the effects. In three primary
fractions (F1 to F3) non-polar aliphatics, non-polar aromatic compounds and more polar
compounds were separated. Fraction F2, co-eluting with non-polar polyaromatic compounds
(PACs) including polycyclic aromatic hydrocarbons (PAHs) gave highest AhR-agonistic
effects and, thus, were sub-fractionated into seven secondary fractions (F2-1 to F2-7).
Fractions F2-1, co-eluting with PCBs and PCDD/Fs, did not cause AhR-agonist activities.
F2-2 to F2-4 containing PACs of less than 16 aromatic C-atoms produced minor activities.
Highest inductions were detected with fractions F2-5 to F2-7, containing compounds of more
than 16 aromatic C-atoms (bio-TEQs up to approximately 4,500 pg/g).
Concentrations and relative potencies (REPs) of priority EPA-PAHs allowed the calculation
of chemical toxicity equivalent concentrations (chem-TEQ values). Based on the chem-TEQs,
EPA-PAHs explained between 16 and 58 % of crude extract bio-TEQs from both rivers.
Whereas fractions F2-1 to F2-4 indicated no biological activities, EPA-PAHs in fraction F2-5
to F2-7 accounted for 2 to > 100 % of AhR-related activities.
4.2 Introduction
There is a general agreement that sediment- and particle-bound substances play an important
role for the water quality in aquatic systems. Sediments can act as sinks for pollutants in the
river system which are, thus, abstracted from the water column and thereby become less or
non-available for aquatic organisms. In fact, worldwide, rivers are loaded with a multitude of
particle-bound biologically active and toxic contaminants as a legacy of the industrial past.
Further, sediments can turn into sources of contaminants, when deposited materials are eroded
and remobilized as a consequence of, e.g., increased discharge and sheer-stress during flood
events (Gerbersdorf et al. 2007b). Therefore, remobilization of sediments was also in the
focus of some studies in the context of flood events (Hilscherova et al. 2007, Zonta et al.
2005). However, with climate change, the impact by floods received further attention, as an
increase of extreme weather conditions, e.g. intense rain, and subsequent extreme floods are
predicted in certain regions (Ikeda et al. 2005, Kleinen & Petschel-Held 2007). Further,
eroded sediments can be transported and dislocated downstream and on inundated areas in
floods as suspended particulate matter (SPM). Hazards of remobilized and contaminated
sediments can be evaluated sampling suspended matter in floods (Stachel et al. 2004).
In this study, suspended particulate matter (SPM) sampled in parallel flood events – caused
by storm precipitation – at the rivers Neckar and Rhine (Germany) were assessed using the
concept of effect-directed analysis (EDA). EDA has been shown to be a powerful tool for
4.3 Materials and methods
82
toxicant identification in complex environmental samples (Brack et al. 2005, Sundberg et al.
2005). In a previous study applying the concept of EDA, strongly persistent polychlorinated
biphenyls (PCBs) as well as polychlorinated dioxins and furans (PCDD/Fs) were shown to be
only responsible for a minor portion (< 10 %) of the total AhR-mediated activities (Wölz et
al. 2008). Thus, the present study aimed at identifying less persistent substances causing
activating Ah receptor-related activity.
Polycyclic aromatic hydrocarbons PAHs can act as mutagens or even carcinogens and are,
hence, of general interest (Chen & White 2004). The present study put major emphasis on the
16 EPA-PAHs (US-EPA, Laboratory Test Protocol Number 610). However, hundreds of
other PAHs are present in the environment, which may affect aquatic organisms in various
ways, although they have not been registered as priority contaminants (Biselli et al. 2005,
Neff et al. 2005).
Thus, the major aims of this study were (1) to identify fractions causing AhR-related effects
in suspended materials collected during a flood event and (2) to determine the contribution of
EPA-PAHs to the overall AhR-mediated activity.
4.3 Material and Methods
4.3.1 Chemicals used
Unless stated otherwise, all chemicals used were provided by Sigma-Aldrich (Deisenhofen,
Germany).
4.3.2 Sampling and preparation
Suspended particulate matters (SPM) were collected with a SPM-trap installed on a floating
bridge of a power station at Heidelberg, Germany and transferred in light-safe glass bottles at
4 °C to the laboratory (methodology cf. Hollert et al. 2000). Rhine samples were taken at the
water quality monitoring station at Worms, Germany. Along the pillars of the bridge, four
pumps were installed to provide a continuous water supply from four lanes across the river. In
the laboratory, SPM were frozen at -20 °C, samples were freeze-dried as early as possible
(beta 1 - 8 K; Christ, Osterode, Germany) and stored at 4 °C in darkness until analysis.
Aliquots of 20 g of freeze-dried SPM samples were extracted with 200 ml dichloromethane
using Soxhlet extraction thimbles (Schleicher & Schuell, Dassel, Germany), and for 14 h at 810 cycles per hour. The solvent was reduced in volume, and residues were evaporated close to
dryness under a gentle N2-stream. Residues were re-dissolved in 1 ml n-hexane and stored at 20°C until fractionation. As process controls, empty extraction thimbles were subjected to
extraction and processed in two parallel experiments.
4.3.3
Fractionation
SPM used for fractionation were selected on the basis of maximum AhR-related activities
found in a previous study (Wölz et al. 2008): Neckar A and B were sampled on January 14,
2004 (2 pm) and January 15, 2004 (8 am): Rhine A is a sample collected over an extended
4.3 Materials and methods
83
period from November 2003 to February 2004, including the January flood event, whereas
Rhine B was sampled exclusively over the flood period from January 15 to 19, 2004.
Fractionation was performed using a recently developed, two-step procedure with some
modifications (Brack et al. 2005). In order to separate and characterize AhR-agonists in crude
extracts a two step procedure was applied starting with open column chromatography on
alumina followed by normal-phase high performance liquid chromatography (HPLC) on a
nitro-phenyl phase (Fig. 1).
Suspended particulate matter
Fractionation acc. to polarity
F1: non polar
aliphatic compounds
(Al2O3) gravity column
F2: non-polar
aromatic compounds
Fractionation acc. to molecular size
F3: more polar
compounds
(NO2) column; NP-HPLC
Fig. 1 Fractionation procedure for AhR-agonists in suspended particulate matter including a first
separation step according to polarity providing three primary fractions and a second step for non-polar
aromatic compounds, according to increasing numbers of aromatic rings (i.e., molecular size; seven
sub-fractions: F2-1 to F2-7). In fraction F2-1, PCBs and PCDD/Fs may have been co-eluted.
SPM − Suspended particulate matter, NP-HPLC − normal-phase high-performance-liquidchromatography, F − fraction
4.3.4 Primary Fractionation
Crude extracts were fractionated in open glass columns with a diameter of 3 cm using 90 g of
alumina (activity 1, ICN, Biomedicals GmbH, Eschwege, Germany) deactivated with 4.5 %
of water per column as a stationary phase (1999). Compounds were eluted with solvents of
increasing polarity: fractions F1 containing non-polar aliphatic compounds were eluted using
75 ml of n-hexane; fractions F2 characterized by non-polar polycyclic aromatic compounds
were eluted with 250 ml n-hexane/dichloromethane (90/10, v/v); and fractions F3 containing
more polar compounds were eluted with 250 ml of dichloromethane. All primary fractions
were evaporated to dryness and re-dissolved in 1 ml dimethylsulfoxid (DMSO),
corresponding to a final concentration of 20 g sediment equivalent/ml DMSO.
4.3 Materials and methods
84
4.3.5 Secondary fractionation
All fractions F2 were separated into seven sub-fractions (F2-1 to F2-7) according to
increasing numbers of aromatic C-rings using normal-phase high performance liquid
chromatography (NP-HPLC) on a stainless steel column (21 x 250 mm) packed with nitro
phenyl propyl silica (5 µm Nucleosil 100-5 NO2, Macherey and Nagel, Düren, Germany) with
a pore diameter of 0.1 nm. An isocratic solvent-mixture of n-hexane and dichloromethane
(95/5, v/v) was used as mobile phase at a temperature of 10 °C and a flow rate of 10 ml/min.
According to Brack et al. (Brack et al. 2003a), secondary fractions are typically characterized
by the following model compounds. F2.1: diaromatic compounds like polychlorinated
biphenyls, dibenzo-p-dioxins, dibenzofurans and naphthalenes as well as the non-chlorinated
parent compounds; F2.2: PAHs with MW 152-166, acenaphthylene and fluorene; F2.3: PAHs
with MW 178, e.g. anthracene and phenanthrene; F2.4: PAHs with MW 202, e.g. pyrene and
fluoranthene; F2.5: PAHs with MW 226-228, e.g. benzo[ghi]fluoranthene and chrysene; F2.6:
PAHs with MW 252, e.g. benzo[a]pyrene and perylene and F2.7: PAHs with MW > 276, e.g.
anthanthrene and benzo[ghi]perylene. The elution was monitored with a diode array detector
at 250 nm. Fractions were stored at 4°C in darkness until required.
4.3.6 PAH analysis
Gas chromatographic separation of the compounds of all fractions was performed on an
Agilent 6890 gas chromatograph (Agilent Technologies, Waldbronn, Germany), equipped
with a 30 m x 0.25 µm film HP-5MS fused silica capillary column (Agilent Technologies).
Helium was used as carrier gas at a constant flow of 1 ml/min. 1 µl of each fraction was
injected in split/splitless mode at 50 °C for 2 min and ramped to 130 °C at 20 °C /min and
then to 320 °C at 4 °C/min. The injector and transfer line temperature were 220 °C and
310 °C, respectively. For quantification of PAHs, fractions were diluted at a factor of 15 and
injected subsequently. Identification and quantification of compounds were carried out on a
mass spectrometer (model 5973N, Agilent Technologies). All fractions were measured in
scan and sim mode. Mass spectal data were collected in the mass range of 50 to 600 amu at
full scan mode and a scan rate of approx. 2 scan/s. Electron impact (EI) at 70 eV were
performed. The 16 EPA-PAHs were identified and quantified by using 6 standard solutions of
different concentrations in the range grom 5 to 1,000 pg/µl (Dr. Ehrenstorfer GmbH,
Augsburg, Germany) by single ion monitoring. The detection limit was 5 pg/µl. Some other
standards of PAHs (Dr. Ehrenstorfer) were measured only for identification of these
compounds. For further identification of unknown compounds by low resolution the spectra
libraries NIST/EPA/NIH Mass Spectral Library (NIST 98) and Wiley/NBS Registry of Mass
Spectral Data, 6th Ed. were used.
4.3.7 EROD induction assay
Method according to chapter 3.3.10.
4.4 Results
85
4.3.8 Bio-TEQ values
Computation of bio-TEQs according to chapter 3.3.11.
4.3.9 Chem-TEQ values
Computation of chem-TEQs according to chapter 3.3.12.
4.4 Results
4.4.1 AhR-agonist activities in primary fractions
In fig. 2, AhR-agonist activities caused by primary fractions are compared to crude SPM
extract inductions. Primary fractions 1 (F1) of each sample, containing non-polar aliphatic
compounds, were shown to contain no AhR-inducing substances. Highest activities were
associated with fractions F2, containing polycyclic aromatic compounds followed by F3,
containing more polar compounds. Bio-TEQs of crude extracts and primary fractions both
indicated a significant increase of activity in sample Neckar B, which was sampled at the peak
of discharge during the flood in January 2004. In the Rhine samples, bio-TEQs reflected a
lower AhR-agonist potency in Rhine B, if compared to the long-term sample Rhine A.
Highest AhR-agonist activities were determined for Neckar B crude extract
(bio-TEQ = 6,620 pg/g).
Total bio-TEQs for the crude extracts and the sum of combined primary fractions of each
SPM were significantly different. At the Neckar River, primary fractions outranged total
activities in the Soxhlet extract about two-times of Neckar A and 19 % of Neckar B. In
contrast, bio-TEQs of crude extracts and added primary fractions were almost congruently in
Rhine A, but 44 % decreased for primary fractions in Rhine B.
0
0
0
0
1
E
C
%
9
1
+
0
0
0
8
F2
F3
0
0
0
6
%
6
-
%
3
8
+
0
0
0
4
]
g
/
g
p
[
Q
E
T
o
i
B
%
4
4
-
0
0
0
2
B
e
n
i
h
R
A
e
n
i
h
R
B
r
a
k
c
e
N
A
r
a
k
c
e
N
0
Fig. 2 AhR-mediated activities of crude extracts and each primary fractions (F1 to F3) based on
EROD inductions in RTL-W1 cells are given as bio-TEQs. F1, containing non polar aliphatic
substances did not cause any AhR-agonist activity. Numbers in percent give the share of added F1 to
F3 to the crude extracts induction. Bio TEQs are given as means of n = 3 independent experiments. CE
− crude extracts
4.4 Results
86
4.4.2 Distribution of activities among secondary PAH fractions
NP-HPLC was used to fractionate F2 fractions according to their number of C-atoms in
aromatic rings and, further, to identify effective secondary fractions (Fig. 3). Fractions F2-1
and F2-2 of Neckar A caused no detectable Ah receptor-agonist activity. Bio-TEQs were low
in fractions F2-2 to F2-4 (bio-TEQ < 143 pg/g). Highest AhR-mediated activities were
determined in F2-5 to F2-7 and a maximum bio-TEQ of 4,600 pg/g in F2-5 of Neckar B.
However, except for F2-5 of Neckar B, fractions F2-6 gave the highest inducing potential
among secondary fractions.
4600
Neckar A
Neckar B
Rhine A
Rhine B
Bio-TEQ [pg/g]
4400
1000
500
0
** ** *
F 2-1
F 2-2
F 2-3
F 2-4
F 2-5
F 2-6
F 2-7
Fig. 3 Bio-TEQs of secondary fractions F2-1 to F2-7, containing non polar PAHs, determined with
SPM samples from the rivers Neckar and Rhine. * − no AhR-mediated activities detected
4.4.3 Quantification of EPA-PAHs
Chemical analysis was applied to determine concentrations of the 16 EPA-PAHs that are
given with relative toxic potencies (REPs) in tab. 1. Chemical analysis indicated a
predominance of PAHs with molecular weights of 202 and more (≥ F 2-4). These compounds
also showed high REP values and, accordingly, caused maximum AhR-agonistic activities in
the EROD assay. Total PAH concentrations were highest in Neckar B (4,920 µg/kg) and
lowest in Rhine B (1,020 µg/kg). Concentrations were increased following the maximum
discharge in Neckar B, with the exception of indeno[1,2,3-cd]pyrene, dibenzo[a,h]anthracene
and benzo[ghi]perylene. In contrast, Rhine samples indicated a decrease of concentrations for
each compound after the peak discharge.
4.4 Results
87
Tab.1 EPA-PAHs are shown in the order of elution in the NP-HPLC fractionation and according
fractions (Brack 1999), as well as substance concentrations. REP values are given as specifically
determined for used RTL-W1 cells by Bols et al. (1999).
Compound concentration [µg/kg]
Neckar A Neckar B Rhine A Rhine B
8.6
15.2
7.5
1.6
0.2
0.3
0.0
0.0
11.9
12.8
6.8
2.5
0,4
25,1
10,4
5,4
244.5
290.5
141.0
60.6
37.0
41.2
19.2
15.1
680.0
844.0
404.7
226.5
504.9
547.0
293.4
163.7
EPA-PAH compounds
REP values Fraction
naphthalene
acenapthylene
acenaphthene
fluorene
phenanthrene
anthracene
fluoranthene
pyrene
n.i.
n.i.
n.i.
n.i.
n.i.
n.i.
n.i.
n.i.
benzo[a]anthracene
0.043 x 10-3
F2-5
308.1
518.5
162.0
84.4
0.047 x 10
-3
F2-5
390.4
636.1
211.4
101.0
0.193 x 10
-3
F2-6
404.0
507.6
146.3
87.9
benzo[k]+[j]fluoranthene 1.039 x 10
-3
F2-6
356.6
436.7
161.3
60,4
benzo[a]pyrene
0.302 x 10
-3
F2-6
449.1
610.6
163.0
91.9
indeno[1,2,3-cd]pyrene
0.278 x 10
-3
F2-7
291.2
240.0
96.1
55.8
dibenzo[a,h]anthracene
-3
F2-7
51.4
27.7
17.5
13.3
F2-7
262.0
170.7
83.9
51.9
chrysene + triphenylene
benzo[b]fluoranthene
benzo[ghi]perylene
Sum of EPA-PAHs
0.35 x 10
F2-1
F2-1
F2-1
F2-2
F2-3
F2-3
F2-4
F2-4
1.039 x 10
-3
4000
4920
1920
1020
F2-1 to F2-7 − secondary fractions no. 1 to no. 7, n. i. − not inducing
4.4.4 Contributions of EPA-PAHs to determined AhR-agonist activity
Finally, contributions of chemically analyzed EPA-PAHs to the Ah receptor-agonist activities
in crude extracts and secondary fractions were determined comparing biologically and
chemically derived TEQs (Fig. 4). Whereas EPA-PAHs analyzed in Neckar A, sampled
before the peak of discharge, caused 58 % of the crude extract activity, EPA-PAHs in
Neckar B, sampled shortly after the peak of discharge, were less dominant and contributed
only 16 % to the crude extracts bio-TEQ. River Rhine SPM extracts gave EPA-PAH
concentrations that were equal to Neckar B. In the flood sample Rhine B, priority EPA-PAHs
were twofold increased compared to Rhine A and equaled half the contribution of Neckar A.
Since EPA-PAHs eluting in fraction F2-1 to F2-4 are considered as non-inducing AhR-related
activities, chem-TEQs and contributions of priority PAHs could not be calculated. EPA-PAHs
did induce AhR-associated effects- in each of the fractions F2-5 to F2-7 and contributed at
rates between 2 and 96 % to the respective fraction bio-TEQs. F2-7 of Neckar A gave more
than 100 % activity of EPA-PAHs to the biological activity.
4.5 Discussion
104
Neckar A
58 %
2%
58 %
103
7%
103
13 %
22 %
137 %
102
102
*
101
Bio-/Chem-TEQ [pg/g]
Rhine A
16 %
Bio-TEQ
Chem-TEQ
*
*
*
Neckar B
16 %
101
*
104
14 %
*
*
*
Rhine B
31 %
103
27 %
103
n.d.
Bio-/Chem-TEQ [pg/g]
104
104
88
4%
46 %
92 %
96 %
102
102
*
101
CE
*
* *
F2.1 F2.2 F2.3 F2.4 F2.5 F2.6 F2.7
Increasing No. of aromatic C-atoms
*
101
CE
* *
*
F2.1 F2.2 F2.3 F2.4 F2.5 F2.6 F2.7
Increasing No. of aromatic C-atoms
Fig. 4 AhR-mediated activities of crude extracts (CE) and secondary fractions (F2-1 to F2-7) induced
by Neckar and Rhine rivers SPM are given as bio-TEQs determined in the EROD assay (black bars).
EPA-PAH concentrations are given as chem-TEQs (grey bars). Contributions of EPA-PAHs of each
fraction are given in percent. n.d. − no toxic effect/bio-TEQ determined, * − EPA-PAHs in this
fraction are not EROD inducing with RTL-W1 cells
4.5 Discussion
4.5.1 Active fractions
Effect-directed analysis and a two-step fractionation allowed to determine and to compare
shares of eluted compounds in fractions of each SPM sample. Non-polar aliphatic compounds
in primary fractions F1 did not cause AhR-mediated activities in any of the samples, which
corroborates conclusions drawn from other studies on Ah receptor-mediated activities (Brack
et al. 2002b, Engwall et al. 1997). Fractions F2 showed the highest inducing potentials,
followed by significantly increased activities in fractions F3. Thus, non-polar and often
assumed PAHs (F2) and more polar compounds (F3) could be identified as major inducers of
AhR-mediated activities in river SPM. These findings are in agreement with other studies on
sediments, which next to non-polar polychlorinated compounds also indicated an induction
potency of more polar compounds in the EROD assay (Keiter et al. 2008, Kleman et al.
1992).
However, site-specific differences were detected, and primary fractions of SPM from each
river showed characteristic ratios of total and summed up bio-TEQs. In detail, activities
4.5 Discussion
89
caused by crude extracts of both Neckar samples were significantly lower than summed up
primary fractions F1 to F3, in particular in Neckar A. In contrast, only Rhine B indicated a
clearly decreased induction comparing bio-TEQs of the crude extract and added F1 to F3.
EROD inductions with more than 100 % in fractions compared to crude extracts may be
caused by retention of humic substances in the fractionation procedure. Thus, previously
antagonistically acting humic substances may be separated and agonists can display their
complete activity, respectively.
Since fractions F2 indicated the highest activities among all SPM assessed, secondary
fractionation was focused on corresponding non-polar aromatic fractions and highest
activities were determined in fractions with more than 16 aromatic C-atoms (F2-5 to F2-7). In
SPM of both rivers, there was no EROD inducing potency in F2-1fractions. Thereby, a higher
importance of less persistent compounds was obvious. Minor AhR-activities were determined
in fractions with PAHs of lower molecular weights. These are more likely degraded, due to
preferred physico/chemical- and bio-degradation (Cerniglia 1992). In contrast, higher
molecular compounds are more persistent to biochemical processes. These compounds
provide dense pi-electron clouds inhibiting nucleophilic substitutions and, accordingly, tend
to accumulate in environmental compartments (Johnsen et al. 2005). Fractions containing
higher molecular PAH compounds were also shown to induce the highest AhR-mediated
activities in earlier studies on AhR-agonistic activity (Boxall & Maltby 1995, Maltby et al.
1995a). Furthermore, they cause various toxic effects in aquatic organisms (Johnsen et al.
2005, Villeneuve et al. 1997) and, therefore, are of elevated interest in ecotoxicology.
4.5.2 Evaluation of prioritized compounds
Chemical analysis was applied to identify compounds causing AhR-agonistic activity to
aquatic organisms. Usually, analysis is focused on few and a priori selected compounds
which are considered to be of high priority, e. g. the 16 EPA-PAHs. Hence, other compounds
and substance classes are regarded as less relevant and are not among the commonly
evaluated contaminants. In this study, most secondary fractions indicated elevated AhRagonist activities. Nevertheless, bio-TEQs determined could often not be explained by
chem-TEQs calculated from the concentrations of the analyzed 16 EPA-PAH. In fractions
with PAHs of low molecular weight, only non-priority PAHs caused the total biological
activity. Low molecular EPA-PAHs can even be excluded from an evaluation of Ah receptoragonistic compounds, since they are not active with respect to AhR-related processes (Bols et
al. 1999) and, thus, chem-TEQs cannot be determined. Compounds in fractions with higher
molecular weights were consistently more potent inducers and showed significantly increased
bio-TEQs. However, for several fractions, chem-TEQs of priority PAHs often explained only
minor portions of the effects, whereas non-priority compounds caused were higher inducers.
This finding emphasizes the potential contribution of non-priority pollutants to environmental
hazards. In fact, in numerous earlier studies, it turned out that so far unknown and usually not
analyzed non-priority PAHs were of higher relevance (Barron et al. 2004, Brack et al. 2005,
Maltby et al. 1995a). Thus, an exclusive focus on prioritized pollutants may result in
4.6 Conclusion
90
inadequate assessment of environmental samples, and analysis of a broader range of
compounds needs to be considered.
Though EPA-PAHs are not necessarily major effective compounds, many studies provide
data and allow a ranking with respect to the PAH-related burden of the SPM analyzed in the
present study. However, it has to be mentioned that very little data are available for suspended
particulate matter, so far. In fact, there is hardly any literature on EPA-PAH concentrations in
SPM during the course of a flood. Some studies worked on PAH contamination in sediments
with a focus on EPA-PAHs and determined concentrations between 0.008 x 10-3 and 8.7 x 103
µg/kg (El Nemr et al. 2007, Shen et al. 2008), while studies with more PAHs (including
EPA-PAHs) varied between 0.01 x 10-3 to 25 x 10-3 µg/kg (Gaspare et al. 2009, Grundl et al.
2003). There are only few internationally published studies on EPA PAHs in suspended
particles available. One of the few studies worked on river near Beijing and reported of
1.33 x 10-3 to 28 x 10-3 µg/kg (Shen et al. 2008).
For an attempt of SPM classification the ATV scheme as recommended by Ahlf et al.
(2002b). The German ATV classification system was established for evaluation of dredged
sediments. EPA-PAH concentrations of this study (1.02 x 10-3 to 4.9 x 10-3 µg/kg) are ranked
in the ATV quality classes II - III (1 x 10-3 to 10 x 10-3 µg/kg) among 6 possible classes (with
class VI being worst). Following this classification system, the quality goal would be reached
with class II reflected by EPA-PAH concentrations of 1 x 10-3 to 4 x 10-3 µg/kg and, thus, the
detected concentrations in this study indicate slightly increased compound concentrations.
Ranking the findings of the present study, EPA-PAH concentrations of Neckar and Rhine
SPM can be rated elevated, although maximum concentrations found in the other studies were
not reached.
PAHs were listed as priority hazardous compounds in water by the Water Framework
Directive (WFD) of the European Union (Annex X), focusing on concentrations in water.
Meanwhile the WFD regulation implemented evaluation of particle-bound pollutants in
sediment, whereas EQS are non-existing, so far. Further, the International Commission for the
Protection of the Rhine (ICPR) listed namely benzo[b]fluoranthene, benzo[k]fluoranthene,
benzo[ghi]perylene, indeno[1,2,3-cd]pyrene and benzo[a]pyrene as substances of concern,
emphasizing the elevated significance of this compound class to reach environmental quality
aims. Hazard assessment of PAHs in the context of the WFD and other approaches with
environmental quality standards shall account for the processes and impact of flood events
and the fate of particle-bound substances in SPM.
4.6 Conclusions
Effect-directed analysis using two steps of fractionation has been proved to be a meaningful
tool to identify compound categories and active PAH fractions causing AhR-related effects in
suspended particular matter (SPM) collected during a flood event. In the assessment of AhRrelated processes in SPM samples, non-polar aliphatic compounds may be excluded, since
they do not possess any AhR-inducing potential. PAHs were determined as the class with
highest inducing potential and should, thus, be in the focus of AhR-related analyses on
4.6 Conclusion
91
contaminations of rivers. In most SPM fractions, however, priority EPA-PAHs contributed
only to a minor extent to the determined Ah receptor-mediated activities. Thus, so far nonprioritized PAH gave the higher inducing potential. Further, focus should be on high
molecular weight PAHs with more than 16 aromatic C-atoms, since they could be identified
as the compounds with the highest inducing potential. Since they are also more resistant to
degradation, such high molecular weight PAHs are of primary ecotoxicological concern.
Acknowledgements
The authors want to thank Mrs. Sperreuter for assistance with fractionation methods and also
express their thanks to Drs. Niels C. Bols and Lucy Lee (University of Waterloo, Canada) for
providing RTL-W1 cells.
4.7 References
92
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dibenzo-p-dioxins, polychlorinated dibenzofurans (PCDD/F) and dioxin-like PCB in suspended
particulate matter (SPM), sediment and fish. Water Sci Technol 50: 309-316
Sundberg H, Ishaq R, Akerman G, Tjarnlund U, Zebuhr Y, Linderoth M, Broman D, Balk L (2005):
A bio-effect directed fractionation study for toxicological and chemical characterization of
organic compounds in bottom sediment. Toxicol Sci 84: 63-72
Villeneuve DL, Crunkilton RL, DeVita WM (1997): Aryl hydrocarbon receptor-mediated toxic
potency of dissolved lipophilic organic contaminants collected from Lincoln Creek, Milwaukee,
Wisconsin, USA, to PLHC-1 (Poeciliopsis lucida) fish hepatoma cells. Environ Toxicol Chem
16: 977-984
Wölz J, Engwall M, Maletz S, Olsman H, Van Bavel B, Kammann U, Klempt M, Braunbeck T,
Hollert H (2008): Changes in toxicity and Ah receptor agonist activity of suspended particulate
matter during flood events at the rivers Neckar and Rhine − A mass balance approach using in
vitro methods and chemical analysis. Environ Sci Pollut Res 15: 536-553
Zonta R, Collavini F, Zaggia L, Zuliani A (2005): The effect of floods on the transport of suspended
sediments and contaminants: A case study from the estuary of the Dese River (Venice Lagoon,
Italy). Environ Int 31: 948-58
95
Section B
Impact of flood SPM on floodplains
and linked conflicts of interests
96
97
Chapter 5
5 Flood Retention and Drinking Water Supply −
Preventing Conflict of Interests
—
The RIMAX Joint Research Project HoT
M. Maier1, D. Kühlers1*, H.-J. Brauch2, M. Fleig2, D. Maier2, G. H. Jirka3, U. Mohrlok3,
E. Bethge3, H. H. Bernhart4, B. Lehmann4, G. Hillebrand4, J. Woelz5, H. Hollert5
1
Stadtwerke Karlsruhe GmbH, Daxlander Str. 72, 76127 Karlsruhe, Germany
2
DVGW-Technologiezentrum Wasser (TZW), Karlsruher Straße 84, 76139 Karlsruhe, Germany
3
Institute for Hydromechanics, Universität Karlsruhe (TH), Kaiserstraße 12, 76128 Karlsruhe,
Germany
4
Institute for Water and River Basin Management, Universität Karlsruhe (TH), Kaiserstraße 12,
76128 Karlsruhe, Germany
5
Department of Zoology, University of Heidelberg, Im Neuenheimer Feld 230, 69120 Heidelberg,
Germany
J Soils Sediments 6 (2) 113-114 (2006); DOI: 10.1065/jss2006.05.157
98
5.1 Background / 5.2 Aim of the project
99
5.1 Background
Diverse studies were able to identify a toxicological risk potential of suspended particles at
high water (Brack et al. 2002, Hollert et al. 2000, 2003, Oetken et al. 2005). However, there
are significant scientific deficiencies with regard to the influence of extreme floods, in
particular on the extraction of drinking water within areas which are inundated at high water.
Numerous studies indeed show high contaminant loads of surface water samples, in part, and
deposited and suspended sediments at differing water levels (Breitung 1999, Brauch et. al.
2001, Foerstner & Westrich 2005, Hollert et al. 2000, 2005, Kosmehl 2004, Maier et. al.
1997, LfU 1996a). However, there is only one pilot study so far with regard to the
toxicological risk potential of suspended particulate matter during flood events on the
production of drinking water. This study indicates that suspended particulate matter during
flood events cause an increase of the (eco-)toxicological hazard potential (in several biotests
and chemical analyses) of near surface soil samples in a riparian area which is frequently
inundated as compared to rarely inundated regions (Ulrich et al. 2002). Present knowledge
does not provide an answer to the question of whether either the contaminants which are
partially sorbed to the soil are eluted and negatively influence the groundwater and drinking
water, respectively, or are reduced or adsorbed along their passage through the unsaturated
zone.
Conflicts of interests are to be expected for virtually all major rivers in Germany. On the one
hand, retention areas have to be provided to minimize the risks associated with extreme flood
events. On the other hand, the groundwater and bank filtrate of many riparian areas are used
for the production of drinking water. Along the river Rhine, between the cities of Basel and
Duisburg alone, there are 15 sites where projected retention basins and water protection areas
overlap (IKSR, IAWR 1998). Water suppliers – who provide drinking water directly or
indirectly taken from the Rhine for more than 20 million people – are concerned about the
potentially increased risk of pollution of the groundwater resource by the establishment of a
retention area in the vicinity of their water extraction facilities: Organic pollutants could be a
danger to the extraction of drinking water by means of the retained water and the transported,
suspended particulate matter at flood events (LfU 1996b). Besides a general degradation of
the quality of the groundwater by nearby flood plains, the actual operation of water collection
facilities may be endangered over longer periods, especially during extreme flood events.
5.2 Aim of the joint research project
Within the project, which is supported by the German Federal Ministry of Education and
Research (BMBF), the dominant processes and mechanisms along the transport path from
flood wave via retention area and groundwater to the waterworks are investigated. On the one
hand, it is aspired to estimate whether contaminants and micro-organisms are able to migrate
from the river into the aquifer and a nearby waterworks under the specific conditions of
extreme floods. On the other hand, it is investigated whether substances and micro-organisms
5.3 Framework
100
may reach the nearby waterworks during a regular operation of the retention basin, which is
necessary for providing the retention area during extreme flood events.
In this project, the transport paths from the flood wave to the nearby waterworks are regarded
as a multi-barrier system. The first barrier is the transport of contaminants and microorganisms into the retention area. The second barrier is the unsaturated zone with its transport
and retardation mechanisms. The third barrier is the flow and transport behavior within the
saturated zone.
On the basis of the achieved knowledge, strategies are to be established in order to minimize
mutual impairments of flood retention and drinking water supply. These strategies will be
summarized to a guideline, which highlights and helps to prevent or minimize the
predominant majority of present and future conflicts between flood management and drinking
water supply by providing a corresponding package of measures.
5.3 Framework of investigation
The following studies are carried out within the joint research project:
• Analysis of chemical and toxicological characteristics of the water quality data and studies
on suspended and deposited sediments depending on the spatiotemporal development and
the discharge situation. The results of this work package will be published as a literature
review in JSS.
• Chemical and toxicological testing of sediment and water samples taken from the river
Rhine at diverse water levels and depending on the method used for collecting the
suspended particulate matter using chemical and bioanalytical methods (e.g., Gustavson et
al. 2004, Klee et al. 2004, Reifferscheid et al. 2005, Seiler et al. 2006)
• Identification of unknown contaminants with biological impact by means of effect-directed
analyses (cooperation with the Helmholtz Centre for Environmental Research Leipzig,
Dr. Werner Brack; cf. Brack et al. 2005).
• Estimation of particle retention in retention areas by hydrodynamic modeling and
approaches of the pelit research.
• Chemical and toxicological field investigations for different soil horizons at the project
study area Bellenkopf/Rappenwoert (projected retention area at the upper Rhine, river
kilometer 354.5 to 359.5) on the differing loads at frequently inundated and not inundated
regions. Laboratory studies to achieve parameters of several characteristic compounds on
their behavior during elution and microbial decomposition.
• Determination of hydraulic soil characteristics in retention areas and modeling of the
transport processes in the unsaturated zone.
• Chemical and toxicological investigations at several groundwater observation wells for
documenting the spatio-temporal change of the contaminant load.
• Numerical groundwater modeling for the determination of conditions on which
contaminants transported into the retention area may reach a nearby waterworks.
The projected retention area Bellenkopf/Rappenwoert (retention volume of 14 Mio. m3) near
Karlsruhe is used for the field studies. A major part of this area is situated within a drinking
5.4 Structure of the project
101
water protection area, which has been created for a projected waterworks. At the model site,
there are regions which are already irregularly inundated at present and are therefore suited
for the investigations in the unsaturated zone.
As there are partly sparse investigations (known to us) carried out or published concerning the
topic of the project, the project collaborators are grateful for references to publications and
also gray literature on chemical and toxicological loads of suspended and deposited sediments
in the context of the literature study ([email protected], [email protected]).
5.4 Structure of the joint research project
The joint research project is a cooperation of five partners from four organizations:
• The Stadtwerke Karlsruhe GmbH is the project coordinator and operating company of the
projected waterworks within the investigation area. It is responsible for the numerical
modeling of the groundwater flow.
• The DVGW-Technologiezentrum Wasser (TZW) Karlsruhe carries out the chemical
analyses of the soil and suspended sediment samples gathered in the field, and,
furthermore, investigates the behavior of the contaminants in the retention area by
comprehensive laboratory studies.
• The Institute for Water and River Basin Management, Universität Karlsruhe (TH) is
responsible for simulations as well as field and laboratory studies on the suspended load
transported into the retention basin.
• The Institute for Hydromechanics, Universität Karlsruhe (TH) carries out laboratory
studies and field investigations to determine the hydraulic characteristics of the
unsaturated zone in the study area and numerical modeling of transport and
transformation of characteristic compounds.
• At the Heidelberg Institute of Zoology (HIZ), toxicological studies and effect-directed
analyses are used to determine the biological hazard potential of suspended particles from
the flood wave up to within the unsaturated zone and to identify unknown groups of
contaminants.
5.5 References
102
5.5 References
Brack W, Altenburger R, Dorusch F, Hubert A, Moeder M, Morgenstern P, Moschütz S, Mothes S,
Schirmer K, Wennrich R, Wenzel K-D, Schürmann G (2002): Hochwasser 2002 - Chemische
und toxische Belastung überschwemmter Gemeinden im Raum Bitterfeld. Z Umweltchem
Oekotox 14: 213-220
Brack W, Erdinger L, Schirmer K, Hollert H (2005): Identification of mutagenicity and ERODinducing potency in aquatic sediments. Environ Toxicol Chem 24: 2445-2458
Brauch H-J, Lucas M, Sacher F (2001): Untersuchungen zum Vorkommen von Xenobiotika in
Schwebstoffen und Sedimenten Baden-Württembergs. Series 'Oberirdische Gewässer,
Gewässeroekologie', Landesanstalt für Umweltschutz Baden-Württemberg, vol. 67
Breitung V (1999): Organische Schadstoffbelastung in Schwebstoffen des Rheins während
Hochwasserwellen. Hydrologie und Wasserwirtschaft 43: 17-21
Foerstner U, Westrich B (2005): BMBF coordinated research project SEDYMO (2002 - 2006):
sediment dynamics and pollutant mobility in river basins. J Soils Sed 5: 134-138
Grunewald K, Unger C, Brauch HJ, Schmidt W (2004): Elbehochwasser 2002 - Ein Rückblick Schadstoffbelastung von Schlamm- und Sedimentproben im Raum Dresden (Sachsen).
Z Umweltchem Oekotox 16: 7-14
Gustavsson L, Klee N, Olsmann H, Hollert H, Engwall M (2004): Fate of Ah receptor agonists during
biological treatment of an industrial sludge containing explosives and pharmaceutical residues.
Environ Sci Pollut Res 11: 379-387
Hollert H, Dürr M, Erdinger L, Braunbeck T (2000): Cytotoxicity of settling particulate matter (SPM)
and sediments of the Neckar River (Germany) during a winter flood. Environ Toxicol Chem 19:
528-534
Hollert H, Haag I, Dürr M, Wetterauer B, Holtey-Weber R, Kern U, Westrich B, Färber H, Erdinger L,
Braunbeck T (2003): Untersuchungen zum oekotoxikologischen Schaedigungspotenzial und
Erosionsrisiko von kontaminierten Sedimenten in staugeregelten Flüssen. Z Umweltchem
Oekotox 15: 5-12
Hollert H, Dürr M, Holtey-Weber R, Islinger M, Brack W, Färber H, Erdinger L, Braunbeck T (2005):
Endocrine disruption of water and sediment extracts in a non-radioactive Dot Blot/RNAse
protection assay using isolated hepatocytes of rainbow trout − Deficiencies between
bioanalytical effectiveness and chemically determined concentrations and how to explain them.
Environ Sci Pollut Res 12: 347-360
Klee N, Gustavsson L, Kosmehl T, Engwall M, Erdinger L, Braunbeck T, Hollert H (2004): Toxicity
and genotoxicity in an industrial sewage sludge containing nitro- and amino-aromatic
compounds during treatment in bioreactors under different oxygen regimes. Environ Sci Pollut
Res 11: 313-320
Kosmehl T, Krebs F, Manz W, Erdinger L, Braunbeck T, Hollert H (2004): Comparative genotoxicity
testing of Rhine River sediment extracts using the permanent cell lines RTG-2 and RTL-W1 in
the comet assay and Ames assay. J Soils Sed 4: 84-94
Landesanstalt für Umweltschutz Baden-Württemberg (1995): Schadstofftransport bei Hochwasser Neckar, Rhein und Donau im Januar. Handbuch Wasser 2, vol. 23
Landesanstalt für Umweltschutz Baden-Wuerttemberg (1996a): Auswirkungen von Überflutungen auf
flussnahe Wasserwerke − Auswertung von Literatur und Betriebserfahrungen von Wasserwerken. Materialien zum Integrierten Rheinprogramm, vol. 6
Maier D, Maier M, Fleig M (1997): Schwebstoffuntersuchungen im Rhein. 29. AWBR-Jahresbericht
1996
Oetken M, Stachel B, Pfenninger M, Oehlmann J (2005): Impact of a flood disaster on sediment
toxicity in a major river system – The Elbe flood 2002 as a case study. Environ Pollut 134:
87-95
5.5 References
103
Reifferscheid G, Arndt C, Schmid C (2005): Further development of the betalactamase MutaGen
assay and evaluation by comparison with Ames Fluctuation tests and the Umu test. Environ Mol
Mutagen 46: 126-39
Seiler TB, Rastall AC, Leist E, Erdinger L, Braunbeck T, Hollert H (2006): Membrane Dialysis
Extraction (MDE): A novel approach for extracting toxicologically relevant hydrophobic
organic compounds from soils and sediments for assessment in biotests. J Soils Sed 6: 20-29
Ulrich M, Schulze T, Leist E, Glaß B, Maier M, Maier D, Braunbeck T, Hollert H (2002):
Oekotoxikologische Untersuchung von Sedimenten und Schwebstoffen: Abschaetzung des
Gefährdungspotenzials für Trinkwasser und Korrelation verschiedener Expositionspfade
(Acetonischer Extrakt, Natives Sediment) im Bakterienkontakttest und Fischeitest.
Z Umweltchem Oekotox 14: 132-137
5.5 References
104
105
Chapter 6
6 Impact of suspended particulate matter sampled
at the river Rhine with respect to operation of
retention basins and drinking water safety
J. Wölz1, M. Fleig2, T. Schulze3, S. Maletz1, U. Lübcke-von Varel3, G. Reifferscheid4,
D. Kühlers5, T. Braunbeck6, W. Brack3, H. Hollert1
1
Department of Ecosystem Analysis, Institute for Environmental Research, RWTH Aachen
University, Worringerweg 1, 52074 Aachen, Germany
2
DVGW-Water Technology Center (TZW), Chemical Analysis Department, Karlsruher Strasse 84,
76139, Karlsruhe, Germany
3
UFZ Helmholtz Centre for Environmental Research, Department of Effect-Directed Analyses,
Permoserstrasse 15, 04318 Leipzig, Germany
4
German Federal Institute for Hydrology, Am Mainzer Tor 1, D-56068 Koblenz, Germany
5
Stadtwerke Karlsruhe GmbH (SWK), Karlsruhe, Germany
6
Department of Zoology, Aquatic Toxicology and Ecology Section, University of Heidelberg, Im
Neuenheimer Feld 230, 69120 Heidelberg, Germany
To be published in Journal of Soils and Sediments
106
6.1 Abstract
107
6.1 Abstract
This study investigated suspended particulate matter (SPM) sampled at the river Rhine
barrage of Iffezheim, Germany. SPM were collected within the RIMAX-HoT joint research
project (2005 - 2009) that worked on the question whether flood management may conflict
with drinking water supply, since for example projected retention basins often overlap with
water protection areas. To answer this question, SPM were sampled periodically throughout
the year 2006 and more frequently in the course of a flood event with a recurrence interval of
10 years in August 2007.
GC-MS analysis was used to determine concentrations of polychlorinated biphenyls (PCBs)
as well as those of the historical river Rhine contaminant hexachlorobenzene (HCB). PCB
concentrations remained more or less constant in 2006 (30 µg/kg) as well as during the flood
(maximum 51 µg/kg). In contrast, the sediment contaminant HCB was constantly detectable
in 2006, but concentrations were clearly increasing in the August flood (maximum:
110 µg/kg).
SPM crude extracts were further assessed in order to determine dioxin-like and aryl hydrocarbon receptor (AhR)-mediated activities using the EROD assay and RTL-W1 cells form
rainbow trout (Oncorrynchus mykiss). EROD induction, given in biological toxicity
equivalent concentrations (bio-TEQs), showed elevated contamination levels in 2006 with
bio-TEQs between 1,159 pg/g and 6,639 pg/g. Further, flood SPM showed a maximum bioTEQ of 6,141 pg/g in the course of the flood. Further, mutagenic potentials were determined
using the Ames Fluctuation assay with the bacterial strains TA98 (frameshift mutation) and
TA100 (base pair substitution). Crude extracts indicated no significantly increased mutagenic
activity with the Ames Fluctuation assay, but caused high maximum induction factors in the
Comet assay (IFmax = 13).
Since hazard potentials were increased with SPM, effect-directed analysis (EDA) was used to
determine effective compound classes in flood SPM; target analysis was applied to identify
shares of EPA-PAHs to total biological effects. Fractionation showed that PAH fractions were
highly EROD inducing. However, EPA-PAHs contributed to less than 1% to the overall
biological activity. Fractions containing more polar to polar compounds gave highest
inductions, at least with SPM sampled after the flood peak. Mutagenic activities of fractions
was increased with SPM sampled after the flood peak, reflected by IFmax = 14.7 with the
bacterial strain TA 98 without metabolic activation by S9 supplement (rat liver homogenate).
With respect to the anticipated conflict of interests between flood retention and drinking
waters supply, it was shown that contaminant concentrations and biological activity were
clearly increased with a flood recurrence interval of 10 years. The retention basins in question
will be operated with recurrence intervals of 100 years. Since these floods can be assumed to
cause definitely higher contaminant (re-)mobilization a considerable compound deposition on
soils of flooded basins has to be assumed.
6.2 Introduction
108
6.2 Introduction
Assessment of particle-bound pollutants in suspended particulate matter (SPM) is of high
relevance for a sound understanding of processes and hazard potentials caused by flood
events. Flood water causes considerable physical damages to inundated sites and goods as
well as further to human and environmental health, since it is loaded with contaminants of
concern (Euripidou & Murray 2004). These compounds mainly originate from sediment
erosion, remobilization and subsequent translocation and redistribution (Koethe 2003). Since
sediments serve as sinks, but also as important secondary contaminant sources, increasing
contaminant loads are expected with more extreme floods in the near future (Heise &
Foerstner 2006). In this context, SPM was recognized as the carrier of contaminants and
hazard potentials (Schulze et al. 2007).
At present, extreme flood events such as the Elbe flood in 2002 are still hydrological outliers
causing considerable economic and ecological damage with minor recurrence intervals (Ikeda
et al. 2005, Klok & Kraak 2008). Nevertheless, as a consequence of climate change, these
events are expected to increase in frequency and intensity in many regions worldwide
(Change 2007). Furthermore, it is assumed that changes in precipitation may even be
amplified in river runoff (Chiew & McMahon 2002), and there is evidence that the magnitude
of peak flows increases (Middelkoop et al. 2001). Thus, recurrence intervals of floods
comparable to that of the Elbe in 2002 will become shorter (Bronstert 2003).
Under specific conditions, flood impact can cause conflict of interests when flood
management and further interests, e.g. drinking water supply, are concerned. In this context,
strategies to manage flood impact implement the operation of retention areas with higher
retention volumes than available at present (Disse & Engel 2001, Hooijer et al. 2004).
However, required retention basins can, e.g., overlap with water protection areas that are
needed for the operation of waterworks (Maier et al. 2006). The resulting conflict of interests
appears by now alone at 15 sites along the river Rhine between Basel (Switzerland) and
Duisburg, Germany). 20 millions of the 50 millions of people who live in the Rhine
watershed today drink treated Rhine water which in most cases is produced from riverbank
filtration (ICPR − International Commission for the Protection of the Rhine 2009). Water
suppliers are concerned about the potentially increased risk of pollution of groundwater
resources by the establishment of retention areas in the vicinity of their water extraction
facilities.
To date, many studies have shown high contaminant loads of sediments and suspended
particulate matter in the context of floods (Hollert et al. 2003, Oetken et al. 2005, Hilscherova
et al. 2007, Wölz et al. 2008). Further, many studies showed elevated contamination of
floodplain soils (Hilscherova et al. 2007, Pies et al. 2007, Yang et al. 2008). However, so far
there is only one pilot study with regard to the (eco-)toxicological hazard potentials of
sediments, suspended particulate matter and floodplain soils to the production of drinking
water, indicating an increase of the (eco-)toxicological hazard potential of near-surface soil
samples at inundated sites (Ulrich et al. 2002). Since elevated impact was determined in this
preliminary study, an interdisciplinary follow up project was initiated to investigate this
hazard potential in more detail. The results, detailed in this study, are part of the outcomes of
6.3 Material and methods
109
that joint research project 'Flood retention and drinking water supply – Preventing conflict of
interests' (RIMAX-HoT, Maier et al. 2006, Kühlers et al. 2009). The project worked on the
possible conflict of interests at the planned retention area Bellenkopf-Rappenwoert and a
nearby projected waterworks, Kastenwoert, both located next to Karlsruhe, Germany.
In this first part of the study, outcomes of ecotoxicological exposure assessment are presented
with respect to contamination of SPM. Therefore, results are detailed for SPM that was
sampled once at monthly intervals in 2006, and more frequently during a flood event in
August 2007 with a recurrence interval of 10 years. Samples were investigated with in vitro
biotests. Dioxin like and aryl hydrocarbon receptor (AhR)-mediated activities were
determined with the EROD assay and the rainbow trout liver cell line RTL-W1 (Lee et al.
1993). Further, mutagenic potentials were assessed with the Ames Fluctuation assay and the
two bacterial tester strains TA 98 and TA 100. Whereas, many compounds were chemically
analyzed, concentrations of PCBs and HCB were highest by far, and, thus, only these
compounds are regarded in this study. A recently developed method of effect-directed
analysis (Lübcke-von Varel et al. 2008) was applied to receive further insight into
contaminant loads in SPM sampled in the context of the flood event assessed. The 18 factions
obtained were assessed using the biotests listed above to identify effective compound classes
and chemical analysis was focused on EPA-PAHs with higher molecular weight, since lowmolecular compounds are not EROD inducers (Bols et al. 1999).
Thus, the present study aimed
(a) to identify in vitro hazard potentials and variations of SPM sampled throughout a year
and with SPM sampled in a flood event,
(b) to apply effect-directed analysis to identify effective compound classes,
(c) to chemically analyze concentrations and pattern of selected compounds in total SPM
and fractions and
(d) to project the results obtained on the assumed conflict of interest between water
retention and drinking water production.
6.3 Material and methods
6.3.1 Chemicals used
Subsequent providers of chemicals will only be named if other than Sigma-Aldrich
(Deisenhofen, Germany). Chemicals were at least reagent grade.
6.3.2 SPM sampling
In this study SPM was collected in 2006 using a continuous-flow centrifuge that was installed
just above the hydro power plant at the river Rhine barrage of Iffezheim, Germany
(for location, see fig. 1) at a depth of 0.8 m according to the method described by Babarowski
(2005). The Padberg Z61 (Padberg, Lahr, Germany) centrifuge type gives a flow rate of
900 L/h, 17,000 U/min and was run for 4 to 6 h.
6.3 Material and methods
110
N
France
Germany
Fig. 1 Location of the continuous-flow centrifuge and the passive sedimentation boxes at the Rhine
barrage of Iffezheim, Germany (circle). River flow direction is shown by light grey arrows; dashed
line gives the road across the river.
Further, SPM was sampled at the same site with higher frequency in the course of a flood
event with a recurrence interval of 10 years in August 2007 (Tab. 1 and fig. 2) using two
passive sedimentation boxes (Schulze et al. 2007). SPM was transferred to glass bottles,
protected from light and transported at 4 °C. SPM were then treated according to DIN 38414,
part 22. Samples were freeze-dried in two steps using a BETA 2 - 16 (Christ, Osterode,
Germany). Initially, SPM were dried for two days at 0.6 to 1 mbar and a temperature of 20 °C
to 25 °C. Subsequently, SPM were post-dried for two days and at least 0.001 mbar to lower
the residual moisture to < 0.5 %. SPM were than sieved at a mesh size of 600 µm for 15 min
using a Bandelin Sonorex RK 255 H ultrasound bath (Schalltech GmbH, MoerfeldenWalldorf, Germany). SPM were stored at 4 °C in darkness until extraction.
4000
4
3
5
***
3
Discharge [m /s]
3500
3000
2
2500
6
*
2000
1
1500
*
7
*
8
*
*
1000
0 2 4 6 8 10 12 14 16 18 20 22 24 26 28 30
Sampling times in August 2007
Fig. 2 Discharge at the gauge of Maxau, Germany, close to the sampling site and sampling periods in the flood
course. * − sampling times of SPM in August 2007
6.3 Material and methods
111
Tab. 1 Sampling times of SPM collected at the river Rhine barrage of Iffezheim, Germany;
fractionated samples are marked (X).
No. Sampling from
[date]
[time]
Sampling to
[date]
[time]
Fractionated
samples
1
14.07.07
12:00 a.m.
31.07.07
12:00 a.m.
2
31.07.07
12:00 a.m.
09.08.07
21:00 p.m.
3
09.08.07
21:00 p.m. 10.08.07
10:40 a.m.
4
10.08.07
10:40 a.m.
10.08.07
12:00 a.m.
●
5
10.08.07
12:00 a.m.
11.08.07
14:50 p.m.
●
6
11.08.07
14:50 p.m. 14.08.07
09:00 a.m.
●
7
14.08.07
09:00 a.m.
17.08.07
14:00 p.m.
8
17.08.07
14:00 p.m. 31.08.07
12:00 a.m.
●
6.3.3 Preparation of crude extracts
10 g of each freeze-dried SPM was weighed in 200 ml extraction thimbles (Schleicher &
Schuell, Dassel, Germany), stoppered with glass wool, placed in 400 ml Soxhlet extractors
and extracted with 250 ml acetone at 8 to 10 cycles/h for 14 h. The solvent was reduced in
volume and residues were evaporated under a gentle N2-stream close to dryness. Residues
were re-dissolved in dimethylsulfoxide (DMSO) and stored at 20 °C until biotesting. Empty
extraction thimbles were subjected to extraction and processed in two parallel experiments to
serve as process controls.
6.3.4 Clean-up of extracts and automated fractionation
10 g of each freeze-dried SPM was Soxhlet extracted as detailed above using a
dichloromethane (DCM):acetone (3:1; v/v) solvent mixture, reduced in volume, evaporated
under a gentle N2-stream and re-dissolved in n-hexane:acetone (7:3; v/v). Accelerated
membrane-assisted clean-up (AMAC) was used for purification of SPM extracts (Streck et al.
2008). Briefly, 1 ml extract with a concentration of 10 g SPM equivalent/ml was transferred
to dialysis membranes (low density polyethylene, 80 µm thickness; Polymer-Synthese-Werk,
Rheinberg, Germany) and dialyzed using an ASE 200 device (Dionex, Sunnyvale, CA) with a
mixture of DCM:acetone (3:1, v/v). The solvents used, temperature, pressure, number and
duration of cycles was chosen as described previously (Lübcke-von Varel 2008). Extracts
were collected in ASE glass vials closed by PTFE-coated screw caps, reduced in volume,
evaporated under a gentle N2-stream and re-dissolved in n-hexane:DCM (9:1; v/v) to a final
concentration of 20 g/ml for subsequent fractionation.
AMAC-purified extracts were fractionated using an automated fractionation method (Lübckevon Varel 2008). Initially, compounds of the AMAP extracts are loaded on three types of
columns: More polar to polar compounds are trapped on a cyanopropyl (CN) silica column
with n-hexane as mobile phase. Non-polar substances are flushed using a nitrophenylpropyl-
6.3 Material and methods
112
silica (NO) column and porous graphitized carbon (PGC) as stationary phase. Flushing of NO
and PGC phase with n-hexane continues eluting the remaining chlorinated diaromatic
substances from the NO to the PGC column.
Subsequent, sequential fractionation starts to elute compounds from each column. First of all,
chlorinated diaromatic compounds are separated of PGC using n-hexane and toluene as
mobile phase. Compounds trapped on the NO phase are successively eluted with
n-hexane:DCM (95:5; v/v). Finally, n-hexane, DCM and acetonitrile are used to elute
substances on the CN column. Fractions were collected in glass vessels, reduced in volume,
evaporated under a gentle N2-stream and re-dissolved in n-hexane (for GC-MS) and DMSO
(for biotesting) to a final concentration of 10 g/ml. Model compounds for each fraction are
detailed by Lübcke-von Varel (2008).
6.3.5 Chemical analysis of HCB and PCBs
The analyzed PCBs included: #28, #52, #101, #118, #138, #153, #170, #180 and #194.
Analysis was performed with a Perkin Elmer Autosystem XL (Waltham, Massachusetts,
USA) equipped with 63Ni electron-capture detector (ECD). The two columns used for
analysis were: Column A (CLP, 30 m x 0.5 mm x 0.32 µm; Restek Corp., Bellefonte, PA,
USA) and Column B (DB5, 30 m x 0.25 mm x 0.32 µm; J&W Scientific, Folsom, CA, USA).
The analysis conditions were: initial column temperature 60 °C (1 min), increased at
20 °C/min to 180 °C, then increased at 3 °C/min to 207 °C and at 1.5 °C to 260 °C that were
finally hold for 5 min. The carrier gas was helium. The injector temperature was 50 °C,
300 °C/min to 270 °C and the volume injected in splitless mode was 4 µl. The detector
temperature was 310 °C. As internal standard 25 µl of TCX/P209, 1 ng/µl, were added to the
sample prior to analysis. In addition to a blank sample with each set of samples (five to ten), a
process control treated like the samples was analyzed.
6.3.6 GC-MS analysis for PAHs
GC-MS analysis was carried out on a HP 6890 GC coupled to a HP MSD 5973 (Agilent, Palo
Alto, USA), equipped with a 30 m x 0.25 mm I.D. x 0.25 µm film HP-5 MS fused capillary
silica column, a 5 m pre-column (Agilent J&W, Folsom, USA) and a splitless injector with
deactivated glass wool. Chromatographic conditions were as follows: 280 °C injector
temperature, 1 µl pulsed splitless injection at oven temperature of 60 °C (1 min isotherm),
then programmed at 30 K /min to 150 °C, at 6 K/min to 186 °C and finally at 4 K/min to
280 °C (16.5 min isotherm). Carrier gas velocity (Helium 5.0, Air Liquide, Boehlen,
Germany) was 1.3 ml/min at constant flow. The MS was operated in electron impact
ionization mode (EI+, 70 eV) with a source temperature of 230 °C scanning from 30 to
500 amu (full scan mode) or single ion monitoring (SIM) for quantification. Target analytes
were quantified using an external calibration in single ion monitoring (SIM). The results were
corrected with an internal standard containing deutered PAH (Mix 35, Promochem, Wesel,
Germany).
6.3 Material and methods
113
6.3.7 EROD-induction assay
Induction of 7-ethoxyresorufin-o-deethylase (EROD) was measured in the CYP1A-expressing
cell line RTL-W1 (Lee et al. 1993) according to the method of Gustavsson et al. (2004) with
the modifications given by Keiter et al. (2008). Cells were seeded in 96-well plates (TPP,
Trasadingen, Switzerland) and allowed to grow to 100 % confluence for 72 h. Subsequently,
the medium was removed and the cells were exposed for 72 h to the SPM extracts diluted in
medium using eight dilutions with six replicates each as well as to the standards. Maximum
DMSO concentration was 0.1 % since DMSO causes cytotoxicity at concentrations higher
than 2 to 3 % in the well (Wölz et al. 2008). The positive control 2,3,7,8tetrachlorodibenzo-p-dioxin (TCDD; Promochem, Wesel, Germany) was serially diluted to
give a final concentration range of 3.13 to 100 pM on two separate rows of each plate.
Exposure was terminated by removing the growth medium and freezing at 70 °C to lyse the
cells.
7-ethoxyresorufin was added to each well as exogenous substrate and incubated in the dark at
room temperature for 10 min. Subsequently, NADPH was supplemented to start the
de-ethylation of the exogenous substrate and plates were incubated for another 10 min. The
reaction was stopped by adding fluorescamine dissolved in acetonitrile. EROD activity was
measured fluorometrically after another 15 min using a GENios plate reader (Tecan,
Crailsheim, Germany; excitation 544 nm, emission 590 nm). Protein was determined
fluorometrically using the fluorescamine method (excitation 355 nm, emission 590 nm;
Lorenzen & Kennedy 1993, Kennedy & Jones 1994). The concentration-response curves for
EROD induction in the RTL-W1 bioassay were computed by non-linear regression using
GraphPad Prism 4 (GraphPad, San Diego, USA) and classic sigmoid or Boltzmann curves as
model equations (Seiler et al. 2006, Olsman et al. 2007). The enzyme-inducing potential of
the samples was converted to biological toxic equivalents (bio-TEQs) as described below.
6.3.8 Bio-TEQ values
Ah receptor agonist activities were determined as EC25 values of each sample and were
normalized to the positive control 2,3,7,8-TCDD as biological toxicity equivalent
concentrations (bio-TEQs; cf. Wölz et al. 2008). Bio-TEQs were calculated as given in eq. 1
given as mean values of n = 3 independent biotests. TCDD-EC25 were determined with each
test plate and mean values were used for the calculating of bio-TEQ values. Subsequently,
bio-TEQs with concentrations in pg TCDD/g of SEQ will be given as pg/g.
Eq. 1: Bio-TEQ [pg TCDD / g SEQ] = TCDD-EC25 [pg TCDD/ml] / Sample-EC25 [g SEQ/ml]
6.3.9 Ames Fluctuation assay
The Ames Fluctuation assay is a modification of the plate incorporation Ames test (Maron &
Ames 1983) according to the method described by Reifferscheid et al. (2005). In contrast to
the classic test, exposure was in liquid medium on 384-well microtitre plates. Mutagenic
activity of SPM was determined with the two tester strains TA 98 (frameshift mutation) and
6.3 Material and methods
114
TA 100 (base pair substitution) as detailed by Maron & Ames (1983). Bacteria were cultured
overnight in Oxoid Nutrient Broth No. 2 and ampicillin (50 µg/ml) at 37 °C ± 1 °C in
a shaking water bath for not more than 10 h. Densities of the overnight inoculum were
computed as formazine attenuation units (FAU) by relating measured optical densities
(λ = 595 nm) to a standard (10 g/L hexamethylenetetramine, 1 g/L hydrazinesulfate; equals
1,800 FAU) according to the method described by (Hawe & Friess 2008). For testing
overnight cultures were adjusted to 1,800 FAU (TA 98) and 450 FAU (TA 100).
Subsequent adjustment, bacteria were pre-incubated with exposure medium, containing low
concentrations of histidine (6.45 µM per well), in 24-well microtiter plates (TPP) for 90 min
at 37 °C to allow some cell divisions. Pre-incubated bacteria were 6-fold diluted in histidine
deficient reversion indicator medium, containing bromocresol purple. Bacteria were
distributed into 384-well plates (TPP) with 48 wells per replicate (controls and sample
dilutions) for 48 h at 37 °C. Only reversed bacteria recover growth in minimal medium.
Acidification by metabolic activity causes a definite switch of bromocresol from purple to
yellow in the well. Wells that indicated reversions were counted. For the evaluation of
metabolic activation, rat liver homogenate S9-fraction (RCC Rossdorf, Germany) from
phenobarbital/ß-naphthoflavon-treated mice (protein concentration: 30.5 mg/ml S9) was
added in a buffer mixture to each well.
For each test ± S9 negative and positive controls were used as validity control. Tests were
valid when mean values of spontaneous revertants in negative controls counted for 0 to ≤ 5
per 48 wells (TA 98) and > 0 and ≤ 10 per 48 wells (TA 100) at all testing conditions with
both strains ± S9. Positive controls were valid when no. of revertants were ≥ 25 per 48 wells
as mean values for both bacterial strains ± S9 at all testing conditions. DMSO was added as
solvent/negative control (maximum of 0.1 % per well). Positive controls were
4-nitro-o-phenylenediamine (20 nM per well) for TA 98 strain without S9, nitrofurantoin
(1.67 nM per well) for TA 100 without S9 and 2-aminoanthracene for TA 98 and TA 100
with S9 treatment (0.87 nM per well).
Ames et al. (1975) used the rule that twofold induction versus the negative control indicate
statistical significant mutagenic activity of a sample for the plate incorporation assay (Cariello
& Piegorsch 1996). However, the Ames Fluctuation assay showed a low number of
spontaneous reversions in the negative control and relatively high standard deviations. Futher,
test replicates were low in number. Thus, the twofold rule may not be used. In contrast,
Fisher`s Exact Binomial test that for low numbers of tests was chosen. Mutagenic activity was
considered statistically significant when p < 0.05. This statistical method is also planned to be
used in the recently developed ISO norm (International Organization for Standardization) of
the Ames Fluctuation assay. Fisher`s Exact test allows to calculate NOEC values (no
observed effect level/concentration). While NOEC values provide information on effects with
respect to concentrations, intensities of effects are not addressed. Thus, in addition maximum
induction factors (IFmax) were computed, that give the induction of the highest inducing
sample concentration, referred to the negative control induction.
6.4 Results
115
6.4 Results
6.4.1 SPM sampled in 2006
Whereas various compounds were analyzed (e.g. HCHs, DDT and metabolites), elevated
concentrations were only determined for HCB and selected PCBs and showed concentrations
of 7.4 to 29 µg/kg (HCB) and 4.7 to 28 µg/kg (PCBs) as given in fig. 3a. These compounds
showed minor variations in concentration throughout the year 2006. Seasonal or
discharge-dependent influences could not be observed. Further on, SPM were assessed with
respect to Ah receptor-mediated activities and showed a bio-TEQ range of 1,160 to
6,640 pg/g (Fig. 3b). In contrast to PCB and HCB concentrations bio-TEQs of SPM showed a
seasonal variation with highest inductions in June and the following month. These elevated
activities were not correlated with discharge. SPM sampled in early and winter month of 2006
indicated comparably lower inductions.
1500
1000
3000
6000
2000
4000
3
3
30
Bio-TEQ [pg/g]
2000
Bio-TEQ
Water level
Discharge [m /s]
2500
60
b
3000
Discharge [m /s]
HCB
sum PCB
Water level
a
Concentration [µg/kg]
8000
3500
90
1000
2000
500
0
0
1
2
3
4
5
6
7
8
Month in 2006
9 10 11 12
n.a.
0
1
2
n.a.
3
4
5
6
7
8
0
9 10 11 12
Month in 2006
Fig. 3 (a) HCB and PCB concentrations determined with SPM sampled in 2006 using a centrifuge at the Rhine
barrage of Iffezheim, Germany, in the context of the water level at Maxau, Germany, close to Iffezheim and (b)
AhR-mediated activity with the same SPM samples determined with n = 3. n.a. − not assessed
6.4.2 SPM sampled in the context of the flood event in August 2007
In accordance to the data presented for SPM sample in 2006, HCB and PCBs remained the
highest concentrated compounds with SPM sample at end of July and in August 2007.
Concentrations of HCB and PCBs as determined in the timeframe of the flood event are given
in fig. 4a. HCB concentrations were more than twofold increased in the first flood sample
(August 9, 2007, 21.00 p.m.) compared to the concentrations determined before the flood at
the end of July (July 31, 2007, 12.00 a.m.). Maximum concentrations of 110 µg/kg were
measured at the peak discharge of the flood. HCB concentrations decreased clearly after the
flood peak. However, SPM sampled subsequent indicated elevated concentrations that were
higher than were about twofold increased compared to the SPM of end of July. In contrast,
PCBs indicated only an increase at the beginning of the flood (67 µg/kg). Subsequent sampled
SPM indicated lower concentrations (5 to 25 µg/kg) and showed no relation to the flood
discharge.
Measured Ah receptor mediated activities, given as bio-TEQs, indicated a clear-cut increase
of activity in accordance to the increasing discharge (Fig. 4b). TEQs indicated decreasing
6.4 Results
116
AhR-agonist activities about one day after the flood peak (August 11, 2007, 14:50 p.m.) in
accordance to HCB analysis. However, the maximum bio-TEQ was measured with the
following sample (6,140 pg/g). SPM sampled subsequent indicated decreased, but still high
bio-TEQs, that were comparable to end of July SPM.
900
60
600
30
300
0
0
:00 1:00 0:40 2:00 4:50 9:00 4:00 5:00
1
1
0
1
1
1
2
12
7, .07, .07, .07, .07, .07, .07, .07,
0
.
.07 9.08 0.08 0.08 1.08 4.08 7.08 1.08
1
3
1
1
1
1
1
0
3
Sampling times within flood in August 2007
Water level
Bio-TEQ
6000
900
4000
600
2000
300
Water level [cm]
Concentration [µg/kg]
b
Water level [cm]
90
Water level
HCB
Sum of PCBs
1200
8000
1200
a
Bio-TEQ [pg/g]
120
0
0
:00 1:00 0:40 2:00 4:50 9:00 4:00 5:00
12
2
1
1
1
0
1
1
7, .07, .07, .07, .07, .07, .07, .07,
0
.
.07 .08 .08 .08 .08 4.08 7.08 1.08
31 09 10 10 11
1
1
3
Sampling times within flood in August 2007
Fig. 4 (a) HCB and PCB concentrations for SPM of August 2007, sampled at the river Rhine barrage
of Iffezheim (Germany) using a sediment trap. (b) Ah receptor-mediated activities for the
corresponding SPM sample, given as bio-TEQ values in pg/g (n = 3).
6.4.3 Identification of effective fractions
For a more profound analysis and identification of effective compound classes, EDA was
applied, providing 18 distinct fractions (Fig. 5). Fractions F1 to F4, containing for example
PCBs and PCDD/Fs and fractions F5 to F7 with PAHs of ≤ 4 aromatic rings indicated minor
dioxin-like and AhR-agonist potentials. Significantly increased TEQs were determined with
each fraction F8 to F11. Fraction F12 mostly containing mononitro-PAHs was less inducing,
while fraction F13, e.g. containing chinone, hydroxy-PAHs, was highest inducing.
Fraction 14, containing e.g. (hydroxy-)quinones, keto-, dinitro-, hydroxy-PAHs, and
N-heterocycles with rising polarity, gave minor bio-TEQs until the flood peak but increased
activities thereafter. Fractions F15 to F18, containing e.g. 2 hydroxyanthraquinone, showed
decreasing but nevertheless elevated bio-TEQs.
In order to determine contributions of compound categories to the total effect each
18 fractions were primarily added giving 11,800 pg/g (August 11, 2007) to 17,300 pg/g
(August 14, 2007). Further, fractions containing PAHs (F5 to F12) and fractions containing
more polar to polar compounds (F13 to F18) were added. First of all, added bio-TEQs of all
18 fractions showed a 3- to 7-fold increase of enzyme inductions compared to crude SPM
extracts with all SPM assessed. Added PAH fractions were the highest inducing compound
category with SPM sampled with increasing discharge at August 9, 2007. With each other
sample, sums of bio-TEQs were higher with fractions containing more polar to polar
compounds and a maximum bio-TEQ of 10,012 pg/g.
6.4 Results
Bio-TEQ [pg/g]
15000
10000
5000
0
20000
Bio-TEQ [pg/g]
20000
09.08.2007, 21:00 a.m.
***
15000
10000
5000
0
** **
*
C A P M 1 2 3 4 5 6 7 8 9 101112131415161718
Fractions
Fractions
20000
11.08.2007, 14:50 p.m.
15000
10000
5000
0
10.08.2007, 12:00 a.m.
C A P M 1 2 3 4 5 6 7 8 9 101112131415161718
Bio-TEQ [pg/g]
Bio-TEQ [pg/g]
20000
117
** *
* *
14.08.2007, 9:00 a.m.
15000
10000
5000
**
*
*
C A P M 1 2 3 4 5 6 7 8 9 1011121314151617
C A P M 1 2 3 4 5 6 7 8 9 1011121314151617
Fractions
Fractions
Fig. 5 Ah receptor mediated activity, given as bio-TEQ values for SPM crude extracts (C), added
fractions F1 to F18 (A), added PAH fractions F5 to F12 (P), added fractions with more polar to polar
compounds F13 to F18 (M) as well as for each single fraction (n = 3). * − No EROD induction
detected
6.4.4 Mutagenic potentials of fractions
Mutagenic activity was measured with each SPM sample and with fractions. SPM sampled in
2006 indicated no significant mutagenic potentials. However, fractions of SPM sampled in
August 14, 2007 at 12 a.m. caused significantly elevated effects. Fisher`s Exact Binominal
test showed significant NOEC values ≤ maximum concentration with fractions containing
more polar to polar compounds (Tab. 2).
Fraction F15 revealed the highest mutagenic potential with TA 98 without S9 metabolism and
a NOEC < 2.08 mg/ml, the lowest concentration assessed. Thus, elevated potentials were
caused by compounds that induce frameshift mutations. Further, mutagenic activity was
highly increased with TA 100 without S9 metabolism in F13, showing that compounds
causing base pair mutations were highest concentrated in this fraction.
Further, maximum induction factors were computed for fractions that were determined to
show significant elevated mutagenic potentials. With respect to IFmax highest mutagenic
activity was determined accordingly in F15 and TA 98 without S9 metabolism (IFmax = 14.7).
6.5 Discussion
118
Tab. 2 Mutagenic potential of SPM fractions of August 14, 2007, 12 a.m. in the Ames Fluctuation
assay using bacterial strains TA 98 and TA 100, determined with n = 1 and 48 replica per test.
Mutagenic potentials are given as NOEC value and maximum induction factor (IFmax).
Fraction
no.
1
…
13
14
15
16
17
18
Induction factor (IFmax)
NOEC [mg/ml]
Fraction
TA98 TA98 TA100 TA100
TA 98 TA 98 TA100 TA100
no.
-S9
+S9
-S9
+S9
-S9
+ S9
-S9
+S9
*
*
*
*
1
*
*
*
*
…
…
…
…
…
…
…
…
…
*
16.67
4.17
*
13
*
3
3.7
*
*
*
*
*
14
*
*
*
*
< 2.08
*
*
*
15
14.7
*
*
*
8.33
8.33
*
16.67
16
4.3
1.7
*
2.0
*
*
*
*
17
*
*
*
*
*
16.67
*
33.33
18
*
5.3
*
9.3
Maximum concentration in test: 66.67 mg/ml, lowest concentration in test: 2.08 mg/ml,
± S9 − Metabolic activation using rat liver homogenate of the S9 fraction in the liver centrifugate,
* − NOEC > 66.67 mg SPM equivalent/ml test medium and no IFmax determined
6.5 Discussion
6.5.1 Chemical loads of crude extracts
Chemical analysis showed that PCBs were detectable at minor rates in SPM sampled in 2006
and less conspicuous in the flood of August 2007. Concentrations in 2006 gave a mean of
11.7 ± 8.6 µg/kg and in the August flood a mean of 16.9 ± 16.5 µg/kg. Thus, PCB
concentrations were comparable to SPM of a flood in January 2004 at the Rhine (recurrence
interval of 2 years) and maximum concentrations of 32 µg/kg (Wölz et al. 2008). Ranking
these findings with other studies that investigated river sediments (0 to 339 µg/kg) indicated
concentrations in SPM to be comparably low (Mai et al. 2002, Zhang et al. 2004, Samara et
al. 2006, Zhang et al. 2007).
In contrast, HCB showed concentrations of 4.8 to 85 µg/kg (median = 16 µg/kg) in SPM of
2006 and 24 to 110 µg/kg (median = 53 µg/kg) in SPM sampled in the August flood, thus,
being 3.3 fold increased in the flood comparing medians. Ulrich et al. (2002) determined
HCB concentrations with a maximum of 203 µg/kg in SPM sampled in the fish ladder at the
barrage of Iffezheim, thus, being twofold increased compared to the present study. Other
studies measured HCB concentrations of 220 µg/kg in sediments at the barrage of Iffezheim
at a depth of 0.2 to 1.2 m and about 40 µg/kg near to surface (Alcock et al. 2003). Using the
Chemistry-Toxicity Test (CTT) approach (Heise et al. 2004, Heise & Foerstner 2006), the
action level for HCB (= 20 µg/kg) is clearly exceeded. Thus, e.g., dumping of Rotterdam port
sediment at sea would no longer be allowed and deposition at specified dumps causing
considerably increased costs would be necessary (Netzband 2007). However, HCB in
detected concentrations is perturbing as this compound is classified as 'substance of concern',
6.5 Discussion
119
since it frequently exceeds regulatory criteria for suspended matter (Heise & Foerstner 2006).
Further, due to its persistence, HCB is listed as one of the 'dirty dozen' in the Stockholm
Treaty on Persistent Organic Pollutants (POPs; SC 2004) and as a priority hazardous
substance in the Water Framework Directive (WFD; EC 2008). Thus, HCB is a major
hazardous compound and has to be included in measure strategies, since a successful
management will take influence on the decision whether a good chemical status is obtained in
water, sediment and biota according to the WFD aims (Coquery et al. 2005, Förstner 2008).
Elevated compound loads indicate a risk of compound introduction in retention basins and
contamination of flooded soils.
6.5.2 Biological hazard potential in crude extracts
Ah receptor agonists were elevated and strongly varying in 2006 with comparable inductions
in the August flood. Whereas, the reasons of the increased effects in 2006 remain unclear, so
far, it is evident that elevated EROD inductions can be caused by incidents other than flood
events. However, the clear difference between both elevated EROD inductions is the time
frame. Floods cause among others rapidly increasing contaminant (re-)mobilization and
exposure, whereas in 2006 effects seemed to increase more slowly, beginning with SPM
sampled in May, but lasted for month and, thus, show elevated long-term contamination.
In a previous study SPM sampled in a flood in January 2004 with a recurrence interval of
2 years showed highest bio-TEQ = 2,300 pg/g, whereas SPM sampled in the winter month
(November 2003 to February 2004) induced bio-TEQs = 3,700 pg/g (Wölz et al. 2008). In the
present study, bio TEQs were about 2.7-fold increased compared to theses maximum values.
Thus, higher impacts are indicated through more intensive floods. Koh et al. (2004)
determined maximum bio-TEQs of 1,500 pg/g in sediment of the Hyeongsan River, Korea,
using H4IIE-luc cells. Hilscherova et al. (2003) used the same cell line and found bio-TEQ of
1,860 pg/g in sediment of the Tittabawassea River, Michigan, USA. Further, Hollert et al.
(2002) used RTL-W1 cells and the EROD induction assay to assess sediments of the
catchment area of the river Neckar, Germany, and determined bio-TEQs of about 1,000 pg/g.
Comparing these bio-TEQs to SPM sampled in the present study underlines increased AhRinducing potentials and, thus, elevated hazard potentials. In accordance to the detailed results
on chemicals, AhR-agonists indicate an increased load of inducing particle-bound
compounds, and, accordingly, an impact to inundated sites, such as retention basins.
6.5.3 AhR-agonists and mutagenic potential in fractions
Whereas increased EROD inductions were determined with SPM crude extracts, active
compounds were not identified so far. Thus, an automated EDA method was used to reduce
the complexity of each sample and to identify inducing fractions and target compounds.
Itemized biotests and chemical target analysis showed that fractions containing PAH caused
increased effects. However, fractions containing more polar to polar compounds were
identified to be inducing highest. Chemical analysis was performed with respect to so called
priority EPA-PAHs (EPA, Laboratory Test Protocol Number 610) with more than four
6.6 Conclusions
120
aromatic rings and gave minor EPA-PAH concentrations of 5.2 to 50.9 µg/kg. Chem-TEQs
were calculated as products of compound concentrations and cell line specific toxicity factors
that were determined relative to the reference (cf. Olsman et al. 2007). Chem-TEQ values
equaled far less than 1 % of the bio TEQs (therefore, data not shown in detail). Thus, other
non-priority chemicals were causing effects in PAH fractions. In general, these findings are in
accordance with other studies that worked on PAH contaminations in sediments (Brack et al.
2002, Barron et al. 2004, Brack et al. 2005) and flood SPM (Wölz et al. 2008).
Next to AhR-agonists mutagenic potentials of SPM crude extracts and fractions were
assessed. Significantly elevated bacterial reversions were not determined with crude SPM
extracts but flood SPM fractions were mutagenic with SPM sampled after the flood peak.
Sediment and SPM extracts of a flood with a recurrence interval of 1 year at the Neckar,
Germany, were also not inducing as determined with the Ames Plate Incorporation assay in
another study (Hollert et al. 2000). However, SPM sampled in a flood with a recurrence
interval of 15 to 20 years at the Neckar was shown to cause IFmax = 3.2 (Hollert et al. 2003).
In the present study, significant inductions were only detected in fractions of SPM sampled
after the flood peak and highest inductions were caused in fraction F14 containing more polar
compounds (IFmax = 14.7). Elevated mutagenic potencies of compounds eluted in these
fractions had been shown before (Schuetzle et al. 1981, Kataoka et al. 2000, Eisentraeger et
al. 2008).
Further, crude SPM extracts of 2006 were also investigated using the Comet assay as detailed
by Singh et al. (1988) in the modification of Schnurstein & Braunbeck (2001; details not
shown). These SPM extracts showed up to 13-fold increased IFmax. Kosmehl et al. (2004)
assessed sediment cores from the river Rhine, Germany, and gave IFmax = 90.5 with RTG 2
cells and IFmax = 47.0 with RTL W1 cells. Ranking these findings with the present study
indicates lower, but nevertheless elevated mutagenic potentials with the SPM assessed. Since
in a subsequent study Kosmehl et al. (2006) showed that mutagenically active compounds of
the sediments were bioavailable in principle, using a novel contact assay with zebrafish
(Danio rerio). With respect to the assumed conflict of interests, these findings indicate that
translocation of particles of floods into retention basin may result in deposition of highly toxic
compounds on soil.
6.6 Conclusions
The investigation of SPM sampled continuously over months and, with higher frequency, in
times of flood events, allows evaluating dioxin-like and Ah receptor agonist activities as well
as mutagenic potentials. Investigation of chemical loads and biological activities over months
allows detection of variations in activity throughout the year and should be taken into
consideration in the evaluation of activities that are determined on small temporal scales in
floods.
Whereas AhR-mediated activities can be assumed to be highly increased in floods, further
influences might lead to comparably elevated hazard potentials. However, AhR-agonists are
6.6 Conclusion
121
highly active throughout the year and, thus, particle-bound contaminants have to be addressed
for evaluation.
In particular, particle-bound tracer compounds typical of each catchment area may be used to
evaluate contaminant loads as mentioned recently by Brack et al. (2009). At the Rhine, HCB
is a known and omnipresent pollutant that acts as such a compound. Since HCB is a particlebound compound in sediments, elevated concentrations in flood SPM act as indicator for
sediment deposition. However, lower concentrations are bound to SPM and detectable
throughout the year. In contrast, PCB contaminations seem to be less correlated with sediment
remobilization in floods; concentrations are constant over long observation periods.
Automated fractionation methods can be used to identify classes of effective compounds in
highly inducing samples. Further, applied target analysis allows identifying concentrations
and shares of analyzed compounds to the overall biological activity. Percentages of priority
compounds, even when minor, provide valuable information since low shares indicate that
other novel compounds are more relevant. Thus, more polar to polar compounds should be
investigated with elevated emphasis in future studies that work on hazard potentials of
contaminant loads in floods, in particular with respect to floods with high recurrence
intervals. Dioxin-like and AhR-mediated, as well as mutagenic activity are valuable endpoints
to determine hazard potentials, since concerning activities in crude extracts and fractions with
more polar compounds have repeatedly often been shown recently.
With respect to conflict of interests between flood management and drinking water supply,
pollution of flooded areas as e.g. retention basin can be assumed (Hilscherova et al. 2007).
Even more so, since the addressed retention basins will be operated only for floods with
recurrence intervals of 100 years and higher. A further hazard to the aquifer and drinking
water resources cannot be derived from these results. Thus, further investigations of RIMAXHoT focused on floodplain soil and groundwater contamination at a site designated for the
operation of a retention basin but also protected as a drinking water protection area.
Acknowledgements
The authors would like to express their thanks to Drs. Niels C. Bols and Lucy Lee (University
of Waterloo, Canada) for providing RTL-W1 cells. We thank Kerstin Winkens, Anne
Schneider, Susanne Miller, Conny Bernecker, Ulrike Diehl for assistance with conducting
biotests. We also thank the Federal Ministry of Education and Research (BMBF), Germany,
for supporting the RIMAX-HoT project within the RIMAX joint No. 02WH0691
6.7 References
122
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to a DNA-binding form by food-borne heterocyclic amines. Carcinogenesis 13: 1619-1624
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Maltby L, Boxall ABA, Forrow DM, Calow P, Betton CI (1995): The effects of motorway runoff on
freshwater ecosystems: 2. Identifying major toxicants. Environ Toxicol Chem 14: 1093-1101
NATO/CCMS (1998): Pilot study on internal information exchange on dioxins and related
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complex mixtures of dioxins and related compounds. Report no. 176, 26 pp.
Neff JM, Stout SA, Gunster DG (2005): Ecological risk assessment of polycyclic aromatic
hydrocarbons in sediments: Identifying sources and ecological hazard. Integr Environ Assess
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Shen Q, Wang KY, Zhang W, Zhang SC, Wang XJ (2008): Characterization and sources of PAHs in
an urban river system in Beijing, China. Environ Geochem Health 31: 453-462
Stachel B, Gotz R, Herrmann T, Kruger F, Knoth W, Papke O (2004): The Elbe flood in August 2002
− Occurrence of polychlorinated dibenzo-p-dioxins, polychlorinated dibenzofurans (PCDD/F)
and dioxin-like PCB in suspended particulate matter (SPM), sediment and fish. Water Sci
Technol 50: 309-316
Sundberg H, Ishaq R, Akerman G, Tjarnlund U, Zebuhr Y, Linderoth M (2005): A bio-effect directed
fractionation study for toxicological and chemical characterization of organic compounds in
bottom sediment. Toxicol Sci 84: 63-72
Villeneuve DL, Crunkilton RL, DeVita WM (1997): Aryl hydrocarbon receptor-mediated toxic
potency of dissolved lipophilic organic contaminants collected from Lincoln Creek, Milwaukee,
Wisconsin, USA, to PLHC-1 (Poeciliopsis lucida) fish hepatoma cells. Environ Toxicol Chem
16: 977-984
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Ah receptor agonist activity of suspended particulate matter during flood events at the rivers
Neckar and Rhine - A mass balance approach using in vitro methods and chemical analysis.
Environ Sci Pollut Res 15: 536-553
Zonta R, Collavini F, Zaggia L, Zuliani A (2005): The effect of floods on the transport of suspended
sediments and contaminants: A case study from the estuary of the Dese River (Venice Lagoon,
Italy). Environ Int 31: 948-58
125
Chapter 7
7 Pollution of riparian areas in consequence of
inundation by extreme flooding
J. Woelz1, T. Schulze2, M. Fleig3, G. Reifferscheid4, U. Lübcke von-Varel2,
W. Brack2, H. Hollert1
Department of Ecosystem Analysis, Institute for Environmental Research, RWTH Aachen
University, Worringerweg 1, 52074 Aachen, Germany
2 UFZ Helmholtz Centre for Environmental Research, Department of Effect-Directed Analyses,
Permoserstrasse 15, 04318 Leipzig, Germany
3
DVGW-Water Technology Center (TZW), Chemical Analysis Department, Karlsruher Strasse 84,
76139 Karlsruhe, Germany
4
German Federal Institute for Hydrology, Am Mainzer Tor 1, D-56068 Koblenz, Germany
1
To be published in Journal of Soils and Sediments
126
7.1 Abstract
127
7.1 Abstract
In this study, soil was sampled at inundated and non-inundated sites within a projected
retention basin that is planned to be operated with floods of recurrence intervals greater than
or equal to 100 years. This basin overlaps with a water protection area that is essential for a
projected nearby waterworks. The detailed investigations are part of the RIMAX-HoT joint
research project (2005 - 2009) that assessed the conflict of interests between flood
management and drinking water supply.
Sampled soil cores were cut into distinct layers and investigated using in vitro biotests.
Dioxin-like and AhR-mediated EROD enzyme inductions (Cytochrome P450 monooxygenase) were assessed using the fibroblast-like RTL-W1 cell line from rainbow trout
(Oncorhynchus mykiss). Mutagenic potentials were assessed with the Ames Fluctuation assay
and the tester strains TA 98 and TA 100 of Salmonella typhimurium bacteria. While
mutagenic activity was not detected in soil layers, elevated EROD inductions were measured
in topsoil, that were decreasing in deeper layers. However, one site - a ground swale - was
determined to be highly inducing as reflected by a biological equivalent concentration
(bio-TEQ) of about 41,000 pg/g. Chemical analysis with respect to HCB (0.049 mg/kg),
EPA-PAHs (39 mg/kg) and selected PCBs (0.19 mg/kg) gave relative increases at this site.
Further, chemical loads and biological activities were determined to be increased at least
down to 90 cm subsurface.
The highly polluted topsoil layer was chosen for fractionation using a recently developed
automated effect-directed fractionation method that was used to identify effective compound
categories. Fractions containing PAHs were determined to cause the bulk of EROD-induction
as reflected by an added fraction bio-TEQ of 32,000 pg/g (of a total bio-TEQ = 43,000 pg/g
of all added fractions). Further, fractions containing moderately polar and polar compounds
caused elevated inductions (bio-TEQ ≈ 8,200 pg/g). Although crude extracts were not
mutagenic single fractions showed heterogeneous and elevated potentials that were computed
as NOEC values (No observed effect concentration) and maximum induction factors (IFmax).
With most fractions, tester strain TA 98 (frameshift mutation) showed significantly reduced
NOECs independent of S9 (rat liver homogenate) metabolic activation. TA 100 indicated only
some few fractions to cause base pair substitution. In accordance to the EROD assay,
fractions containing PAHs, moderately polar and polar compounds caused elevated mutagenic
activity. However, the latter compounds were more toxic and showed NOECs down to 0.03
mg/ml and IFmax = 29.
Relating these findings to the assumed conflict of interests, an impact to the aquifer and, thus,
drinking water resources cannot be excluded. Since ground sampled in a swale was shown to
be highly polluted far below surface, compounds may be less retained and passage to the
aquifer is facilitated at those sites. Further, the elevated load of mutagenic moderately polar
and polar compounds may easily pass the unsaturated soil zone. Introduced to the aquifer,
these compounds may represent a threat to groundwater quality. Thus, further research with
7.2 Introduction
128
respect to groundwater contamination and hazardous compounds will assist to evaluate the
risk of aquifer pollution.
7.2 Introduction
Floodplain soil often is loaded with many contaminants as a consequence of inundation
during flood events (Zonta et al. 2005). In general, floods cause increasing sediment erosion
in accordance with water discharge. Erosion can reach deep and remobilize in particular
highly loaded (older) sediment layers (Hollert et al. 2007a, Stronkhorst & van Hattum 2003).
Following erosion sediment can be translocated as suspended particulate matter (SPM) and,
thus, be displaced at any inundated site. Therefore, next to downstream river sections, eroded
matter primarily affects floodplains, depositing in a considerable amount since currents are
lower at flat river banks that are usually abundantly covered with vegetation (Jeffries et al.
2003).
SPM are initial matter for soil genesis and appear to be an important nutrient source (Wassen
et al. 2002). However, remobilized sediments are also hazardous since in the river they act as
both sinks and important secondary sources of contaminants introduced into the aquatic
environment (Foerstner 2004, Kosmehl et al. 2004). Thus, matter deposited on floodplains
provides a potential to (highly) contaminate affected sites, in particular with intensified
erosion during extreme floods (Weber et al. 2008b). Typical contaminants at the river Rhine
are polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs) and as a
special case of the Rhine catchment hexachlorobenzene (HCB; Heise & Foerstner 2006, Klok
& Kraak 2008, Wölz et al. 2008).
Deposited matter preferably accumulates in surface depressions and water basins that, thus,
often contain the highest contaminant loads (Asselman & Middelkoop 1995). Following
floods, these contaminants can impact adjacent areas by wind drift but significant amounts
remain on floodplains (Baborowski et al. 2007). These may be retained in the topsoil layers
that provide humic compounds and clay minerals. However, floodplain soils are not a uniform
matrix, but are highly heterogeneous geosorbents composed of various sized grains, which
have different origin, formation, and physicochemical properties (Yang et al. 2008).
The findings presented in this study are part of the joint research project 'Flood retention and
drinking water supply − Preventing conflicts of interests' (RIMAX-HoT, Maier et al. 2006,
Kühlers 2009). This project aimed to identify potential conflicts of interests at the projected
retention basin Bellenkopf-Rappenwoert and a nearby planned waterworks Kastenwoert, both
located next to Karlsruhe, Germany. Chemical loads and hazard potentials of SPM were
previously detailed by Woelz et al. (2009a) and elevated PAH and HCB concentrations as
well as elevated AhR-agonist activities and mutagenic potencies were determined in a flood
in August 2007. Thus, the present study aimed to assess whether elevated compound
concentrations and biological activities can be determined in soil cores sampled at inundated
sites compared to non-inundated sites that are located behind a levee. Chemical analysis was
used to identify loads of PAHs, PCBs and HCB. 7-ethoxyresorufin-o-deethylase (EROD)
induction assay and Ames Fluctuation assay showed biological hazard potentials with respect
7.3 Materials and methods
129
to in vitro biotest systems. Since elevated compound concentrations and Ah-mediated
activities (computed as bio-TEQs) were measured and an automated fractionation procedure
(Lübcke-von Varel et al. 2008) was used to identify effective compound classes. Target
analysis showed shares of so-called priority EPA-PAHs (defined by the United States
Environmental Protection Agency, US-EPA) to the overall EROD induction. Accordingly,
fractions were assessed with the Ames Fluctuation assay and mutagenic potencies were
detected as caused by compounds in each fraction.
Thus, the present study aimed to
(a) measure chemical loads and biological responses in soil core layers from inundated
and non inundated sites,
(b) to use an automated fractionation procedure to identify effective fractions and shares
of target analytes, and
(c) to determine whether inundated sites have an impact in contrast to non-inundated sites.
7.3 Materials and methods
7.3.1 Chemicals used
Provider of chemicals used in this study will only be listed if other than Sigma-Aldrich,
Deisenhofen, Germany. Chemicals were at least reagent grade and LiChrosolv grade for
fractionation.
7.3.2 Soil sampling
In this study soil was sampled at August 22/23, 2006, at the projected retention basin
Bellenkopf-Rappenwoert. Soil was sampled at six locations in the basin area (north, middle,
south), with three of them sampled at inundated sites close to the river and three at sites
behind a levee that were, thus, not influenced by flooding (Fig. 1).
D
N
F
M
S
inundated foreland
non-inundated hinterland
Fig. 1 Location of the projected retention basin Bellenkopf-Rappenwoert near Karlsruhe, Germany.
Inundated foreland and non-inundated hinterland are separated by a levee (straight black line). Soil
was sampled in the north (N), middle (M) and south (S). Grey lines and filled areas give water courses
and basins in the basin. Black arrows show the river Rhine flow direction. D − Germany, F − France
7.3 Materials and methods
130
Soil was sampled from surface down to a depth of 90 cm using a viscoplastic standard
stainless steel soil corer according to Dr. Pürckhauer (diameter: 28 mm; Schierholz et al.
2000) and a maximum drilling depth of 1,000 mm. Each sample was further separated into
three sub-samples of 30 cm (0 - 30 cm, 30 - 60 cm, 60 - 90 cm). Samples were transferred to
glass bottles, transported at 4 °C and protected from light. Samples were shock-frozen at
-30 °C and freeze-dried on an Alpha 1 - 4 freeze-drier (Christ, Osterode, Germany) at -40 °C
and 0.1 mbar as fast as possible and stored at 4 °C in darkness until extraction.
7.3.3 Soil extraction for assessment of total samples
Soil extraction for assessment of total samples
Total soil samples were treated according to DIN 38414, part 22. Samples were freeze-dried
in two steps using a BETA 2 - 16 (Christ, Osterode, Germany). Initially, soil was dried for
two days at 0.6 to 1 mbar and a temperature of 20 °C to 25 °C. Subsequently, SPM were postdried for two days and at least 0.001 mbar to lower the residual moisture below 0.5 %. Soil
was than sieved at a mesh size of 600 µm for 15 min using an ultrasound bath type Bandelin
Sonorex RK 255 H (Schalltech GmbH, Mörfelden-Walldorf, Germany).
10 g of each freeze-dried soil sample were weighed in 200 ml extraction thimbles (Schleicher
& Schuell, Dassel, Germany) stoppered with glass wool, placed in 400 ml Soxhlet extractors
and extracted with 250 ml dichloromethane (Sigma-Aldrich, Deisenhofen, Germany) for 14 h
at 8 - 10 cycles per hour according to the method given by Hollert et al. (2000). The solvent
was reduced in volume and residues were evaporated under a gentle N2-stream. Residues
were re-dissolved in 1 ml n-hexane and stored at -20 °C until fractionation. Empty extraction
thimbles were subjected to extraction and processed in two parallel experiments to serve as
process controls.
7.3.4 Soil extraction and clean-up for fractionation
10 g of each freeze-dried soil layer was Soxhlet-extracted as detailed above using
dichloromethane (DCM):acetone (3:1; v/v) solvent mixture, reduced in volume, evaporated
under gentle N2-stream and re-dissolved in n-hexane:acetone (7:3; v/v). Further, a recently
developed membrane-assisted clean-up step (AMAC) technique was used for purification of
soil extracts according to the protocol by Streck et al. (2008). For this end, 1 ml extract with a
concentration of 10 g soil equivalent/ml was transferred to polyethylene dialysis membranes
and extracted using an ASE 200 device (Dionex, Sunnyvale, CA). Detailed extraction
conditions are provided by Lübcke-von Varel (2008). Extracts were sampled in ASE glass
vials closed by PTFE-coated screw caps. Extracts were reduced in volume, evaporated under
a gentle N2-stream and re-dissolved in n-hexane:DCM (9:1; v/v) to a final concentration of
10 g/ml for subsequent fractionation.
7.3.5 Automated fractionation procedure
Method according to chapter 6.3.4
7.4 Results
131
7.3.6 GC-MS analysis of fractions
GC-MS analysis were carried out on a HP 6890 GC coupled to a HP MSD 5973 (Agilent,
Palo Alto, USA), equipped with a 30 m x 0.25 mm I.D. x 0.25 µm film HP-5 MS fused
capillary silica column, a 5 m pre-column (Agilent J&W, Folsom, USA) and a splitless
injector with deactivated glass wool. Chromatographic conditions were as follows: 280 °C
injector temperature, 1 µl pulsed splitless injection at oven temperature of 60 °C (1 min
isotherm), then programmed at 30 K/min to 150 °C, at 6 K/min to 186 °C and finally at
4 K/min to 280 °C (16.5 min isotherm). Carrier gas velocity (Helium 5.0, Air Liquide,
Boehlen, Germany) was 1.3 ml/min at constant flow. The MS was operated in electron impact
ionization mode (EI+, 70 eV) with a source temperature of 230 °C scanning from 30 to
500 amu (full-scan mode) or single ion monitoring (SIM) for quantification. Target analytes
were quantified using an external calibration in single ion monitoring (SIM). The results were
corrected with an internal standard containing deutered PAH (Mix 35, Promochem, Wesel,
Germany).
7.3.7 EROD induction assay
Method according to chapter 6.3.7.
7.3.8 Bio-TEQ values
Computing of bio-TEQs according to chapter 6.3.8.
7.3.9 Ames Fluctuation assay
Method according to chapter 6.3.9.
7.4 Results
7.4.1 AhR-mediated activities and identified compounds
Soil sampled at recently inundated sites and at non-inundated sites was assessed in 30 cm
layers down to a depth of 90 cm with respect to PAHs, PCBs and HCB EROD induction
(Fig. 2). At each location, the highest concentrations of EPA-PAHs could be identified in the
top soil layer (0 - 30 cm). Highest load was measured with 39 mg/kg in the topsoil of the
northern inundated foreland (NF). Further, PCBs (0.19 mg/kg) as well as HCB (0.049 mg/kg)
indicated this sample to be highest contaminated site. Deeper soil layers at this site gave
decreasing compound concentrations and only EPA-PAH concentrations were elevated below
60 cm depth. At each other site, PAHs were the only detectable substances and measured in
topsoil above a depth of 30 cm. Concentrations were equal and ranged between 0.083 and
0.127 mg/kg in the inundated foreland and 0.07 and 0.142 mg/kg in the non-inundated area
behind the levee.
In accordance to the chemical analysis, EROD inductions indicated the NF soil extract as
sample with the highest Ah-receptor inducing potencies and maximum induction was
7.4 Results
132
determined in the topsoil layer with a bio-TEQ of 43,000 pg/g. In contrast to the NF site,
AhR-agonists were less active at the other inundated sites shown with mean concentrations
among all of 153 ± 0.7 pg/g and at sites behind the levee with means of 129 ± 77 pg/g. Lower
soil layers showed decreasing Bio-TEQs, at least compared to the topsoil at each site
investigated.
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Fig. 2 Concentrations of chemically analyzed HCB, PCBs and PAHs as well as bionalytically
determined bio TEQs are shown for distinct soil layers (0 - 30, 30 - 60, 60 - 90 cm), allowing the
comparison of samples from the north (N), middle (M) and south (S) of the inundated foreland (F) and
the non-inundated hinterland (H) which are separated by a levee.
7.4 Results
133
7.4.2 EROD inducing potential by soil fractions
In order to identify active compound categories in the EROD assay, the highest inducing soil
layer (0 - 30 cm at the inundated north site) was selected for effect-directed analysis (Fig. 3).
Fractions F1 to F4 containing PCBs and PCDD/Fs indicated no or minor activities. In
contrast, PAH fractions (F5 to F11) were highly inducing and showed highest bio TEQ
(= 13 x 103 pg/g) in fraction 10, containing PAHs with 6 aromatic rings (e.g. indeno[1,2,3-cd]
pyrene). Fraction 12 containing mainly mononitro-PAHs gave negligible inductions. In
contrast, F14 to F17 with more polar compounds indicated elevated activities, but about
4-fold lower than PAH fractions. Induction of F18 with most polar compounds induced
negligible.
50000
Bio-TEQ
Crude extract
Added fractions F1 to F18
Added PAH fractions F6 to F12
Added more polar to polar fractions F13 to F18
NF
Bio-TEQ [pg/g]
40000
30000
20000
10000
30
0
***
*
C A P M 1 2 3 4 5 6 7 8 9 101112131415161718
Fractions
Fig. 3 EROD induction given as bio-TEQs determined with HPLC fractions of the topsoil layer
sampled at the NF site. n.d. − no bio-TEQ determined
7.4.3 Mutagenic potential of individual fractions
Total soil extracts of each site caused no mutagenic activity with the Ames Fluctuation assay.
In contrast, fractions of the NF sample were investigated and showed elevated potentials.
Significantly decreased NOECs and increased IFmax were measured in all soil fractions except
for fraction F8 containing PAHs with four aromatic rings (Tab. 1). Highest potentials were
determined in fraction F17 and TA 98 and S9 metabolism (0.03 mg dry soil equivalent per ml
test medium) containing more polar compounds (e.g. 2-hydroxyanthraquinone). For all
fractions, highest potentials were determined in the Ames strain TA98 without S9
metabolism. However, fractions treated with TA98 and metabolic activation showed higher
mutagenicity. Tester strain TA 100 indicated fractions to be less active. Lower activities were
determined in the approach without exogenous S9 supplementation. Fractions were nonactive in the Ames tests with S9, except F16 that was the highest inducing fraction with
TA100 as well as one of the highest inducing fractions at all.
7.5 Discussion
134
Tab. 1 Mutagenic activity of HPLC factions determined in the Ames Fluctuation assay with the
bacterial tester strains TA 98 and TA 100 with and without adding exogenous S9 supplement for
metabolic activation of the NF soil sample. IFmax were only computed for fractions with significantly
reduced NOECs. Data are given as no observed effect concentration in mg soil equivalent per ml test
medium and as maximum induction factor.
NOEC [mg/ml]
Fraction
TA98 TA98
No.
-S9
+S9
1
*
1.04
2
*
16.67
3
8.34
8.34
4
8.34
*
5
*
2.08
6
4.17
4.17
7
*
16.68
8
*
*
9
1.04
1.04
10
4.17
2.08
11
2.08
2.08
12
2.08
2.08
13
*
2.08
14
2.08
*
15
*
*
16
2.08
*
17
0.07
0.03
18
0.13
0.26
Maximum induction factor (IFmax)
TA100 TA100 TA 98 TA 98 TA100 TA100
-S9
+S9
-S9
+ S9
-S9
+S9
*
*
*
4.3
*
*
*
*
*
6.7
*
*
*
*
6.7
3.9
*
*
*
*
5.7
*
*
*
*
*
*
7.1
*
*
16.67
*
3.6
5.7
8.0
*
*
*
*
3.3
*
*
*
*
*
*
*
*
*
*
16.1
13.3
*
*
*
*
8.6
19.0
*
*
16.67
*
4.3
12.3
3.3
*
16.67
*
4.3
1.9
2.4
*
*
*
*
2.9
*
*
*
*
6.7
*
*
*
16.67
*
*
*
4.3
*
*
1.04
16.7
*
*
*
*
*
19.0
29.0
*
*
*
*
19.0
14.3
*
*
* − NOEC > 33.33 mg SEQ/ml
7.5 Discussion
7.5.1 Chemical contamination of crude extracts
In order to determine contaminations in soil of inundated and non-inundated sites soil layers
were assessed at different locations of the projected retention area. Chemical analysis
indicated that soil sampled at the inundated area in the north was highest polluted with respect
to EPA-PAHs, PCBs and HCB. Since that site is a ground swale it acts as an accumulation
basin for SPM, especially during flooding and may explain elevated compound concentrations
in the soil core.
EPA-PAHs were determined in each top soil layer at inundated and non-inundated sites.
However, the NF soil core showed elevated concentrations down to a depth of 90 cm, which
was comparable to topsoil layer concentrations at any other site. Along with PCBs and HCB,
these findings show the high pollution at site and indicate that contaminants were translocated
deep into the unsaturated zone. EPA-PAH concentrations measured of about 40 mg/kg were
comparable to maximum concentrations of 20 mg/kg that were measured by Hilscherova et al.
7.5 Discussion
135
(2007) in floodplain soil of the rivers Morava and DÍevnice, Czech Republic, and its
tributaries after an extreme flood event with a recurrence interval of 100 years. In contrast,
concentrations measured were minor, if compared to 2,600 mg/kg EPA-PAHs as determined
by Eom et al. (2007) in soil of a highly PAH contaminated site at a former coke oven plant.
PCBs were detectable at NF down to a depth of 60 cm with highest concentrations
(0.19 mg/kg) in the topsoil. Measured concentrations were in accordance to the above-named
study of Hilscherova et al. (2007) that gave about 0.1 mg/kg in the floodplain soil measured
before the flood event at one site. Thus, soil of both catchment areas, the Rhine and the
Morava and DÍevnice rivers indicated comparable EPA-PAH loads following an extreme
flood, whereas PCB concentrations were usually lower, if compared to NF, but were
increased compared to any other site assessed in the present study.
HCB concentrations can be used as a tracer to indicate SPM deposition following inundation,
since this compound is a specific contaminant of the river Rhine basin and, thus, elevated
concentrations should not be detectable at non-inundated sites. Whereas HCB was highest
concentrated in topsoil and 0.05 mg/kg were measured, SPM sampled in a flood in August
2007 with a recurrence interval of 10 years gave a maximum of 0.11 mg/kg at the peak of the
flood (Wölz et al. 2009). Thus, HCB was equally concentrated in the topsoil layer sampled of
the swale and with flood SPM. This may indicate that HCB concentration in the soil layer
originate from inundation, since HCB is less degradable and, therefore, tends to accumulate
(Heise & Foerstner 2006, Isensee et al. 1976). Further, HCB measured in sediments sampled
at a depth of 0.2 to 1.2 m at the river Rhine barrage of Iffezheim showed concentrations of
0.22 µg/kg and 0.04 µg/kg in surface sediments (Alcock et al. 1998). Thus, concentrations of
HCB were comparable to sediments of the barrage, SPM sampled at the peak of a flood with a
recurrence interval of 10 years and soil from the projected retention basin that is located
downstream the barrage. This may furthermore indicate that soil concentrations are due to
inundation and deposition of HCB loaded SPM. With respect to regulatory thresholds, HCB
concentrations at this highest contaminated site among all assessed sites are still below, e.g.,
children playgrounds activity levels of 4 mg/kg as mentioned in the German Soil
Conservation Act (1999). Nevertheless, these concentrations highlight the hazard of deposited
particle-bound contaminants at inundated sites.
In the pilot study of this project, Ulrich et al. (2002) assessed some topsoil layers with respect
to EPA-PAHs and HCB that had been influenced by inundation of some weeks prior to
sampling. EPA-PAH concentrations varied between 0.19 to 0.76 mg/kg at rarely inundated
sites and 0.37 to 1.64 mg/kg at frequently inundated sites. Accordingly, HCB concentrations
were determined and were measured with < 0.001 to 0.002 mg/kg (rarely inundated sites) and
0.015 to 0.053 mg/kg (frequently inundated sites). Presence of HCB in soil of the pilot study
that was not detected in the present study might be due to the previous inundation. However,
HCB concentrations were equal to the highly polluted NF site in this study. In contrast,
EPA-PAHs were about 50-fold higher concentrated compared to the pilot study. These
findings show elevated differences of contamination levels in time and space that might also
be due to the high soil heterogeneity at the projected retention basin. Thus, an evaluation of
pollution levels may only be secure with measurements from many carefully selected sites.
7.5 Discussion
136
It can be stated that HCB and EPA-PAH concentrations can be elevated at defined sites in
floodplains and retention basins. HCB and higher molecular PAHs might be minor dissolved
or translocated due to water passage through the soil. However, frequent translocation
processes at highly polluted sites with deep reaching contaminations may pose a hazard to the
aquifer and drinking water resources over time.
7.5.2 Biological hazard potentials by crude extracts
Each sampling site indicated elevated AhR-agonist activities at least with topsoil layers and
the maximum EROD induction, in accordance to chemical analysis, was determined with the
foreland site NF. Maximum inductions were determined with bio-TEQ of 43,000 pg/g.
Anderson et al. (2009) investigated soil sampled at a PAH contaminated site and determined a
maximum bio-TEQ of about 45,000 pg/g using the luciferase gene expression (CALUX)
assay. Since different test systems and cell lines were used, that TEQ may not directly be
compared to the maximum TEQ determined in the present study. However, a comparison
indicates the high pollution at the NF site. Lower bio-TEQ of about 10,000 pg/g were
determined by Keiter et al. (2008) assessing sediment from the Danube river. Further, SPM
sampled in a flood with a recurrence interval of two years at the river Rhine showed bio-TEQ
of about 2,300 pg/g and SPM of a flood at the river Neckar with the same recurrence interval
showed maximum bio-TEQ of 8,300 pg/g (Wölz et al. 2008). TEQ values at the NF site are
elevated compared to the specified sediment and SPM concentrations and, thus, cannot be
explained completely by particle deposition on floodplain soils. However, published studies
providing data on EROD inducing sediment and SPM samples are rare and detailed bio-TEQs
may potentially act as snap-shots that do not provide information on effective TEQ ranges.
Fractions containing PAHs were determined to contribute highest to the overall biological
activity of the sample, in particular the fractions containing compounds of four to six aromatic
rings (F8 to F10). Target analysis to determine contributions of EPA-PAHs to the overall
biological effect showed that far less than 1 % of the overall biological activity could be
explained with theses priority compounds (therefore, data not detailed). These contributions
were surprisingly low since other studies showed that EPA-PAHs contributed at least to an
extent of some percent to the overall biological effect. Further, these compounds were once
prioritized since they were highly concentrated in environmental compartments. Although the
NF site showed the relatively highest EPA-PAH concentrations of all assessed soils, total
concentrations were low and contributions to EROD induction were minor. Thus, other so far
unknown and non-priority PAHs were responsible for the detected EROD inductions and
mutagenic potentials in PAH fractions (Brack et al. 2005, Wölz et al. 2008).
7.5.3 Identification of active fractions
Since EPA-PAHs concentrations and bio-TEQs were found to be highest concentrated with
the NF site topsoil layer, this sample was used for an automated fractionation procedure to
identify EROD inducing fractions and compound classes.
7.5 Discussion
137
Pattern of EROD-inducing fractions were in line with previously published investigations of
suspended particulate matter in the RIMAX-HoT project, using the same automated
fractionation method (Wölz et al. RIMAX-SPM 2009). With respect to PAH containing
fractions, findings were further in accordance to SPM sampled at the Rhine in another study
(Wölz et al. 2009b) and fractionated using a precursor fractionation method (Brack et al.
2003a). Effect pattern comparable to other studies at the river Rhine may advice that the
highly polluted NF site soil is influenced by (frequent) inundation and, further, that particlebound contaminants were deposited at site. However, inducing potencies among the fractions
varied between the different investigations. The highest bio-TEQ (= 6.500 pg/g) was
determined by Woelz et al. (2009b) at the flood peak, being about 7-fold less concentrated
than maximum NF soil concentrations in this study. Thus, this site is relatively high
contaminated what might be due to the accumulative item since the site is a ground swale.
Further, more polar compounds (e.g. (hydroxyl-)quinones, keto-, dinitro-, hydroxy-PAHs, and
N-heterocycles with rising polarity, 2-hydroxyanthraquinone) were determined to cause
elevated EROD inductions. More polar to polar compounds were determined to cause
elevated effects in some studies investigating sediments and SPM in the recent years (Keiter
et al. 2008, Wölz et al. 2009b). Thus, these compounds which are often given less attenuation
should be set in the focus of upcoming research into environmental pollution. Even more so,
since higher polarity indicates that compounds are more likely to be dissolved, and, thus, are
better bioavailable in the aquatic environment than non-polar PCBs, PCDD/Fs and most
PAHs. Investigations of more polar and polar compounds were not in the focus of this study,
and, thus, effective compounds may only be discussed as model compounds. Likewise,
Petrovi et al. (2003) discussed emerging contaminants such as surfactant degradates,
pharmaceuticals and polar pesticides.
Next to EROD induction mutagenic potentials of crude extracts from each sediment site and
layer as well as of the fractions from the NF site were assessed. However, in contrast to
fractions crude extracts showed no significant effects. This indicates the removal of masking
or inhibiting compounds during the fractionation procedure. Following fractionation, masking
compounds might be separated in parts or completely in (therefore) non-active fractions
(Brack et al. 2005). Although cytotoxicity was not quantified by photometrical means, optical
inspection advised at least no viewable cytotoxic effects that would turn out as less turbid
well bottoms compared to the negative control. Thus, cytotoxicity of the crude extract versus
the bacteria strains may be excluded as a reason of masking (Chenon et al. 2003). Using the
example of the NF site, the soil layers showed no significant reversion, but most fractions
indicated elevated or highly increased mutagenic activity at least with the tester strain TA 98.
Thus, in the crude extract antagonistic processes – inhibiting frameshift mutations – may have
inhibited DNA interferences, since cytotoxic effects were not observed. Mutagenically active
fractions containing PAHs, more polar and polar compounds were shown to be mutagenic in
other studies before (Fernandez et al. 1992, Thomas et al. 2002).
Fractions were shown to commonly cause increased reversion rates and highest maximum
induction factors with tester strain TA 98, whereas TA 100 indicated only some active
fractions and minor induction factors. Thus, most fractions caused frameshift mutations.
7.6 Conclusions
138
Further, most fractions showed significant mutagenic activity with TA98 and direct as well as
with indirect S9 treatment. For these fractions, mutagenic potentials were comparable with
respect to NOECs and IFmax. However, there is no clear trend towards direct or indirect
mutagenic activity. It may only be stated that mutagenic potentials were increased in fractions
containing PAHs that are well known inducers of mutagenic effects (Brack et al. 2005, Perez
et al. 2003, White 2002) as well as in fractions with moderately polar and polar substances
(Marvin & Hewitt 2007, Villalobos-Pietrini et al. 2007).
At least elevated mutagenic potentials in the latter fractions might be hazardous in the context
of the dispute between flood retention and drinking water supply. Compounds with polar
characteristics are more likely to be solved and transported in water and, thus, might more
easily pass the unsaturated zone and reach the aquifer. However, the investigations of total
soil and fractions were carried out using organic extracts of soil matter and, thus, e.g.,
bioavailability has not been addressed. Nevertheless, independent of availability to organisms
high contaminations were determined and these have to be considered in risk evaluation in the
named conflict of interests.
7.6 Conclusions
Remobilized and highly contaminated sediments may be translocated to floodplains and
preferably deposit in ground swales that act as accumulation basins for pollutants. Thus,
topsoil layers of inundated sites can show elevated contaminant loads and biological
responses with in vitro biotests. Typical contaminants of rivers and pattern of biological
responses in fractions can be used as tracers to indicate pollution as a consequence of
flooding.
To evaluate contamination and contaminant translocation at site, soil should be investigated
as a core with distinct layers since topsoil samples provide only limited information of surface
pollution. This is of relevance since soils may be addressed as stable long-term memories of
contamination levels and patterns. Thus, contaminated soils should be considered as
important secondary sources of pollutants, in particular in floodplains with facilitated
compound translocation due to flooding. In this context, elevated chemical loads and
biological responses, in particular in deeper soil layers, indicate that contaminants may more
easily pass the unsaturated zone with its retardation mechanisms in the topsoil.
Elevated EROD inductions and maximum mutagenic inductions were determined at least with
fractions containing PAHs, moderately polar and polar compounds. At least, PAHs are more
likely to be dissolved and can more easily be translocated through the unsaturated zone into
the aquifer. Thus, with respect to the assumed conflict of interests between retention basins
and water protection areas, impacts through contaminant translocation to the unsaturated zone
of the aquifer cannot be excluded with present investigations. Since the addressed retention
basins will be operated in floods with recurrence intervals of greater than or equal to
100 years, considerably more sediment erosion and contaminant remobilization will be likely
to take place. Elevated deposition or transfer of contaminants might impacts soils and aquifers
due to stored and akinetic water. This instance should be addressed carefully in planning the
7.6 Conclusions
139
morphology and operation of retention basins. Furthermore, with respect to
(eco-)toxicological impacts, the hazard potential of flood events and inundation to the aquifer
may assist to evaluate risks towards water quality and, thus, were in the focus of another part
of the RIMAX-HoT project.
Acknowledgements
The authors would like to express their thanks to Drs. Niels C. Bols and Lucy Lee (University
of Waterloo, Canada) for providing RTL-W1 cells. We want furthermore to express our
thanks to Georg Reifferscheid (Federal Institute of Hydrology, Koblenz, Germany).
We thank Kerstin Winkens, Anne Schneider, Susanne Miller, Conny Bernecker, Ulrike Diehl
for assistance with conducting biotests. We also thank the Federal Ministry of Education and
Research (BMBF), Germany, for supporting the RIMAX-HoT project within the RIMAX
joint, No. 02WH0693.
7.7 References
140
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Maier M, Kuehlers D, Brauch H-J, Fleig M, Maier D, Jirka GH, Mohrlock U, Bethge E, Bernhart HH,
Lehmann B, Hillebrand G, Wölz J, Hollert H (2006): Flood retention and drinking water supply
- Preventing conflict of interests. J Soils Sediments 6: 113-114
Maron DM, Ames BN (1983): Revised methods for the Salmonella mutagenicity test. Mutat Res 113:
173-215
Marvin CH, Hewitt LM (2007): Analytical methods in bioassay-directed investigations of
mutagenicity of air particulate material. Mutat Res 636: 4-35
Perez S, Reifferscheid G, Eichhorn P, Barcelo D (2003): Assessment of the mutagenic potency of
sewage sludges contaminated with polycyclic aromatic hydrocarbons by an Ames Fluctuation
assay. Environ Toxicol Chem 22: 2576-2584
Petrovi M, Gonzalez S, Barceló D (2003): Analysis and removal of emerging contaminants in
wastewater and drinking water Trends Anal Chem 22: 685-696
Reifferscheid G, Arndt C, Schmid C (2005): Further development of the beta-lactamase MutaGen
assay and evaluation by comparison with Ames fluctuation tests and the umu test. Environ Mol
Mutagen 46: 126-139
RIMAX-HoT joint research project (2005-2009): http://rimax-hot.ifh.uni-karlsruhe.de/
Schierholz I, Schäfer D, Kolle O (2000): The Weiherbach data set: An experimental data set for
pesticide model testing on the field scale. Agricul Water Manage 44: 43-61
Streck HG, Schulze T, Brack W (2008): Accelerated membrane-assisted clean-up as a tool for the
clean-up of extracts from biological tissues. J Chromatogr A 1196-1197: 33-40
7.7 References
142
Stronkhorst J, van Hattum B (2003): Contaminants of concern in Dutch marine harbor sediments.
Arch Environ Contam Toxicol 45: 306-316
Thomas KV, Balaam J, Barnard N, Dyer R, Jones C, Lavender J, McHugh M (2002): Characterisation
of potentially genotoxic compounds in sediments collected from United Kingdom estuaries.
Chemosphere 49: 247-258
Ulrich M, Schulze T, Leist E, Glaß B, Maier M, Maier D, Braunbeck T, Hollert H (2002):
Ökotoxikologische Untersuchung von Sedimenten und Schwebstoffen - Abschätzung des
Gefährdungspotenzials für Trinkwasser und Korrelation verschiedener Expositionspfade
(Acetonischer Extrakt, Natives Sediment) im Bakterienkontakttest und Fischeitest.
Z Umweltchem Ökotox 14: 132-137
Villalobos-Pietrini R, Hernandez-Mena L, Amador-Munoz O, Munive-Colin Z, Bravo-Cabrera JL,
Gomez-Arroyo S, Frias-Villegas A, Waliszewski S, Ramirez-Pulido J, Ortiz-Muniz R (2007):
Biodirected mutagenic chemical assay of PM(10) extractable organic matter in Southwest
Mexico City. Mutat Res 634: 192-204
Wassen MJ, Peeters WHM, Venterink HO (2002): Patterns in vegetation, hydrology, and nutrient
availability in an undisturbed river floodplain in Poland. Plant Ecol 165: 27-43
Weber R et al. (2008): Dioxin- and POP-contaminated sites - Contemporary and future relevance and
challenges. Environ Sci Pollut Res 15: 363-393
White PA (2002): The genotoxicity of priority polycyclic aromatic hydrocarbons in complex mixtures.
Mutat Res 515: 85-98
Wölz J, Engwall M, Maletz S, Olsman H, Van Bavel B, Kammann U, Klempt M, Braunbeck T,
Hollert H (2008): Changes in toxicity and Ah receptor agonist activity of suspended particulate
matter during flood events at the rivers Neckar and Rhine - A mass balance approach using in
vitro methods and chemical analysis. Environ Sci Pollut Res 15: 536-553
Wölz et al. RIMAX-SPM (2009a): The RIMAX-HoT project: 1. Impacts of suspended particulate
matter sampled at the river Rhine with respect to operation of retention basins and drinking
water safety. In preparation
Wölz J, Brack W, Möhlenkamp C, Claus E, Braunbeck TH, H (2009b): Identification of dioxin-like
activity in suspended particulate matter of floods using effect directed analysis. Submitted to Sci
Tot Environ
Yang Y, Ligouis B, Pies C, Grathwohl P, Hofmann T (2008): Occurrence of coal and coal-derived
particle-bound polycyclic aromatic hydrocarbons (PAHs) in a river floodplain soil. Environ
Pollut 151: 121-129
Zonta R, Collavini F, Zaggia L, Zuliani A (2005): The effect of floods on the transport of suspended
sediments and contaminants: A case study from the estuary of the Dese River (Venice Lagoon,
Italy). Environ Int 31: 948-58
143
Chapter 8
8 Contaminant entry into and transport in the
saturated groundwater zone subsequent to
extreme flood events
J. Wölz1, K. Grosshans1, M. Fleig2, G. Streck3, T. Schulze3, A. Rastall4, L. Erdinger4,
W. Brack3, D. Kühlers5, T. Braunbeck6, H. Hollert1
1
Department of Ecosystem Analysis, Institute for Environmental Research, RWTH Aachen
University, Worringerweg 1, 52074 Aachen, Germany
2
DVGW-Water Technology Center (TZW), Chemical Analysis Department, Karlsruher Strasse 84,
76139, Karlsruhe, Germany
3
UFZ Helmholtz Centre for Environmental Research, Department of Effect-Directed Analyses,
Permoserstrasse 15, 04318 Leipzig, Germany
4
Institute of Hygiene and Medical Microbiology, University of Heidelberg, INF 324, 69120
Heidelberg,
5
Stadtwerke Karlsruhe GmbH (SWK), Karlsruhe, Germany
6
Department of Zoology, Aquatic Toxicology and Ecology Section, University of Heidelberg, Im
Neuenheimer Feld 230, 69120 Heidelberg, Germany
To be published in Journal of Soils and Sediments
144
8.1 Abstract / 8.2 Introduction
145
8.1 Abstract
In this study, the yeast estrogen screen (YES) assay with Saccharomyces cerevisiae was used
to determine the estrogenic potential of solid phase-extracted water samples from three
groundwater wells in a projected retention basin near Karlsruhe, Germany. Further, fractions
derived from a recently developed fractionation procedure of a highly polluted soil as well as
of suspended particulate matter (SPM), sampled during a flood event at the river Rhine at the
barrage of Iffezheim, Germany, was assessed for endocrine activity. Target analysis was
applied to identify effective compounds. Estrogenic activities of each sample were expressed
as 17ß-estradiol equivalent concentrations (E2-EQ). Groundwater was sampled between
June 2006 and January 2008 and more frequently following a flood event with a recurrence
interval of 10 years in August 2007. Well no. 1 located closest to the river Rhine showed
elevated concentrations of the river trace compound carbamazepine (CBZ) at all sampling
times. Subsequent to the August flood, concentrations were also elevated in well no. 2.
Further, E2-EQ indicated a flood dependent increase of estrogenic activity in the month
following the flood at well no. 1. Groundwater sampled at well no. 2 showed increasing
E2-EQ with a delay of some weeks. Heterogeneous but also highest inductions
(6.7 ng E2-EQ/L) were detected at well no. 3 with the longest distance to the Rhine.
Since elevated YES activities were determined in groundwater, and since the hypothesis has
been stated that translocation of particle bound pollutants may influence the ground water, an
automated fractionation method and target analysis were used to identify effective compound
classes and single compounds of SPM and soil extracts. Fractions F13 to F18 containing more
polar compounds (e.g. (hydroxyl-)quinones, keto- , dinitro- , hydroxy-PAHs, N-heterocycles,
hydroxyanthraquinone), caused elevated endocrine activities in the soil sample and the SPM.
Further, fractions F6 to F12 containing PAHs showed minor effects for the fractionated soil
sample and F4 induced significantly with SPM 2 sampled after the flood peak. Added
fractions E2-EQ gave comparable activities for soil (2 ng E2-EQ/g) and SPM (0.9 and
2.3 ng E2-EQ/g). Target analysis identified minor concentrations of active compounds and,
thus, other non-analyzed substances were effective. This indicates the need of further
investigations with respect to more polar and polar compounds and their potential hazard to
drinking water.
This study was part of a project (RIMAX-HoT) that aimed at the identification/
characterization of the possible conflict of interests between flood management (retention
basins) and drinking water supply (waterworks close to the basin). With respect to the
question, whether the operation of retention basins increases the risk of contaminant
introduction into the aquifer (drinking water resource), the results presented document a
significant hazard potential.
8.2 Introduction
Ground water contamination in riparian areas is (still) of increasing concern in many regions
worldwide, since they often are used as drinking water resources (Levin et al. 2002). Once the
8.2 Introduction
146
aquifer is contaminated, residues may remain for long time. Further, the movement of
groundwater is difficult to monitor and there are substantial time lags between emissions and
detection of chemicals (Finch et al. 2007). Whereas groundwater contamination may result
from mineralization or other natural processes, it is usually attributed to waste disposal
practices and industrial and agricultural activities (Böhlke 2002, Naik et al. 2007). These
activities may continuously impact the groundwater quality. Nevertheless, further events such
as floods have the potential to heavily pollute the aquifer in riparian and inundated areas. At
least in floods with higher recurrence intervals, thus being more hazardous, considerable
amounts of sediments may be eroded (Hollert et al. 2007a). These suspended sediments
contribute to suspended particulate matter (SPM), which is translocated downstream or to
flooded sites along the rivers. Remobilization of sediments impacts inundated areas since
sediment acts as sink and important secondary source of contaminants (Brils 2008, Kosmehl
et al. 2004). Whereas particle-bound compounds deposit and consolidate most time of the
year in rivers, they may become available by erosion during flood events. Thus, they can be
translocated and deposited at inundated sites. Further, groundwater is recharged by inundating
flood water that infiltrates the soil at floodplains and indicates the potential of mass transfer
through the unsaturated zone in the saturated zone of the aquifer (Brouyere et al. 2004,
Kazamaa et al. 2007). Contaminant introduction into the aquifer is of elevated interest since
major streams as the river Rhine and its aquifer are important in terms of drinking water
abstraction and, thus, are in the focus of scientific investigations (Schwarzbauer & Heim
2005). In the European Union, the Water Framework Directive (WFD; 2000/60/EC)
constituted a general set of subjects of protection and objectives to achieve a 'good water
status' for all waters, by 2015. This includes the protection of groundwater resources and,
thus, the WFD demands measures to ensure the progressive reduction of groundwater
pollution and to prevent its further pollution (Borja et al. 2004).
The presented study details results of the joint research project 'Flood retention and drinking
water supply – Preventing conflicts of interests (RIMAX-HoT)' as introduced by Maier et al.
(2006). The project aimed to characterize the possible conflict of interests at the planned
retention area Bellenkopf-Rappenwoert and a projected nearby waterworks Kastenwoert, both
located next to Karlsruhe, Germany. This part of the study focused on the assessment of
endocrine effects in groundwater and soil as well as suspended particulate matter (SPM)
samples, which were taken during the project. Estrogenic activity was determined using the
Yeast Estrogen Screen (YES) assay with transgenic bakery yeast Saccharomyces cerevisiae,
containing the human estrogen receptor (hER).
Groundwater was sampled over a period of 2 years at three groundwater wells in the projected
retention basin. Since elevated YES activities were determined in groundwater, and since the
hypothesis has been stated that translocation of particle-bound pollutants may influence the
ground water, an automated fractionation method and target analysis were used to identify
effective compound classes and single compounds of SPM and soil extracts. The soil
originated from a site which proved highly contaminated in a previous study (Woelz et al.
2009a). Furthermore, SPM sampled in the course of a flood with a recurrence of 10 years,
showing elevated loads and toxic effects, was assessed (Woelz et al. 2009b).
8.3 Material and methods
147
This last study within the RIMAX-HoT project aimed
•
to assess endocrine inducing potentials in groundwater samples of the model retention
basin,
•
to investigate agonist activities in fractions of soil (sampled at the basin) and SPM
(sampled at the barrage of Iffezheim close to the basin),
•
to use target analysis to possibly identify effective compounds and
•
to conclude how these findings can assist to resolve the conflict between floods to
groundwater quality and drinking water safety.
8.3 Materials and methods
8.3.1 Chemicals used
Subsequent provider of chemicals will only be listed if other than from Sigma-Aldrich,
Deisenhofen, Germany. Chemicals were at least reagent grade.
8.3.2 Sampling and Preparation
Groundwater was sampled at three wells situated at a diagonal within the projected retention
basin Bellenkopf-Rappenwoert and located behind a levee (Fig. 1) at several times over a
period of 2 years (Tab. 1). Well 1 is situated directly behind a levee and in a distance of
250 m to the Rhine. Distances between well no. 1 and no. 2 as well as between well no. 2
and no. 3 were about 500 m. Wells had a depth of about 30 m. Before sampling groundwater,
each well was pumped dry (MP1; Grundfos, Grödig, Austria) and allowed to refill again to
update the sampled groundwater with 2.5 m3/h. Water was pumped into 2 L brown glass
bottles, transported to a cooling chamber and stored at 4 °C until extraction.
D
F
R
1
2
3
B
B
Groundwater wells
Soil sampling site
Fig. 1 Location and scheme of the projected retention basin Bellenkopf-Rappenwoert near Karlsruhe,
Germany. Grey lines and filled areas give water courses/resources in the basin. F − France,
D − Germany, B − Bellenkopf, R − Rappenwoert
In order to investigate whether elevated estrogenic activities in groundwater samples were
caused by contaminants of flood water, SPM that was sampled in the course of a flood with a
8.3 Material and methods
148
recurrence interval of 10 years as well as a soil sample of an inundated site were additionally
assessed as detailed by Wölz et al. (2009 a,b). SPM 1 was sampled between 10.40 a.m. and
12.00 a.m. at August 10, 2007 before the flood peak and SPM 2 was sampled between
12.00 a.m. at August 10, 2007, and 14.50 p.m. at August 11, 2007 after the flood peak. The
soil was sampled at August 22/23, 2006 in the inundated foreland of the projected retention
basin (shown in Fig. 1).
Tab. 1 Sampling schedule for ground water collection at wells numbered according to fig. 1.
Date
26.06.2006
27.08.2006
30.05.2007
13.08.2007
16.08.2007
22.08.2007
31.08.2007
10.09.2007
11.10.2007
19.11.2007
17.12.2007
14.01.2008
Groundwater wells
1
2
3
x
x
x
x
x
x
x
x
x
x
x
x
x n.a. n.a.
x
x
x
x n.a. n.a.
x n.a. n.a.
x
x
x
x
x
x
x
x
x
x
x
x
n.a. − not assessed
8.3.3 Water extraction
For extraction, water samples were filtered over 0.1 mm glass fibre filters (type C5,
MembraPure, Bodenheim, Germany), acidified with concentrated H2SO4 to pH 2.0, divided
into two 1 L samples and extracted using reverse phase C18 solid phase extraction columns
(RP-C18 SPE; 1 g; Bakerbond, J.T. Baker, Deventer, The Netherlands), which had been
conditioned with 3 x 3 ml n-hexane, 3 x 3 ml acetone as well as 1 x 3 ml deionised water
according to a protocol by Spengler et al. (2001) and Rastall et al. (2004). After extraction,
columns were centrifuged for 10 min at 2000 g and dried under a nitrogen stream. Elution
was carried out with 2 x 5 ml acetone. Eluted samples were blown close to dryness under a
nitrogen stream, reconstituted in 2.5 ml dimethylsulfoxide (DMSO) and stored at 4 °C until
chemical and biological analysis. Deionised water and tap water were treated and extracted
according to the same protocol and were used as process controls.
An automated fractionation method, bioanalytical investigations and target analysis were used
to identify effective compound classes and single compounds of SPM and soil extracts.
SPM 1 was sampled at the peak of a flood event (August 10, 2007, 12 p.m.) and SPM 2 after
the flood peak (August 11, 2007, 14:50 p.m.). The soil was sampled at an inundated and
highly contaminated site in the north of the projected retention basin in order to serve as a
worst case scenario in respect to hazard potential to the ground water (see fig. 1).
8.3 Material and methods
149
8.3.4 Automated fractionation of SPM and soil samples
The method according to chapter 6.3.4.
8.3.5 Chemical analysis – Carbamazepine (CBZ)
The HPLC (HP-1100, Agilent Technologies, Palo Alto, CA, USA) was equipped with a mass
spectrometer API 2000 (PE-SCIEX; Waltham, Massachusetts, USA), separation column
(250 x 2 mm, 5 µm C18 Luna; Phenomenex, Torrance, CA, USA), solid phase material SDB1
Bakerbond (JT Baker, Devender, The Netherlands) and a UV detector (UV-2201, Shanghai).
The analysis conditions were: initial column temperature 60 °C (1 min), increased at
20 °C/min to 180 °C, then increased at 3 °C/min to 207 °C and at 1.5 °C to 260 °C that were
finally hold for 5 min. The carrier gas was helium. The injector temperature was 50 °C,
300 °C/min to 270 °C and the volume injected in splitless mode was 4 µl. The detector
temperature was 310°C. Detection limit was 10 ng/L.
8.3.6 Method for the instrumental analysis of estrogenic compounds
Chemical analysis included four compound classes: Natural and synthetic steroids, nonyl- and
octylphenol, bisphenol A and musk compounds. Prior to analysis a derivatization of steroidal
compounds was inevitable for a proper determination with gas chromatography (Streck 2009).
All other compounds were processed without derivatization. Derivatization was done
modifying a method proposed by Labadie and Budzinski (2005). Briefly, all samples were
dried under a gentle stream of N2 and then re-dissolved in 50 µl of a mixture of MSTFA
(N-methyl-N-(trimethylsilyl)-trifluoroacetamide; purity > 97 %), 0.6 % 2-mercaptoethanol
(Merck, Darmstadt, Germany) and 1 mg ammonia iodide (Riedel-de Haën, Seelze, Germany).
The samples were heated for 40 minutes at 65 °C. After cooling to room temperature, the
samples were dried again with N2 to dryness and finally re-dissolved in 500 µL toluene.
Analysis was achieved using an Agilent HP6890 capillary column gas chromatograph
equipped with a HP5973 mass selective detector used in electron impact mode (70 eV). The
compounds were separated on an HP5-MS capillary column (length 30 m; inner diameter
0.32 mm; stationary phase thickness 0.25 µm). Helium was the carrier gas. The injector was
kept at a temperature of 250 °C, the interface between the gas chromatograph and the ion
source at 280 °C, while the ion source itself obtained a temperature of 250 °C. All injections
were done in splitless mode with a volume of 1 µl. The oven was programmed as follows:
60 °C for one minute, increasing at 30 K/min to 150 °C, continuing with a gradient of
6 K/min to 186 °C, then increasing at 4 K/min to the final temperature of 280 °C, which was
kept for 6.5 minutes. The analysis was conducted in SIM mode. Concentrations of the target
analytes were calculated using an external calibration and corrected by means of the injection
standards. Limit of detection was 7.5 mg/L.
8.3.7 Yeast Estogen Screen (YES) assay
Recombinant yeast cells (Saccharomyces cerevisiae) stably transfected with the gene for the
human estrogen receptor (hER) and containing expression plasmids carrying strong promoter
8.4 Results
150
sequences and the lac-Z (β-galactosidase) reporter gene (Routledge & Sumpter 1996) were
used to investigate estrogen-like activity of suspended SPM, soil and groundwater. The
YEAST Screen assay was carried out according to Routledge and Sumpter (1996) using a
slightly modified version of the procedure described by Rastall et al. (2004) and Keiter et al.
(2006). Briefly, 100 µl aliquots of extracted groundwater, soil and SPM were serially diluted
along alternate rows of a 96 well microtitre plate. A 17β-estradiol (E2) positive control was
added to a separate row and accordingly serially diluted to give a final concentration range of
1.0 x 10-8 to 4.8 x 10-12 M. 100 µl of the ethanol vehicle were then added to each vacant well
and the ethanol in all 96 wells allowed to evaporate. 50 ml of YES assay medium containing
500 µl of a 1.65 x 10-2 M aqueous solution of the chromogenic substrate chlorophenol-redβ-D-galactopyranoside (CPRG) and 4.0 x 107 recombinant yeast cells were then prepared and
200 µl transferred to each well. The plates were sealed and incubated at 32 °C for 72 h.
Estrogenic potentials were subsequently determined photometrically at 540 nm following the
conversion of the CPRG from yellow to red by β-galactosidase secreted into the growth
medium in response to the presence of hER agonists in the sample. With each test distilled
water was used as a negative control and, further, process controls were tested accordingly.
Significant activities compared to the negative control were determined using the 99 %
confidence interval.
Activities were computed as estrogen equivalent concentrations (E2-EQ) by normalizing
estrogen activities of samples or fractions to the natural estrogen 17ß-estradiol (Hollert et al.
2005, Tan et al. 2007). E2-EQs are calculated as the quotient of the EC50 (effective
concentration that produces 50% of the maximum effect level) and is given either in the
unit ng E2/L (subsequent ng/L) of groundwater or as ng E2/g Soil or SPM (subsequent ng/g):
Eq. 1 E2-EQ = EC50 [17ß-estradiol] / EC50 [sample]
8.4 Results
8.4.1 Investigation of groundwater samples
The groundwater sampled at three sites in a diagonal to the river Rhine indicated elevated
concentrations of river key contaminants such as the pharmaceutical CBZ (Fig. 2). Increased
but heterogeneous concentrations were measured at well no. 1, which is closest to the river
Rhine (distance: 250 m). Well no. 2 showed elevated concentrations, in particular following
August 2007. In contrast, CBZ was not detectable in water of well no. 3 with the longest
distance to the river Rhine.
With respect to endocrine activity, statistical significant but nevertheless negligible low
effects were determined with water sampled at well no. 1 and no. 2, whereas well no. 3
showed moderately elevated endocrine effects in 2006 (Fig. 2). Effects were low during the
following samplings, but showed elevated endocrine activities following the flood event of
August 2007. In particular, well no. 1 indicated increasing endocrine effects following the
flood and showed a decreasing tendency after August. The maximum E2-EQ was determined
with 2.9 ng/L (August 10, 2007). Likewise, well no. 2 showed a clearly increasing activity
8.4 Results
151
following the flood, however, with a delay and a maximum of E2-EQ = 2.6 ng/L
(September 10, 2007). Groundwater of well no. 3 caused endocrine activities in the YES
assay, which showed approximately the activity pattern of well no. 2 (rPearson = 0.80).
However, endocrine effects at well no. 3 were heterogeneous and not clearly increased in the
time after the flood. Well no. 3 showed the maximum E2-EQ of all samples assessed
(= 6.67 ng/L).
7
a
Well 1
Well 2
Well 3
40
6
5
EEQ [ng/L]
Cabamazepine [µg/L]
50
30
b
Well 1
Well 2
Well 3
4
3
2
20
1
26
.0
6
27 .06
.0
8
30 .06
.0
5
13 .07
.0
8
16 .07
.0
8.
22 07
.0
8
31 .07
.0
8
10 .07
.0
9
11 .07
.1
0
19 .07
.1
1
17 .07
.1
2
14 .07
.0
1.
08
*
Dates of sampling
0
* *
*
26
.0
6
27 .06
.0
8
30 .06
.0
5.
13 07
.0
8
16 .07
.0
8
22 .07
.0
8
31 .07
.0
8
10 .07
.0
9.
11 07
.1
0
19 .07
.1
1
17 .07
.1
2
14 .07
.0
1.
08
* *
10
Dates of sampling
Fig. 2 (a) Carbamazepine (CBZ) concentration and (b) estrogenic activities at each well and sampling
time with groundwater of the projected retention basin Bellenkopf-Rappenwoert. Endocrine activities
are given as E2-EQs in ng/L and were calculated from one assay with two measurements. Significant
endocrine activity was determined using the 99% confidence interval. * − wells no. 2 and 3 not
assessed
8.4.2 Estrogenic activity in individual fractions and target analysis
Suspended particulate matter no. 1 (SPM 1) caused endocrine activities in fractions
F14 to F17 containing more polar compounds (such as N-heterocycles) and maximum
activities were detected in F15 giving E2-EQ = 1.44 ng/L (Fig. 3). Addition of activities in
single fractions equaled E2-EQ = 2.3 ng/L. Extracts of SPM 2 caused lower activities in
fractions with more polar compounds. Nevertheless, highest endocrine effectiveness was
determined in fraction F15 (E2-EQ = 0.5 ng/L). Further, in contrast to SPM 2, fraction F4
caused (minor) endocrine activity (E2-EQ = 0.31 ng/L). Added single fractions of SPM 2
equaled E2-EQ = 0.9 ng/L. The soil surface layer (0 - 30 cm depth) indicated lower endocrine
activities in F6 to F8 (Fig. 3). Fractions containing the more polar compounds (F13 to 17)
caused elevated endocrine activities and maximum induction was determined in fraction F15
(E2-EQ = 0.7 ng/L). The total equivalent concentration of all fractions was E2-EQ = 2 ng/L.
8.4 Results
152
1,6
SPM 1
1,4
EEQ [ng/L]
1,2
1,0
0,8
0,6
0,4
0,2
0,0
* * * * * * * * * * * * * *
*
1,6
SPM 2
1,4
EEQ [ng/L]
1,2
1,0
0,8
0,6
0,4
0,2
0,0
1,6
* * * *
* * * * * * * *
*
*
Soil
1,4
EEQ [ng/L]
1,2
1,0
0,8
0,6
0,4
0,2
0,0
* * * * * *
0
2
4
** * *
6
8
10
12
*
14
16
18
Fractions
Fig. 2 Endocrine activities in HPLC-fractions of SPM sampled during a flood event in August 2007
(recurrence interval of 10 years) as well as of a topsoil layer (0-30 cm) of a site with elevated
contamination. Endocrine activity is given as E2-EQ in ng/L and was calculated from one assay with
two measurements. Significant endocrine activity was determined using the 99% confidence interval.
* − fractions not inducing
8.4.3 Target analysis in fractions
In order to identify effective compounds, target analysis was conducted; data are provided in
tab. 2. Further compounds were evaluated but not detected above detection limit (such as
estrone, nonylphenol, trichlosan). Concentrations of each compound were lower than 1 ng/kg.
Highest concentrations were determined for amberonne in fraction F14 (0.86 ng/kg; SPM 1)
8.5 Discussion
153
and octylphenol in F17 (0.8 ng/kg; SPM 2). Amberonne and galaxoide were highest
concentrated in F14 of the soil sample. However, 2,6-diisopropylnaphthaline and octylphenol
were present in each fraction, and, thus total concentrations, at least of octylphenol were
highest concentrated. Further, fractions F14 showed the highest added compound
concentrations (maximum of 2.2 ng/kg in SPM 1). REP values are shown when available
from literature and were used to compute chem-TEQ values (chem-TEQ [ng/g] = compound
concentration [ng/g] x REP; detailed by Woelz et al. 2008, ESPR). However, chem-TEQs
explained far less than 1 % contribution to the overall endocrine activity.
Tab 2 Concentrations of endocrine compounds in HPLC fractions of suspended particulate matter
(SPM 1 and 2) and of a soil layer sample from the northern inundated foreland (0 - 30 cm). REP
values are shown when provided in other studies. Different REPs are given for bisphenol A, thus, the
lowest2 and highest values3 are shown.
ng/kg
1
Benzophenone
OTNE (Amberonne)
2,6Diisopropylnaphthaline
Galaxolide (HHCB)
Tonalide (AHTN)
Diphenylsulfone
Octylphenol
Bisphenol A (2,3)
Added concentrations
1
SPM 1
REP
SPM 2
Soil
F14
F15
F16
F17
F14
F15
F16
F17
F14
F15
F16
F17
2*10
n.a.
0.18
n.d.
n.d.
n.d.
0.21
n.d.
n.d.
n.d.
0.10
n.d.
n.d.
n.d.
0.86
n.d.
n.d.
n.d.
0.53
n.d.
n.d.
n.d.
0.27
n.d.
n.d.
n.d.
n.a.
n.a.
n.a.
n.a.
0.02
0.04
0.02
0.02
0.04
0.02
n.d.
0.02
0.03
0.02
0.01
0.03
0.62
n.d.
n.d.
n.d.
0.36
n.d.
n.d.
n.d.
0.18
n.d.
n.d.
n.d.
0.21
n.d.
n.d.
n.d.
0.14
n.d.
n.d.
n.d.
0.05
n.d.
n.d.
n.d.
n.d.
0.27
n.d.
n.d.
n.d.
0.34
n.d.
n.d.
n.d.
0.15
n.d.
n.d.
0.31
0.66
0.55
0.54
0.52
0.24
0.65
0.80
0.09
0.12
0.05
0.07
2.5*10 2.63*10-4
n.d.
n.d.
n.d.
0.56
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
/
2.20
0.97
0.57
1.12
1.80
0.60
0.65
0.82
0.72
0.29
0.06
0.10
-6
-5
Han et al. 2002, 2 Vinggaard et al. (2000), 3 Routledge and Sumpter (1996), n.a. − not available
8.5 Discussion
8.5.1 Carbamazepine as a tracer for riverine contamination
In order to determine groundwater contamination by compounds of the river Rhine CBZ was
determined since this compound is a typical trace contaminant at the river Rhine for
anthropogenic impact (Schwarzbauer & Heim 2005). Groundwater sampled at three wells in
the projected retention area was shown to cause elevated endocrine activity independent of the
distance to the river. However, estrogenic activities were was quite heterogeneous among the
wells. Oellers et al. (2001) determined CBZ concentrations between 30 and 250 ng/L in water
samples from the Lake Greifen (Greifensee) and its main tributaries Aa and Aabach,
Switzerland. Liebig et al. (2006) measured 454 ng/L in German surface waters and Herberer
et al. (2002) detected 25 to 1,075 ng/L in waterway samples of Berlin, Germany. Thus,
ranking the CBZ concentration of this study (14 to 47 µg/L) indicates minor concentrations in
groundwater of the aquifer than detected in many surface waters in Central Europe. Further,
CBZ was shown to be a drinking water contaminant that was measured with maximum
8.5 Discussion
154
concentrations of 24 and 258 ng/L in finished drinking water of Canada and the United States,
respectively (Jones et al. 2005, Stackelberg et al. 2004).
Since the pharmaceutical carbamazepine is an anthropogenic contaminant, in particular of the
aquatic environment, detection in groundwater indicates an interaction of river water with
aquifer layers. In this study, CBZ indicates the interaction of Rhine water with the aquifer
throughout the year and changes in consequence of the August flood.
Total concentrations at the sites varied two- to threefold over the sampling time frame.
Nevertheless, concentrations were lowest in August 2007 and highest inductions were
measured in the weeks and month after the flood event in August 2007. These findings
indicate that groundwater velocity, at least in this time, was directed to the groundwater
well 1 and compounds were transported with a lack of time due to slow groundwater velocity.
However, CBZ concentrations with about 40 ng/L were also measured at the first sampling in
June 2006. Thus, CBZ contents seem to be also under the control of factors other than flood.
Since the water level was high in June and precipitation was minor in this month, it may be
assumed that river water was intruding into the aquifer, causing the measured concentrations.
Further, well no. 2 showed elevated CBZ concentrations, in particular from October to
December 2007. Although this well is located at a distance of about 750 m to the Rhine, mass
transfer through the aquifer seems to be possible, in particular, since CBZ may only be
introduced in the groundwater via groundwater velocity. This assumption is further supported
by the findings of another subproject in RIMAX-HoT indicated a hydraulic conductivity of
1.5 x 10-3 m/s (Kühlers et al. 2009). Well no. 3 seemed not to be influenced by river Rhine
water, since CBZ was not detected.
8.5.2 Estrogenic activities in the groundwater
Endocrine activities over time at each well advised that CBZ may be used as a tracer
compound for river water interaction with the groundwater. However, CBZ is no endocrine
active compound and its specific distribution does not necessarily indicate how far other
active compounds may be translocated in the aquifer. In particular, the pattern of endocrine
effects at each well over the time and among the wells is independent of the presence of the
tracer compound. Well no. 1 showed low activities before the flood, but were immediately
increased following the flood. Subsequent samples showed steadily decreasing activities. This
well that is closest to the river was, thus, intensively interfering or reset by river Rhine water,
at least within days. This finding is in accordance to the CBZ analysis results.
Further, endocrine activities were increasing in well no. 2 with a delay of days. However, the
exact time of beginning increase might have been missed, since this well was first sampled
about 10 days after the flood. A flood dependent increase of activity can be assumed, since
CBZ concentrations increased, but showed a longer delay. CBZ with a log KOW of 2.45
(Wiegel et al. 2004) might be translocated more slowly compared to compounds with lower
KOW being more polar and more easily and faster introduced in backland aquifer.
Water sampled at well no. 3 was conspicuous and gave elevated effects with most samples.
Further, highest inductions were determined with groundwater extracts of this well even
though it is the farthest away from to the river water. CBZ concentrations were not
8.5 Discussion
155
significantly increased at any sampling time at this well, at least until the end of sampling.
Estrogenic effects were very heterogeneous and, in addition, with the distance and CBT
results, it may be assumed that activities at this well were not influenced by river water or
flooding, but by another contaminant source. In fact, the Federbach creek, which flows close
to well no. 3, is known to be highly polluted by organic compounds (Obrdlik & Fuchs 1991),
and can, therefore, be expected to be the source of contamination.
The increased estrogenic activity at the projected retention basin was obvious comparing
E2-EQs with groundwater that was sampled at further drinking water protection areas in the
region of Karlsruhe (E2-EQ = 0.18 to 0.38 ng/L; details not shown). Ranking the endocrine
activities of the present study indicates that effects were in the lower range of river water with
0.3 to 19.4 ng E2-EQ/L (Pawlowski et al. 2004, Vermeirssen et al. 2005) and sewage treatment plant effluents showing 1 to 68 ng E2-EQ/L (Peck et al. 2004, Tan et al. 2007). Further,
effects can be discussed with other studies investigating estrogenic activity of groundwater
samples. Ancke-Hahn et al. (2009) determined E2-EQ between 0.63 and 2.48 ng/L in the area
areas of communities in South Africa. Braeken and van der Bruggen (2009) reported of
E2-EQ = 19.85 ng/L in groundwater. Thus, E2-EQs of the present study were comparable to
the first study and lower compared to the latter one; nevertheless, indicating an elevated
endocrine activity in the groundwater sample that was comparable to other sites. Further, a
flood event with a recurrence interval of 10 years was shown to have the potential for
groundwater contamination until far (some hundred meters) in the non-inundated hinterland.
And additionally, according to the objectives of the WFD groundwater and drinking water
resources should not be endocrine active. Thus, the recent outcomes indicate a potential
conflict with this European directive.
8.5.3 Active fractions and target analysis
Automated fractionation clearly showed that more polar compounds were major contributors
to the estrogenic activity of SPM and soil extract fractions. Fraction 15 was shown to cause
the maximum effect with each sample. Schlenk et al. (2005) used EDA for investigation of
marine sediments and also found fractions containing more polar compounds to cause the
bulk of estrogen activity. SPM 1 sampled before the flood peak of the August flood 2007
(recurrence interval of 10 years) showed more than twofold increased estrogenic effects
compared to SPM 2 sampled shortly after the flood peak. Thus, estrogenic active compounds
seem to be more concentrated with increasing discharge but quickly decreased after maximum
discharge. However, equivalent concentrations of this study are relatively low compared to
sediment E2-EQs determined in other studies investigating endocrine effectiveness of
sediments, SPM and soils using in vitro assays.
The added activities of single fractions in this study (0.9 to 2.3 ng/L) were minor compared to
E2-EQ values reported in the literature as shown by Oh et al. (2000) and Hashimoto et al.
(2005) with E2-EQ between 3.4 and 70 ng/L.
Total inductions of SPM 1 and the frequently inundated soil sample showed comparable
E2-EQ values that might indicate equal effective compounds. In order to identify endocrine
active effective compounds target analysis was applied. Measured compounds and computed
8.6 Conclusions
156
chem-TEQs only minor explained a small portion of E2-EQs determined in the fractions.
However, part of the endocrine compounds known show activities at concentrations of
nanogram per liter, thus, close to the detection limit of many compounds. Highly active
compounds such as estrone could not be quantified in this target analysis, but compounds like
estrone are very potent inducers and E2-EQs determined with concentrations just below
detection limits could explain most of determined activities determined. Since measurements
may simply not be sensitive enough to detect the addressed compounds, the possibility of
concentrations below detection limit should at least be kept in mind (Hollert et al. 2005).
8.6 Conclusions
The YES assay is suitable to identify estrogenic activities in groundwater samples as well as
in fractions of soil and SPM samples. Chemical analyses of tracer compounds and biological
activity can be used to provide information on the range of interactions between river and
aquifers and on changes over time and space. Significant compound introduction into the
aquifer can be expected at sampling sites that are very close to inundated areas. Elevated
effects may also be caused with delay at sites with some hundred meters distance to the river.
However, elevated tracer compound concentrations or endocrine activities caused by floods
may not be expected in the distant hinterland.
Automated fractionation procedures can be used to identify compound categories of effective
substances and allows to focus further research. Since more polar compounds are meanwhile
known to be important inducers in different biotests, they are of high relevance for endocrine
activity. However, further compound classes such as PAHs need to be given more emphasis.
Target chemical analysis may assist in identifying effective compounds. Nevertheless,
YES-inducing compounds can be numerous and analysis needs a broad adjustment to cover
these substances.
Focusing on areas of conflict between retention basins to manage future flood impacts and
waterworks to manage the provision of drinking water, a definite advice cannot be provided.
However, bank filtrate was shown to influence groundwater far into the hinterland not only
during floods, but also in times with average discharge. Thus, flooding of retention areas may
intensify the risk of compound introduction into the saturated zone. Further research is needed
to elucidate on the particular question if compounds are more easily translocated to the
aquifer at inundated sites.
Acknowledgement
The authors would like to express their thanks to Dr. Routledge (Brunel University , UK). We
further thank Katharina Graf for assistance with conducting biotests. We also thank the
Federal Ministry of Education and Research (BMBF), Germany, for supporting the
RIMAX-HoT project within the RIMAX joint No. 02WH0691.
8.7 References
157
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160
161
Chapter 9
9 Contaminant transport to public water supply
wells via flood water retention areas
D. Kühlers1, E. Bethge2, G. Hillebrand3, H. Hollert4, M. Fleig5, B. Lehmann3,
D. Maier6, M. Maier1, U. Mohrlok2 and J. Wölz4
1
Stadtwerke Karlsruhe GmbH (SWK), Karlsruhe, Germany
2
Institute for Hydromechanics (IfH), Karlsruhe, Germany
3
Institute for Water and River Basin Management (IWG), Karlsruhe, Germany
4
Institute for Environmental Research (IfU), Aachen, Germany
5
DVGW-Technologiezentrum Wasser (TZW), Karlsruhe, Germany
6
Heinrich-Sontheimer-Laboratorium (HSL), Karlsruhe, Germany
Nat Hazards Earth Syst Sci (2009) 9: 1047-1058
162
9.1 Abstract / 9.2 Introduction
163
9.1 Abstract
The essential processes and mechanisms of the transport of contaminants from a river to a
well field via a flood water retention area are presented. The transport is conceptualized as a
succession of three phases:
(1) Contaminant entry into the retention area
(2) Passage through the soil zone and
(3) Transport with the groundwater flow
Depending on the conditions of a given location and on the properties of the contaminants of
interest, processes within each transport phase may reduce the concentration of the
contaminants at the well field. For the Kastenwoert-Rappenwoert study area, the results of the
described processes are shown by chemical and ecotoxicological analyses as well as by
numerical modeling. Based on the results of the analyses, it is predicted that some
contaminants in the study area will be completely detained along the transport path, while
others will be transported as far as the well field, although in significantly reduced
concentrations.
9.2 Introduction
Along many rivers, flood retention areas have to be built to protect downstream settlements
against the impacts of extreme flooding. In these floodplains, riparian aquifers are often used
for drinking water production. Consequently, the proximity of retention areas to drinking
water production wells may lead to conflicts of interest. Drinking water providers are
concerned that river water, which often bears elevated loads of inorganic and organic particlebound and dissolved contaminants, could be directed through the retention areas toward
production wells, decreasing the groundwater quality at the municipal well fields. To predict
the magnitude of the effect, an in-depth understanding of the processes and mechanisms
active along the transport path of the contaminants is crucial.
9.2.1 Contaminant transport
The transport of the contaminants from the river to a water department’s well field via a
retention area occurs in three consecutive phases:
(1) entry into the retention area
(2) passage through the soil zone and
(3) transport with the groundwater flow
For these phases, corresponding conceptual “compartments” can be identified, all of which
are depicted schematically in fig. 1. Depending on the conditions at a given site and on the
properties of the contaminants of interest, each of the compartments, the retention area, the
soil zone and the aquifer, may act as a barrier, reducing the concentration of the contaminants.
9.2 Introduction
164
Fig 1: Contaminant transport from a river to a well field via a flood water retention area
In the first transportation phase, contaminants are conveyed into the retention area and
retained within. The transport of dissolved substances may be assumed to be identical to that
of conservative tracer compounds for which no reduction in concentration occurs at this stage.
Significant concentration effects, however, are observed for contaminants adsorbed to
suspended sediments. The transport of sorbed contaminants is strongly dependent on the
highly complex depositional processes of sediments within the retention area, which lead to
an accumulation at particular locations. During the second transport phase, the contaminants
pass from the surface of the retention area through the soil zone and into the aquifer. The soil
zone reduces the contaminant mass entering the groundwater primarily through the following
mechanisms: reduction of the percolation rate, retardation of the contaminants in the soil
matrix and microbiological degradation.
The third phase consists of the transport of the contaminants with the groundwater flow. The
concentration of contaminants at the production wells can be reduced during transport through
the aquifer by the processes of advective transport, dilution, retardation and degradation.
9.2.2 Kastenwoert-Rappenwoert study area
A site-specific examination of the contaminant transport processes and mechanisms presented
in the previous section was carried out at the Kastenwoert-Rappenwoert study area, located
south of Karlsruhe, Germany. The investigation area is situated on the eastern bank of the
Rhine River close to the point where the German Federal States of Baden-Wuerttemberg and
Rhineland Palatinate border each other and France. It is located within the Upper Rhine
Graben, a sedimentary rift basin that is bounded by the raised shoulders of the Vosges
Uplands of France to the west and the Black Forest to the east. The study area consists
primarily of forest with limited agricultural use and small expanses of open water. The soil at
the site is the product of changing flow and sedimentation conditions during the development
of the floodplain. There are significant differences between more elevated areas, where sandy
soil predominates and the lower-lying beds of infilled abandoned river channels with their
finer sediments. Due to the topography of the retention area, the soil zone thickness varies
between approximately 1 and 4 m. In general, the clay and organic carbon content of the soil
decreases with depth along the soil profiles (Bechler and Hofmann, 1996). The aquifer is
composed entirely of fluvial sediments of varying thickness. In the study area, the uppermost
9.2 Introduction
165
aquifer, from which the groundwater withdrawal is planned, extends to a depth of about 30 m
and is composed of highly permeable sediments with an average hydraulic conductivity of
about 1.5 x 10-3 m/s. The Rhine River has an average annual discharge of 1,250 m3/s near the
study area (HVZ, 2008) and its watershed encompasses the entire surrounding region.
However, water levels in the Rhine are largely independent of local precipitation because the
Rhine derives its water primarily from the Alps. When water levels in the Rhine are high, they
have a damming effect on groundwater draining to the river from the surrounding aquifer.
Under the influence of the local reversal in the groundwater gradient, the groundwater flows
northward along secondary streams until it eventually empties into the Rhine further
downstream. Contaminants found in significant concentrations in the Rhine near the study
area include highly adsorptive organic compounds like HCB, PAH and PCB (Maier et al.,
1998, Kosmehl et al. 2004, 2007), as well as highly mobile organic compounds like EDTA,
pharmaceutical residues and X-ray contrast agents (Fleig et al., 2008). The study area,
depicted in Fig. 2, contains the planned “Bellenkopf/Rappenwoert” flood water retention area.
With an areal extent of 5.1 x 106 m2 the retention area will provide a flood water storage
volume of 14 x 106 m3. The retention area is one component of the Federal State BadenWuerttemberg 'Integrated Rhine Program (IRP)', which includes 13 flood water retention
areas along the Rhine River between Basel, Switzerland, and Mannheim, Germany. The
Bellenkopf/Rappenwoert area has, however, also been designated as part of the wellhead
protection zone for the Kastenwoert well field (maximum extraction of 7.4 million m3/a)
which the local water department plans to build adjacent to the planned retention area site.
Planning for both projects is nearing completion. Extensive chemical and ecotoxicological
analyses of the suspended load in the river, soil and groundwater were conducted to determine
the status quo in the study area in each of the presented compartments. Using this baseline,
numerical models were applied to predict the interactive effect of the projects, neither of
which has yet been realized. Together the results of the analyses and the modeling were used
to characterize the three compartments of the transport path.
Fig 2: The Kastenwoert-Rappenwoert study area
9.3 Phase 1: Entry of contaminants in the retention area
166
9.3 Phase 1: Entry of contaminants into the retention area
9.3.1 Characterization
The transport of contaminants into the retention area and their retention constitute the first
phase of the contaminant transport. The contaminants reach the retention areas either in
dissolved form or adsorbed to suspended sediments. Substances dissolved in the water of the
flood wave are transported like a conservative tracer compound. The suspended sediments,
that often hold adsorbed contaminants, behave differently depending on their size and weight.
Therefore, to study and evaluate the first phase of the contaminant transport, in-depth
knowledge about the advective and diffusive transport mechanisms as well as the deposition
of suspended sediments within the retention area is crucial.
A series of field measurements were conducted near the study area to get an impression of
deposition patterns on the inundated floodplains. Sediment traps consisting of artificial turf
mats were used to determine the amount of sediments deposited during two flood events, a
smaller one in September 2006 and a larger one in August 2007. As an example, Fig. 3 shows
a cross-section of the floodplain with the approximated peak water levels for both flood
events and the average amounts of sediment deposited at three locations of increasing
distance from the river main channel. Firstly, Fig. 3 illustrates that higher flood events yield
higher sediment deposits, but no direct proportionality is found. Furthermore, comparing
different values of the same flood event, higher inundation heights increase the amount of
deposited sediments. This effect becomes particularly evident for low relative inundation
heights.
Fig 3: Deposited sediments of two flood events at a cross-section of the floodplain of the Rhine River
near Neuburgweier, Germany. The peak flood height is depicted for both events.
9.3 Phase 1: Entry of contaminants in the retention area
167
The corresponding grain size distribution of the deposits is depicted in fig. 4. The average
grain size decreases with increasing distance from the river main channel. This is due to the
reduced transport capacity of the overbank flow. Coarser sediment is deposited soon after
reaching the overbank area. Finer sediments are more uniformly distributed throughout the
floodplain. (This is supported by additional data which are not presented here.) On the other
hand, the comparison of the two flood events shows that the grain size distribution at one
location does not vary strongly for different flood events. The main difference can be found
with the clayey fraction. For the 2007 flood event, the proportion of deposited clay sediment
is higher than for the 2006 flood event, probably due to the longer duration of floodplain
inundation (about 7 days in 2007 instead of about 2 days in 2006) rather than to the higher
flood level.
Fig 4: Grain size distribution of deposited sediments of two flood events at a cross-section of the
floodplain of the Rhine River near Neuburgweier, Germany. The peak flood height is depicted for
both events. The bars show the proportions of the respective grain size classes.
Based on the field study and on an additional literature review (cf. e.g. Asselman and
Middelkoop, 1995, He and Walling, 1997, Howard, 1992, Walling and Bradley, 1989), the
following conclusions on the most important influencing factors of the deposition patterns of
suspended sediments on floodplains can be drawn:
• Local suspended sediment concentrations significantly influence the amount of
deposited sediment. High concentrations lead to high deposition rates
• Grain sizes influence suspended load and settling velocities. Coarser sediments deposit
on the floodplain close to the river channel. Finer fractions are more uniformly
distributed within the suspension across the floodplain
• The total deposition is directly proportional to the duration of the specific flood event
for near-steady conditions. For unsteady conditions, high local inundation heights at a
9.3 Phase 1: Entry of contaminants in the retention area
168
given site generally correspond to long inundation periods. Thus, high local
inundation heights usually correlate with increased deposition. Trapping effects of
flood water in local depression areas also substantially increase the amount of
deposited sediments
• Local flow velocities influence the deposition rates by reducing the probability of
sediment deposition for high velocities. However, there is evidence that a definite
critical value above which no deposition occurs does not exist in reality as was
believed in the past (Haralampides et al., 2003, Krishnappan and Engel, 1997,
Krone,1993, Kuijper et al., 1991).
9.3.2 Chemical analysis
To predict the concentration of contaminants that could be extracted at the planned well field,
it is important to identify the substances that are present in the river water. A comprehensive
data set measured by governmental institutions and water suppliers were analyzed to get an
overview of the current water quality situation in the Upper River Rhine. Special attention
was given to measurements taken during high flow events. The most important groups of
chemical substances regularly found dissolved in water samples from the Upper River Rhine
are:
•
•
•
•
•
Total and surrogate parameters with organic carbon (DOC), organic sulphur (AOS) or
organic halogens (AOX)
Complexing agents (NTA, EDTA, DTPA)
Pesticides and their metabolites (N,N-dimethylsulfamide) and
Xenobiotic organic compounds (chlorinated benzenes, MTBE)
Most of these substances are polar and often quite persistent, so it can be assumed that they
potentially infiltrate into the groundwater. Organic substances with a high KOW, like
Hexachlorbezene (HCB), polychlorinated biphenyls (PCB) and polycyclic aromatic
hydrocarbons (PAH), are almost exclusively found adsorbed to suspended matter. Therefore,
suspended sediments were also analyzed. Suspended sediments were collected on deposition
with turf mats as described in sect. 2.1 during the flood event in August 2007 at various
distances from the Rhine River. After lyophilization the amount of different adsorbed organic
compounds were determined.
From the bank of the Rhine River towards the dyke the amount of the indicator compound
HCB decreases by more than 90 %. The concentrations of other compounds decrease by half
at locations further away from the river. Only AOX and the carbon fraction increase with
distance from the Rhine River.
9.3.3 Ecotoxicological analysis
Suspended sediment are known to provide huge surfaces and binding sites for organic and
inorganic compounds (Hollert et al., 2007, Kosmehl et al., 2007). Furthermore, most rivers
are (highly) loaded with contaminants of various sources, at least as a legacy of the past
9.3 Phase 1: Entry of contaminants in the retention area
169
(Stronkhorst and van Hattum, 2003). While sediments may act as contaminant sinks under
normal hydrological conditions, (extreme) flood events cause sediment erosion and, thus,
sediments and contaminants may be released into the water column (Hilscherova et al., 2007,
Hollert et al., 2000, 2003, Wölz et al., 2008). Subsequently, they may pose a threat to
organisms in the aquatic and inundated terrestrial environment, but also to human interests
and health, e.g. regarding drinking water supply (Maier et al., 2006). Thus, the recent
ecotoxicological study focused on the evaluation of impacts by particle-bound organic
compounds. In order to evaluate the risk of particle translocation from the river on inundated
sites, e.g. retention areas, a battery of in vitro biotests comprising several ecotoxicological
endpoints was applied. For instance, the acute toxic Neutral Red retention assay (cytotoxicity;
Babich and Borenfreund, 1992, Klee at al., 2004, Keiter et al., 2006) and the mechanism
specific 7-ethoxyresorufin-o-deethylase (EROD) assay (Ah receptor-mediated activity;
Behrens et al., 1998, Keiter et al., 2008, Kennedy et al., 1996, Lorenzen and Kennedy, 1993)
were used. Under normal hydrological conditions, some samples of suspended sediments
indicated temporarily elevated cytotoxic effects, while most samples were only minimally
toxic. On the other hand, EROD inductions were mostly increased and some samples clearly
indicated significant effects that could not be correlated with any other parameter. The
sediments collected with turf mats during the flood event in August 2007 were also examined
using effect-directed analysis. Combining fractionation methods (in cooperation with
Dr. Werner Brack, Helmholtz Centre for Environmental Research Leipzig), cell-based in vitro
biotests and chemical analysis were applied to identify effect causing compounds. Fractions
containing PAHs and more polar to polar compounds were found to be most toxic (Brack et
al, 2005). In conclusion, based on the fact of particle translocation in the retention area,
a contamination of the inundated site by primarily bound compounds has to be considered,
unless measures may diminish particle entry or guarantee short residence times.
9.3.4 Modelling
In order to determine the deposition patterns within the planned retention area, input data
about hydrologic, hydraulic and sediment parameters is necessary. Water levels and local
flow velocities were determined by a 2D hydrodynamic numerical model for different flood
scenarios. The output data from this model was converted into raster-based datasets. Several
deposition models from the literature with different approaches were tested with the field data
from the investigated flood events and then applied to the retention area. The results will not
be presented in this article, but some essential findings should be mentioned:
•
Advection appears to be a significant process for the floodplain in question, mainly
due to the dominant flow through topographic depressions, side channels or ditches on
the floodplain, which are activated during flood events.
•
Pure diffusion models, which do not account for any advection across the floodplain,
significantly underestimated the sediment deposited at greater distances from the main
river channel (e.g., Pizzuto, 1987). Side channels could be incorporated in the model
as additional sediment sources to at least partly compensate for this effect.
9.4 Phase 2: Passage through the soil zone
170
•
Each of the investigated models uses several empirical parameters, which greatly
influence not only the amount of deposited sediments but also the deposition pattern.
The determination of those empirical parameters is challenging, a factor which must
be taken into account when evaluating the results of the simulations.
•
While hydraulic parameters may be determined with sufficient accuracy, reliable
information on sediment characteristics during flood events are difficult to obtain.
This should be considered, as model results prove to be highly sensitive to variations
of model input parameters (especially water levels and sediment concentrations).
9.4 Phase 2: Passage through the soil zone
9.4.1 Characterization
As described in the previous section contaminants can reach the ground surface of the
retention area in dissolved form or sorbed to settled flood water sediments. Because of the
filtering effect of the soil matrix, particle-bound contaminants are mainly retained at the soil
surface. Therefore, a considerable mass flux of dissolved contaminants to the aquifer is more
likely. The infiltration of dissolved organic contaminants into the soil zone is determined by
the seepage rate. Preferential pathways (macropores) can allow soil water and solutes to
bypass the soil matrix and travel quickly into deeper soil regions (Beven and Germann, 1982;
Wang and Narasimhan,1985).
Within the soil zone compartment, the contaminant mass flux into the aquifer is reduced by
three processes. First, the infiltration of contaminants into the soil zone is decreased.
Contaminated sediments are mainly retained at the soil surface, and additionally the
infiltration of dissolved contaminants is decreased by the low permeable floodplain soil.
Secondly, the transport of infiltrated contaminants within the soil zone is retarded by
adsorption to the soil matrix. As a third effect, microbes act to degrade the infiltrated
contaminant mass. The mechanisms can influence each other. A low infiltration rate reduces
the contaminant input into the soil zone and slows down the transport velocity towards the
aquifer. This, in turn increases the travelling time of contaminants, thereby supporting the
microbial degradation.
The input of dissolved organic contaminants into the soil zone of the retention area is
determined by the infiltration rate of the flood water. This infiltration generates water flow
through the soil zone which is affected by several changes in the hydraulic conditions during
a flood event. For the soil zone, three consecutive stages with different hydraulic conditions
can be identified during a flooding period:
• Stage 1 (S1): Infiltration of flood water into unsaturated soil
• Stage 2 (S2): Infiltration of flood water under saturated conditions
• Stage 3 (S3): Drainage of soil after the flood event
9.4 Phase 2: Passage through the soil zone
171
The essential factors controlling the mass input into the soil zone are the hydraulic properties
of the soil (average hydraulic conductivity) and the characteristics of the flood event (flooding
height and duration). The hydraulic gradient between surface and groundwater plays a major
role for these processes. It evolves towards lower values during the flooding period and shows
large spatial variability depending on the elevation of the groundwater table and the thickness
of the soil zone. Additionally, steep hydraulic gradients often occur at the inland bank of a
retention area during a flood. Within a highly conductive soil zone, the resulting seepage rate
will yield a strong mass input into the soil zone there.
The mass output to the aquifer is controlled by the transport velocity within the soil and the
storage capacity of the soil zone. The transport velocity is controlled by the average seepage
velocity and the retardation by sorption to the soil matrix. The sorption of organic
contaminants to the soil matrix is related to the soil organic carbon content and the
octanol:water partitioning coefficient. The storage capacity of the floodplain soil is a function
of those two properties and the soil thickness. The highly variable properties of the floodplain
soils and the hydraulic conditions within the retention area have to be taken into account to
calculate the contaminant mass flux through the soil zone to the aquifer.
9.4.2 Chemical analysis
Soil was sampled to a depth of up to 90 cm at two locations. The first site, located directly
along a Rhine backwater, is periodically inundated. The second, situated 15 m inland is only
inundated during extreme flood events. The concentrations of HCB, PCB and PAH at the site
near the river were five times higher than the values from the second site and unlike at other
sites increase with depth. The backwater site, an in filled oxbow, was filled with contaminated
sediment over time. Decreases in sediment contaminant loads in recent decades led to the
observed concentration depth profile. The observed increase is therefore clearly not to be
explained by transport processes in the soil zone. In conclusion, the repeated deposition of
contaminated river sediments at frequently inundated sites is confirmed by the results of the
chemical analyses.
The remobilisation of sediment-sorbed pollutants was examined in batch experiments with a
soil-to-water-ratio of one to ten. Only a few chemicals (HCB, 1,2,3-trichlorbezene) showed
even small reductions.
In another series of experiments, soil from the study area was used to fill small columns
through which water containing a known concentration of an organic pollutant was
percolated. The outflow concentration was determined at short time intervals. Most
contaminants broke through very rapidly, within one to two days, appearing at the column
outlet at the same concentration as at the input. The experiments thus demonstrated that not
all pollutants could be retained by the soil or degraded by microorganisms.
9.4.3 Ecotoxicological analysis
In sect. 2.3 it was shown that contaminants, once entering the retention area and passing the
first compartment, increase the contamination risk of the following compartments soil and
9.4 Phase 2: Passage through the soil zone
172
groundwater. Thus, the toxic effects and displacement of particle-bound compounds were
assessed comparing soil samples of periodically inundated sites, including those described in
sect. 3.2, and non-inundated sites at the planned retention area Bellenkopf/Rappenwoert.
The assessment of soil profiles with a depth of 90 cm, separated in sub-samples, allowed the
determination of total effects and the effect course in each profile by applying the in vitro
biotests mentioned above. While cytotoxic effects were minor at nearly all sites, EROD
activities were elevated at some sites and significantly increased in surface depressions,
containing accumulated deposited sediments. In addition, the Ames Fluctuation assay (Perez
et al., 2003) was applied and indicated considerable mutagenic effects, according to the highly
effective samples identified in the EROD assay. At each site with exception of the backwater
site described in sect. 3.2, it could be shown that toxic effects decreased from the surface to
the lowest soil layer assessed. Commonly, these findings are due to the high contents of
humic compounds in the upper soil layer, providing huge binding surfaces. In particular at
river banks, lower soil layers consist of large grained compounds as gravel and, thus, provide
significantly less binding sites (Fernandez-Galvarez 2007), resulting in a decrease in toxic
effects.
Concluding, at least in depressions at periodically inundated sites, toxic effects were
comparable to effects detected in the first compartment. Thus, SPMs displaced in (extreme)
flood events and deposited on floodplains accumulate and act as contaminant sinks that may
subsequently turn into sources, potentially releasing hazardous compounds into the soil. In
absence of mechanisms to reduce contaminant deposition on floodplains/retention areas these
aspects should be considered.
9.4.4 Modeling
To calculate the transport of dissolved organic contaminants through the soil zone of the
retention area to the aquifer, a one-dimensional contaminant transport model (FWinf) was
developed (Bethge and Mohrlok, 2008). The transport processes during a flood event were
described using a time efficient analytical modeling approach. The relatively short computing
time allowed the mass flux to the aquifer to be calculated for a large number of locations, thus
accounting for the spatial variability of the floodplain soils.
In the model FWinf, the soil zone was represented with a two-layered soil profile, where a
loamy topsoil overlies a sandy subsoil. The presence of macropores is considered for the
topsoil only. To calculate the transport of the infiltrated contaminants through the soil zone,
advective transport of the contaminants with the soil water movement is considered. The
bypass flow through macropores in the topsoil was taken into account by a simple macropore
bundle approach (Chen and Wagenet, 1992). The retarding effect of the contaminant sorption
to the soil matrix was simulated using a retardation factor, whereas contaminant mass
degradation was implemented with a first order kinetic model.
The mass balance model FWinf was applied to the planned flood water retention area
“Bellenkopf/Rappenwoert”. Field infiltration experiments have been conducted to estimate
the macropore porosity of the topsoil. Soil samples were taken from the field to determine the
saturated conductivity and water retention parameters. Additionally, the organic carbon
9.4 Phase 2: Passage through the soil zone
173
content was determined at different depths. The soil map and land use categories (forestry,
agricultural, open water bodies) were used to regionalize these data over the floodplain and to
outline simulation units that were used to calculate the one dimensional contaminant transport
within the soil zone.
To calculate the mass flux through the soil zone a flooding scenario with a 20 year recurrence
interval was chosen. For the properties of the contaminant in the flood water an average
sorptivity (log KOW = 0.3 m3/kg) and degradation rate (λ = 1.1 x 10-7 L/s) were chosen
(similar to the pharmaceutical carbamazepine). The concentration in the flood water was set
to 1.0 x 10-3 kg/m3.
In fig. 5 the areal distribution of the contaminant mass flux to the aquifer is shown for the
retention area Bellenkopf/Rappenwoert. Near the inland embankment drainage measures to
control the inland groundwater level establish a large hydraulic gradient between surface
water and groundwater head. As a result almost 99 % of the mass output from the soil zone
into the groundwater occurs within 200 m of the inland dykes. The highest contaminant mass
fluxes are found where high seepage rates coincide with small contaminant storage capacity
of the soil zone (small thickness of the soil zone, low organic carbon content).
In fig. 6 the calculated overall mass balance for the soil zone of the study area is shown for
the different flow stages of the seepage rate (Sect. 3.1). Mass input (Min) is observed during
the flow stages S1 and S2, mass flux into the groundwater (Mout) takes place during flow
stages S2 and S3. The highest mass input and output are observed in flow stages S2. For the
given contaminant properties the mass storage in the soil (Msoil) and mass degradation (Mdeg)
are small compared to the mass inflow and outflow.
Fig 5: Areal distribution of the calculated
contaminant
Fig 6: Calculated average contaminant mass
flux to the aquifer mass fluxes during the infiltration
phases.
9.5 Phase 3: Groundwater flow
174
9.5 Phase 3: Groundwater flow
9.5.1 Characterization
The third phase of the contaminant transport to the well field of a water department is the
transport with the groundwater flow. A reduction of concentrations in the production wells
occurs as the result of mechanisms of advective transport, dilution, retardation and
degradation.
The contaminants are transported primarily advectively with the groundwater flow. Therefore,
even if they enter the groundwater, they can only reach the production wells if the
groundwater gradient in the extremely transient flow field is oriented towards the production
wells for a sufficiently long time. If instead the groundwater gradient is directed primarily
towards the river, the appearance of a significant concentration of contaminants in the
production wells can be excluded. Therefore, water that infiltrated into the aquifer during
flooding of the retention area, can, in most cases, only reach the well field if the well field’s
zone of contribution overlaps the retention area under average hydrologic conditions. If there
is no overlap, the groundwater quality at the well field will be unaffected by the retention area
in most cases.
Fig. 7 depicts an idealized well field with a catchment area which overlaps with the flood
water retention area and even extends to the river at average hydraulic conditions. In such a
case, even under average hydraulic conditions, the groundwater extracted at some production
wells will contain a significant percentage of infiltrated river water. When the retention area is
flooded with river water, the gradient between the retention area and the well field will
increase and thereby directly increase the percentage of river water at the municipal wells.
Fig. 7: River water in the aquifer before flooding of the retention area. of the retention area.
Furthermore, during flooding, river water will infiltrate in the aquifer mainly at the edge of
the retention area, as discussed in sect. 3.4 and depicted in fig. 8. After flooding, this
infiltrated water will be transported by the normal groundwater flow to the production wells
9.5 Phase 3: Groundwater flow
175
and will eventually reach them, as much as years after the flooding event, depending on the
velocity of the groundwater flow.
Fig. 8: River water in the aquifer after flooding
The concentration of substances from the river that reach the pumping wells will be reduced
by dilution. First, it must be noted that the volume of the river water entering the aquifer
during flooding of the retention area will generally be much smaller than the volume of
groundwater already in the aquifer. On its way towards the extraction well, the water and the
substances it carries with it will be mixed with the groundwater by dispersive processes,
reducing the concentrations of contaminants significantly. An additional dilution process
takes place in the withdrawal wells and in the water treatment plant. The pumping wells of the
water works draw groundwater from both the retention area side and the inland side. Under
average hydraulic conditions, when the retentions area is not flooded, considerably more
water from inland reaches the pumping wells because of the low groundwater gradient
between the well field and the stagnation point, which lies in direction of the retention area.
Furthermore, several of the withdrawal wells often extract no infiltrated river water (Fig. 7
and fig. 8). Both mechanisms strongly reduce the concentrations of infiltrated substances from
the retention area at the water works.
Retardation and degradation can further reduce contaminant concentrations at the well field,
but both mechanisms are highly dependent on the properties of the substance of interest.
Certain substances, such as many X-ray contrast agents, will neither be retarded nor degraded
in an aquifer.
9.5.2 Chemical analysis
The analysis of ground water samples collected following the flood event in August 2007
revealed changes in concentrations of chemical compounds directly associated with river
water. In particular, increases in the concentrations of complexing agents and fuel additives as
well as dissolved oxygen, pharmaceutical products and a pesticide metabolite were observed.
9.5 Phase 3: Groundwater flow
176
The analyses confirmed that inputs of river water to the aquifer occurred not only through
infiltration at the river bank, but also at a backwater site. The tests established that dissolved
organic contaminants were able to pass through the soil zone and enter the aquifer. The ability
of microbial organisms to degrade organic compounds strongly influences the amount of
pollutants that reach the drinking water wells. Because the planned well field will
substantially alter the groundwater flow regime, it was impossible to examine changes in
contaminant concentrations in field studies. Therefore laboratory experiments were carried
out.
In a series of tests, defined amounts of typical pollutants were added to Rhine water which
was then circulated in a closed experimental plant. The tests showed that the concentrations of
some pharmaceuticals and organic pollutants were reduced by less than 50 % even after thirty
days and were therefore defined as substances of concern for drinking water suppliers. It must
be concluded, that transport through the aquifer cannot protect the well field against some
persistent organic pollutants.
9.5.3 Ecotoxicological analysis
While sec. 2.3 and 3.3 introduced the background and outcomes regarding contaminant entry
with deposition/displacement through the soil, the focus here is to determine whether
contaminants may pass through the soil layer and enter the drinking water aquifer.
Neither Ah receptor agonist (EROD assay) nor mutagenic activity (Ames Fluctuation assay)
indicated significantly increased activities, nor could cytotoxicity be determined. However,
endocrine activity mediated through the human estrogen receptor (ER) in the YES assay
(Routledge and Sumpter, 1996) could be measured, showing elevated endocrine effectiveness
for some of the samples and considerable fluctuations at different sampling times.
Nevertheless, the data reflected a time-effect correlation, at least, with groundwater sampled
at short intervals following the more intense Rhine flood (recurrence interval
approximately10 years) in August 2007. So far, there are no data available which would
indicate a flood dependent increase; and maybe only a temporal coherence is given.
As endocrine active compounds seemed to be the most relevant groundwater contaminants at
the site, effect-directed analysis (Lübcke-von Varel et al., 2008) was applied to identify
relevant inducers. Furthermore, fractionation was performed with flood SPMs and highly
effective soil. Endocrine activity was measured in each sample, at least constituting to a
considerable extent to effects in groundwater. Additional chemical analysis to identify
effective compounds is underway.
In conclusion, previously effective compound categories and ecotoxicological endpoints were
no longer relevant in the groundwater, while endocrine activity was detectable. Furthermore,
activities were measured in sample fractions of the other compartments. Thus, although the
origin of the measured ER agonist activities remains so far unclear, considering the complete
ecotoxicological assessment an impact of particle-bound compounds to drinking water supply
cannot be excluded. Further research has to focus on the identification of effective compounds
and will be completed following this project.
9.5 Phase 3: Groundwater flow
177
9.5.4 Modeling
For the Kastenwoert study area a numerical groundwater flow and transport model was
constructed to predict the influence of the planned retention area on the percentage of Rhine
water in the groundwater extracted at the well field. The 3-dimensional finite element model
was constructed using the groundwater modeling program FEFLOW from DHI-WASY. The
model area covers an 11 km stretch of the eastern Upper Rhine valley (a total of approx.
120 km2). The large model area was necessary in order to include the whole catchment area of
the water work. The spatial discretization of the model, in particular near the Rhine, had to be
very fine in order to ensure stable mass transport calculations. In total, the groundwater model
contained 529,060 elements (75,580 per layer) and 304,904 nodes (38,113 per slice). A
simulation time of 46 years was chosen (from January 1960 to December 2005), the longest
time period for which data were available. This made it possible to examine long-term
developments. The lengths of the time steps, which varied between a few minutes to 7 days,
were automatically selected by the program system by the predictor-corrector scheme
(AB/TR).
In order to prepare the model for transport simulation, the Dirichlet boundary condition
simulating the Rhine River was assigned a mass concentration of 100 mg/L. The water
infiltrated through the retention area was given the same value. Thereby the percentage of the
Rhine water in the aquifer during the transport simulation could very simply be observed and
illustrated. As some of the substances in the river water, for example X-ray contrast agents,
are neither degraded nor retarded in the groundwater flow, these processes were not included
in the transport simulation. On a standard PC, the calculation of a complete transport scenario
took about 3 weeks.
Fig. 9 illustrates the ratio of Rhine water to groundwater in the planned well field as
calculated with the numerical groundwater model described here. The red line illustrates the
percentage of Rhine water in a scenario in which the municipal wells start pumping in 1960
and pump consistently and steadily for the next 46 years at a combined rate of
7.4 million m3/a. The retention area is not active in this scenario. The percentage of Rhine
water in the extracted groundwater lies between 5 and 10 % in this scenario. The green line
shows the fraction of Rhine water in the planned water works in a scenario in which both the
wells and the retention area are brought online in 1960. Consistent with historical Rhine water
levels measured during that period, about 50 flood events, some of them quite small, were
simulated across the retention area. The repeated flooding of the retention area over the course
of the simulation period raises the percentage of Rhine water in the extracted groundwater by
about 5 to 10 %. Following the most significant flood event of the observation period, which
occurred in 1999, the percentage of Rhine water rises by as much as 15 % to compose about
one quarter of the extracted groundwater.
The simulation results were entirely consistent with the mechanisms described in sect. 4.1.
During the flood events, peaks in Rhine water input are observed and repeated flooding of the
retention area leads to an overall increase of the percentage of Rhine water in the pumping
wells of the water works in the long term. Due to dilution effects, the probability is very high
9.6 Discussion and conclusion
178
that the percentage of Rhine water extracted at the water work will remain far less than 50 %,
despite the proximity of the retention area.
9.6 Discussion and conclusion
The transport of contaminants from a river to a well field via a flood water retention area is
affected by the processes and mechanisms occurring in the three compartments: Retention
area, soil zone and aquifer.
The first compartment is chiefly characterized by the sedimentation of suspended matter
within the flood water retention area. Generally, the concentrations of dissolved contaminants
are unaltered by the processes in the first compartment. Contaminants sorbed to suspended
matter, on the other hand, are significantly affected, as they will be deposited with the
suspended matter.
The chemical and ecotoxicological analysis of suspended matter in the Rhine River show
significant presence of contaminants. The same contaminants and ecotoxicological effects
could be detected in the upper part of the soil layer of several periodically inundated sites,
with intensities varying from site to site. Thereby, the result of the variations in the highly
complex depositional processes of the sediment within the retention area could be
demonstrated. Numerical models can be applied successfully to improve the prediction of
deposition patterns, which in turn define locations within the retention area where
contaminants will accumulate.
The second compartment is the soil zone. Its effect on the contaminant mass flux to the
aquifer is mainly determined by the properties of the soil and the contaminant as well as the
flood event characteristics. The variability of soil properties and hydraulic conditions within
the retention area leads to distinct differences of the soil zone’s ability to prevent a
contaminant input into the aquifer. Applying the contaminant transport model FWinf on the
Kastenwoert-Rappenwoert study area, the highest contaminant mass fluxes of dissolved
organic contaminants into the aquifer are calculated for low lying soils in the near dyke area
with a low organic carbon fraction.
Presently, only a small section of the floodplains directly adjacent to the Rhine River is
inundated during floods. The chemical and ecotoxicological analysis of the areas subject to
inundation have revealed high concentrations of sorptive contaminants and significant
ecotoxicological effects in the upper part of the soil layer, while the lower part of the soil zone
and the groundwater were free of sorptive contaminants and any significant ecotoxicological
effects. It may, therefore, be concluded that the storage capacity of the floodplain soil is
sufficient for the current mass input to completely retain sorptive contaminants. The
contaminants still found in the groundwater are known to be highly persistent and mobile and
thereby had evidently been able to pass through the soil zone or river bank without being
significantly affected. However, these substances were found to show no ecotoxicological
effects.
In the third compartment, the contaminants are transported with the groundwater to the well
field. In most cases contaminants from the river can only reach the production wells if the
9.6 Discussion and conclusion
179
catchment area of the well field at average hydraulic conditions overlaps with the flood water
retention area. In this case, the concentrations of the contaminants are significantly reduced on
their way in the aquifer by dilution, and sometimes additionally by microbiologic degradation
and retardation.
Regarding the low flow velocities of the groundwater in the study area, it has to be concluded
that the organic contaminants recently found in the groundwater have partially been present in
the aquifer for several years. Therefore it must be assumed that, given the groundwater flow
calculated by the numerical aquifer simulation, these substances would be able to reach the
planned well field even after very long transit times.
Using numerical modeling of the study area, it was predicted that some contaminants found in
the Rhine River, including HCB, PAHs or PCBs, which are mostly adsorbed to suspended
matter, will probably be completely retained within the transport path, while others, like
EDTA or X-ray contrast agents, which are highly mobile and persistent, will most likely
travel to the well field, although in significantly reduced concentrations.
Acknowledgements
This paper presents results from the HoT joint research project, which received financial
support from the German Federal Ministry of Education and Research (BMBF) and was part
of the German national research program "Risk Management of Extreme Flood Events"
(RIMAX).
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183
Chapter 10
10 Conclusions of the study
184
10. Conclusions of the study
185
In the light of the preceding chapters, the entire study has been successful in investigating the
repercussions of flooding events on the degree of hazard potential of water-borne and
sediment-borne contaminants in river systems. The study broadly employed effect-directed
analysis (EDA) to determine which contaminants are bound to SPM and are responsible for
the toxicological effects. This methodology uses a combined approach of in vitro biotests,
chemical analysis and fractionation methods. Used biotests demonstrated that cytotoxicity,
dioxin-like/Ah receptor (AhR)-mediated and endocrine activities, as well as mutagenic
potentials, were increased with many samples assessed. Furthermore, the combination of
fractionation methods to separate persistent, less persistent and more polar compounds,
respectively, with chemical analysis and biotesting were shown to be adequate tools to
identify effective compound classes in distinct fractions.
The first part of the study investigated on contaminants bound to suspended particulate matter
(SPM) sampled in flood events. Results revealed that less persistent polycyclic aromatic
hydrocarbons (PAHs) and more polar compounds are the main inducers of biological
responses. These results allow to direct future investigations to more relevant and higher
effective compounds.
The appraisal of environmental contamination is normally limited to a priori selected
compounds. This is considered a comprehensible strategy since chemical analysis cannot
account for any possible contaminant. However, such scheme of preselecting toxicants poses
the risk of oversighting hazard potentials that can be caused by non-prioritized compounds, as
shown with most samples and fractions in this study. Evaluation is considered a failure when
the analyzed contaminants only reflect a few percent of the total biological activities while
disregarding the majority of effective compounds. In order to avoid false evaluation,
investigations should be designed on a broader perspective. When chemical analysis cannot
explain biological effects, extended fractionation methods can assist to identify responsible
compounds or at least compound classes. Accordingly, biological analysis has to include
different significant endpoints to reflect given impacts of contaminant loads.
However, flood SPM indicated elevated chemical loads and biological hazard potentials.
Thus, the question arises whether river contaminants impact, e.g. inundated sites. In
particular, areas that are especially chosen as retention basins and which are built temporarily
to protect downstream sites from extreme floods with recurrence intervals of more than or
equal 100 years should be given crucial attention. When these retention basins are operated
and flood water is kept over periods of days to weeks, particles and bound contaminants can
deposite on large scale on the basin ground. Furthermore, flood water can infiltrate the
unsaturated zone and infiltrate the aquifer that often is used as drinking water resource. Thus,
conflict of interests is expected. This aspect was investigated in the second part of the present
study at a model site where conflict ensues.
Investigations showed that flood SPM is highly loaded with contaminants and causes
maximum effects at the peak of discharge. Further research was carried out on soil sampled at
inundated and non-inundated sites and was shown to reveal contamination of sites due to
particle deposition. Results indicated that soil has to be sampled at several sites; different
geomorphological structures such as ground swales have to be included in the assessment
10. Conclusions of the study
186
since they accumulate river pollutants. Likewise, soil has to be sampled both at the surface
and deeper layers since proximity of contaminants to aquifer provides important information
on potential pollution hazards to drinking water resources.
Accomplishment of EDA, again, provided useful information on the classes of effective
compounds and the relative contributions to the total biological effects. In particular PAHs of
higher molecular weight can be expected to be very effective. Furthermore, more polar
compounds can be expected to be highly effective. Chemical analysis of priority compounds,
again, was less appropriate to explain biological effects. Both results underline the need to
broadly analyze complex environmental samples, and, in particular, to include further and
recently determined substances of concern, e.g. heterocycles. Thus, hazard identification and
assessment of contaminants (e.g. in drinking water resources) should also be directed towards
these compounds.
Another aspect of the study was to assess if river contaminants infiltrate the aquifer during a
flood event. Typical tracer compounds such as the pharmaceutical carbamazepine can be
analyzed in order to determine how deep river water encroaches in the aquifer. In parallel,
biological effects have to be measured since tracer concentration and effects are not
necessarily connected to each other, in particular, since infiltration depth differs between
(effective) compounds. Thus, test design, applied biotests and chemical analysis have to be
selected carefully and should be angled broader to competently identify the effective
toxicants.
Evaluation of SPM during flood events and the consequent particle deposition at inundated
sites revealed an elevated contamination load and a concomitant increase of biological effects.
Since even aquifer is infiltrated far distant in the hinterland, impacts to drinking water
resources appear to be possible. In this study, investigations were accomplished with samples
of a flood with a recurrence interval of up to 10 years. Underlying conflict of interests will be
intensified when retention basins are operated with floods and with a recurrence interval of
greater than or equal to 100 years. Thus, inundated sites, retention basins, and aquifers will be
impacted by even higher contaminant loads. Recent and future flood management strategies
have to cope with factors other than just the physical impact of extreme floods.
187
Chapter 11
11 List of references
188
11. List of Referencess
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Papers in ISI-listed journals (Impact factor 2007 in brackets)
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RTL-W1-Zellen mit Hilfe des Comet-Assays. Altex 21: 162-163 (ISI: 1.3)
Other publications
Hollert H, Ernst M, Ahlf, W , Duerr M , Erdinger L , Grund S , Keiter S , Kosmehl T , Seiler T-B ,
Wölz J, Braunbeck T (2009): Strategien zur Sedimentbewertung – Ein Überblick.
Z Umweltchem Oekotox 21,160-176 ; DOI 10.1007/s12302-009-0045-5
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Sedimentbelastung des Tietê River in Brasilien. Z Umweltchem Oekotox 18, 70
Maier M, Kuehlers D, Brauch H-J, Fleig M, Maier D, Jirka GH, Mohrlock U, Bethge E, Bernhart H H,
Lehmann B, Hillebrand G, Wölz JH, H. (2005): RIMAX-Verbundprojekt HoT – Spannungsfeld
Hochwasserrückhaltung und Trinkwasserversorgung: Vermeidung von Nutzungskonflikten.
Z Umweltchem Oekotox 17, 248-249
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In press / accepted / in revision / submitted
Wölz J, Brack W, Möhlenkamp C, Claus E, Braunbeck TH, H (2009): Identification of dioxin-like
activity in suspended particulate matter of floods using effect directed analysis. Submitted to Sci
Tot Envrion
Wölz et al. (2009): The RIMAX-HoT project: 1. Suspended particulate matter as a transport path for
particle bound substances in the river Rhine and influences of extreme flood events. To be
submitted to J Soils Sediments
Wölz et al. (2009): The RIMAX-HoT project: 2. Pollution of riparian areas in consequence of
inundation by extreme flooding. To be submitted to J Soils Sediments
Wölz et al. (2009): The RIMAX-HoT project: 3. Contaminant entry and transport in the saturated
groundwater zone in the context of extreme flood events. To be submitted to J Soils Sediments
Cofalla C, Müller E, Roger S, Wölz J, Hollert H, Schüttrumpf H (2009): Report: Workshop on
'Sediment dynamics in the context of hydro mechanical and ecotoxicological considerations',
Workshop held at the Guesthouse of the University of Aachen, Aachen. In preparation
Book chapter
Hollert H, Duerr M, Haag I, Wölz J, Hilscherova K, Blaha L, Gerbersdorf S (2007): Influence of
hydrodynamics on sediment ecotoxicity. In: Westrich B, Foerstner U (Eds.), Sediment
hydraulics and pollutant mobility in rivers. Springer-Verlag, Heidelberg, pp. 401-416
Diploma thesis
Wölz J (2005): Entwicklung von Fraktionierungsmethoden zur Untersuchung des Dioxin-ähnlichen
Potenzials von Hochwasserschwebstoffen, Ruprecht-Karls University, Heidelberg, 95 pp.
Oral presentations
Wölz J, Großhans K, Luebcke-Von Varel U, Schulze T , Maletz S, Bragenheim L, Brack W, Hollert H
(2008): Oekotoxikologische Untersuchungen und Effekt-dirigierte Analysen zur Bewertung des
Konfliktpotentials zwischen der Errichtung von Retentionsräumen und der Sicherstellung der
Trinkwasserversorgung. 13th Annual SETAC-GLB meeting, Frankfurt am Main, Germany
Wölz, J , Maier, M , Kühlers, D , Fleig, M , Mohrlock, U , Bethge, E , Lehmann, B , Hillebrand,
G (2008): Flood retention and drinking water supply – Preventing conflicts of interests. 17th
Annual SETAC Europe meeting, Warsaw, Poland
Mariani C F, Pompeo M L M, Rocha P S, Zimmer H, Zielke H, Woelz J, Erdinger L H, Hollert H
(2008): Integrated assessment of sediments from Rio Grande Reservoir: Biotests preliminary
results. 10th Brasilian Ecotoxicology Congress, Bento Gonçalves, Brasilia
Wölz J, Maier M, Kühlers D, Brauch H -J, Fleig M, Maier D, Jirka G H, Mohrlock U, Bethge E,
Bernhart H H, Lehmann B, Hillebrand G, Hollert H (2007): Spannungsfeld Hochwasserrückhaltung und Trinkwassergewinnung – Vermeidung von Nutzungskonflikten. 12th Annual
SETAC-GLB meeting, Leipzig, Germany
11. List of Referencess
191
Wölz J, Bragenheim L, Maletz S, Fleig M, Maier M, Kühlers D, Maier D, Braunbeck T & Hollert H
(2006): Spannungsfeld Hochwasserrückhaltung und Trinkwassergewinnung – Vermeidung von
Nutzungskonflikten (RIMAX-HoT). 11th Annual SETAC-GLB meeting, Landau, Germany
Wölz J, Olsman H, Hagberg J, Brack W, Moehlenkamp C, van Bavel B, Engwall M, Claus E, Manz
W, Braunbeck T, Hollert H (2006): Effect-directed analysis of Ah-receptor agonists in
suspended matters from flood events. 16th Annual SETAC Europe meeting, The Hague, The
Netherlands
Wölz J, Olsman H, Hagberg J, Brack W, Möhlenkamp C, van Bavel B, Engwall M, Claus E, Manz W,
Braunbeck T & Hollert H (2005): Bioassay-geleitete Fraktionierung und Identifikation von AhRezeptor Agonisten in Hochwasserschwebstoffen. 10th Annual SETAC-GLB meeting, Basel,
Switzerland
Wölz J (2004): Entwicklung von Fraktionierungsmethoden zur Untersuchung des Dioxin-ähnlichen
Potentials von Hochwasserschwebstoffen. Federal Institute of Hydrology (BfG), Koblenz,
Germany
Kosmehl T, Wölz J, Garke V, Krebs F, Erdinger L, Braunbeck T & Hollert H (2003): Vergleichende
gentoxische Untersuchungen von Sedimenten des Rheins mit RTG-2- und RTL-W1-Zellen mit
Hilfe des Comet-Assays. 8th Annual SETAC-GLB meeting, Heidelberg, Germany
Posters
Brinkmann M, Hudjetz S, Cofalla C, Roger S, Kammann U, Giesy J, Hecker M, Schüttrumpf M,
Lennartz G, Wölz J, Hollert H (2009): Searching for the (eco-)toxicological relevance of
sediment remobilisation and transport during flood events in rivers - Part B: Biomarker
approach. 14th Annual SETAC GLB meeting, Weihenstephan, Germany
Hudjetz H et al. (2009): Searching for the (eco)toxicological relevance of sediment remobilization and
transport during flood events in rivers. Part A: instrumental chemical analysis and effectdirected analysis (EDA), 14th Annual SETAC GLB meeting, Weihenstephan, Germany
Schneider A J, Brinkmann M, Gerstner A, Wölz J, Heger S, Weber R, Engwall M, Seiler T-B and
Hollert H (2009): Assessment of dioxin-like toxicity in soils contaminated by a chloralkali
process and a leblanc factory. 29th International Symposium on Halogenated Persistent Organic
Pollutants, Beijing, China
Brinkmann M, Hudjetz S, Cofalla C, Roger S, Kammann U, Giesy J, Hecker M, Schüttrumpf M,
Lennartz G, Wölz J, Hollert H (2009): Searching for the (eco-)toxicological relevance of
sediment remobilisation and transport during flood events in rivers - Part B: Biomarker
approach. 19th Annual SETAC Europe meeting, Goeteborg, Sweden
Hudjetz H et al. (2009): Searching for the (eco)toxicological relevance of sediment remobilization and
transport during flood events in rivers. Part A: instrumental chemical analysis and effectdirected analysis (EDA), 19th Annual SETAC Europe meeting, Goeteborg, Sweden
Wölz J, Hudjetz S, Brinkmann M, Hecker M, Kammann U, Lennartz G, Hollert H (2009):
Floodsearch. Annual RWTH Transparent meeting, RWTH Aachen University, Aachen,
Germany
Hallare A V, Seiler T-B, Schulze T, Wölz J, Brack W & Hollert H (2008): 1.Effect-directed analyses
of mutagens and potentially cytochrome P4501A inducers in contaminated sediments from
Laguna Lake, Philippines, SEKt Conference in Koblenz.
11. List of Referencess
192
Ernst M, Gerstner A, Heger S, Rnkovic S, Perovic A, Erdinger L, Wölz J, Braunbeck T, Hollert H &
Seiler T-B (2008): Assessing the sediment quality of Lake Skadar/Shkodra. 13th Annual
SETAC-GLB meeting, Frankfurt am Main, Germany
Mariani C F, Pompeo M L M, Rocha P S, Zimmer H, Zielke H, Wölz J, Erdinger L, Hollert H (2008):
Integrated assessment of sediments from Rio Grande Reservoir: Biotests preliminary results,
10th Brasilian Ecotoxicology Congress, Bento Gonçalves, Brasilia
Wölz J, Maier M, Kühlers D, Fleig M, Mohrlock U, Bethge E, Lehmann B, Hillebrand G, Hollert H
(2008): Flood retention and drinking water supply – Preventing conflicts of interests.
5th SETAC World Congress, Sydney, Australia
Gerstner A, Ernst M, Rnkovic S, Perovic A, Erdinger L, Wölz J, Braunbeck T, Hollert H & Seiler T-B
(2008): Assessing the sediment quality of Lake Skhodra. 17th Annual SETAC Europe meeting,
Warsaw, Poland
Maletz S, Wölz J, Maier M, Braunbeck T, Hollert H (2007): Ökotoxikologische Charakterisierung von
Schwebstoffen aus dem Rhein. 12th Annual SETAC-GLB meeting, Leipzig, Germany
Rocha P, Brack W, Jurajda P, Ondracková M, Wölz J, Seiler T-B, Kosmehl T, Braunbeck T, Storch V,
Hollert H (2007): Assessment of ecotoxicological risks and hazard factors of contaminated
sediments from European freshwater ecosystems. International conference "Risk Assessment in
European River basins – State of the art and future challenges”, Leipzig, Germany
Rocha P, Jurajda P, Ondracková M, Wölz J, Seiler T-B, Kosmehl T, Storch V, Braunbeck T, Brack W,
Hollert H (2007): Assessment of ecotoxicological risks and hazard factors of contaminated
sediments from European freshwater ecosystems. 12th Annual SETAC-GLB meeting, Leipzig,
Germany
Wölz J, Maletz S, Bernecker C, Kühlers D, Fleig M, Rastall A, Erdinger L, Braunbeck T, Hollert H
(2007): Spannungsfeld Hochwasserrückhaltung und Trinkwassergewinnung – Vermeidung von
Nutzungskonflikten. AquaBase Workshop, Aachen, Germany
Wölz J, Maletz S, Bernecker C, Kühlers D, Fleig M, Rastall A, Erdinger L, Braunbeck T, Hollert H
(2007): Spannungsfeld Hochwasserrückhaltung und Trinkwassergewinnung – Vermeidung von
Nutzungskonflikten. RiskBase Workshop, Leipzig, Germany
Wölz J, Maletz S, Bernecker C, Kühlers D, Fleig M, Rastall A, Erdinger L, Braunbeck T, Hollert H
(2007): Spannungsfeld Hochwasserrückhaltung und Trinkwassergewinnung – Vermeidung von
Nutzungskonflikten. 12th Annual SETAC-GLB meeting, Leipzig, Germany
Wölz J, Bragenheim L, Maletz S, Bernecker C, Kühlers D, Fleig M, Rastall A, Erdinger L,
Braunbeck T, Hollert H (2007): Evaluation of the ecotoxicological hazard potential of extreme
flood events to drinking water supply. 17th Annual SETAC Europe meeting, Porto, Portugal
Wölz J, Bragenheim L, Maletz S, Fleig M, Maier M, Kühlers D, Maier D, Braunbeck T, Hollert H
(2006): Oekotoxikologische Abschätzung des Schädigungspotenzials von extremen
Hochwasserereignissen für die Trinkwassergewinnung (RIMAX-HoT). 11th Annual
SETAC-GLB meeting, Landau, Gemany
Wölz J, Brack W, Olsman H, Hagberg J, Möhlenkamp C, van Bavel B, Claus E, Manz W, Engwall M,
Braunbeck T & Hollert H (2006): Effekt-dirigierte Analyse von AhR-Agonisten in
Hochwasserschwebstoffen. 11th Annual SETAC-GLB meeting, Landau, Germany
11. List of Referencess
193
Rocha P S, Keiter S, Seiler T-B, Wölz J, Kosmehl T, Braunbeck T, Storch V, Hollert H (2006):
Weight-of-evidence study to assess sediment contamination in Tietê River, Brazil. 11th Annual
SETAC-GLB meeting, Landau, Germany
Borck D, Wölz J, Hollert H, Witt G (2006): Mobilität, Bioverfügbarkeit und toxische Effekte
polyzyklischer aromatischer Kohlenwasserstoffe (PAK) in Sedimentkernen der westlichen
Ostsee (Mecklenburger Bucht). 11th Annual SETAC-GLB meeting, Landau, Germany
Rocha P S, Pompeô M, Brandimatre A L, Mariani C F, Keiter S, Wölz J, Seiler T-B, Kosmehl T,
Boettcher M, Braunbeck T, Storch V & Hollert H (2006): Integrated assessment of sediment
contamination in Tietê River, Brazil. 16th Annual SETAC Europe meeting, The Hague, The
Netherlands
Wölz J, Olsman H, Hagberg J, Brack W, Möhlenkamp C, van Bavel B, Engwall M, Claus C, Manz W,
Braunbeck T & Hollert H (2006): Effect-directed analysis of Ah-receptor agonists in suspended
matters from flood events. 16th Annual SETAC Europe meeting, The Hague, The Netherlands
Maier M, Kuehlers D, Brauch H-J, Fleig M, Maier D, Jirka G H, Mohrlock U, Bethge E, Bernhart H
H, Lehmann B, Hillebrand G, Wölz J, Hollert H (2005): Spannungsfeld Hochwasserrückhaltung
und Trinkwassergewinnung – Vermeidung von Nutzungskonflikten. RIMAX Kick-OffMeeting, Potsdam, Germany
Perovic A, Nikcevic S, Bushati N, Wölz J, Seiler T-B, Sundic D, Erdinger L, Hollert H (2005): An
evaluation of results from monitoring and eco-toxicity testing of the Skadar/Shkodra Lake by
triad approach. 15th Annual SETAC Europe, Lille, France
Wölz J, Olsman H, Rastall A, Erdinger L, Reifferscheid G, Manz W, van Bavel B, Engwall M,
Braunbeck T & Hollert H (2004): Ermittlung der Oekotoxikologischen Belastung von
Hochwasserschwebstoffen mit Fraktionierungsmethoden und Biotests. 2nd Joint annual
SETAC-GLB und GDCh meeting, Aachen, Germany
Kosmehl T, Wölz J, Garke V, Krebs F, Erdinger L, Braunbeck T, Hollert H (2003): Vergleichende
genotoxische Untersuchungen von Sedimenten des Rheins mit RTG-2- und RTL-W1-Zellen mit
Hilfe des Comet-Assays. 8th Annual SETAC-GLB meeting, Heidelberg, Germany
Hinger T, Garke V, Wölz J, Krebs F, Engwall M, Braunbeck T & Hollert H (2002):
Biotransformationsenzyme und Lipidperoxidation als Biomarker für die Untersuchung von
Sedimentextrakten aus dem Einzugsgebiet des Rheins: Vergleich zweier In vitro-Testsysteme
mit der Zelllinie RTL-W1 aus der Regenbogenforelle und embryonaler Organkulturen aus
Hühnerlebern. 1st Joint annual SETAC-GLB and GDCh meeting, Braunschweig, Germany
11. List of Referencess
194
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