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Daphnid population responses to pesticides

Pieters, B.J.

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Download date: 03 Oct 2020

Daphnid population responses

to pesticides

Academisch proefschrift

ter verkrijging van de graad van doctor aan de Universiteit van Amsterdam

op gezag van de Rector Magnificus prof. dr. J.W. Zwemmer ten overstaan van een door het college voor promoties ingestelde commissie,

in het openbaar te verdedigen in de Aula der Universiteit op donderdag 26 april 2007, te 14.00 uur

door Barry Johan Pieters geboren te Haarlem

Promotiecommissie Promotores Prof. dr. W. Admiraal Prof. dr. M. Liess Co-promotor Dr. M.H.S. Kraak Overige leden Dr. A.M. Breure Dr. T.C.M. Brock Prof dr. A.J. Hendriks Prof. dr. R.W.P.M. Laane Prof. dr. M.W. Sabelis Prof. dr. N.M. van Straalen Dr. P. de Voogt Faculteit der Natuurwetenschappen, Wiskunde en Informatica Daphnid population responses to pesticides Barry J. Pieters Proefschrift Universiteit van Amsterdam, FNWI, 2007 ISBN 978-90-76894-68-3

This study was financed by the Centre for Environmental Research (UFZ) and conducted at the Department of System Ecotoxicology, UFZ, Leipzig, Germany. The printing of this thesis was done by Ponsen & Looijen B.V. (http://www.p-l.nl) and financially supported by: Laboratory for Ecological Risk Assessment (LER) at the National Institute for Public Health and the Environment (RIVM) and the University of Amsterdam Cover artwork: front; Daphnia pulex under a microscope with polarized light, G.P. Matthews (http://www.gpmatthews.nildram.co.uk), back; picture from Bio-Rad Cell Science (http://www.cellscience.bio-rad.com) Cover design: Dick de Zwart and Barry Pieters

you're keeping in step, in the line got your chin held high and you feel just fine because you do, what you're told but inside your heart it is black and it's hollow and it's cold just how deep do you believe? will you bite the hand that feeds? will you chew until it bleeds? can you get up off your knees? are you brave enough to see? do you want to change it? what if this whole crusade's a charade and behind it all there's a price to be paid for the blood on which we dine justified in the name of the holy and the divine so naive to keep holding on to what I want to believe I can see, but I keep holding on and on and on and on… Nine Inch Nails, 2005 Voor mijn ouders en Eugénie…to bring you my love…

Contents Chapter 1 9 General introduction Chapter 2 33 Influence of food limitation on the effects of fenvalerate pulse exposure on the life history and population growth rate of Daphnia magna Chapter 3 49 Maternal nutritional state determines the sensitivity of Daphnia magna offspring to short-term fenvalerate exposure Chapter 4 67 Modeling responses of Daphnia magna to pesticide pulse exposure under varying food conditions: intrinsic versus apparent sensitivity Chapter 5 83 Population developmental stage determines the recovery potential of Daphnia magna populations after fenvalerate application Chapter 6 99 Concluding remarks

Summary 108

Samenvatting 110

Acknowledgements 114

Publications list 116

Curriculum Vitae 117

9

Chapter 1

General introduction

Pesticide effects on different levels of biological organization Intensive application of agrochemicals has resulted in a wide-spread contamination of surface waters with pesticides (Capri & Trevisan 1998). Pesticides, a highly diverse group of compounds used to combat pest animals, plants and fungi, form therefore one of the most important groups of chemical stressors in freshwater ecosystems (Gliwicz & Guisande 1986, Cooper 1993, Hanazato 2001, Liess & Ohe 2005). Modern pesticides have been designed to avoid some of the deleterious effects of the first generation of pesticides that had a long residence time in the environment and accumulated in the foodchain (Heckman 1981). However, compounds currently applied in agriculture are biologically very active and their specificity could not be enhanced to such an extent that effects on non-target species can be avoided (Hill et al. 1994, Schulz et al. 2002). In addition, improved methods of application cannot avoid unintended spreading through spray-drift, runoff (see Figure 1), drainage or atmospheric deposition (Heckman 1982, Baughman et al. 1989, Mian & Mulla 1992, Liess et al. 1999). Consequently, acute mortality of invertebrate species due to pesticide contamination of streams has been observed (Leonard et al. 2000, Schulz 2004), as well as effects on the structure and functioning of natural communities (Helgen et al. 1988, Hatakeyama et al. 1990, Liess & Schulz 1999).

The environmental problems caused by pesticide use prompted the European Union to adopt the Uniform Principles, a registration procedure for accepting pesticides on the European market (EU 1997). This registration procedure prescribes that no unacceptable impacts on the viability of exposed organisms may occur under field conditions. This raises the question how to assess these ‘unacceptable impacts under field conditions’. Ecosystems harbour large numbers of species and the effects of pesticides may show up at different levels of

Chapter 1

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biological organization, i.e. individuals of a species, populations of species and consortia of species forming communities with trophic links. At these higher levels of biological integration, complex food web interactions (indirect effects) may confound a simple assessment of direct pesticide effects (Preston & Snell 2001, Preston 2002). Direct effects may depend on the concentrations reached (Naddy et al. 2000), the duration of the exposure (Reynaldi & Liess 2005) and the mode of action of the chemical (Barata & Baird 2000). Direct effects of pesticide exposure on sensitive species include acute reductions in survival (Barry et al. 1995), or effects on sublethal endpoints such as consumption (Day & Kaushik 1987a), growth (Hooper et al. 2005), age at first reproduction (Antunes et al. 2004) and offspring production (Hosmer et al. 1998). When these individual life-history characteristics are affected, population growth rates are also impaired (Forbes & Calow 2002, Sibly & Hone 2002). Effects of pesticides on population dynamics have been well studied for zooplankton in ponds and enclosures (Kaushik et al. 1985, Day et al. 1987, Yasuno et al. 1988). Medina et al. (2004) studied the effects of a single pesticide application on marine plankton communities in mesocosms and observed reductions in zooplankton density and biodiversity. Zooplankton density recovered after dissipation of the pesticide, but zooplankton community structure remained altered as copepods were killed and numbers of rotifers strongly increased. These indirect effects through trophic cascades in communities are common (Fleeger et al. 2003) and they have been argued to have similar or greater influence on species abundance than direct effects (Lampert et al. 1989). For example, the most frequently reported community response to pesticides is a change from dominance by large zooplankton species to dominance by small zooplankters through relieved competition (Havens & Hanazato 1993, Hanazato 1998). Pesticide exposure may also affect swimming behaviour which consequently alters predator-prey relationships (Dodson et al. 1995), or decrease filtration rates, reducing grazing pressure by zooplankton on phytoplankton, with consequences for species abundances and community composition (Ferrando & Andreu 1993, Bengtsson et al. 2004).

Environmental risk assessment of pesticides aims to quantify and predict effects at the ecosystem level under field conditions (Chapman 1995, Campbell et al. 1999). Although community studies for risk assessment have been performed in situ and in aquatic microcosms and mesocosms (Graney et al. 1994, Van den Brink et al. 1995 & 1996), the necessary time, costs and logistics make it often

General introduction

11

impossible to evaluate even a limited number of agrochemicals. Thus risk assessment still relies heavily on standardized laboratory ecotoxicity tests using individuals where effects are merely measured on simple endpoints such as survival and reproduction (OECD 1998) rather than on demographic endpoints, and is therefore criticized for lacking environmental realism (Kimball & Levin 1985). One concern is that small, statistically undetectable effects on one or more individual life-history traits may be magnified into large impacts on population dynamics (Halbach et al. 1983). The population growth rate, which integrates complex interactions between individual life-history traits, has therefore been proposed as a more relevant measure for ecological impact (Forbes & Calow 1999). Thus, there is an urgent need for incorporating demographic endpoints in ecotoxicological effect assessments in laboratory toxicity tests (Bechmann 1994, Kammenga et al. 1996, Calow et al. 1997, Walthall & Stark 1997). This thesis focuses therefore on population responses to pesticides.

Figure 1. Runoff event occurring on arable land after heavy rainfall, a typical entry route for pesticides into the aquatic environment (picture from UFZ). Population responses to pesticides In environmental risk assessment, population responses to toxicant exposure are often inferred from observations at the individual level, which are then translated to the population level by integrating toxic effects on individual life history traits into effects on population growth rate (Sibly 1996, Calow et al. 1997). Exposure to a toxic substance may affect population growth rates by reducing age-specific

Chapter 1

12

survival probabilities (Sibly 1996) and consumption (Baird et al. 1990), and thus increase generation time and reduce offspring production. Reproduction rates could also be affected as a result of resources being diverted from reproduction to repair or detoxification mechanisms (Baird et al. 1990). But to what extent do effects on life-history variables contribute to impairments of population growth rate? Sibly (Sibly 1996), for example, reanalyzed the data from Daniels & Allan (Daniels & Allan 1981) and demonstrated that toxicant-induced changes in survival and reproduction contributed equally to changes in population growth rate. Kammenga et al. (Kammenga et al. 1996) showed that although maturation time was relatively insensitive to toxicant exposure, small changes in this trait had a large influence on population growth rate; in contrast, the length of the reproductive period was most sensitive to toxicant exposure, but changes in this trait had no detectable influence on population growth rate. Thus, any investigation on the dynamics and/or sensitivity of populations to pesticides should address the question how effects of pesticides on individual life-history traits affect population growth.

To determine population responses to pesticides, toxicity tests may be performed with real populations with interacting individuals instead of testing individuals. However, such demographic bioassays are almost always initiated with individuals having the same life-stage, age and size, unlike natural populations that harbor individuals of all life stages (Stark & Wennergren 1995). Age-dependent differences in sensitivity have been demonstrated (Stuijfzand et al. 2000, Hanazato & Hirokawa 2001), explaining earlier contrasting observations on effects on daphnid populations at a different stage of development (Hanazato & Yasuno 1990, Stark & Banken 1999, Hanazato & Hirokawa 2004). Moreover, populations may include several generations and effects of pesticides on parents may propagate to non-exposed offspring (Sánchez et al. 1999, Villarroel et al. 2000). Using real populations also allows for intraspecific interaction, which may affect individual life-history traits (Van Buskirk 1987). It is well known that life-history parameters are modified by competition (Matveev & Gabriel 1994) and crowding (Seitz 1984, Begon 1985, Cox et al. 1992, Guisande 1993, Burns 1995, Cleuvers et al. 1997). Intraspecific interactions mostly lead to inhibition of individuals (Power 1997, Liess 2002), although sometimes stimulation has been reported (Bertness & Leonard 1997). Competition and crowding often lead to food shortage and food level has therefore been the focal point of many ecotoxicity studies (Grant 1998, Sibly 1999, Sibly et al. 2000, Forbes et al. 2001,

General introduction

13

Barata et al. 2002, Moe et al. 2002, Hooper et al. 2003). When food is severely limited the rate of net population growth drops to zero and the population is in steady state: i.e., the carrying capacity is reached and the rates of multiplication and death are similar. It has been suggested that exposing a density-limited population to a toxicant could reduce the intensity of density dependence, by removing a fraction of the population. This occurs if the toxicant kills some individuals resulting in a higher food level per capita for the survivors and to a stimulated growth (Postma et al. 1994). In exponentially growing populations such a feed back does not exist and thus toxicant exposure would have a stronger impact on exponentially growing populations than on density-limited populations.

It is concluded that heterogeneous populations consisting of individuals differing in development stage under control of density limitation are essential new subjects for toxicity testing. However, the variability of population responses to pesticides may be increased in such experiments, and therefore the present thesis compares cohorts of individuals with mixed populations of Daphnia magna. Variable environment Historically ecosystems and populations have been characterized as being static, tightly organized and highly self-regulating systems of interacting populations that possess a definable equilibrium state (Odum 1971). Much ecological research focused on determining conditions under which this equilibrium would be stable or unstable, and on characterizing the responses of populations that made them move away from the equilibrium (Lewontin 1969, May 1973). Ecologists perceive populations and ecosystems nowadays from a dynamic rather than a static perspective. Zooplankton populations, for instance, have indeed been indicated to show strong cycles in abundance and composition (McCauley & Murdoch 1987). These fluctuations often result from variable biotic and a-biotic environmental conditions, which now have been accepted in ecotoxicology as a common aspect of natural aquatic ecosystems (Heugens et al. 2001). Factors such as temperature and nutritional state have been demonstrated to modify toxicity of pollutants to animals living under conditions close to their environmental tolerance limits. Toxicity is usually enhanced by increasing temperature (Heugens et al. 2003) and decreasing food availability (Koivisto et al. 1992, Barry et al.

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1995, Heugens et al. 2006). Therefore food limitation is one of the environmental variables incorporated in the present research.

Application of pesticides and accidental emissions cause short pulsed exposures to toxicants. Pesticides reach surface waters following periods of spray drift, surface runoff, or drain flow (Williams et al. 1995, Kreuger 1998, Schulz et al. 1998, Liess et al. 1999). These pulses may range from a few minutes to several hours whereafter concentrations decrease rapidly because of adsorption and degradation processes and the permanent water renewal in for example streams (Kreutzweiser & Sibley 1991). Standard laboratory toxicity tests applying continuous exposure scenarios may therefore not be the appropriate tool to investigate the toxicity of short-term pulsed or intermittent exposures of aquatic organisms to pesticides (Parsons & Surgeoner 1991a & b) and may therefore not predict realistic environmental exposure effects (Pascoe & Shazili 1986). Hence, episodic pollution has been addressed in several aquatic ecotoxicity studies (McCahon & Pascoe 1990, Handy 1994, Naddy et al. 2000, Naddy & Klaine 2001, Reinert et al. 2002, Diamond et al. 2006). Some of these studies have demonstrated the importance of including a post-exposure observation period to record the response of the test population (Hansen & Kawatski 1976, Wright 1976). By incorporating post-exposure periods, delayed effects on invertebrate populations following only a brief pesticide exposure have been observed (Beketov & Liess 2005, Forbes & Cold 2005), demonstrating that standard toxicity tests may not adequately evaluate latent, delayed or cumulative effects that can occur after the exposure period. Thus, the incorporation of pulsed- and post-exposure tests into environmental risk assessment for pesticides is essential (Brent & Herricks 1998).

Toxicant-induced perturbations of populations with subsequent recovery appear to occur frequently in the field (Niemi et al. 1990, Wallace 1990, Strauss 1991). The factors that influence the recovery rates of populations after strong pesticide disturbances are numerous and complex, and may depend on variables such as the species in question or the specific life-history trait affected (Sherratt et al. 1999). However, in-depth knowledge of the mechanisms and relationships between affected life-history characteristics and recovery is still lacking, although these variables are crucial for predicting recovery in mathematical models (Barnthouse 2004).

Thus, environmental conditions may vary strongly in the field modifying the responses of organisms to pesticides. The underlying mechanisms are often

General introduction

15

poorly understood and most risk assessment procedures use toxicity tests in which organisms are subjected to toxicants under constant and favorable experimental conditions. In reality however, population responses to toxicants are defined by a large number of variables that regulate biological processes ranging from the activity at the biochemical receptor site to the dynamics of multi-generation populations. So far studies have addressed parts of this range, but it is still a challenge to integrate these variables into a single concept as attempted in this thesis. To this purpose the present project investigated the effects of varying food conditions and included short-term exposure regimes as well as extended post-exposure observation periods to examine effects propagation and recovery. Aim and objectives of the present thesis The present thesis addresses population responses to pesticides, incorporating the many variables involved: individual size and age, variable food regime and recovery from temporal exposure. The objectives of this study are:

1. To determine the joint effects of food limitation and pesticide pulse exposure on life-history traits and population growth rate of Daphnia magna cohorts.

2. To integrate food limitation and pulse exposure to pesticides in a process-based model for an improved understanding of the dynamics of population responses to pesticides.

3. To verify the results obtained for cohorts with observations on mixed Daphnia magna populations in different stages of their development.

Test species The freshwater filter feeder Daphnia magna (Figure 2, left) was selected as test organism in this study. Daphnia is frequently used as a standard test organism in laboratory ecotoxicity studies due to its parthenogenetic reproduction which excludes genetic variation. Other benefits include its ease of handling, rapid culturing and short generation times, and a generally high sensitivity to toxicants (Adema 1978). Much is known about the biology of the genus (Ebert 1991, Ebert 1992). Daphnia sheds its carapace by moulting in order to grow. At 20 °C, Daphnia moults approximately every 2-3 days. Daphnia carrys its eggs and

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embryos in a transparent brood pouch on its back until they hatch. The first brood of neonates is released at the age of 6-10 days, just before moulting. Shortly after moulting, new eggs are deposited in the brood pouch, so the incubation period almost equals the intermoult period (Ebert 1994). Well-fed adult females can reach a length of 5 mm while the maximum life-span of a daphnid is approximately 2-3 months.

In contrast to our laboratory cultures, daphnids in the field reproduce by cyclic parthenogenesis that is characterized by one or more parthenogenetic generations (i.e. diploid females produce diploid offspring), alternating with sexual reproduction (Figure 2, right). Cyclic parthenogenesis gives Daphnia the advantage of rapid recruitment, but also the advantage of maintaining genetic diversity by sexual reproduction (Maynard Smith 1986). In spring Daphnia populations usually consist of females which produce offspring in numerous broods under high food conditions. At the end of the growing season asexual females are cued by declines in food quantity and quality (Slobodkin 1954), crowding (Larsson 1991), or a decrease in photoperiod (Stimpfl 1971) to produce males and sexual females. Sexual females then produce two meiotic eggs which after fertilization will sink to the bottom and persist throughout the winter (i.e., resting eggs), with females hatching again in spring time when conditions become favorable again.

The genus Daphnia (Crustacea; Cladocera) plays a key role in pelagic food webs of temperate regions. Daphnia feeds on algae, fungi, cyanobacteria, bacteria, protozoa, detritus and silt particles (Burns 1969) using a filtering apparatus. When Daphnia is present in lakes and ponds, they are often the most important grazers on phytoplankton (Gliwicz & Lampert 1990). Hence their occurrence modifies the density, seasonal succession and species composition of phytoplankton. In turn Daphnia is an essential food source for carnivorous invertebrates and fish (Reede 1998). Consequently it is under strong top-down control. Adaptive phenotypic plasticity is common in Daphnia and ensures the future existence of populations. For example, individual life-history traits such as size at first reproduction may change under size-selective fish predation (Lampert 1993, Barata et al. 2001). Phenotypic variation can also be expressed as differences in morphology and physiology (Jacobs 1987), or in behaviour (De Meester, 1995).

Food availability strongly modifies life history traits of Daphnia and it induces trade-offs between traits which guarantee survival under varying food

General introduction

17

conditions. Daphnia allocate energy to reproduction in such a way that the sum of present and future reproduction is maximized (Stearns 1976, Stearns 1992). Limiting food conditions induce a shift in reproductive strategy from quantity to quality (Tessier & Consolatti 1991), characterized by increases in lipid content (Tessier et al. 1983) and egg and neonate size (Glazier 1992, Gliwicz & Guisande 1992), but fewer offspring. Food shortage during the life cycle causes reductions in somatic growth rate and reproduction, and delays the onset of reproduction. Due to this plasticity in traits in response to varying food concentrations, Daphnia is a suitable test organism to be used in the present research on population responses to pesticides under varying food conditions. Figure 2. Morphology of Daphnia magna (left) and its parthenogenetic life cycle (right) under field conditions (from Freeman & Bracegirdle 1971). Test compound The pyrethroid insecticide fenvalerate (CAS: 51630-58-1) was chosen as model pesticide in the present study (Figure 3). Pyrethroid insecticides have been used in agriculture for more than 30 years to control insect pests in a range of crops. They account for approximately one-fourth of the worldwide insecticide market (Casida & Quistad 1998). Pyrethroid insecticides have a low toxicity to mammals and birds (LD50s generally > 1.0 mg Kg-1), but are highly toxic to terrestrial insects and non-target aquatic arthropods (LC50s generally < 10 µg L-1; Clark et al. 1989). Pyrethroids normally occur in surface waters in a pulsed manner following periods of spray drift, surface runoff, or drain flow (Kreuger 1998, Liess & Schulz 1999). This makes fenvalerate a very relevant compound for studying the

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persistence of toxic effects following the end of a pulse exposure, and for investigating recovery. Having a high octanol-water partition coefficient (Log Kow = 6.2 at 20 °C) it has a relatively high potential for bioaccumulation into organisms. Pyrethroid insecticides are not very persistent in the environment; the half-life time in water is generally less than 4 days. Although pyrethroids may bioconcentrate in organisms, depuration also occurs, whereas bioaccumulation through the food chain is not a significant route for intoxication (Solomon et al. 2001). Although vertebrate species such as fish are capable to actively biodegrade pyrethroids (Ohkawa et al. 1980, Haya 1989), metabolization of pyrethroids by invertebrate species such as Daphnia does virtually not occur (Ohkawa et al. 1980).

The pyrethroid class of insecticides was derived from natural compounds (the pyrethrins) isolated from the plant genus Chrysanthemum (Casida 1980). Natural pyrethrins do have insecticidal activity and they are inherently unstable when exposed to light. Therefore, the pyrethrin structure was modified to produce more stable compounds that retained the desirable insecticidal and toxicological properties (Valentine 1990). All pyrethroids contain several common features: an acid moiety, a central ester bond, and an alcohol moiety. Pyrethroid chemistry and action are classified as type 1 or type 2 (Solomon et al. 2001), depending on the alcohol constituent. Fenvalerate (Figure 3) belongs to the type 2 class and specifically contains an a-cyano-3-phenoxybenzyl alcohol and an altered acid portion of the molecule by including a phenyl ring, which increases insecticidal activity considerably (Adelsbach & Tjeerdema 2003). Fenvalerate has a primary mechanism of action at the pre-synaptic membrane that involves increased release of synaptic vesicles through an effect on voltage-dependent calcium channels (Clark & Brooks 1989). In cladocerans this causes loss of coordination (Christensen et al. 2005), immobilization (Reynaldi et al. 2006), and therefore reduction in feeding rates (Day & Kaushik 1987a, McWilliam & Baird 2002).

Fenvalerate has been put on the list of priority compounds by the Dutch government. The ad hoc Maximum Permissible Concentration (MPC) of fenvalerate adopted in The Netherlands is 4.08 µg L-1. However, peak concentrations in agricultural streams have been measured up to 6.2 µg L-1 (Liess & Schulz 1999), whereas a NOECecosystem threshold value of 0.01 µg L-1 has been observed (Day et al. 1987). Invertebrates are far more susceptible to fenvalerate than fish, while oysters, snails, molluscs and algae and are relatively insensitive (Solomon et al. 2001, Anderson 1982). Fenvalerate has been shown to affect

General introduction

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cladoceran survival, growth, maturation time, reproduction and consequently population growth rate in laboratory toxicity tests (Day & Kaushik 1987b). Application of fenvalerate in outdoor mesocosm studies resulted in long-term disturbances of the structure of macroinvertebrate communities (Day et al. 1987, Woin, 1998), Heckmann & Friberg 2005). Fenvalerate exposure associated to agricultural runoff in the field has likewise been indicated to affect aquatic community structure and functioning (Baughman et al. 1989, Schulz & Liess 1999). Figure 3. The molecular structure of the pyrethroid insecticide fenvalerate. Outline of the present thesis This thesis departs from analyzing the influence of food limitation on the effects of fenvalerate pulse exposure on the life history and population growth rate (r) of Daphnia magna cohorts (Chapter 2). Neonates were subjected to a 24-h pesticide pulse exposure under high and a low food condition and were monitored for 21 days. Given the joint effects of food limitation and pesticide exposure on neonates, the question was raised what the effect of maternal food limitation on population responses to pesticide exposure would be. Chapter 3 describes the results of transgenerational life-cycle experiments where the effects of varying maternal food level on subsequent offspring performance and sensitivity to a pulsed fenvalerate exposure were studied. Daphnia cultures were maintained at a high food level and a low food level and the produced offspring was subsequently used in life-table response experiments under a low food condition. It remained unclear in the Chapters 2 & 3 whether the food-dependent fenvalerate toxicity was due to differences in accumulation kinetics and resource allocation, or results from a changed intrinsic sensitivity of daphnids. Therefore, it is hypothesized in Chapter 4 that, if food level only affects resource allocation and accumulation kinetics of a toxicant, then the intrinsic sensitivity to this toxicant should be the same for all food regimes. This hypothesis was investigated using the mathematical model DEBtox (Kooijman & Bedaux 1996), based on the theory of

Chapter 1

20

Dynamic Energy Budgets, re-examining the results of the Chapters 2 & 3. Chapter 5 verifies the results obtained for cohorts with observations on mixed Daphnia magna populations in different stages of their development. Single doses of fenvalerate were applied to populations growing exponentially and at carrying capacity. Population stage-depending differences in sensitivity and dynamics of recovery were analyzed. The concluding remarks (Chapter 6) discuss the main findings of this thesis and review how the many variables involved jointly determine population responses to pesticides. References Adelsbach, T.L., and R.S. Tjeerdema. 2003. Chemistry and fate of fenvalerate and

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Chapter 2

Influence of food limitation on the effects of

fenvalerate pulse exposureon the life-history and population growth rate of Daphnia magna

B.J. Pieters, A. Paschke, S. Reynaldi, M.H.S. Kraak, W. Admiraal, and M. Liess.

Environmental Toxicology & Chemistry (2005) 24: 2254-2259. Abstract Laboratory ecotoxicity tests may not adequately evaluate the effects of pesticides, because they often do not include more environmentally relevant conditions, such as pulsed toxicant exposures and low food conditions. Therefore, we tested the effects of a pulse of the pyrethroid insecticide fenvalerate (FV) on the life-history and population growth rate (r) of the cladoceran Daphnia magna. The daphnids were subjected to a 24 h pesticide pulse exposure (0.03, 0.1, 0.3, 0.6, 1.0, and 3.2 mg L-1) under high and low food conditions and were monitored for 21 days. Chemical analysis showed that at t = 1 h, the nominal FV concentrations were reduced by 50 to 66%. Fenvalerate decreased survival and growth in the week following pulse exposure. Age at first reproduction increased, with consequent adverse effects on cumulative reproduction per living female and, therefore, on r. Thus, a short-term exposure of FV caused a long-term reduction on r as a result of increased mortality and a delay in development. Low food conditions exacerbated the effects of the FV exposure on juvenile survival and growth during the first week. This caused a much stronger reduction in r under low food conditions. We concluded that a pulsed FV exposure may result in long-term reduction of r that can be predicted only with more environmentally relevant toxicity tests, as described in the present study.

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Introduction

Laboratory ecotoxicity tests are generally performed under high food conditions in order to obtain high control survival (Buikema et al. 1980, Snell & Moffat, 1999). In the field, however, populations may experience periods of low nutritional supply (Neill 1978, Tessier 1981, McCauley et al. 1990), even in eutrophic waters when populations have grown to their carrying capacity (Kooijman & Metz 1984). Starved individuals have fewer resources available for physiological defense against stressors, and therefore may be more sensitive to toxicants (Sibly 1999). Indeed, several studies demonstrated an increase in toxicity of xenobiotics under limiting food conditions (Winner et al. 1977, Enserink et al. 1995, Klüttgen et al. 1996, Rose et al. 2002).

A second limitation is that organisms in laboratory ecotoxicity tests are usually continuously exposed to toxicants (Naddy et al. 2000). This may be in contrast with populations in the field that may be exposed in an episodic manner. Pesticide application in agriculture, for example, results in short-term pulse concentrations due to spray drift, drain flow, or edge-of-field runoff (Baughman et al. 1989, Reynaldi & Liess 2005). Hence, test designs with continuous exposure and no post-exposure observation period may lack environmental realism because they do not include latent effects or the recovery of the survivors once the pulse exposure has ceased.

The influence of low food conditions on short-term pesticide exposure has received little attention. Therefore, the aim of the present study was to evaluate the influence of food level on the effects of a 24 h pesticide exposure on the life-history characteristics and the instantaneous rate of population growth (r) of the model test species Daphnia magna. To this purpose, we performed life-table response experiments which are commonly used in the assessment of potential risks of toxicants (Caswell 1996, Liess et al. 1999). We were especially interested in examining the extent to which low food levels affected life-history characteristics and ultimately population growth rate in the period after the pesticide pulse exposure had ended. Since r integrates effects on juvenile and adult survival, reproduction, and age at first reproduction, this parameter has been proposed as a more relevant measure to predict consequences of toxicants on the population level (Levin et al. 1996).

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35

The pyrethroid insecticide fenvalerate (FV) was chosen as model pesticide. It is widely used for the protection of agricultural crops due to its high efficacy and generally low mammalian toxicity. Fenvalerate has regularly been detected in the aquatic environment (Forbes & Calow 1999), and peak concentrations of 0.1 to 6.2 µg L-1 have been measured in streams and estuaries (Baughman et al. 1989, Reynaldi & Liess 2005, Kreuger 1998, Scott et al. 1999). A high sensitivity has, however, been found for nontarget organisms such as aquatic macroinvertebrates (Anderson 1982). Also, long-term effects in the structures of macroinvertebrate communities after a FV pulse exposure have been observed in outdoor mesocosm studies (Woin 1998, Liess 2002) and in the field (Lies & Schulz 1999). Materials & methods Daphnia magna culture

The experiments were performed with individuals of the cladoceran Daphnia magna Straus (clone B; Bayer, Monheim, Germany). Continuous cultures of 10 adults L-1 in artificial Elendt M7 medium (OECD 1998) were maintained at 20°C ± 1°C, normal light (~1400 Lux), and a 16:8 light:dark photoperiod. The pH was 7.4, conductivity 630 µS cm-1, and the dissolved oxygen level was 7.15 mg L-1. Once a week a new culture was initiated using < 24 h old neonates from a three week old culture. The newly released neonates were removed twice a week, and the medium was renewed three times a week. The daphnids were fed three times a week with a suspension of batch-cultured green algae (Desmodesmus subspicatus). Algae were cultured in algae medium according to Grimme & Boardmann (Grimme & Boardmann 1972) and continuously aerated (3% CO2). The algae were harvested in exponential growth phase, centrifuged, and the pellet was resuspended in M7 medium. The food quantity given as equivalent daily ration was 5.3 × 105 cells mL-1 (0.75 mg carbon L-1). Measurements were performed with a total organic carbon analyzer (Shimatzu TOC-5000 Analyzer, Duisburg, Germany) and a coulter counter (Casy1 counter, Scharfe Systems, Reutlingen, Germany). Fenvalerate exposure & measurement

Fenvalerate, (RS)-α-cyano-3-phenoxybenzyl (RS)-2-(4-chlorophenyl)-3-methylbutyrate (CAS: 51630-58-1), was obtained from Riedel-de Haën, Seelze,

Chapter 2

36

Germany (high-performance liquid chromatography (HPLC) technical grade: 99.9% purity). Dimethylsulfoxide (DMSO) was used as carrier solvent and obtained from Merck, Darmstadt, Germany (HPLC technical grade: 99.8% purity). The maximum amount of DMSO present in the test medium was 0.0003% (volume/volume). Our previous research did not show any detectable effect on D. magna at this DMSO concentration (unpublished results). Exposure consisted of seven different nominal concentrations for both food regimes (control, 0.03, 0.1, 0.3, 0.6, 1.0, and 3.2 µg L-1). Actual exposure concentrations were determined in triplicate for both treatments at t = 1 h before food application. Accurate detection was limited to the concentrations 0.6, 1.0, and 3.2 µg L-1. Additionally, a 1.0 µg L-1 test solution (n = 3) was measured at t = 24 h. Samples were measured by solid-phase extraction of 1 L volumes with C18 columns (Baker, Phillipsburg, NJ, USA), followed by gas chromatography–electron capture detection (gas chromatograph HP 5990, Series II, Hewlett-Packard, Avondale, PA, USA) and confirmed with gas chromatography–mass spectrometry (negative chemical ionization, Varian 3400 gas chromatograph, Varian, Walnut Creek, CA, USA) with an HP 7673 autosampler (Hewlett-Packard) that was directly capillary-coupled to the quadruple mass spectrometer SSQ 700 (Finnigan, Bremen, Germany). Analytical measurements (± standard deviations) of the 0.6, 1.0, and 3.2 µg L-1 test solutions at t = 1 h showed a reduction in the nominal concentrations of 50 to 66% (0.20 ± 0.00, 0.43 ± 0.12, and 1.47 ± 0.28 µg L-1, respectively). The actual concentration of the 1.0 µg L-1 test solution decreased to 0.37 ± 0.06 µg L-1 after 24 h. Nominal concentrations are given in the following sections. Life-table response experiments

The standard 21-d Daphnia reproduction test (OECD 1998) was adjusted to study the influence of low food levels on the effects of a FV pulse exposure on the life-history characteristics of D. magna. The life-table response experiments were carried out with two varying food levels: high and low. A rangefinding reproduction test with various food levels was conducted to establish the high food level. The guidelines of the Organization for Economic Cooperation and Development provide a minimum cumulative reproduction criterion of 60 neonates per living female after 21 d. The high food level used in this study (equivalent daily ration of 5.3 × 105 cells mL-1; 0.75 mg carbon L-1) resulted in 64 neonates per living female after 21 d. An approximately three times lower food

Influence of food limitation

37

level (equivalent daily ration of 1.5 × 105 cells mL-1; 0.21 mg carbon L-1) was chosen as the low food level that allowed females to produce 24 neonates after 21 d.

The life-table response experiments were initiated with newly released neonates (< 24 h) from a three-week-old culture, with a mean body length of 0.75 mm (± 0.05 standard deviation; n = 40). Each treatment consisted of 20 individuals incubated in 20 glass beakers containing 80 mL of test medium. The daphnids were exposed to FV for 24 h and were fed with the respective food levels of both experiments during the exposure in order to prevent food stress. The neonates were rinsed after 24 h and gently transferred into uncontaminated M7 medium. Feeding and renewing of the media was done three times a week. Other experimental conditions were similar to those described for the culture. On a daily basis the following life-history characteristics of the daphnids were recorded: Survival (defined as swimming and immobilized animals) and the number of living neonates produced. Newly released offspring were removed. From these daily-observed life-history characteristics, age at first reproduction, mean brood number, mean brood size, and cumulative reproduction per living female at day 21 were determined. The population growth rate was calculated from the integration of the age-specific data on survival and fecundity probabilities. The r values were calculated iteratively from the Euler/Lotka equation (1) (Lotka 1913, Euler 1970):

1= lxmxe

− rx

X=0

Ω

∑ (1)

Where r = per capita rate of increase for the population d-1, x = age class (days; 1, 2, 3…Ω), Ω = oldest age class in the population (21 d in the present study), lx = probability of surviving at age x, and mx = neonates per mother at age x.

Interactive calculations were performed in order to determine r values according to Equation 2. Uncertainties were estimated from jackknife pseudovalues according to Meyer et al. (Meyer et al. 1986).

To determine somatic growth of the daphnids during the test period, body lengths (distance from the middle of the eye to the base of the tail) were measured at day 7, 14, and 21. Measurements were performed using a Leica MS 5 Microscope equipped with a Leica DFC 300 F Digital Camera, and a Leica KL 1500 LCD light source (Leica Microsystems, Solms, Germany).

Chapter 2

38

Statistical analysis Data for all parameters were tested for normality (Kolmogorov-Smirnov test)

and homoscedasticity (Levene’s test). When these conditions were not met, data were log(x+1) transformed. Controls of both food levels were compared with each other using two-tailed t tests. Fenvalerate treatments within food levels were compared with the corresponding control using one-way analysis of variance (ANOVA; type 3 sums of squares), followed by Dunnett’s post-hoc test. In order to assess the contribution of FV concentration and food level to the observed variation, a two-way ANOVA was performed. For survival as a function of time, significant differences between FV treatments and the corresponding control were analyzed by conducting pairwise comparisons using Gehan-Wilcoxon Survival Analysis. An alpha level of 0.05 was used for all statistical tests and three significance levels were adopted ( p < 0.001, p < 0.01, p < 0.05). RESULTS Survival

No mortality occurred in the controls during the experiments. The 24 h exposure to FV strongly impaired survival (Figure 1). Mortality generally occurred between day 2 to 7 after the pulse exposure and complete mortality at both food levels was observed at 3.2 µg L-1. However, the Lowest Observed Effect Concentration (LOEC) for the low food treatment was 0.6 µg L-1, whereas for the high food treatment it was 1.0 µg L-1 (Figure 1). This indicated that low food levels aggravated the effects of FV on survival of D. magna. Body length

Low food conditions caused significant (p < 0.001) smaller control body lengths at day 7, 14, and 21 compared to the high food level (Figure 2). Reduced growth after the short-term FV exposure was only apparent at day 7. However, the LOEC value for the low food treatment was at 0.3 µg L-1, whereas it was 1.0 µg L-1 for the high food treatment (Figure 2). Two-way ANOVA demonstrated that at day 7 the statistical interaction term between the parameters FV and food was significant in explaining the observed variation in body length (p < 0.05). This indicated that the low food conditions significantly exacerbate the effect of FV on somatic growth of D. magna in the first week. Complete recovery of body

Influence of food limitation

39

length to the corresponding controls occurred at day 14 at both food levels, and no significant differences compared to the corresponding controls were observed. Figure 1. Survival of Daphnia magna expressed as a function of time after pulse exposure to fenvalerate (µg L-1) under high and low food conditions. Asterisks denote significant differences compared to the corresponding controls (Gehan-Wilcoxon survival analysis: ** p < 0.01, *** p < 0.001). Figure 2. Effects of a pulse exposure to fenvalerate (µg L-1) on the body length of Daphnia magna under high and low food conditions. Error bars indicate ± 95% confidence intervals and asterisks denote significant differences compared to the corresponding controls (analysis of variance: ** p < 0.01, *** p < 0.001). Reproduction

Low food conditions significantly (p < 0.001) increased age at first reproduction (Figure 3) and decreased mean brood number, mean brood size, and

high food

days after exposure0 7 14 21

0

20

40

60

80

100

control0.030.10.30.61.03.2

surv

ival

(% in

itial

anim

als)

***

**

high food

days after exposure0 7 14 21

0

20

40

60

80

100

control0.030.10.30.61.03.2

surv

ival

(% in

itial

anim

als)

******

****

0 7 14 210

20

40

60

80

100

***

***

**

high food

days after exposure0 7 14 21

0

20

40

60

80

100

******

******

**

high food

days after exposure

high food

fenvalerate concentration (µg L-1)0.03 0.1 0.3 0.6 1.0

body

leng

th(m

m)

1.5

2.0

2.5

3.0

3.5

4.0

4.5

5.0

Day 7Day 14Day 21

control

**

high food

fenvalerate concentration (µg L-1)0.03 0.1 0.3 0.6 1.0

body

leng

th(m

m)

1.5

2.0

2.5

3.0

3.5

4.0

4.5

5.0

Day 7Day 14Day 21

control

****

0.03 0.1 0.3 0.6 1.01.5

2.0

2.5

3.0

3.5

4.0

4.5

5.0

control

low food

******

***

fenvalerate concentration (µg L-1)0.03 0.1 0.3 0.6 1.0

1.5

2.0

2.5

3.0

3.5

4.0

4.5

5.0

control

low food

************

******

fenvalerate concentration (µg L-1)

Chapter 2

40

cumulative reproduction per living female at day 21 (Table 1) in comparison to the high food controls. Also, the 24 h FV exposure affected several of these reproductive traits. Both age at first reproduction and mean brood number were affected and showed LOEC values for the low food treatment at 0.3 µg L-1, whereas these were 0.6 µg L-1 for the high food treatment (Figure 3 and Table 1). Figure 3. Effects of a pulse exposure to fenvalerate (µg L-1) on the age at first reproduction of Daphnia magna under high and low food conditions. Error bars indicate ± 95% confidence intervals and asterisks denote significant differences compared to the corresponding controls (analysis of variance: ** p < 0.01,.*** p < 0.001). Fenvalerate caused no significant reductions in mean brood size, in contrast to the low food treatment which showed a significant increase at 1.0 µg L-1 (Table 1). The cumulative reproduction after 21 d was strongly impaired after FV exposure, although identical LOEC values were exhibited at 0.6 µg L-1 for both food treatments (Table 1). Two-way ANOVA demonstrated no significant (p > 0.05) interactions between the parameters FV and food for all reproductive traits. This indicated that low food conditions did not exacerbate the effects on reproductive traits caused by the short-term FV exposure, although this contradicted with the differing LOEC values for age at first reproduction, mean brood number, and mean brood size. Closer examination revealed that this discrepancy could be explained by the relatively small differences in percent changes in inhibition between food treatments and their overlapping confidence intervals.

0.03 0.1 0.3 0.6 1.0

age

at fi

rstr

epro

duct

ion

(day

s)

7

8

9

10

11

12

13

14 high foodlow food

controlfenvalerate concentration (µg L-1)

******

***

*****

0.03 0.1 0.3 0.6 1.0

age

at fi

rstr

epro

duct

ion

(day

s)

7

8

9

10

11

12

13

14 high foodlow food

controlfenvalerate concentration (µg L-1)

************

******

**********

Influence of food limitation

41

Table 1. Mean brood number (BN), mean brood size (BS), and cumulative reproduction per living female at day 21 (RPR) of Daphnia magna exposed to Fenvalerate (FV) under high and low food conditionsa. food regime FV (µg L-1) BN BS RPR high control 4.86 ± 0.34 12.4 ± 0.6 64.5 ± 3.6 0.03 4.70 ± 0.27 13.3 ± 0.6 61.7 ± 3.5 0.1 4.79 ± 0.27 12.6 ± 1.1 59.9 ± 6.5 0.3 4.57 ± 0.30 12.6 ± 1.0 57.4 ± 5.0 0.6 4.33 ± 0.39* 12.2 ± 0.6 52.5 ± 5.2** 1.0 4.30 ± 0.41** 12.6 ± 1.5 53.7 ± 6.1* low control 4.26 ± 0.27 5.72 ± 0.27 23.7 ± 1.0 0.03 4.15 ± 0.27 5.69 ± 0.35 23.6 ± 1.0 0.1 3.89 ± 0.22 5.75 ± 0.32 22.2 ± 1.4 0.3 3.69 ± 0.26** 5.78 ± 0.52 21.2 ± 2.1 0.6 3.36 ± 0.34** 5.86 ± 0.41 19.6 ± 1.7* 1.0 2.83 ± 0.43*** 6.97 ± 0.70** 19.7 ± 3.1*

aValues are presented as the mean ± 95% confidence intervals. Asterisks denote significant differences compared to the corresponding controls (analysis of variance; * p < 0.05, ** p < 0.01, *** p < 0.001). Population growth rate

The population growth rate in the control (Figure 4) was significantly lower (p < 0.001) at the lower food level than at the high food level. Fenvalerate pulse exposure strongly affected r calculated for the 21 d post treatment period. At day 21, the LOEC value for the high food treatment was 0.6 µg L-1, whereas it was at 0.3 µg L-1 for the low food level (Figure 4). Two-way ANOVA demonstrated that at day 21 the statistical interaction term between the parameters FV and food was significant in explaining the observed variation of r (p < 0.001). This indicated that the effects of FV on r were enhanced at the low food level. Discussion The present study revealed that a 24 h FV pulse exposure affected population growth rate of Daphnia magna during a subsequent 21 day period, and that low

Chapter 2

42

food levels exacerbated this long-term adverse effect of the pesticide pulse. The effects of FV on r can be traced back to impairments in the life-history characteristics mortality, growth, and reproduction. Despite the brief FV exposure of only 24 h, FV-induced mortality still occurred up to day 7 and even resulted in complete mortality of the individuals at the highest test concentration. Figure 4. Effects of a pulse exposure to fenvalerate (µg L-1) on the population growth rate (r) of Daphnia magna during a 21 d period, under high and low food conditions. Error bars indicate ± 95% confidence intervals and asterisks denote significant differences compared to the corresponding controls (analysis of variance: *** p < 0.001). Similar long-term effects (> 13 d) on survival after a 1 h exposure to the FV isomer esfenvalerate have been demonstrated for juvenile Gammarus pulex (Cold & Forbes 2004). A period with decreased growth (until day 7) caused by the FV exposure was followed by a period of rapid growth leading to an uniform body length by day 14. Delayed development after FV exposure has been observed previously (Liess 2002). This is likely the consequence of either immobilization, leading to reduced filtration and assimilation rates (Day & Kaushik 1987a), or due to changed resource allocation caused by activation of detoxification systems. Effects of FV on mean brood size were not demonstrated, while reductions in the mean brood number and cumulative reproduction per living female were the result of the delay in age at first reproduction. Hence, the decrease in r due to the FV pulse exposure could be explained by delays in the individual life-history characteristics that were affected.

0.03 0.1 0.3 0.6 1.00.10

0.15

0.20

0.25

0.30

0.35

high foodlow food

control

popu

latio

ngr

owth

rat

e (d

-1)

fenvalerate concentration (µg L-1)

******

***

***

***

0.03 0.1 0.3 0.6 1.00.10

0.15

0.20

0.25

0.30

0.35

high foodlow food

control

popu

latio

ngr

owth

rat

e (d

-1)

fenvalerate concentration (µg L-1)

************

******

******

******

Influence of food limitation

43

Population growth rate in the control treatment at the low food level declined as a result of a delay in age at first reproduction, a reduced number of broods per time, a smaller brood size, and a lower cumulative reproduction per living female. Especially the delay in age at first reproduction may have affected r, since population theory states that r depends mainly on the reproductive success of the younger individuals, and reproduction in older individuals is of less importance (Kammenga et al. 1996).

Low food conditions exacerbated the effects of the short-term FV exposure on r. Several investigations conclude that limiting food conditions increases the toxicity of xenobiotics, as summarized in the review of Heugens et al. (Heugens et al. 2001). A decreased survival of daphnids at low nutritional supply has been demonstrated earlier for continuous exposure to metals (Koivisto et al. 1992), pesticides (Barry et al. 1995), and metabolites of pesticides such as 3, 4-dichloroaniline (Kooijman & Metz 1984). The stronger reduction in growth under low food conditions matches well with findings of studies on metals (Enserink et al. 1995, Chandini 1989) but contradicts the observations made in the study of Barry et al. (Barry et al. 1995), who investigated the effects of several food concentrations on the toxicity of esfenvalerate to Daphnia carinata. A plausible explanation for this discrepancy may be that Barry et al. compared two algae concentrations that differed only slightly. Smaller body lengths usually cause an increase in age at first reproduction (Taylor & Aiken 1985), but this allometric relationship was, remarkably, not demonstrated in our results, since no interaction between FV and food level was shown in the age at first reproduction. Daphnids were apparently able to recover rapidly after day 7, as was already indicated by the complete recovery in body length at day 14. Since no increased sensitivities to FV at low food levels were shown for age at first reproduction and all reproductive endpoints, increased juvenile mortality must have been mainly responsible for the stronger reduction in r. The delay in population development was still apparent after 21 d, and effects on populations may therefore fade away after even longer periods of time (> 21 d).

The most common explanation given for increased sensitivities at low food supply is that individuals have fewer resources available for physiological defence against stressors. In addition, sorption of toxicants to different concentrations of food particles may either decrease or increase bioavailability of FV, depending on the route of uptake (Heugens et al. 2001). This may have played a role in the present study as well, since the lipophilic nature of FV may have caused sorption

Chapter 2

44

of FV to algae. Algal cells with absorbed FV are not considered the main route of uptake in Daphnia and do not contribute to an increase in bioaccumulation (Day & Kaushik 1987b), and dissipation through the surface area of Daphnia may therefore be a more relevant route of uptake. The fraction of FV bound to algal cells is likely to increase with increasing algal cells, and therefore decreasing the waterborne FV concentrations. Hence, the high food levels may have induced a decreased bioaccumulation of FV. Although the relative contribution of the increased sensitivity of the test organisms and the increased bioavailability of the test compound cannot be fully distinguished, it does not influence one of the main outcomes of the present study, i.e., that recovery from FV pulse exposure is delayed by food limiting conditions.

The incorporation of more environmentally relevant conditions in our experiments indicated that low food conditions may aggravate the effects of a brief FV exposure on the population growth rate of Daphnia magna even after 21 d, which was predominantly the result of an enhanced juvenile mortality. We also showed that tests with constant toxicant exposure may not adequately evaluate the effects of a FV pulse exposure on Daphnia magna due to the persistence of lethal and sublethal effects after the cessation of exposure, and the exhibited recovery trends in, e.g., growth.

Acknowledgements

The authors would like to thank I. Ränker (UFZ) for technical assistance in the laboratory. References Anderson, R.L. 1982. Toxicity of fenvalerate and permethrin to several non-target

aquatic invertebrates. Environ. Entomol. 11: 1251-1257. Barry M.J., D.C. Logan, J.T. Ahokas, and D.A. Holdway. 1995. Effect of algal

food concentration on toxicity of two agricultural pesticides to Daphnia carinata. Ecotoxicol. Environ Saf. 32: 273-279.

Baughman D.S., D.W. Moore, and G.I. Scott. 1989. A comparison and evaluation of field and laboratory toxicity tests with fenvalerate on an estuarine crustacean. Environ. Toxicol. Chem. 8: 417-429.

Buikema, A.L. Jr., J.G. Geiger, and D.R. Lee. 1980. American Society for Testing and Materials. Daphnia toxicity tests. STP 715. In Aquatic Invertebrate Bioassays. Philadelphia, PA, pp 48-69.

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Caswell, H. 1996. Demography meets ecotoxicology: Untangling the population level effects of toxic substances. In Newman, M.C., and C.H. Jagoe, Eds., Ecotoxicology: A Hierarchical Treatment. Lewis, Boca Raton, FL, USA, pp 255-292.

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Day, K., and N.K. Kaushik. 1987b. The adsorption of fenvalerate to laboratory glassware and the algae Chlamydomonas reinhardii, and its effect on uptake of the pesticide by Daphnia galeata mendotae. Aquat. Toxicol. 10: 131-142.

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Klüttgen, B., N .Kuntz, and H.T. Ratte. 1996. Combined effects of 3,4-dichloroaniline and food concentration on life-table data of two related cladocerans, Daphnia magna and Ceriodaphnia quadrangula. Chemosphere 32: 2015-2028.

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Taylor, B.E., and S.C. Aiken. 1985. Effects of food limitation on growth and reproduction of Daphnia. Archiv für Hydrobiologie 21: 285-295.

Tessier, A.J. 1981. The comparative population regulation of two planktonic Cladocera (Holopedium bibberum and Daphnia Catawba). Ph.D. thesis. University of Pennsylvania, Philadelphia, PA, USA.

Winner, R.W., T. Keeling, R. Yeager, and M.P. Farrell. 1977. Effect of food type on the acute and chronic toxicity of copper to Daphnia magna. Freshw. Biol. 7: 343-349.

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49

Chapter 3

Maternal nutritional state determines the sensitivity of Daphnia magna offspring to short-term fenvalerate

exposure

B.J. Pieters, and M. Liess. Aquatic Toxicology (2006) 76: 268-277. Abstract Daphnia populations in the field suffer periodic natural stress conditions such as low food levels. It is known that at low nutritional supply, Daphnia produces fewer but larger offspring which are acutely less sensitive to chemical stress. We hypothesized that the change in the reproductive strategy may also alter the chronic sensitivity of offspring to pesticides which usually occur in the field in a pulsed manner. We therefore investigated the influence of maternal food quantity of Daphnia magna on (1) offspring size and successive performance, and (2) sensitivity to a 24 h exposure to the pyrethroid insecticide Fenvalerate. Daphnia cultures were maintained at a high and low food level. Produced offspring were subsequently used in life-table response experiments which were performed under low food conditions. Results showed that low maternal food conditions (compared to high) increased the offspring size at time of birth, reduced age at first reproduction, and increased reproductive output, which jointly enhanced offspring fitness as estimated by the population growth rate (r). Fenvalerate exposure in combination with low maternal food levels caused (compared to high) a strong decrease in acute sensitivity of neonates, which was generally also observed for chronic endpoints. Hence, reduction in population growth rate resulting from short-term pesticide exposure appeared to be less strong when compared to daphnids originating from high fed mothers. Our findings illustrate the importance of considering more environmentally relevant conditions (i.e., low maternal food level, short-term exposure) when linking effects observed in laboratory tests to potential risks in the field.

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Introduction Food supply is one of the key factors governing zooplankton dynamics (Murdoch & McCauley 1985, McCauley & Murdoch, 1987). Low food conditions periodically occur in the aquatic environment (Buikema et al. 1980, McCauley et al. 1990) and may alter the sensitivity of Daphnia to pollutants (Biesinger & Christensen 1972, Koivisto et al. 1992, Chandini 1989, Antunes et al. 2003). Despite these documented changes in susceptibility, low food conditions are only scarcely incorporated in laboratory ecotoxicity tests (Snell & Moffat 1992).

Daphnia maintains a high reproductive rate when food is abundant (Stearns 1976), but low food conditions induce a shift in reproductive strategy from quantity to quality (Tessier & Consolatti 1991). This adaptive response is thought to ensure the future survival of the population in a highly variable environment. The enhanced maternal investment in offspring can be characterized by increases in lipid content (Tessier et al. 1983), or egg and neonate size (Glazier 1992, Gliwicz & Guisande 1992). Since the successive offspring performance is thereby altered, several studies have focused on the optimization of culture conditions in order to supply uniform daphnids for standard laboratory ecotoxicity tests (Cox et al. 1992, Naylor et al. 1992a, Martinez-Jeronimo et al. 2000). Although low postnatal food conditions may increase the vulnerability of Daphnia to toxicants (Heugens et al. 2001, Pieters et al. 2005), low maternal food supply appears to decrease the sensitivity of offspring to stresses such as starvation (Lynch & Ennis 1983, Cowgill et al. 1984, Gliwicz & Guisande 1992), and toxicants. Indeed, one to three times decreased acute sensitivities to toxicants have been reported at low maternal food supply (Baird et al. 1989, Enserink et al. 1990). Other studies on the contrary have observed rather indecisive results regarding acute offspring sensitivity (Naylor et al. 1992b, Vigano 1993). Variations in maternal nutrition have nonetheless been suggested as one explanation for the poor interlaboratory agreement often seen in ring tests (Cowgill 1987, Baird et al. 1989).

Chronic toxicity in combination with differing maternal food levels has gained far less attention. To our knowledge, only Enserink et al. (1993) investigated the chronic toxicity of metal exposure to Daphnia, but could not determine a clear relationship between maternal food level and the sensitivity of its progeny. Furthermore, most standard toxicity tests do not incorporate short-term toxicant exposures followed by post-exposure observation periods (Naddy et al. 2000). This is especially relevant for pesticides, since their application in

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51

agriculture results in short-term exposures due to spray drift, drain flow, or surface runoff (Baughman et al. 1989, Liess et al. 1999). Test designs with continuous pesticide exposure may therefore lack environmental realism because they do not include latent effects nor allow recovery after the cessation of the exposure.

Hence, there is an urgent need to examine the combined effects of short-term toxicant exposure and varying maternal food levels on chronic toxicity to Daphnia. We therefore investigated the influence of maternal nutritional state of Daphnia magna Straus on its offspring size and successive performance, and its acute and chronic sensitivity to a 24 h exposure to the pyrethroid insecticide Fenvalerate (FV). To this purpose, 21-day life-table response experiments were performed using neonates originating from cultures reared at either high or low food levels. The offspring were subsequently maintained under low food conditions after the initial 24 h FV exposure. Responses of survival, growth, and reproductive traits were examined. Effects on these life-history traits were ultimately integrated and evaluated in the fitness parameter population growth rate (r).

Pesticides are a relevant stressor in freshwater systems (Liess & Ohe 2005). Fenvalerate is widely used for the protection of agricultural crops due to its high efficacy and generally low mammalian toxicity. Fenvalerate has regularly been detected in the aquatic environment (Chandler 1990), whereas short-term peak concentrations of 0.8 to 6.2 µg L-1 were measured in agricultural streams (Liess & Schulz 1999, Liess et al. 1999, Scott et al. 1999). A high sensitivity to FV has been found for non-target organisms such as aquatic macroinvertebrates (Anderson 1982). Also long-term effects in the structures of macroinvertebrate communities after a FV pulse exposure have been observed in an outdoor mesocosm study (Woin 1998), and in the field (Liess & Schulz 1999). Materials & Methods Daphnia magna culture

The experiments were performed with individuals of the cladoceran Daphnia magna Straus (clone B, Bayer, Monheim, Germany). Continuous cultures of 20 adults in 2 L artificial Elendt M7 medium (OECD 1998) were maintained at 20°C ± 1°C, normal light (~1400 Lux) and a 16:8 light:dark photoperiod. The pH was ~7.4, conductivity ~600 µS cm-1 and the dissolved oxygen level was ~7.1 mg

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L-1. Once a week a new culture was initiated using < 24 h old neonates from a three week old culture. The newly released neonates were removed twice a week, and the medium was renewed three times a week. Feeding of the animals occurred three times a week by means of a suspension of batch-cultured green algae (Desmodesmus subspicatus). Algae were cultured in algae medium according to Grimme & Boardmann (1972), continuously aerated (3% CO2), and harvested in exponential growth phase. Quantification was performed with a total organic carbon analyzer (Shimatzu TOC-5000 Analyzer, Duisburg, Germany) and a coulter counter (Casy1 counter, Scharfe Systems, Reutlingen, Germany).

Two life-table experiments were simultaneously performed using offspring from two maternal cohorts cultured at two different food levels: ‘high’ and ‘low’. Animals maintained at the high food level received equivalent daily rations of 0.75 mg carbon L-1 (5.3 × 105 cells mL-1; 0.07 mg carbon daphnid-1 day-1), whereas animals cultured at the low food level received equivalent daily rations of 0.21 mg carbon L-1 (1.5 × 105 cells mL-1; 0.02 mg carbon daphnid-1 day-1). Daphnids were reared for two generations at these food levels to insure adaptation to these food regimes. Chronic tests

An adjusted 21-day Daphnia reproduction test (OECD 1998) was adopted in this study. The two life-table response experiments were carried out under low food conditions similar to the low food daphnid culture (0.21 mg carbon L-1; 1.5 × 105 cells mL-1).

The life-table response experiments were initiated with newly released neonates (< 24 h) from three week old cultures. Twenty neonates per FV exposure treatment were put individually into glass beakers each containing 80 mL of test medium. The daphnids were exposed to FV for 24 h and received the low feeding regime during the exposure. The neonates were rinsed after 24 h and gently transferred into uncontaminated M7 medium. Further feeding and renewing of the media was done three times a week. Newly released offspring were removed daily. Other experimental conditions were the same to those described for the culture. On a daily basis the following life-history traits of the daphnids were recorded: survival (defined as swimming and immobilized animals) and the number of living neonates produced. From these observed life-history characteristics, age at first reproduction, mean brood number, mean brood size and cumulative reproduction per living female at day 21 were determined. The

Maternal nutritional state

53

population growth rate (r) was calculated from the integration of the age-specific data on survival and fecundity probabilities. The r values were calculated iteratively from the Euler/Lotka equation (1), (Lotka 1913, Euler 1970):

∑Ω

=

−=0X

rxxx eml1 (1)

Where r = per capita rate of increase for the population (d-1), x = age class (days; 1, 2, 3…Ω), Ω = oldest age class in the population (21-days in the present study), lx = probability of surviving at age x, and mx = neonates per mother at age x.

Because this calculation involves a summation over several age classes, r cannot be isolated on one side of the equation to provide a closed-form, algebraic solution. Instead, interactive calculations must be performed in order to determine an r value that satisfies Equation 1. Variance associated with r values for statistical comparison was estimated using the Jackknife-method as described by Meyer et al. (1986).

To determine somatic growth of the daphnids, body lengths (distance from the middle of the eye to the base of the tail) were measured at day 7 and 21. Measurements were performed using a Leica MS 5 Microscope equipped with a Leica DFC 300 F Digital Camera, a Leica KL 1500 LCD light source (Leica Microsystems, Solms, Germany), using the computer program Openlab 3.1.2 for Macintosh (Improvision, Heidelberg, Germany). Fenvalerate exposure & measurement

Fenvalerate (i.e. (RS)-α-cyano-3-phenoxybenzyl (RS)-2-(4-chlorophenyl)-3-methylbutyrate, CAS: 51630-58-1, technical grade (HPLC) = 99.9%), was obtained from Riedel-de Haën, (Seelze, Germany) in powder form. Dimethylsulfoxide (DMSO, technical grade (HPLC) = 99.8%) was used as carrier solvent and obtained from Merck (Darmstadt, Germany). The maximum amount of DMSO present in the test medium was 0.0003% (volume/volume). Eight different nominal concentrations were tested (control, 0.03, 0.1, 0.3, 0.6, 1.0, 3.2 and 10.0 µg L-1). Actual exposure concentrations were determined in triplicate at t = 1 h before food application. Accurate detection was limited to the concentrations 0.6, 1.0, 3.2 and 10.0 µg L-1. The 1.0 µg L-1 test solution (n = 3) was measured additionally at t = 24 h. Samples were measured by solid-phase extraction of 100 mL volumes with C18 columns (Baker, Phillipsburg, NJ, USA). The measurements were performed with gas chromatography–electron capture

Chapter 3

54

detection (gas chromatograph HP 5990, Series II, Hewlett-Packard, Avondale, PA, USA) and confirmed with gas chromatography–mass spectrometry (negative chemical ionization, Varian 3400 gas chromatograph, Varian, Walnut Creek, CA, USA) with an HP 7673 autosampler (Hewlett-Packard, Avondale, PA, USA) that was directly capillary-coupled to the quadruple mass spectrometer SSQ 700 (Finnigan, Bremen, Germany). Analysis of the 0.6, 1.0, 3.2 and 10.0 µg L-1 test solutions at t = 1 h showed actual concentrations (± S.D.) ranging from 33% to 46% of the nominal concentrations (0.20 ± 0.00, 0.43 ± 0.12, 1.47 ± 0.28 and 4.07 ± 0.31 µg L-1, respectively). The actual concentration of the 1.0 µg L-1 test solution further decreased to 0.37 ± 0.06 µg L-1 after 24 h. Nominal concentrations are given in the following sections. Statistical analysis

Data for all parameters were tested for normality (Kolmogorov-Smirnov test) and homoscedasticity (Levene’s test). When these requirements were not met, data were log(x+1) transformed. Controls for all endpoints of both experiments were compared with each other using two-tailed t-tests. Exposure treatments within one maternal food level experiment were compared with the corresponding control using one-way analysis of variance (ANOVA), followed by Dunnett’s post-hoc test. In order to assess the contribution of FV concentration and maternal food ration to the observed variation, a two-way ANOVA was performed. For survival as a function of time, significant differences between FV treatments and the corresponding control were analyzed by conducting pairwise comparisons using Gehan-Wilcoxon survival analysis. Data were only graphically shown for FV concentrations up to 1.0 µg L-1 because above this concentration low survival hampered statistics. All tests were run at a significant probability level of p < 0.05. All analyses were conducted with Statistica 6.1 (StatSoft, Tulsa, USA).

Maternal nutritional state

55

Results Survival

No mortality occurred in the controls during the experiments, but survival decreased strongly with increasing FV concentrations in both experiments (Figure 1). Mortality generally occurred between days 2 and 7 for both experiments after the 24 h Fenvalerate exposure. However, the Lowest Observable Adverse Effect Concentration (LOAEC) differed between both maternal food treatments. The lowest FV concentration that caused a significant (p < 0.05) lower survival compared to the control was 0.6 µg L-1 for the high maternal food treatment, whereas this significance for the low maternal food treatment was at 3.2 µg L-1. Figure 1. Survival of Daphnia magna expressed as a function of time, exposed for 24 h to fenvalerates. Asterisks denote significant differences compared to the corresponding control (Gehan-Wilcoxon survival analysis, p < 0.05).Mind the superposition of the 0.03-0.1 µg L-1 treatments above the controls. Body length

Two exemplary cohorts of newly released neonates from the third week culture were measured directly before the initiation of the experiments (n = 40). The body length of the neonates from the high maternal food level was significantly (p < 0.05) lower (0.75 mm ± 0.05 S.D.) than the body length of the neonates from the low maternal food level (0.94 mm ± 0.06 S.D.). This difference in size lasted during the course of the experiment, since the control body lengths at day 7 and day 21 of the high maternal food treatment (2.60 mm ± 0.06 C.I. and 3.11 mm ± 0.03 C.I., respectively) were significantly (p < 0.05) smaller than the

low maternal food

days after exposure0 7 14 21

0

20

40

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control0.03 µg L-1

0.1 µg L-1

0.3 µg L-1

0.6 µg L-1

1.0 µg L-1

3.2 µg L-1

10.0 µg L-1

**

surv

ival

(% in

itial

anim

als)

*

*

*

high maternal food

days after exposure0 7 14 21

0

20

40

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100low maternal food

days after exposure0 7 14 21

0

20

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100

control0.03 µg L-1

0.1 µg L-1

0.3 µg L-1

0.6 µg L-1

1.0 µg L-1

3.2 µg L-1

10.0 µg L-1

**

surv

ival

(% in

itial

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*

*

*

high maternal food

days after exposure0 7 14 21

0

20

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Chapter 3

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control body lengths of the low maternal food treatment (3.02 mm ± 0.06 C.I. and 3.53 mm ± 0.03 C.I., respectively). Body length at day 7 decreased strongly with increasing FV concentrations (Figure 2) and showed significant (p < 0.05) reductions compared to the corresponding controls at similar concentrations (0.3-1.0 µg L-1) for both maternal food treatments. Despite the identical LOAEC values, percent reductions in body lengths for the high maternal food treatment were stronger between 0.3-1.0 µg L-1 compared to the low maternal food treatment. This increase in sensitivity was confirmed by two-way ANOVA which demonstrated that the interaction between FV and maternal food ration at day 7 was significant (Table 1). Recovery in body length, however, occurred from day 7 onwards and at day 21 the body length was only significantly (p < 0.05) smaller compared to the corresponding control at 1.0 µg L-1 for the low maternal food treatment. As a result, two-way ANOVA showed no significant interaction between FV and maternal food level for body length at day 21 (Table 1). Figure 2. Effects of a 24 h exposure to fenvalerate on body length (shown as percentage of mean control) of Daphnia magna at day 7 and day 21 for the low maternal food treatment () and the high maternal food treatment (). Error bars indicate ± 95% confidence intervals and asterisks denote significant differences compared to the corresponding controls according one-way ANOVA (Dunnett’s post hoc test, p < 0.05). Reproductive endpoints

The controls of the low maternal food treatment showed compared to the controls of high maternal food treatment a significantly (p < 0.05) decreased age at first reproduction (8.42 days ± 0.21 C.I. and 9.21 days ± 0.32 C.I., respectively), and a significantly (p < 0.05) higher cumulative reproduction (53.90 ± 1.88 C.I. and 23.70 ± 0.96 C.I., respectively). Exposure to FV (Figures 3AB)

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fenvalerate concentration (µg L-1)0.03 0.1 0.3 0.6 1.0control

Fenvalerate concentration (µg L-1)0.03 0.1 0.3 0.6 1.0control

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day 7 day 2150

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**

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day 7 day 21

Maternal nutritional state

57

caused in the low maternal food treatment a significant (p < 0.05) increase in age at first reproduction (0.6-1.0 µg L-1) and a significant (p < 0.05) reduction in the cumulative reproduction (1.0 µg L-1), while the cumulative reproduction demonstrated also significant increases (p < 0.05) at 0.1-0.3 µg L-1. Fenvalerate exposure in the high maternal food treatment caused a significant (p < 0.05) increase in age at first reproduction (0.3-1.0 µg L-1) and a reduction in cumulative reproduction (0.6-1.0 µg L-1).

Figure 3. Effects of a 24 h exposure to fenvalerate on the age at first reproduction (A) and cumulative reproduction at day 21 (B) of Daphnia magna for the low maternal food treatment () and the high maternal food treatment (). Error bars indicate ± 95% confidence intervals and asterisks denote significant differences compared to the corresponding controls according one-way ANOVA (Dunnett’s post hoc test, p < 0.05). Mind the different scales on the y-axis. Two-way ANOVA showed that the interaction between FV and maternal food level was significant for both reproductive endpoints (Table 1), which indicated that maternal food level influenced offspring responses to FV exposure. In particular, age at first reproduction of the high maternal food treatment showed an increased sensitivity reflected by the lower LOAEC values and stronger inhibitions. This increased sensitivity was also demonstrated at intermediate FV concentrations (0.1-0.6 µg L-1) for cumulative reproduction. The opposite was observed however at 1.0 µg L-1 where stronger reductions occurred in the low maternal food treatment.

AFR

(% m

ean

cont

rol)

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fenvalerate concentration (µg L-1)0.03 0.1 0.3 0.6 1.0control

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(% m

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fenvalerate concentration (µg L-1)0.03 0.1 0.3 0.6 1.0control

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fenvalerate concentration (µg L-1)0.03 0.1 0.3 0.6 1.0control

RPR

(% m

ean

cont

rol)

**

*

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**

*

*

*

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(% m

ean

cont

rol)

90

100

110

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fenvalerate concentration (µg L-1)0.03 0.1 0.3 0.6 1.0control

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fenvalerate concentration (µg L-1)0.03 0.1 0.3 0.6 1.0control

RPR

(% m

ean

cont

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**

*

**

**

*

*

*

A B

Chapter 3

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Table 1. Summary table of the two-way analyses of variance applied to the two life-table response experiments with Daphnia magna (FV - fenvalerate; d.f. - degrees of freedom; p - probability).

parameter source of variation F d.f. p

body length day 7 food FV food × FV

4.48 68.98 5.44

1 5 5

< 0.001 0.029 < 0.001

body length day 21 food FV food × FV

4.62 3.74 0.55

1 5 5

0.003 0.033 0.742

age at first reproduction food FV food × FV

18.87 49.78 3.16

1 5 5

< 0.001 < 0.001 0.009

cumulative reproduction day 21

food FV food × FV

14.39 26.48 11.12

1 5 5

< 0.001 < 0.001 < 0.001

population growth rate food FV food × FV

41.44 153.34 8.45

1 5 5

< 0.001 < 0.001 < 0.001

Population growth rate (r)

The control r value at day 21 of the high maternal food treatment (0.23 ± 0.01 C.I.) was significantly (p < 0.05) lower than the control of the low maternal food treatment (0.32 ± 0.01 C.I.). Exposure to FV strongly affected r calculated for the 21 days post treatment period. At day 21, significantly (p < 0.05) lower r values compared to corresponding controls were demonstrated at 0.3-1.0 µg L-1 for both maternal food treatments (Figure 4). Although LOAEC values were again equal, percent reductions in r for the high maternal food treatment was considerably stronger between 0.3-1.0 µg L-1 compared to the low maternal food treatment. This enhanced sensitivity was confirmed by two-way ANOVA which showed that the interaction between FV concentration and maternal food level was significant (Table 1).

Maternal nutritional state

59

Figure 4. Effects of a 24 h exposure to fenvalerate on the population growth rate (r) of Daphnia magna during a subsequent 21 d period for the low maternal food treatment () and the high maternal food treatment () Results are shown as percentage of mean control. Error bars indicate ± 95% confidence intervals and asterisks denote significant differences compared to the corresponding controls according one-way ANOVA (Dunnett’s post hoc test, p < 0.05). Discussion Offspring size & performance

The present study was carried out to investigate whether low maternal food conditions induce a change in the reproductive strategy of Daphnia magna, i.e., from quantity to quality, and therefore cause modifications in the acute and chronic sensitivity of its offspring to a FV pulse exposure.

Our results clearly demonstrated that mothers reared at low food levels produced larger offspring. This shift in reproductive strategy supports the observations of others (Cox et al. 1992, Glazier 1992, Guisande & Gliwicz 1992). The difference in offspring size consequently influenced the cumulative reproduction at day 21. Larger body lengths usually cause decreases in age at first reproduction (Porter et al. 1983, Tillmann & Lampert 1984, Taylor & Aiken 1985) and larger brood sizes (Lampert 1993), which therefore increase the cumulative reproduction and brood number. These allometric relationships between growth and reproductive traits support our results and have collectively

fenvalerate concentration (µg L-1)0.03 0.1 0.3 0.6 1.0

r21(%

mea

n co

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l)

40

50

60

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80

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100

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*control

*

*

**

*

fenvalerate concentration (µg L-1)0.03 0.1 0.3 0.6 1.0

r21(%

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*

*

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attributed to the enhancement in offspring fitness of the low maternal food treatment as estimated by the population growth rate. Acute toxicity

The increased neonatal body length in the low maternal food treatment resulted in an approximately five times reduced acute sensitivity to FV. This response pattern is in agreement with other studies showing lower acute sensitivity of Daphnia magna offspring produced by mothers experiencing food limitation (Baird et al. 1989, Enserink et al. 1990). The decreased acute toxicity to FV may be explained in several ways. One explanation may be that the lower surface-to-volume ratio of the neonates originating from the low fed mothers may have caused a decreased bioaccumulation of the lipophilic FV and consequently a diminished response to the toxicant. Another explanation may be that the larger neonates produced by low fed mothers attained higher lipid stores which enabled the offspring to allocate more energy reserves to detoxification mechanisms. Sublethal FV concentrations are also known to cause immobilization through interference with the nerve system (Clark & Brooks 1989), causing reduced filtration and algal assimilation rates in Daphnia (Day & Kaushik 1987). Hence, the toxicant may have acted as a pseudo food stressor itself and larger neonates with higher energy reserves may have been able to cope better with this reduced food uptake. Chronic toxicity

Also for chronic endpoints the maternal food level strongly affected the responses of Daphnia magna to FV. Growth in the first week was impaired, but less strong for the neonates of the low maternal food treatment. Delays in development after a short-term FV exposure have already been observed (Liess & Schulz 1996), and are likely the direct result of feeding impairment (Day & Kaushik 1987). The daphnids were nevertheless able to recover from the short-term FV exposure and showed body lengths similar or close to controls at day 21. The FV induced delay in development caused an increased age at first reproduction with a lower sensitivity to FV in the low food treatment. A similar decreased sensitivity at low maternal food levels was demonstrated for cumulative reproduction at intermediate FV concentrations, but the opposite was observed at 1.0 µg L-1. Although we do not have a clear explanation for this, unexpected modifications in trade-off mechanisms have been observed earlier under low food

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conditions (Enserink et al. 1995) and may have influenced the susceptibility of Daphnia to FV in this range of concentration.

It is concluded that the stronger effects of FV on the population growth rate in the high maternal food treatment were caused by the stronger effects of FV on reproductive endpoints (regardless whether positive or negative responses) and by the considerably higher mortality observed during the first week compared to the low maternal food treatment. Ecological implications & conclusions

Planktonic cladocerans such as Daphnia have to respond to highly variable food levels in the field. Depending on the nutritional state of the mothers, investments in progeny are allocated in such a way that the highest future survival (i.e., fitness) of the population is insured. Maternal trade-offs between the size and number of offspring has been previously observed for Daphnia depending on age (Boersma 1997a), phenotype (Sakwinska 2004), social environment (Cleuvers et al. 1997), and resource quantity (Ebert 1993, Boersma 1997b) and quality (Brett 1993). Information on the long-term persistence of maternal effects on Daphnia is on the contrary extremely rare. The decrease in vulnerability to FV of offspring originating from low fed mothers corresponds well with a study on severly food limited Daphnia pulex populations exposed to a single dose of Carbaryl (Hanazato & Hirokawa 2004). Moreover, the stronger effects of FV on r of individuals originating from high fed mothers may have consequences for field populations, since reductions greater than 30% have been predicted to cause the extinction of rotifer populations on the long term (Snell & Serra 2000). However, relating maternal effects on individual life-history traits observed in laboratory toxicity studies to potential population effects in the field is notoriously problematic, especially since experimental conditions may not exactly mimic the field situation. The introduction of pesticides via spray drift or runoff from agricultural fields into rivers often lead to short-term exposure durations much shorter than the 24 h used in this study (Williams et al. 1995), whereas limiting food conditions in the field may reach concentrations as low as 0.05 mg carbon L-

1 (Murdoch et al. 1998). Nevertheless, the present study has shown how effects of maternal food level may persist in their offspring affecting the acute and chronic sensitivity after a short-term pesticide exposure.

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Acknowledgements The authors thank M.H.S. Kraak and W. Admiraal for valuable discussions and comments on

the manuscript. References Anderson, R.L. 1982. Toxicity of fenvalerate and permethrin to several non-target

aquatic invertebrates. Environ. Entomol. 11: 1251-1257. Antunes, S.C., B.B. Castro, and F. Goncalves. 2003. Effect of food level on the

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Baird, D.J., I. Barber, M. Bradley, P. Calow, and A.M.V.M. Soares. 1989. The Daphnia bioassay: a critique. Hydrobiologia 188/189: 403-406.

Baughman, D.S., D.W. Moore, and G.I. Scott. 1989. A comparison and evaluation of field and laboratory toxicity tests with fenvalerate on an estuarine crustacean. Environ. Toxicol. Chem. 8: 417-429.

Biesinger, K.E., and G.M. Christensen. 1972. Effects of various metals on survival, growth, reproduction and metabolism of Daphnia magna. J. Fish. Res. 29: 1691-1700.

Boersma, M. 1997a. Offspring size and parental fitness in Daphnia magna. Evol. Ecol. 11: 439-450.

Boersma, M. 1997b. Offspring size in Daphnia: does it pay to be overweight? Hydrobiologia 360: 79-88.

Brett, M.T. 1993. Resource quality effects in Daphnia longispina offspring fitness. J. Plankton Res. 15: 403-412.

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67

Chapter 4

Modeling responses of Daphnia magna to pesticide pulse exposure under varying food conditions: intrinsic versus

apparent sensitivity

B.J. Pieters, T. Jager, M.H.S. Kraak, and W. Admiraal. Ecotoxicology (2006) 15: 601-608.

Abstract Recent studies showed that limiting food conditions resulted in either increased or decreased sensitivity of Daphnia magna to toxicants. It remained unclear whether these contrasting food-dependent alterations in toxicity resulted from differences in intrinsic sensitivity of the daphnids or from changes in toxicokinetics and resource allocation. It is hypothesized here that, if food level only affects accumulation kinetics and resource allocation, then the intrinsic sensitivity to this toxicant should be the same for all food regimes. This hypothesis was investigated using the DEBtox model, which is based on the theory of Dynamic Energy Budgets. We examined results of two recently conducted life-cycle studies on the combined effects of food level and a pulsed exposure to the pyrethroid insecticide fenvalerate (FV) on Daphnia magna. The model described the effects of the time-varying exposure well, and indicated that when the animals did not die from exposure to FV, full reversibility of toxic effects was possible, allowing a complete recovery. Results revealed furthermore that the data from both studies could be described by the same NECs for survival and assimilation, killing rate and tolerance concentration (132 (49.2-228) × 10-6 µg L-1, 0 (0-1.18 × 10-5) µg L-1, 74.4 (55.6-96.4) L (µg d)-1 and 5.39 (2.72-18.5) × 10-3 µg L-1, respectively). It is therefore concluded that food-dependent FV toxicity can be explained by altered toxicokinetics and resource allocation, but not by changes in the intrinsic sensitivity of the daphnids. This study implies that the effect of pesticide application in the field depends on the trophic state of the receiving water body, but also that full recovery of survivors is possible after FV application.

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68

Introduction Food shortage affects the physiological condition of organisms and may therefore increase their sensitivity to toxicants (Heugens et al. 2001). Many studies have indeed reported an increased sensitivity to a wide range of pollutants of invertebrates grown under limiting food conditions (Chandini 1989, Enserink et al. 1995, Barry et al. 1995, Klüttgen et al. 1996, Rose et al. 2002). However, in the common test species Daphnia, low food supply causes a reproductive trade-off resulting in fewer, but larger sized offspring (Glazier 1992, Gliwicz & Guisande 1992), which were found to have a lower sensitivity to toxicants (Baird et al. 1989, Enserink et al. 1990). Thus limiting (maternal) food conditions resulted in both increased and decreased sensitivity to toxicants. In all these studies however, increased or decreased sensitivity was assessed as a change in a crude summary statistic such as the LC50 or EC50 at a certain time point. However, a change in EC50 does not necessarily reflect an effect on the ‘intrinsic sensitivity’ of the test organism (In this paper, we define intrinsic sensitivity as the relationship between the concentration at the target receptor (e.g., tissue concentrations) and the physiological process that is affected (e.g., assimilation of energy from food)). At the same intrinsic sensitivity, apparent sensitivity to organic chemicals may differ with food level, because of 1) differences in bioavailability, causing the chemical to reach the target to a lesser extent (Barry et al. 1995), 2) alterations in toxicokinetics, because food affects growth and thereby the uptake and elimination kinetics (Day & Kaushik 1987a), and 3) physiological interactions. This last factor requires some elaboration. At low food levels, maintenance takes up a larger proportion of the total energy budget, because unlike growth and reproduction, maintenance costs cannot be simply decreased by the organism. As a result, the same percentage change in feeding rate or maintenance costs due to a toxicant will lead to a larger effect on growth and reproduction at low food levels (Kooijman & Metz 1984). Few studies have actually tested these explanations and the aim of the present study was therefore to analyze whether the contrasting results for the effect of food on sensitivity of daphnids to toxicants were actually the result of a changed intrinsic sensitivity of the test organisms. For this purpose, we examined results of two recently conducted life-cycle studies on Daphnia magna testing the combined effects of food level and a pulsed exposure to the pyrethroid insecticide fenvalerate (FV) (Pieters et al. 2005, Pieters & Liess 2006). Episodic exposure to FV was chosen

Modeling responses of Daphnia magna

69

because a single toxicant dose is especially relevant in agriculture, where pesticide application may result in short-term pulse exposures of aquatic biota due to spray drift, drain flow, or field runoff (Liess et al. 1999). In spite of its relevance, episodic pollution received relatively little attention (McCahon et al. 1991, McCahon & Pascoe 1991, Handy 1994, Naddy et al. 2000, Reinert et al. 2002). Results of our first study (Pieters et al. 2005) showed that limiting postnatal food conditions exacerbated the effects of a 24 h FV pulse exposure. In contrast, Pieters & Liess (2006) demonstrated that daphnids originating from low-fed mothers were less sensitive to the pulsed FV exposure compared to daphnids born from high-fed mothers (Pieters & Liess 2006). It is hypothesized here that food level only affects accumulation kinetics of FV and resource allocation, but that the intrinsic sensitivity to FV was the same for all food regimes. This hypothesis was investigated using the DEBtox model (see Kooijman & Bedaux 1996), which is based on the theory of Dynamic Energy Budgets (DEB, see Kooijman 2000). Because the model explicitly considers toxicokinetics (and the effects of body size on toxicokinetics) and resource allocation, it was possible to incorporate time-varying FV concentrations and to distinguish between effects of food level on the uptake and resource allocation on the one hand, and changes in intrinsic sensitivity of the organism to FV on the other. The applicability of DEBtox to time-varying concentrations was shown earlier by Péry et al. (2001) for effects on survival only. Materials & methods Experimental data set

Life-cycle toxicity data from two previous laboratory studies with Daphnia magna (Pieters et al. 2005, Pieters & Liess 2006) were used for this study. In short, 21-day Daphnia reproduction tests (OECD 1998) were conducted in which newly released neonates (< 24 h old) from a laboratory culture were individually transferred into glass beakers containing 80 mL of test medium and pulse exposed for 24 h to several concentrations of FV (control, 0.03, 0.1, 0.3, 0.6, 1.0, 3.2 and 10.0 µg L-1). After the exposure, offspring were transferred to uncontaminated medium and life-history traits (survival, growth, reproduction) were subsequently monitored during 21 days. Survival and the number of living neonates produced were recorded on a daily basis, and the body lengths of neonates were measured at birth and on day 7, 14 and 21. Daphnids were fed with a suspension of the

Chapter 4

70

green algae Desmodesmus subspicatus. The first study was initiated with neonates originating from maternal cohorts cultured at a high-food level (equivalent daily ration: 5.3 × 105 cells mL-1; 0.07 mg carbon daphnid-1 day-1). Next, neonates were exposed to FV at an identical high food regime in the ‘high-food’ treatment (n = 15), while neonates in the ‘low-food’ treatment (n = 20) received a low-food level (1.5 × 105 cells mL-1; 0.02 mg carbon daphnid-1 day-1). The second study was initiated with neonates (n = 25) originating from maternal cohorts fed with the respective low-food regime (i.e., ‘maternal low-food’ treatment) and were also exposed to FV at this low-food level. The experimental data set was subjected to the DEBtox model to evaluate whether potential changes in the intrinsic sensitivity of the daphnids and/or changes in FV toxicokinetics determined the influence of food level on FV toxicity. A full description of the model is given in (Kooijman & Bedaux 1996), but here we will address the relevant assumptions, equations and extensions. For variables see also Table 1. Toxicokinetics

Conform the concept of critical body residues (see McCarthy & Mackay 1993), DEBtox describes that a toxicant is taken up from the environment by an organism before it can elicit an effect. The internal concentration determines the organism’s probability to die if this concentration exceeds a certain threshold level. In the present study the internal concentration of FV was deduced from its concentration in the solution (i.e., external concentration) according to a one-compartment model for kinetics (see Heugens et al. 2003). Since measured internal concentrations were not available, scaled internal concentrations cv (i.e., internal concentrations divided by the bioconcentration factor; BCF) were used. This implies that cv is proportional to the true internal concentration, but with the dimensions of an external concentration (cd). The body residues are expected to follow one-compartment kinetics, in which the effects of growth are accounted for:

( ) 3ln ldtdccc

lk

cdtd

VVde

V −−= (1)

This equation has only two unknowns, the elimination rate (ke) and the scaled body length (l), i.e., length as a fraction of the maximum length at abundant food, which occurs twice in this equation. First it is used in the term ke/l, reflecting that the rate of uptake and elimination depends on the surface-to-volume ratio of the

Modeling responses of Daphnia magna

71

organism (a large organism takes more time to reach steady state than a small one). The second term in this equation represents dilution by growth.

The DEBtox software protocol was modified to introduce a time-varying external concentration in Equation 1. Measurements of actual concentrations were available at the start of the pulse exposure, and for the high food treatment after 1 and 24 h (see Pieters et al. 2005). The decrease of actual FV concentration over time was assumed to depend on food level (i.e., algal concentration) due to adsorption of the hydrophobic FV to the algae. Therefore, the declining actual FV concentrations of the low-food experiment were simulated to accommodate sorption to be proportional to the algal density. Effects on survival

In DEBtox, the survival probability of individuals is calculated via the hazard rate (hc). As long as the internal concentration is below a no-effect concentration (NEC), the hazard rate due to toxicant stress is zero; above the NEC, the hazard increases linearly with the concentration above the NEC. Note that the internal concentration and hence the NEC (c0s) are both scaled in the dimensions of an external concentration. The hazard rate is given by:

( )+−= sVc ccbh 0 (2)

The + indicates that the maximum of the expression between brackets and zero is taken. The proportionality constant b is called the killing rate. It is a measure of the effect of the chemical on survival, when the NEC is exceeded. The hazard rate due to toxicant stress is added to a constant background mortality h0 and integrated to yield the survival probability q, for any concentration and any time point:

( )( ) ⎥⎦

⎤⎢⎣

⎡+−= ∫

t

c dhhtq0

0exp)( ττ (3)

Effects on growth & reproduction Substances such as FV cause a loss of coordination and immobilization, and

therefore result in a reduction of filtration rates. Therefore, the physiological mode of action that is most suited to this case is effects on assimilation. The DEBtox equations below were designed to express this mode of action; a stress factor s is defined depending on the scaled internal concentration cv. Equivalent to the hazard rate, the stress factor is zero below the NEC (c0a), and above the NEC it is proportional to the internal concentration. The proportionally constant cT has

Chapter 4

72

the dimensions of an external concentration, and is therefore called the tolerance concentration for assimilation:

( )T

aV

ccc

s +−= 0 (4)

When no toxicant is present and as long food density (f) remains constant, growth of the isomorphically growing daphnids reduces in DEB to the Von Bertalanffy growth curve (see Kooijman 2000). A scaled body length is adopted to simplify the equations. Under abundant food levels, f equals 1, but this value becomes smaller than 1 under limiting food levels. This way, food limitation was included in the model. When the daphnids are exposed to a toxicant affecting assimilation, the scaled body length is given as a function of time and food level by:

[ ]lsfrldtd

B −−= )1( (5)

where rB is the Von Bertalanffy growth rate constant. Both body length and food density control reproduction. The reproduction rate (R) is given as a satiating function of body length (or scaled length l). As long as the scaled body length is below the scaled length at puberty (lp), the reproduction rate is zero. Above the length at puberty, the reproduction rate increases with body size, until it reaches the maximum rate (Rm) when f = 1. For effects on assimilation, the full equation for the reproduction rate at a constant food level is:

( ) ⎥⎦

⎤⎢⎣

⎡−

++

−−= 32

33

11 p

p

m lflfglg

lR

sR (6)

The dimensionless constant g is the energy investment ratio; the ratio of the volume-specific costs for growth and the fraction of the reserves allocated to growth and maintenance. This parameter generally has little influence on the toxicity parameters, and 1 is a reasonable value for Daphnia (see Kooijman & Bedaux 1996). Fitting procedure

The DEBtox model was fitted to the data for all endpoints of the high- and low-food experiments simultaneously (see Jager et al. 2004). This was done by expressing all model fits in terms of likelihood and combining the separate likelihoods into an overall value that can be maximized. The 95% confidence

Modeling responses of Daphnia magna

73

intervals were generated using the profile likelihood (see Meeker & Escobar 1995). The full analysis was programmed in MatLab® Version 7.0 (Mathworks, MA, USA). However, the emerging growth and reproduction parameters for the daphnids could not be incorporated in model runs of the maternal low-food data set. This was likely the result from the fact that low-fed mothers enhanced their maternal investment per offspring, characterized by producing fewer but larger sized neonates with a different life-history. Therefore a different set of values for the growth and reproduction parameters was required for the experiment on maternal low food. For this purpose, the maternal low-food data set was analyzed separately, using the toxicological parameters generated by modeling the other two experiments (see Table 1). Results & discussion Figure 1 presents the model fits for D. magna exposed to different combinations of FV and (maternal) food levels. Table 1 shows the corresponding parameter estimates with their likelihood-based 95% confidence intervals. Mode of action of FV

The growth and reproduction data were well described by the model, assuming a hampered assimilation of food as the mode of action for FV. This assumption corresponds fully with available literature on daphnids and pyrethroids (Day & Kaushik 1987b, Day et al. 1987c, McWilliam & Baird 2002, Christensen et al. 2005, Reynaldi et al. 2006). Due to this effect on assimilation, strong effects were observed on growth curves, which were immediately translated to a delayed onset of reproduction, thereby ultimately affecting the reproduction curves (Figure 1). Reversible effects of FV pulse exposure

Good model fits were obtained for the data, indicating that the model was well able to describe the effects of the time-varying exposure regime, and confirming the link between the estimated internal concentrations and observed effects. Some deviations occurred, especially for survival. However, it must be note that death is an all-or-nothing effect, and because it is treated as a stochastic process, a good comparison between model and data requires many individuals. Because only a limited number of daphnids survived, one or two more deaths

Cha

pter

4

74

Tab

le 1

. Par

amet

er e

stim

ates

of t

he h

igh-

food

(HF)

, low

-food

(LF)

and

the

mat

erna

l low

-food

(MLF

) exp

erim

ent,

with

lik

elih

ood-

base

d 95

% c

onfid

ence

inte

rval

s giv

en b

etw

een

brac

kets

.

desc

riptio

n sy

mbo

l un

it H

F an

d LF

sim

ulta

neou

sly

MLF

sepa

rate

ly

phys

iolo

gica

l par

amet

ers

Von

Ber

tala

nffy

gro

wth

rate

r B

d-1

0.

138

(0.1

32-0

.143

) 0.

243

(0.2

38-0

.256

) in

itial

leng

th

L o

mm

0.

753

(n.e

.) 0.

940

(n.e

.) le

ngth

at f

irst r

epro

duct

ion

L p

mm

2.

38 (2

.32-

2.42

) 2.

84 (2

.79-

2.87

) m

axim

um le

ngth

L m

m

m

4.15

(4.1

2-4.

18)

4.17

(4.1

4-4.

20)

max

imum

repr

oduc

tion

rate

R m

ju

v d-1

7.

24 (6

.90-

7.53

) 17

.3 (1

6.3-

18.6

)

scal

ed fo

od d

ensi

ty

f -

HF:

1.0

(n.e

.) LF

: 0.8

18 (0

.809

-0.8

25)

a

toxi

colo

gica

l par

amet

ers

elim

inat

ion

rate

k e

d-1

2.

37 (1

.33-

10.1

) × 1

0-3

a bl

ank

haza

rd ra

te

h 0

d-1

0.66

5 (0

.231

-2.6

6) ×

10-3

a

NEC

for s

urvi

val

c 0s

µg L

-1

132

(49.

2-22

8) ×

10-6

a

NEC

for a

ssim

ilatio

n c 0

a µg

L-1

0

(0-1

.18 ×

10-5

) a

tole

ranc

e co

ncen

tratio

n c T

µg

L-1

5.

39 (2

.72-

18.5

) × 1

0-3

a ki

lling

rate

b

L (µ

g d)-1

74

.4 (5

5.6-

96.4

) a

max

imum

wat

er so

lubi

lity

- µg

L-1

0.

524

(0.4

73-0

.593

) a

Mod

e of

act

ion

is a

dec

reas

e in

ass

imila

tion

due

to fe

nval

erat

e.

a Va

lue

from

the

high

- and

low

-food

exp

erim

ent w

as u

sed.

N.e

. is n

ot e

stim

ated

.

Modeling responses of Daphnia magna

75

Figure 1. Model fits of data on survival, body length and cumulative reproduction of Daphnia magna following a 24 h pulse exposure to a range of fenvalerate concentrations (µg L-1). Graphs show the high- and low-food experiments (fitted simultaneously), and the maternal low-food experiment (fitted separately). already result in a visual misfit of the model curve. Regarding the deviations for growth and reproduction in the higher exposure treatments, it must be note that these were likely resulting from the limited number of surviving individuals at high FV concentrations (the fitting procedure therefore put little weight on these results). A reduced reproduction towards the end of the experiment was observed in all treatments, most likely due to variations in food availability caused by the high adult grazing pressure.

The survival data showed that FV-induced mortality still occurred after cessation of exposure which illustrated the slow toxicokinetics of FV, reflected by a low estimate of the elimination rate ke (Table 1). The daphnids obviously

high food low food maternal low food

time (days) time (days)time (days)

body

leng

th (m

m)

cum

ulat

ive

repr

oduc

tion

frac

tion

surv

ival

0

0.2

0.4

0.6

0.8

1.0

0

1.0

2.0

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accumulated FV to concentrations above the NEC and excreted it so slowly during the post exposure period that mortality continued in the FV-free period. The higher the exposure concentration, the more FV had been accumulated and hence the longer mortality continued. Due to the combined processes of growth dilution and elimination (see Equation 1), internal concentrations finally did not exceed the NEC anymore and mortality ceased.

The model assumption that effects relate only to the actual internal concentration implies that if an organism does not die, effects are reversible and full recovery should be possible. The good match between data and model fits indeed confirmed this assumption, as illustrated by the similar reproduction rates between FV treatments and controls in the reproduction curves (see Figure 1). Similar findings have been reported by Forbes & Cold (2005) who demonstrated that a short-term exposure to the (S)-acid isomer of fenvalerate, S-fenvalerate, during early larval life of Chironomus riparius had effects on larval survival and developmental rates. However, no lasting effects of S-fenvalerate on egg laying or egg viability were observed for surviving individuals. Sensitivity verus toxicokinetics & resource allocation

It was hypothesized that, if food level only affects accumulation kinetics and resource allocation, then the intrinsic sensitivity to FV should be the same for all food regimes. This was obviously the case, since all data could be described with the same NECs for survival and assimilation, killing rate and tolerance concentration (Table 1). Therefore below we will discuss how food level depending changes in accumulation kinetics and resource allocation may have changed FV toxicity to Daphnia magna.

Differing rates of development of the daphnids in the two food regimes with consequent effects on accumulation rates may have contributed to the increased FV toxicity to daphnids in the low-food experiment compared to the high-food experiment. Previous research indicated that the early lifestages of D. magna were more sensitive to a short-term FV exposure than adults (Day & Kaushik 1987a), and suggested that this was the result of their smaller body size. Similarly, in the present study the higher surface-to-volume ratio of the slower growing neonates in the low-food experiment compared to the high-food experiment, may have caused a higher bioaccumulation of FV (see Equation 1). In addition, as daphnids grow during and especially after the 24 h exposure period, the enhanced toxicity of FV to the daphnids in the low-food experiment may also have been caused by

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their slower growth dilution of the pesticide. This mechanism is taken into account in the one-compartment kinetics (Equation 1), and likely decreased internal FV concentrations of the high-fed daphnids much more rapidly over time. Thus, the effect of food level on growth during as well as after pesticide application was likely an important factor modifying FV toxicity.

Differences in resource allocation may also have contributed to the contrasting sensitivities of daphnids between the high- and low-food experiments. The rationale behind this assumption is that FV effects on assimilation act much stronger under low food conditions due to a relatively higher demand of energy for somatic maintenance, thereby ultimately affecting growth and reproduction stronger than under high-food conditions. This phenomenon has been predicted previously (Jager et al. 2004, Alda Álvarez et al. 2005).

The results of the maternal low-food experiment support the explanation on the interaction of feeding, toxicokinetics and resource allocation given above. A decreased apparent toxicity of FV to daphnids in the maternal low-food experiment agreed with findings of other trans-generational studies using Daphnia exposed to 3, 4-dichloroaniline (Baird et al. 1989) and cadmium (Enserink et al. 1990). Neonate body size at time of birth was considerably larger than in the low-food experiment, decreasing the surface-to-volume dependent accumulation, whereas their growth rate (see Table 1) and hence their growth dilution was much faster. Likewise, the study of Enserink et al. (Enserink et al. 1990) also suggested a causal relationship between body size and Cd toxicokinetics. Conclusions We demonstrated that the process-based model DEBtox could accurately describe our experimental data, and it could easily account for differing food densities and time-varying exposure concentrations. We proved that effects related only to the actual internal concentration (exceedance of the NEC) and therefore were reversible and full recovery was possible, albeit after some delay, because of the slow toxicokinetics of FV. It is concluded that food-dependent FV toxicity can be explained by altered toxicokinetics and resource allocation, and not necessarily by changes in the intrinsic sensitivity of the daphnids. This implies that the effect of pesticide application on field populations of daphnids will depend on the trophic state of the receiving water body, but also on the reproductive state and size of the

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animals. Also a full recovery of survivors and resumption of population growth is possible after a FV pulse. References Alda Álvarez, O., T. Jager, S.A.L.M. Kooijman, and J.E. Kammenga. 2005.

Responses to stress of Caenorhabditis elegans populations with different reproductive strategies. Func. Ecol. 19: 656-664.

Baird, D.J., I. Barber, M.Bradley, P. Calow, and A.M.V.M. Soares. 1989. The Daphnia bioassay: a critique. Hydrobiologia 188/189: 403-406.

Barry, M.J., D.C. Logan, J.T. Ahokas, and D.A. Holdway. 1995. Effect of algal food concentration on toxicity of two agricultural pesticides to Daphnia carinata. Ecotoxicol. Environ. Saf. 32 : 273-279.

Chandini, T. 1989. Survival, growth and reproduction of Daphnia carinata (Crustacea: Cladocera) exposed to chronic cadmium stress at different food (Chlorella) levels. Environ. Pollut. 60 : 29-45.

Christensen, B.T., T.L. Lauridsen, and H.W. Ravn, 2005. A comparison of feeding efficiency and swimming ability of Daphnia magna exposed to cypermethrin. Aquat. Toxicol. 73: 210-220.

Day, K., and N.K. Kaushik. 1987a. An assesment of the chronic toxicity of the synthetic pyrethroid, Fenvalerate, to Daphnia galeata mendotae, using life tables. Environ. Pollut. 44 : 13-26.

Day, K., and N.K. Kaushik. 1987b. Short-term exposure of zooplankton to the synthetic pyrethroid Fenvalerate and its effects on rates of filtration and assimilation of the algae Chlamydomonas reinhardii. Arch. Environ. Contam. Toxicol. 16: 423-432.

Day, K., N.K. Kaushik, and K.R. Solomon. 1987c. Impact of Fenvalerate on freshwater enclosed communities filtration rates in zooplankton. Can. J. Fish. Aquat. Sci. 44: 1714-1728.

Day, K., and N.K. Kaushik. 1987d. The adsorption of fenvalerate to laboratory glassware and the algae Chlamydomonas reinhardii, and its effect on uptake of the pesticide by Daphnia galeata mendotae. Aquat. Toxicol. 10: 131-142.

Day, K.E. 1991. Effects of dissolved organic carbon on accumulation and acute toxicity of Fenvalerate, Deltamethrin and Cyhalothrin to Daphnia magna (Straus). Environ. Toxicol. Chem. 10 : 91-102.

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Enserink, L., W. Luttmer, and H. Maas-Diepeveen. 1990. Reproductive strategy of Daphnia magna affects the sensitivity of its progeny in acute toxicity tests. Aquat. Toxicol. 17: 15-26.

Enserink, E.L., M.J.J. Kerkhofs, C.A.M. Baltus, and J.H. Koeman. 1995. Influence of food quantity and lead exposure on maturation in Daphnia magna: evidence for a trade-off mechanism. Funct. Ecol. 9: 175-185.

Forbes, V.E., and A. Cold. 2005. Effects of the pyrethroid esfenvalerate on life-cycle traits and population dynamics of Chironomus riparius - Importance of exposure scenario. Environ. Toxicol. Chem. 24 : 78-86.

Glazier, D.S. 1992. Effects of food, genotype, and maternal size and age on offspring investment in Daphnia magna. Ecology 73: 910-926.

Gliwicz, Z.M. and C. Guisande. 1992. Family planning in Daphnia: resistance to starvation in offspring born to mothers grown at different food levels. Oecologia 91: 463-467.

Handy, R.D. 1994. Intermittent exposure to aquatic pollutants: assessment, toxicity and sublethal responses in fish and invertebrates Comp. Biochem. Physiol. C Pharmacol. Toxicol. Endocrin. 107: 171-184.

Heugens, E., A. Hendriks, T. Dekker, N.M. Van Straalen, and W. Admiraal. 2001. A review of the effects of multiple stressors on aquatic organisms and analysis of uncertainty factors for use in risk assessment. Crit. Rev. Toxicol. 31: 247-284.

Heugens, E.H.W., T. Jager, R. Creyghton, M.H.S. Kraak, A.J. Hendriks, N.M. Van Straalen, and W. Admiraal. 2003. Temperature-dependent effects of cadmium on Daphnia magna: accumulation versus sensitivity. Environ. Sci. Technol. 37: 2145-2151.

Jager, T., T. Crommentuijn, C.A.M. Van Gestel, and S.A.L.M. Kooijman. 2004. Simultaneous modeling of multiple endpoints in life-cycle toxicity tests. Environ. Sci. Technol. 38: 2894-2900.

Klüttgen, B., N. Kuntz, and H.T. Ratte. 1996. Combined effects of 3,4-Dichloroaniline and food concentration on life-table data of two related Cladocerans, Daphnia magna and Ceriodaphnia quadrangula. Chemosphere 32: 2015-2028.

Kooijman, S.A.L.M., and J.AJ. Metz. 1984. On the dynamics of chemically stressed populations: the deduction op population consequences from effects on individuals. Ecotox. Environ. Saf. 8: 254-274.

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Kooijman, S.A.L.M., and J.J.M. Bedaux. 1996. Analysis of toxicity tests on Daphnia survival and reproduction. Water Res. 30: 1711-1723.

Kooijman, S.A.L.M. 2000. Dynamic energy and mass budgets in biological systems; Cambridge University Press: Cambridge, UK.

McCarthy, J.R. 1983. Role of particulate organic matter in decreasing accumulation of polynuclear aromatic hydrocarbons by Daphnia magna. Arch. Environ. Contam. Toxicol. 12: 559-568.

McCarthy, L.S., and D. Mackay. 1993. Enhancing ecotoxicological modeling and assessment. Body residues and modes of toxic action. Environ. Sci. Technol. 27: 1719-1728.

McCahon, C.P., and D. Pascoe. 1991. Brief-exposure of first and forth instar Chironomus riparius larvae to equivalent assumed doses of Cadmium: Effects on adult emergence. Water, Air, and Soil Pollution. 60: 396-403.

McCahon, C.P., M.J. Poulton, P.C. Thomas, Q. Xu, D. Pascoe, and C. Turner. 1991. Lethal and sub-lethal toxicity of field simulated farm waste episodes to several freshwater invertebrate species. Water Res. 25: 661-671.

McWilliam, R.A., and D.J. Baird. 2002. Postexposure feeding depression: A new toxicity endpoint for use in laboratory studies with Daphnia magna. Environ. Toxicol. Chem. 21: 1198-1205.

Meeker, W.Q., and L.A. Escobar. 1995. Teaching about approximate confidence regions based on maximum likelihood estimation. Am. Stat. 49: 48-53.

Naddy, R.B., K.A. Johnson, and S.J. Klaine. 2000. Response of Daphnia magna to pulsed exposures of Chlorpyrifos. Environ. Toxicol. Chem. 19: 423–431.

OECD. 1997. Guidelines for Testing of Chemicals: Daphnia magna Reproduction Test. Organization for Economic Cooperation and Development (OECD), vol. 211, Paris, France, pp 1-21.

Péry, A.R.R., J.J.M. Bedaux, C. Zonneveld, and S.A.L.M. Kooijman 2001. Analysis with time-varying concentrations. Wat. Res. 35: 3825-3832.

Pieters, B.J., A. Paschke, S. Reynaldi, M.H.S. Kraak, W. Admiraal, and M. Liess. 2005. Influence of food limitation on the effects of Fenvalerate pulse exposure on the life history and population growth rate of Daphnia magna. Environ. Toxicol. Chem. 24: 2254-2259.

Pieters, B.J., and M. Liess. 2006. Maternal nutritional state determines the sensitivity of Daphnia magna offspring to short-term Fenvalerate exposure. Aquat. Toxicol. 76: 268–277.

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Reinert, K.H., J.M. Giddings, and L. Judd. 2002. Effects analysis of time-varying or repeated exposures in aquatic ecological risk assessment of agrochemicals. Environ. Toxicol. Chem. 21: 1977-1992.

Reynaldi, S., S. Duquesne, K. Jung, and M. Liess. 2006. Linking feeding activity and maturation of Daphnia magna following short-term exposure to Fenvalerate. Environ. Toxicol. Chem. 25: 1826-1830.

Rose, R.M., M.J. Warne, and R.P. Lim. 2002. Food concentration affects the life history response of Ceriodaphnia cf. dubia to chemicals with different mechanisms of action. Ecotoxicol. Environ. Saf. 51: 106-114.

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Chapter 5

Population developmental stage determines the recovery potential of Daphnia magna populations after

fenvalerate application

B.J. Pieters, and M. Liess. Environmental Science & Technology (2006) 40: 6157-6162.

Abstract This study investigated the responses of Daphnia magna populations to pulsed exposures of the pyrethroid insecticide fenvalerate applied during an early and a late stage of population development, and analyzed the dynamics of the subsequent recovery. A novel digital observation technique was used to describe the size and numbers of animals. High fenvalerate concentrations caused high mortality rates during exponential population growth as well as during the food limited stationary phase. However, recovery of populations took considerably longer in the stationary phase than in populations growing exponentially. The poor nutritional and reproductive state of food deprived adults was indicated as the main cause of the slow recovery of populations. It is argued that populations operating at the carrying capacity of their environment are vulnerable to toxicant-induced disturbances to an extent not predictable from observations on exponentially growing populations such as are commonly used in ecotoxicology.

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Introduction Food shortage and interspecific interaction have only occasionally been considered in ecotoxicology (Cleuvers et al. 1997, Grant 1998, Barata et al. 2002). Stationary populations of daphnids have been shown to be more vulnerable to toxicants (Sibly 1999). However, other studies have indicated that toxicant-induced mortalities alleviate crowding effects (Postma et al. 1994), and therefore increase the amount of resources per capita for the survivors. Thus overall performance of the animals in the presence of toxicants may increase (Hooper et al. 2003). Nonetheless, toxicant impacts on populations are often assessed at low population density and high food availability, whereas toxicant effects on carrying capacity have less frequently been addressed (Forbes & Calow 1999, Sibly et al. 2000, Liess et al. 2006).

We therefore investigated in this study the effects of a short-term fenvalerate (FV) exposure on population dynamics (density, size structure and recovery time) under density-independent and density-dependent conditions. A single toxicant dose is especially relevant in agriculture, where pesticide application may result in short-term pulse exposures of aquatic biota due to spray drift, drain flow, or field runoff (Liess et al. 1999). These may cause severe population disturbances owing to brief episodes of increased mortality and/or reduced growth, which may be followed by recovery to equilibrium densities.

Laboratory microcosm tests with Daphnia magna Straus populations were performed, in which the test organisms were exposed to single insecticide doses during exponential growth (high food level/low density) and during the stationary phase (low food level/high density), so that they were in different physiological states at the moment of pesticide application. Recovery of the population density and size structure was examined.

Daphnia magna was chosen because it is frequently used in ecotoxicity tests (Baudo 1987), is a key food source in the food-chain of aquatic ecosystems (Shapiro 1980), and has a higher sensitivity compared to many other aquatic invertebrates (Von der Ohe & Liess 2004). The neurotoxic pyrethroid insecticide Fenvalerate was selected as chemical stressor. Fenvalerate exposure causes rapid increases in mortality to daphnids (Pieters et al. 2005). Effects of FV at sublethal concentrations include a loss of coordination and immobilization, reduction of filtration rates (Day and Kaushik 1987), which consequently reduces growth, maturity time and reproductive output of daphnids (Pieters et al. 2005).

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Fenvalerate has regularly been detected in the aquatic environment (Kreuger 1998) and a high sensitivity has been found for non-target organisms such as aquatic macroinvertebrates (Pieters et al. 2005, Anderson 1982). Furthermore, long-term effects on macroinvertebrate species after FV pulse exposure have been observed at the individual level (Liess & Schulz 1996), population level (Liess 2002), and in communities (Woin 1998). Materials & methods Daphnia magna culture

The experiments were performed with the cladoceran Daphnia magna Straus (clone B; Bayer, Monheim, Germany). Continuous cultures of 10 individuals (inds) L-1 in artificial Elendt M7 medium (OECD 1998) were maintained under controlled laboratory conditions at 20 °C, with a photoperiod of 16:8 h light:dark. Light intensity directly above the culture was approximately 1400 Lux. The pH was 7.4, conductivity 630 µS cm-1, and the dissolved oxygen level was 7.15 mg L-

1. Once a week a new culture was initiated using < 24 h-old neonates from a three-week-old culture. The newly released neonates were removed twice a week, and the medium was renewed three times a week. Daphnids were fed three times a week with a suspension of batch-cultured green algae (Desmodesmus subspicatus). Algae were cultured in algae medium according to Grimme and Boardmann (Grimme & Boardmann 1972) and continuously aerated (3% CO2). The algae were harvested in exponential growth phase, centrifuged (3000 rpm for 10 min), and the pellet was resuspended in M7 medium. The food quantity given as equivalent daily ration was 5.3 × 105 cells mL-1 (0.75 mg carbon L-1). Measurements were performed with a total organic carbon analyzer (Shimatzu TOC-5000 Analyzer, Duisburg, Germany) and a coulter counter (Casy1 counter, Scharfe Systems, Reutlingen, Germany).

Experimental design microcosm

Eighty-nine cylindrical glass vessels (Harzkristall, Derenburg, Germany), each containing 4.7 L M7 medium, were prepared at the start of the experiments. Each vessel contained 500 g of washed aquarium gravel (diameter 1.0-2.0 mm) as support for nutrient cycling (nitrifying bacteria) to improve self-purification of the system. Next, cohorts of thirty neonates (< 24 h old) originating from a third brood of the Daphnia culture were introduced into each vessel. Four days after

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inoculation of the vessels, Daphnia population dynamics were recorded every 1-4 d, depending on the density changes, using a digital image analysis technique (see next section for details and results on validation). The medium was aerated three times a day for 15 min via glass tubes (length 18 cm, diameter 3 mm) which were submerged 14 cm below the water surface. A transparent plastic plate was placed on top of each vessel in order to prevent excessive evaporation. Approximately 80% of the medium was renewed every week using an overflow tube with a 200-µm nylon mesh to prevent loss of animals through the outflow. Water quality parameters were measured weekly and did not significantly differ between exposure treatments and experiments (t-test, p > 0.05), and were at levels where no deleterious effects on the daphnids would be expected (Table 1). The animals were fed daily with an algae suspension (1.63 × 105 cells mL-1; 0.23 mg carbon L-

1). Other experimental conditions were similar to those described for the Daphnia culture. When the Daphnia populations showed a log-phase growth with a mean total population density of about 60 inds L-1, forty-three vessels were randomly allocated to test treatments and tested for uniformity (t-test, p > 0.05). Exposure occurred by a single dose of seven FV concentrations (control, 0.03, 0.1, 0.3, 0.6, 1.0, 3.2 and 10.0 µg L-1). The control was replicated seven times, and the FV treatments were replicated six times at the concentrations from 0.03 to 1.0 µg L-1 and three times for 3.2 and 10.0 µg L-1. The day of FV application was defined as day ‘zero’. The remaining forty-six Daphnia populations of the ‘stationary phase’ experiment were allowed to reach an apparent carrying capacity. At a relatively stable mean total population density of about 117 inds L-1 for approximately 10 days, Daphnia populations were distributed and exposed similarly as described above for the ‘exponential growth’ experiment, apart from three additional replicas for the 3.2 µg L-1 treatment.

Next, Daphnia populations were divided into three body-length groups: 0.7-1.0 mm (neonatal), 1.0-1.8 mm (juvenile) and 1.8-4.0 mm (adult). The juvenile and adult size classes were derived from size measurements on animals at first reproduction (Lampert 1993). Periods of recovery of these different size classes were determined and defined as the duration between the lowest density following exposure and the first observation day on which the density deviated by less than 20% of the control. For purposes of standardization, observation days on which the densities were lowest were retrieved from the total population densities (day 2 in the exponential growth experiment and day 3 in the stationary phase experiment). Both experiments ended when no density differences in any size

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class were observed compared to the corresponding controls after recovery from FV application (day 34 in the exponential growth experiment and day 61 in the stationary phase experiment). Table 1. Water-quality parameters presented as mean values (± standard deviation in brackets) calculated for all treatments combined and for the entire duration of the experiments.

parameter exponential growth stationary phase

NH4+ (mg L-1) 0.14 (± 0.03) 0.15 (± 0.03)

NO2- (mg L-1) 0.13 (± 0.03) 0.12 (± 0.02)

dissolved O2 (mg L-1) 7.6 (± 0.1) 7.3 (± 0.3) pH 7.7 (± 0.2) 7.8 (± 0.1) conductivity (µS cm-1) 663 (± 22) 701 (± 39) temperature (°C) 20.3 (± 0.1) 20.5 (± 0.1)

Image analysis technique

An extended version of an image analysis system (Liess et al. 2006) was used for the estimation of abundance and size structure of Daphnia populations. The image analysis system is based on digitized images of Daphnia populations which were post-processed using image analyzing software. Pictures were taken with a high-performance digital camera which was mounted on one end of a plastic rectangular box that was painted black on the inside. The other end of the box was cut to fit the side surface of the cylindrical glass vessel (diameter 20 cm, height 22 cm). Two parallel white fluorescent tubes, each 70 W (~1400 Lux) and situated 10 cm directly above the test vessels, provided illumination. Black pond liner was taped to the back of each vessel to cover 60% of the surface in order to provide contrast between the daphnids and the surrounding water. The analysis is based on three consecutive time-lapsed (~0.4 sec) images (2048 × 1536 pixels, 1/30 sec shutter speed, F 2.8 aperture, ISO 400, focal depth in middle of vessel) which were recorded of swimming Daphnia populations. By subtracting three successive images, all background pixels (i.e., pixels that are constant over time and therefore ‘non-daphnid’) are zeroed. The non-zero pixels in the difference image consists of moving objects only (i.e., ‘daphnids’), resulting in a background-free image of the moving objects in their final positions. The

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resulting image is then further post-processed by standard algorithms of image analysis in which the image is transformed to binary (i.e., black ⁄red) by applying a ‘thresholding’ algorithm, which separates the image into background (black) and objects (red). Objects are then subjected to a segmentation step, which deletes small objects caused by image noise and merges and fills holes in the remaining ones. Finally, in the quantification step, objects in the segmented image are counted and the geometrical property “surface area of individual objects” is recorded. This allowed the estimation of the total number of individuals and the size structure of the Daphnia populations.

Validation & extrapolation methods

Individuals in the vessels obscured each other increasingly at higher densities, so that the actual population size was underestimated by image analysis. Compensation for this superposition was performed by extrapolations from linear regression (Figure 1a). For this purpose, a continuous Daphnia culture was initiated at densities of 10 inds L-1 and fed daily (1.63 × 105 cells mL-1; 0.23 mg carbon L-1). When a mean body length of about 2 mm was reached, cohorts of 25 individuals were successively added to a single test vessel, and recorded with image analysis, until a total number of 700 individuals were reached.

The ‘pixel’ size distribution was transformed to size classes based on body lengths in mm using linear regression (Figure 1b). This was done by measuring on digital pictures the body lengths (distance from the middle of the eye to the base of the tail spine) of individuals (n = 20) belonging to 10 different cohorts of differing ages obtained from the Daphnia culture described above. Next, the 20 individuals of the respective cohorts were transferred into a test vessel, digitally analyzed, and the surface areas of these individuals determined in pixels. Fenvalerate analysis

Fenvalerate, (RS)-α-cyano-3-phenoxybenzyl (RS)-2-(4-chlorophenyl)-3-methylbutyrate, CAS 51630-58-1, was obtained from Riedel-de Haën, Seelze, Germany (High-Performance Liquid Chromatography (HPLC) technical grade; 99.9% purity). Dimethylsulfoxide (DMSO) was used as carrier solvent and obtained from Merck, Darmstadt, Germany (HPLC technical grade; 99.8% purity). A FV stock solution in DMSO was prepared, from which an aqueous FV stock solution was prepared. Nominal test concentrations of 0.03, 0.1, 0.3, 0.6, 1.0, 3.2 and 10.0 µg L-1 were subsequently prepared by adding appropriate

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aliquots of the aqueous FV stock solution to the M7 medium in the vessels. The maximum amount of DMSO present in the most concentrated test solution was 0.00005% (volume/volume). This respective DMSO concentration was also added to the controls. Actual exposure concentrations were determined in triplicate for all treatments in both experiments at t = 1 h. Figure 1. (A) Relationship between log10 image analysis counts and log10 manual counts, and (B) relationship between log10 daphnid surface area (pixels) and log10 daphnid body length (mm). Error bars indicate ± standard deviations. Equations and coefficients of determination (r2, p) of least-square linear regression lines are shown for each relationship. Additional water samples were collected from the 3.2 µg L-1 test solutions (n = 3) at t = 24, 48, 96 and 144 h for both experiments. The test solutions in all vessels were completely renewed with clean M7 medium at day 7. Measurements were performed by means of solid-phase extraction of 100-mL water samples with C18-columns (Baker, Phillipsburg, NJ, USA), followed by chromatography/mass spectrometry applying electron impact ionisation and selected ion monitoring mode (Hewlett Packard 6890 Series GC System with Hewlett Packard 7683 Series Injector and Hewlett Packard 5973 Mass Selective Detector).

Results of chemical analyses are shown in Tables 2 and 3. Statistical analyses demonstrated that the actual FV concentrations at t = 1 h were not significantly different (t-test, p > 0.05) between the two experiments. For the 3.2 µg L-1 FV treatment a strong initial decrease in the actual concentration to 24% was observed after 24 h, followed by a slower decrease to only 4% of the initial concentration at t = 144 h. Also no significant differences (t-test, p >0.05) were shown between the actual FV concentrations of the 3.2 µg L-1 treatment in the two

2.8 Y = 0.14 + 0.91Xr 2 = 0.999p = < 0.001

log10 manual counts

log 1

0im

age

anal

ysis

coun

ts

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log 1

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ea(p

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anal

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coun

ts

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log 1

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experiments over time. All exposure concentrations in the following graphs and tables are given as nominal values. Table 2. Actual fenvalerate (µg L-1) concentrations (mean with ± standard deviation in brackets) measured at t = 1 h and their corresponding nominal concentrations (µg L-1) for both experiments.

exponential growth stationary phase

nominal conc. actual concentration

0.03 n.d. a n.d. a 0.1 0.07 (± 0.02) 0.08 (± 0.01) 0.3 0.24 (± 0.04) 0.26 (± 0.01) 0.6 0.42 (± 0.01) 0.47 (± 0.05) 1.0 0.74 (± 0.02) 0.79 (± 0.01) 3.2 2.27 (± 0.05) 2.07 (± 0.15)

10.0 7.41 (± 0.36) 7.77 (± 0.43) a n.d. = not detected due to high detection limit. Table 3. Time-series measurements of actual fenvalerate (µg L-1) concentrations in the 3.2 µg L-1 treatment (mean with ± standard deviation in brackets) at t = 1, 24, 48, 96 and 144 h and their corresponding nominal concentrations (µg L-1) for both experiments.

exponential growth stationary phase

time (h) actual concentration

1 2.27 (0.05) 2.07 (0.15) 24 0.74 (0.06) 0.83 (0.10) 48 0.49 (0.08) 0.59 (0.12) 96 0.31 (0.03) 0.35 (0.08)

144 0.16 (0.03) 0.13 (0.01)

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Results Development of population structure before FV application

The thirty neonates which were introduced into the vessels showed initially somatic growth, after which the age at first reproduction (8 days in physiological time) was reached at day -4 in the exponential growth experiment and at day -25 in the stationary phase experiment (Figure 2). Due to the start of reproduction, neonatal and juvenile size classes increased while the adult size class showed a slower increase. At the moment of FV application in the exponential growth experiment, the population consisted of 33% neonates, 45% juveniles and 23% adults. Daphnia populations to be used for the stationary phase experiment continued to grow, showed peak densities in the neonatal and juvenile size classes, and subsequently reached a stationary phase at carrying capacity. At the moment of FV application, the stationary phase experiment showed a population size structure consisting of 10% neonates, 41% juveniles and 49% adults.

Figure 2. Development of Daphnia magna populations as a function of time in three body-length groups (0.7-1.0 mm (neonatal), 1.0-1.8 mm (juvenile), and 1.8-4.0 mm (adult)) before FV application during exponential growth (left side) and during the stationary phase (right side). Error bars represent ± standard deviation and arrows indicate the moment of fenvalerate application.

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Population structure after FV application Fenvalerate-induced mortality in the 1.0-10.0 µg L-1 treatments caused rapid

and severe declines in all size classes in both experiments (Figure 3). Complete mortality occurred at 10.0 µg L-1 FV, partial mortality occurred at 1.0-3.2 µg L-1 FV, while generally no effects on survival were observed at 0.03-0.6 µg L-1 FV (not shown in Figure 3). Examination of percent reductions in neonatal, juvenile and adult densities 2 days after FV application demonstrated no differences between the 3.2 µg L-1 treatments of both experiments, whereas only minor differences were observed between both experiments in the 1.0 µg L-1 treatments. Recovery periods were only determined in the case of partially lethal concentrations (1.0 and 3.2 µg L-1 FV; see Table 4).

Recovery during exponential growth started almost instantaneously. Daphnids exposed during the stationary phase, however, showed delays in the onset of recovery since constant periods of low density were observed up to about 10 d at 3.2 µg L-1 FV. Recovery periods therefore varied between the two experiments (Table 4). In general, recovery periods increased with increasing FV concentration, and recovery periods during exponential growth were much shorter than during the stationary phase. Overshoots above controls were observed after recovery at 1.0-3.2 µg L-1 FV in the neonatal and juvenile size classes in both experiments. However, density peaks were relatively higher and stabilisation took longer during the stationary phase than during the exponential growth. The adult size class in the exponential growth experiment also demonstrated overshoots at 1.0-3.2 µg L-1 FV, whereas none were observed in the stationary phase experiment. Table 4. Daphnia magna recovery periods (days) in the three body length groups after exposure to 1.0 and 3.2 µg L-1 fenvalerate during exponential growth and stationary phase.

exponential growth stationary phase

1.0 µg L-1 3.2 µg L-1 1.0 µg L-1 3.2 µg L-1 size class (n = 6) (n = 3) (n = 6) (n = 6)

0.7-1.0 mm 1 4 7 10 1.0-1.8 mm 2 5 7 12 1.8-4.0 mm 5 7 9 35

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Figure 3. Daphnia magna population densities (inds L-1) as a function of time in three body-length groups (0.7-1.0 mm (neonatal), 1.0-1.8 mm (juvenile), and 1.8-4.0 mm (adult)) after FV application during exponential growth (left side) and during the stationary phase (right side). Error bars and 0.03-0.6 µg L-1 FV treatments have been omitted for clarity. Arrows indicate the moment of fenvalerate application.

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Discussion This study demonstrated that populations at carrying capacity exposed to a single FV dose required considerably more time to recover than populations at lower densities that allow for saturating food supply. These differences in recovery may have depended on the nutritional and reproductive state of adult females at the moment of pesticide application. In the short-term these adults may govern the recovery of the populations’ density, while surviving neonates or juveniles still have to reach maturity. Adult individuals in the stationary phase suffered starvation and most likely halted reproduction by ceasing egg production (Bradley et al. 1991a). Recovery may consequently have not occurred directly after FV exposure, since surviving adult individuals first had to go through the time-consuming process of restoring body mass and starting egg development, after which reproduction and hence recovery were finally initiated. The relatively slow reaction of the surviving adults in the stationary phase to the sudden mortality-induced increase in available food shortly after FV application has previously been observed in starvation and re-feeding experiments with Daphnia (Perrin et al. 1990, Bradley et al. 1991b). However, these studies did not use test designs which incorporated population-scale responses in combination with toxicant stress.

The delayed onset of recovery observed during the stationary phase may have additionally been exacerbated by Fenvalerate-induced impairments of filtration and assimilation of algae (Day & Kaushik 1987). The temporal inability to feed may not have affected surviving adults during the exponential growth phase equally strongly because they likely possessed large egg clutches in the brood pouch prior to exposure, ready to be released. Visual observations of individuals in the exponentially growing populations showed indeed that, while mortality occurred in all size classes, adult females simultaneously produced offspring as early as one day after FV application. The ability to reproduce was probably not impaired by toxicant-induced egg mortality, which is consistent with the observation that FV did not affect clutch size (Pieters et al. 2005). Furthermore, FV concentrations decreased markedly within 24 h to non-lethal concentrations, not affecting the newly released neonates. Thus, reproduction occurred soon after exposure and recovery periods were consequently shorter during exponential growth.

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Other studies have also demonstrated the importance of population structure at the time of toxicant exposure (Stark & Banken 1999, Hanazato & Hirokawa 2004). Stark & Banken (1999) stressed likewise the importance of adult arthropods to recovery, but they showed on the contrary that populations consisting of all adults were less susceptible than populations consisting primarily of neonates. However, these terrestrial arthropods were kept in soil containers that provided ample food. We hypothesize that in the present study the food shortage in the stationary phase overrode the ability of the surviving adults to reproduce. Hanazato & Hirokawa (2004) performed a similar study examining the recovery of Daphnia populations after application of a single dose of Carbaryl (acetylcholinesterase inhibitor) at different stages of development (i.e., growing, peak density). They used another species (Daphnia pulex) however, and only one concentration was applied which killed juveniles but not adults. Consistent with our findings the recovery periods were longest during the stationary phase (peak density), but in contrast to our line of argumentation, this was explained by low food-induced increases in sensitivity of adults, and thus by the very small total number of surviving individuals from which population recovery had to begin. Interestingly, differences in sensitivity of individual adults to FV were not apparent in the present experiments. We hypothesize that at highly lethal FV concentrations an ‘all-or-nothing’ effect occurs, preventing the expression of food-dependent sensitivity of adults. More importantly, the individual recovery time of individuals was very different.

In the exponential growth experiment some overshoot of recovering individuals was observed and may be explained by the strongly increased food availability after the FV-induced mortality, causing rapid production of offspring. Overshoots of neonatal and juvenile size classes in the stationary phase experiment were even stronger due to the prolonged periods of low population density. The overshoot in numbers, however, may not fully reflect in an overshoot of biomass which depends very much on the size class distribution.

In summary, our findings indicated that the application of a single dose of Fenvalerate caused similar mortality rates in rapidly developing Daphnia populations and in stationary, food-limited populations. However, the recovery rates of surviving populations differed markedly depending on the growth phase, indicating that food condition and population structure are major variables governing susceptibility of daphnid populations to FV. One important mechanism which may have caused the delayed onset of recovery was that adult individuals

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needed up to approximately 10 days to resume reproduction, possibly as a result of the adaptation to the low food supply. Thus, populations operating at the carrying capacity of their environment may be vulnerable to toxicant-induced disturbances not predictable from observations on exponentially growing populations, which are a common reference for ecotoxicological studies. Acknowledgements

The authors thank W. Admiraal and M.H.S. Kraak for valuable comments and discussions on the manuscript. We also acknowledge A. Paschke, U. Schröter and M. Heinrich for technical assistance with chemical analyses. Finally, we thank S. Wahrendorf for support with the image analysis system and I. Raenker for practical help in the laboratory. References Anderson, R.L. 1982. Toxicity of fenvalerate and permethrin to several non-

target aquatic invertebrates. Environ. Entomol. 11: 1251-1257. Barata, C., D.J. Baird, and A.M.V.M. Soares. 2002. Demographic responses of a

tropical cladoceran to cadmium: effects of food supply and density. Ecol. Appl. 12: 552-564.

Baudo, R. 1987. Ecotoxicological testing with Daphnia magna; Peters, R.H., DeBernardi, R., Eds.; Mem. Ist. Ital. Idrobiol.: Pallanza, pp 509-518.

Bradley, M.C., D.J. Baird, and P. Calow. 1991a. Mechanisms of energy allocation to reproduction in the cladoceran Daphnia magna Straus. Biol. J. Linn. Soc. 44: 325-333.

Bradley, M., N. Perrin, and P. Calow. 1991b. Energy allocation in the cladoceran Daphnia magna Straus, under starvation and refeeding. Oecologia. 86: 414-418.

Cleuvers, M., B. Goser, and H.T. Ratte. 1997. Life-strategy shift by intraspecific interaction in Daphnia magna: change in reproduction from quantity to quality. Oecologia. 110: 337-345.

Day, K., and N.K. Kaushik. 1987. Short-term exposure of zooplankton to the synthetic pyrethroid Fenvalerate and its effects on rates of filtration and assimilation of the algae Chlamydomonas reinhardii. Arch. Environ. Contam. Toxicol. 16: 423-432.

Forbes, V.E., and P. Calow. 1999. Is the per capita rate of increase a good measure of population-level effects in ecotoxicology? Environ. Toxicol. Chem. 18: 1544-1556.

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Grant, A. 1998. Population consequences of chronic toxicity: incorporating density dependence into the analysis of life table response experiments. Ecol. Model. 105: 325-335.

Grimme, L.H., and N.K. Boardmann. 1972. Photochemical activities of particle fraction P1 obtained from the green algae Chlorella fusca. Biochem. Biophys. Res. Commun. 49: 1617-1623.

Hanazato, T., and H. Hirokawa. 2004. Changes in vulnerability of Daphnia to an insecticide application depending on the population phase. Freshw. Biol. 49: 402-409.

Hooper, H.L., R.M. Sibly, S.J. Maund, and T.H. Hutchinson. 2003. The joint effects of larval density and 14C-cypermethrin on the life history and population growth rate of the midge Chironomus riparius. J. Appl. Ecol. 40: 1049-1059.

Kreuger, J. 1998. Pesticides in stream water within an agricultural catchment in southern Sweden, 1990-1996. Sci. Total. Environ. 216: 227-251.

Lampert, W. 1993. Phenotypic plasticity of the size at first reproduction in Daphnia: the importance of maternal size. Ecology. 74: 1455-1466.

Liess, M. 2002. Population response to toxicants is altered by intraspecific interaction. Environ. Toxicol. Chem. 21: 138-142.

Liess, M., B.J. Pieters, and S. Duquesne. 2006. Long-term signal of population disturbance after pulse exposure to an insecticide: rapid recovery of abundance, persistent alteration of structure. Environ. Toxicol. Chem. 25: 1326-1331.

Liess, M., and R. Schulz. 1996. Chronic effects of short-term contamination with the pyrethroid insecticide fenvalerate on the caddisfly Limnephilus lunatus. Hydrobiologia. 324: 99-106.

Liess, M., R. Schulz, M.H.D. Liess, B. Rother, and R. Kreuzig. 1999. Determination of insecticide contamination in agricultural headwater streams. Water Res. 33: 239-247.

OECD, Organization for Economic Cooperation and Development. 1998. OECD guidelines for testing of chemicals: Daphnia magna reproduction test, 211, Paris, France, pp. 1-21.

Perrin, N., M.C. Bradley, and P. Calow. 1990. Plasticity of storage allocation in Daphnia magna. Oikos. 59: 70-74.

Pieters, B.J., A. Paschke, S. Reynáldi, M.H.S. Kraak M.H.S., W. Admiraal, and M. Liess. 2005. Influence of food limitation on the effects of Fenvalerate

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pulse exposure on the life history and population growth rate of Daphnia magna. Environ. Toxicol. Chem. 24: 2254-2259.

Postma, J.F., M.C. Buckert-de Jong, N. Staats, and C. Davids. 1994. Chronic toxicity of cadmium to Chironomus riparius (Diptera: Chironomidae) at different food levels. Arch. Environ. Contam. Toxicol. 26: 143-148.

Shapiro, J. 1980. The importance of trophic-level interactions to the abundance and species composition of algae in lakes. In Hypertrophic Ecosystems; Barica, J., and L.R. Mur, Eds.; Dr Junk: The Netherlands, pp 105-116.

Sibly, R.M. 1999. Efficient experimental designs for studying stress and population density in animal populations. Ecol. Appl. 9: 496-503.

Sibly, R.M.; T.D Williams,and M.B. Jones. 2000. How environmental stress affects density dependence and carrying capacity in a marine copepod. J. Appl. Ecol. 37: 388-397.

Stark, J.D., and J.A.O. Banken. 1999. Importance of population structure at the time of toxicant exposure. Ecotoxicol. Environ. Saf. 42: 282-287.

Von der Ohe, P., and M. Liess. 2004. Relative Sensitivity Distribution (RSD) of aquatic invertebrates to organic and metal compounds. Environ. Toxicol. Chem. 23: 150-156.

Woin, P. 1998. Short- and long-term effects of the pyrethroid insecticide fenvalerate on an invertebrate pond community. Ecotoxicol. Environ. Saf. 41: 137-156.

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Chapter 6

Concluding remarks The aim of the present thesis was to examine population responses to pesticides under variable environmental conditions. To this purpose we investigated the effects of food limitation and included short-term exposure regimes in combination with extended post-exposure observation periods to investigate effect propagation and recovery. The results obtained for cohorts were verified with observations on mixed Daphnia magna populations in different stages of their development. The findings urge to re-consider the current methods to assess population effects of pesticides and have essential implications for pesticide risk assessment as discussed below. Pulse exposure & delayed effects Modern pesticides have been designed to have relatively short half-lives, but meanwhile to be very active. Together with the episodic pesticide application in the field, followed by unintended spreading through spray-drift, runoff, drainage or atmospheric deposition (Heckman 1982), this results in an extreme temporal variability in exposure of non-target organisms. This was the motive for applying short-term exposures as well as extended post-exposure observation periods in the present study. The importance of a post-exposure observation period became immediately clear in the first set of experiments (Pieters et al. 2005). Aspects of delayed and cumulative effects coincided with signs of recovery in a life-history trait specific way: mortality occurred between days 2 and 7 after the fenvalerate pulse exposure, but meanwhile complete recovery of body length occurred at day 14. This raised the question about the net effects after intermediate or long periods of time. Chapter 2 (Pieters et al. 2005) gives a clear answer on this question: at the end of the 21 d exposure period population growth rate at the highest pesticide levels is decreased. Therefore, the present study clearly supports the choice of population growth rate (r) as a relevant measure for ecological impact (Forbes &

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Calow 1999) and confirms that there is an urgent need for incorporating demographic endpoints in laboratory ecotoxicity tests (Bechmann 1994, Kammenga et al. 1996, Calow et al. 1997, Walthall & Stark 1997). Also at odds with standard laboratory toxicity tests, the present thesis showed that a post-exposure observation period is essential to detect delayed and cumulative effects, as well as recovery. For some species a very long observation period of even 84 days may be required (Liess & Schulz 1996). These findings urge to re-consider the current methods for assessing effects of pesticides and require major adjustments of standard operating procedures (SOPs) and guidelines for testing. Adaptation to food limitation Daphnia populations in the field often encounter food limitation (Murdoch & McCauley 1985, McCauley et al. 1988), which may increase their sensitivity to toxicants (Heugens et al. 2001). This prompted us to test the effects of the model pesticide also under food limitation. The experiments demonstrated that low food supply exacerbated the adverse effects of the pesticide, reflected by a reduced population growth rate (Pieters et al. 2005). Cladocerans, however, exhibit a specific reproductive strategy in response to changes in food availability: females produce many small-sized neonates under high food conditions, while only a few large-sized and stress-resistant offspring are born when food is scarse (Cowgill et al. 1985, Enserink 1990). In chapter 3 (Pieters & Liess 2006a) low maternal food conditions indeed increased the offspring body size at time of birth. These larger sized juveniles proved to be less sensitive to the pesticide and reduced effects were observed on chronic endpoints. Hence, counter-intuitively populations adapted to food limitation were less sensitive than well fed animals. Thus the nutritional state of the adults as well as the adaptation of populations to a specific food level affects the response to pesticide exposure. The timing of pesticide application in relation to the actual food level is therefore of crucial importance for the net effect on the population level.

The present study also showed that the available food level affects population structure. It modulated the response to pesticide pulse exposure in a way not demonstrated before. When food is provided to a daphnid population at a fixed rate the population develops exponentially to a high density at which population growth equals mortality rate, and the population is in a steady state: the carrying capacity of the environment is reached. Chapter 5 (Pieters & Liess 2006b) showed

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that pulse exposure to the same pesticide concentration affected recovery of exponentially developing populations and populations stabilized in response to a limited food supply essentially different: development was resumed rapidly in the growing populations, while a prolonged phase of recovery followed the pulse exposure in the stabilized, high density population.

In summary, the present observations advance on Heugens et al. (2001) showing that population responses to the combined factors food and toxicants are not predictable from responses of individual animals. These observations urge a further critical examination of population dynamics as a new and realistic endpoint in toxicity testing.

Testing real populations Populations differing in development stage and age structure proved to respond differently to pesticide exposure (Chapter 5; Pieters & Liess 2006b). This study demonstrated that the age structure of daphnid populations is a key factor for the response to pesticides, and indicated that the nutritional condition of the populations determined recovery potential. Demographic bioassays were almost always initiated with individuals having the same life-stage, age and size, unlike natural populations that harbor individuals of all life stages (Stark & Wennergren 1995). Testing of non-synchronized populations requires that effects are described via the numbers of animals in different size and age classes. Since in classical experiments the counting effort is already quite often a limiting step, new techniques were required. Therefore the optical quantification of size and numbers and the non-destructive nature of the method reported in Chapter 5 are essential for establishing experimental evidence for genuine population effects of pesticides. Applying these new techniques, evidence of population effects non-predictable from cohort studies was provided in Chapter 5 (Pieters & Liess 2006b). However, the potential of these new techniques needs to be further evaluated, generating more experimental data. Some indications can be derived from mesocosm studies that describe the response of communities with cladocerans to pesticides (Graney et al. 1994, Van den Brink et al. 1995, Van den Brink et al. 1996), but so far a rigorous comparison with either the microcosm or the cohort method is lacking.

It is concluded that heterogeneous populations consisting of individuals differing in developmental stage are an important subject of ecotoxicity testing,

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and that such testing is facilitated by rapid non-destructive optical registration of populations. It remains to be established to what extend this new approach could replace cohort testing. Predicting population responses Chapter 4 (Pieters et al. 2006) showed that the process-based model DEBtox could accurately describe our experimental data, and it could easily account for differing food levels and time-varying exposure concentrations. These findings contrast with the current state of population ecotoxicology (Kammenga & Laskowski 2000, Newman 2001, Stark & Banks 2003). In his handbook on population ecotoxicology, Newman (2001) commented critically on the quantification of toxic inhibition: “LC50, EC50, and NOEC data have significant deficiencies as measures of effects on natural populations. All methods based on these metrics share their shortcomings”. And furthermore “A population-based paradigm for assessing ecological risk should replace the currently reigning paradigm based on effects on individuals.” The present study and Jager et al. (2006) recognized the fundamental shortcomings of current effect metrics and argue that these are in fact “apparent values”. The dose of a toxicant relevant to an individual is determined by the rate of uptake and the dilution of body concentrations by growth. It is argued that the intrinsic sensitivity is determined by the interaction of toxicants with the receptor sites of the organism. Any overall response of the organism to an external dose is therefore time dependent and modified by size dependent uptake and excretion rates, and by the feeding and growth rates determining the effective dose at the receptor site. This approach has been confirmed in the present study and therewith also the criticism of Newman (2001) on the traditional metrics EC50, LC50 and NOEC is confirmed. The current practice of ecotoxicity testing follows these individual metrics that are not suitable for extrapolation to populations. In support of Newman (2001) a new approach to risk assessment for populations is needed. However, a change of paradigms from individuals to population may not be essential; rather a proper modeling of time-dependent observations in experiments is required. To this purpose, experiments have to be designed such that they allow for analysis of the data with the appropriate models. Although time-repeated measurements may seem to increase the experimental effort, we argue that the non-destructive optical measurements described in chapter 5 allow repeated measurements over time

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without additional efforts; in fact laborious counting and measuring is avoided. Time-repeated measurements permit proper control of the dynamics of toxicant exposure and the development rate of the test species. Modeling the results of time-repeated measurements will allow derivation of time-independent metrices, which is essential to solve the current bias of the ECx parameters. Implications for pesticide risk assessment The present study showed that the regime of exposure to an insecticide and the time-dependent effect expression are key factors determining the overall effect on a test population. In ecological risk assessment of pesticides and in the admission procedure for new pesticides this time-dependence is not included.

Exposure concentrations are collected from respectively field sampling programs and dispersion models. Concentrations of pesticides in local natural waters are usually recorded in monthly sampling schemes as prescribed under the Water Framework Directive (WDF, Directive 2000/60/EC) in Europe. However, such sampling schemes are not suitable to detect the pulses of pesticides discharged into local waters. Also the calculation of pesticide runoff is subject to great variability caused by the actual weather at the time of spraying, although progress in exposure models has been made in recent years (SANCO 2001). Nevertheless, the evaluation of pesticide concentrations in relation to safety standards is still precarious. Over- or under-estimation of potential population and community effects are likely to occur.

The uncertainties involved with the L(E)Cx parameters propagate into the systems of (1) derivation of safety standards for pesticides as a tool in environmental management and (2) admission of new pesticides on the market. In The Netherlands maximum permissible concentrations (MPC) and negligible concentrations (NC) are used to set standards for pesticides in the aquatic environment (Crommentuijn et al. 1997, Van Vlaardingen & Verbruggen 2006). The MPC is the concentration in the environment at which most species in a hypothetical community are assumed to be protected against the exposure to a specific chemical. The NC is the hundred-fold lower concentration in the environment at which any effect is highly unlikely. Aquatic MPC and NC values for each compound are calculated from the available set of NOEC and/or L(E)Cx data for aquatic test organisms. In addition, the Dutch board for the authorization of pesticides (CTB) uses a tiered approach for risk assessment. In the first tier

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standard test organisms are used such as algae, daphnids, fish and in the case of herbicides also macrophytes. Criteria for pesticides are deduced for each of these species on the basis of the results of standard ecotoxicity tests with assessment (or safety) factors: 0.01 x L(E)C50 for the acute toxicity, and 0.1 x NOEC for the chronic toxicity to Daphnia. It is therefore concluded that in pesticide risk assessment a large amount of uncertainties are involved. Likewise, the extrapolation of the results presented in this thesis to the field is also subject to some degree of uncertainty. Are the responses to pesticide exposure of daphnid populations in the field, which temporarily may also consist of males, similar or not? Can we indiscriminately extrapolate the toxic effects of the pesticide on the standard test species Daphnia to other (macroinvertebrate) species? And would the NOEC values obtained from this study therefore be protective for exposed populations in the field?

It was argued above that standard summary statistics (L(E)Cx,) are of limited use in part because they change with exposure duration (Jager et al. 2006; Chapters 2, 3, and 4). This problem was also recognized at the 1996 OECD workshop in Braunschweig, Germany (OECD, 1998) on the statistical analysis of aquatic ecotoxicity data. A move towards process-based approaches that deal explicitly with exposure time has been realized in the present study. Further information on the field exposure of aquatic biota to pesticides is needed, especially a time resolution of pesticide pulses. The present thesis demonstrated effects of episodic pollution on daphnid populations and showed that there is no fundamental obstacle for including variable exposure doses. Accordingly pesticide risk assessment must be based on more realistic exposure regimes (i.e., episodic) to reflect application of these substances in agriculture and their runoff to surface water. Thus, time explicit exposure and time explicit effect determination are key elements supporting an improved risk assessment of pesticides. References Bechmann, R.K. 1994. Use of life tables and LC50 tests to evaluate chronic and

acute toxicity effects of copper on the marine copepod Tisbe furcata (Baird). Environ. Toxicol. Chem. 13: 1509-1517.

Calow, P., R.M. Sibly, and V. Forbes. 1997. Risk assessment on the basis of simplified life-history scenarios. Environ. Toxicol. Chem. 16: 1983-1989.

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Cowgill, U.M., D.L. Hopkins, I.T. Applegath, S.D. Takahashi, S.D. Brooks, and D.P. Milazzo. 1985. Brood size and neonate weight of Daphnia magna produced by nine diets, pp. 233-244, STP 891. American Society for Testing and Materials. Philadelphia, Pa, USA.

Crommentuijn, T., D.F. Kalf, M.D. Polder, R. Posthumus, and E.J. Van de Plassche. 1997. Maximum permissible concentrations and negligible concentrations for pesticides. National Institute for Public Health and the Environment, Bilthoven, The Netherlands. RIVM Report no 601501 002. 174 pp.

Directive 2000/60/EC of the European Parliament and of the Council of 23 October 2000 establishing a framework for Community action in the field of water policy. OJ L 327, 22.12.2000, pp1-51.

Enserink, L., W. Luttmer, and H. Maas-Diepeveen. 1990. Reproductive strategy of Daphnia magna affects the sensitivity of its progeny in acute toxicity tests. Aquat. Toxicol. 17: 15-26.

Forbes, V.E., and P. Calow. 1999. Is the per capita rate of increase a good measure of population-level effects in ecotoxicology? Environ. Toxicol. Chem. 18: 1544-1556.

Graney, R.L., J.H. Kennedy, and J.H. Rogers. 1994. Aquatic mesocosm studies in ecological research. Boca Raton, FL, USA, Lewis.

Heckman, C.W. 1982. Pesticide effects on aquatic habitats. Environ. Sci. Technol. 16: 48-57.

Heugens, E., A. Hendriks, T. Dekker, N.M. Van Straalen, and W. Amiraal. 2001. A review of the effects of multiple stressors on aquatic organisms and analysis of uncertainty factors for use in risk assessment. Crit. Rev. Toxicol. 31: 247-284.

Jager, T., E.H.W. Heugens, and S.A.L.M. Kooijman. 2006. Making sense of ecotoxicological test results: towards application of process-based models. Ecotoxicology 15: 305-314.

Kammenga, J.E., M. Busschers, N.M. Van Straalen, P.C. Jepson, and J. Bakker. 1996. Stress induced fitness reduction is not determined by the most sensitive life-history trait. Funct. Ecol. 10: 106-111.

Kammenga, J.E., and R. Laskowski, eds. 2000. Demography in Ecotoxicology, pp 297. John Wiley & Sons, Ltd., Chichester.

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Liess, M., and R. Schulz. 1996. Chronic effects of short-term contamination with the pyrethroid insecticide fenvalerate on the caddisfly Limnephilus lunatus. Hydrobiologia 324: 99-106.

McCauley, E., W.W. Murdoch, and S. Watson. 1988. Simple models and variation in plankton densities among lakes. Am. Nat. 132: 383-403.

Murdoch, W.W., and E. McCauley. 1985. Three distinct types of dynamics shown by a single planktonic system. Nature 318: 628-630.

Newman, M.C. 2001. Population Ecotoxicology. John Wiley & Sons Inc. OECD. 1998. Report of the OECD Workshop on Statistical Analysis of Aquatic

Toxicity Data. OECD Environmental Health and Safety Publications 10: 1-133.

Pieters, B.J., T. Jager, M.H.S. Kraak, and W. Admiraal. 2006. Modeling responses of Daphnia magna to pesticide pulse exposure under varying food conditions: intrinsic versus apparent sensitivity. Ecotoxicology 15: 601-608.

Pieters, B.J., and M. Liess. 2006a. Maternal nutritional state determines the sensitivity of Daphnia magna offspring to short-term fenvalerate exposure. Aquat. Toxicol. 76: 268-277.

Pieters, B.J., and M. Liess. 2006b. Population developmental stage determines the recovery potential of Daphnia magna populations after fenvalerate application. Environ. Sci. Technol. 40: 6157-6162.

Pieters, B.J., A. Paschke, S. Reynaldi, M.H.S. Kraak, W. Admiraal, and M. Liess. 2005. Influence of food limitation on the effects of fenvalerate pulse exposure on the life history and population growth rate of Daphnia magna. Environ. Toxicol. Chem. 24: 2254-2259.

[SANCO] Sante des Consommateurs. 2001. FOCUS surface water scenarios in the EU evaluation process under 91/414/EEC. European Commission, Health & Consumer Protection Directorate-General, SANCO/4802/2001 rev. 1. Brussels (BE).

Stark, J.D., and J.E. Banks. 2003. Population effects of pesticides and other toxicants on arthropods. Annu. Rev. Entomol. 48: 505-519.

Stark, J.D., and U. Wennergren. 1995. Can population effects of pesticides be predicted from demographic toxicological studies? J. Econ. Entomol. 88: 1089-1096.

Van den Brink, P.J., E. Van Donk, R. Gylstra, S.J.H. Crum, and T.C.M. Brock. 1995. Effects of chronic low concentrations of the pesticides chlorpyrifos and atrazine in indoor freshwater microcosms. Chemosphere 31: 3181-3200.

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Van den Brink, P.J., R.P.A. Van Wijngaarden, W.G.H. Lucassen, T.C.M. Brock, and P. Leeuwangh. 1996. Effects of the insecticide Dursban 4E (active ingredient chlorpyrifos) in outdoor experimental ditches: II. Invertebrate community responses and recovery. Environ. Toxicol. Chem. 15: 1143-1153.

Van Vlaardingen, P.L.A., and E.M.J. Verbruggen. 2006. Guidance for the derivation of environmental risk limits within the framework of the project ‘International and National Environmental Quality Standards for Substances in The Netherlands’ (INS). National Institute for Public Health and the Environment, Bilthoven, The Netherlands. RIVM Report no. 601 501 031.

Walthall, W.K., and J.D. Stark. 1997. A comparison of acute mortality and population growth rate as endpoints of toxicological effect. Ecotox. Environ. Safe. 37: 45-52.

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Summary The aim of the present thesis was to examine population responses to pesticides under variable environmental conditions. Exposure of non-target organisms to pesticides is characterized by an extreme temporal variability, since modern pesticides have a high biological activity, relatively short half-lifes and are applied in an episodical manner. The response of populations to pesticides is also highly variable, partly due to food level dependent differences in the condition of individuals and the structure of populations. The objectives of this study were therefore to determine the combined effects of food limitation and pesticide pulse exposure on populations of the model species Daphnia magna. A process-based model was applied for a better understanding of the dynamic population responses. In the last set of experiments the responses of cohorts and mixed Daphnia magna populations were compared.

This thesis departed from analyzing the influence of food limitation on the effects of fenvalerate pulse exposure on the life-history and population growth rate (r) of Daphnia magna cohorts (Chapter 2). Neonates were subjected to a 24 h pesticide pulse exposure under a high and a low food condition and were subsequently monitored for 21 days. Chemical analysis showed that after 1 h, the nominal fenvalerate concentrations were reduced by 50 to 66%. Fenvalerate decreased survival and growth in the week following the pulse exposure. Age at first reproduction increased, with consequent adverse effects on cumulative reproduction per living female and, therefore, on r. Thus, a short-term exposure to fenvalerate caused a long-term reduction of r as a result of an increased mortality and a delay in development. Low food conditions exacerbated the effects of the fenvalerate exposure on juvenile survival and growth during the first week. This caused a much stronger reduction in r under low food conditions. We concluded that a pulsed fenvalerate exposure may result in long-term reduction of r that can be predicted only with environmentally relevant toxicity tests.

The combined effects of food limitation and pesticide exposure on neonates triggered questions on the effect of maternal food limitation on population responses to pesticide exposure. Chapter 3 describes the results of transgenerational life-cycle experiments where the effects of varying maternal food level on subsequent offspring performance and sensitivity to a pulsed fenvalerate exposure were studied. At low nutritional supply, Daphnia produces fewer but larger offspring, which are less sensitive to acute chemical stress. It was

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hypothesized that the change in the reproductive strategy may also alter the chronic sensitivity of offspring to pesticide exposure, which are usually applied as a single dose. Therefore the influence of maternal feeding condition on (1) offspring size and successive performance, and (2) sensitivity to a 24 h pesticide exposure was evaluated. Daphnia magna cultures were maintained at a high and a low food level. The two types of offspring were subsequently used in life-table response experiments, which were performed under low food conditions. Results showed that the low maternal food condition (compared to high) increased the offspring size at time of birth, reduced age at first reproduction and increased reproductive output, which jointly enhanced offspring fitness as estimated by the population growth rate (r). Fenvalerate exposure in combination with the low maternal food level caused (compared to high) a strong decrease in acute sensitivity of neonates, which was generally also observed for chronic endpoints. Hence, population growth rate of daphnids originating from poorly fed mothers was less reduced than that of daphnids originating from high fed mothers.

It remained unclear in the chapters 2 & 3 whether the food-dependent fenvalerate toxicity was due to differences in accumulation kinetics and resource allocation, or resulted from a changed intrinsic sensitivity of daphnids. Therefore, it was hypothesized in Chapter 4 that food level only affects resource allocation and accumulation kinetics of a toxicant, while the intrinsic sensitivity to this toxicant remained the same for all food regimes. This hypothesis was investigated using the mathematical model DEBtox, based on the theory of Dynamic Energy Budgets, re-examining the results of the chapters 2 & 3. The model described the effects of the time-varying exposure well, and indicated that when the animals did not die from exposure to fenvalerate, full reversibility of toxic effects was possible, allowing a complete recovery of the populations. Results revealed furthermore that the data obtained from both studies could be described by the same NECs for survival and assimilation, killing rate and tolerance concentration. It is therefore concluded that food-dependent fenvalerate toxicity could be explained by altered toxicokinetics and resource allocation, but not by changes in the intrinsic sensitivity of the daphnids. This study implies that the effect of pesticide application in the field depends on the trophic state of the receiving water body and that full recovery of survivors is possible after fenvalerate application.

Chapter 5 verified the results obtained for cohorts with observations on mixed Daphnia magna populations in different stages of their development. To this

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purpose, we investigated the responses of Daphnia magna populations to pulsed exposures of the pyrethroid insecticide fenvalerate applied during an early and a late stage of population development, and analyzed the dynamics of the subsequent recovery. A novel digital observation technique was used to describe the size and numbers of animals. High fenvalerate concentrations caused high mortality rates during exponential population growth as well as during the food-limited stationary phase. However, recovery of populations took considerably longer in the stationary phase than in populations growing exponentially. The poor nutritional and reproductive state of food-deprived adults was indicated as the main cause of the slow recovery of populations. It was argued that populations operating at the carrying capacity of their environment are vulnerable to toxicant-induced disturbances to an extent not predictable from observations on exponentially growing populations such as are commonly used in ecotoxicology.

In conclusion the present study stresses the use of population growth rate (r) as an essential endpoint for ecotoxicity testing. At odds with standard laboratory toxicity tests, the present thesis showed that a post-exposure observation period is crucial to detect delayed and cumulative effects, as well as recovery. Moreover, this study demonstrated for the first time that the age structure of daphnid populations is a key factor for the response to pesticides, and indicated that the developmental stage of a population determines its recovery potential. It is concluded that heterogeneous populations consisting of individuals differing in developmental stage are an important subject of ecotoxicity testing, and that such testing is facilitated by rapid non-destructive optical registration of populations. The present study showed that the regime of exposure to an insecticide and the time dependent effect expression are key factors determining the overall effect on a test population. Consequently, time explicit exposure and time explicit effect determination are key elements for an improved risk assessment of pesticides. Samenvatting Het doel van deze studie was om de reactie van populaties op blootstelling aan pesticiden onder van nature variërende omstandigheden te onderzoeken. De blootstelling van organismen aan pesticiden wordt gekarakteriseerd door een grote temporele variabiliteit, veroorzaakt door de hoge biologische activiteit van moderne pesticiden, hun relatief korte halfwaardetijd, en hun vaak eenmalige toepassing. De reactie van populaties op blootstelling aan pesticiden is ook zeer

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variabel. Dit wordt mede veroorzaakt door voedselniveau afhankelijke verschillen in de conditie van individuen en de leeftijdsopbouw van de populatie. De doelstelling van deze studie was daarom om de gecombineerde effecten van voedseltekort en kortstondige blootstelling aan pesticiden op populaties van het model testorganisme Daphnia magna te bepalen. Een procesgeoriënteerd model werd gebruikt om beter inzicht te krijgen in de dynamiek van de reacties van populaties op blootstelling aan pesticiden. In de laatste experimenten werden de reacties van cohorten en echte Daphnia magna populaties met elkaar vergeleken.

Het onderzoek begon met de analyse van de invloed van voedseltekort op de effecten van een kortstondige fenvaleraat blootstelling op levenscyclusparameters en populatiegroeisnelheid (r) van Daphnia magna cohorten (hoofdstuk 2). Neonaten werden 24 uur blootgesteld aan fenvaleraat bij een hoog en een laag voedselniveau, waarna hun ontwikkeling gedurende 21 dagen werd gevolgd. Fenvaleraat verminderde de overleving en remde de groei in de eerste week na de blootstelling. De dag waarop de eerste reproductie plaatsvond werd vertraagd, met nadelige gevolgen voor de cumulatieve reproductie per individu, en daarmee ook voor de populatiegroeisnelheid. Een kortstondige blootstelling aan fenvaleraat veroorzaakte dus op langere termijn een lagere populatiegroeisnelheid, door verhoogde sterfte en remming van de ontwikkeling. Voedseltekort verergerde de effecten van fenvaleraat op juveniele overleving en groei gedurende de eerste week en daarmee ook op de populatiegroeisnelheid. Er werd geconcludeerd dat effecten van een kortstondige fenvaleraatblootstelling op de populatiegroeisnelheid op langere termijn alleen voorspeld kunnen worden met behulp van toxiciteitstesten waarin van nature variërende omstandigheden daadwerkelijk getest worden.

De gecombineerde effecten van voedseltekort en pesticiden op Daphnia magna cohorten beschreven in hoofdstuk 2 leidde tot de vraag of maternale voedseltekorten de reactie van populaties op een blootstelling aan pesticiden beinvloeden. Er was reeds bekend dat bij een laag voedselniveau Daphnia minder, maar grotere nakomelingen voortbrengt, die minder gevoelig zijn voor acute chemische stress. Er werd verondersteld dat deze verandering in reproductieve strategie mogelijkerwijs ook de chronische gevoeligheid van de nakomelingen voor kortdurende blootstelling aan pesticiden zou verminderen. Op grond van deze veronderstelling werd in hoofdstuk 3 de invloed onderzocht van maternale voedseltekorten op (1) de grootte van de nakomelingen en hun prestaties, en (2) hun reactie op een 24 uur durende blootstelling aan pesticiden.

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Daphnia magna werd gekweekt bij een hoog en een laag voedselniveau. Nakomelingen van deze twee kweken werden vervolgens gebruikt in levenscyclusexperimenten, die uitgevoerd werden bij het lage voedselniveau. De resultaten lieten zien dat door een laag maternaal voedselniveau de grootte van de net geboren nakomelingen toenam, de dag waarop de eerste reproductie plaatsvond vervroegde en de cumulatieve reproductie toenam, wat leidde tot een hogere populatiegroeisnelheid. Maternaal voedseltekort veroorzaakte een grote vermindering in gevoeligheid van de nakomelingen voor fenvaleraat. Dus de populatiegroeisnelheid van watervlooien afkomstig van slecht doorvoede ouders werd minder geremd door pesticiden dan die van watervlooien afkomstig van goed doorvoede ouders.

Het bleef onduidelijk in de hoofdstukken 2 & 3 of de voedselniveau afhankelijke gevoeligheid voor fenvaleraat werd veroorzaakt door verschillen in accumulatiekinetiek van de toxicant en energieallocatie binnen het organisme, of door een veranderde gevoeligheid van de watervlooien. In hoofdstuk 4 werd daarom aangenomen dat voedselconditie alleen de energieallocatie en accumulatiekinetiek beinvloedt, terwijl de gevoeligheid van de watervlooien voor deze toxicant hetzelfde blijft bij alle voedselniveaus. Deze hypothese werd onderzocht met behulp van het mathematische model DEBtox (gebaseerd op de theorie van Dynamic Energy Budgets), waarmee de uitkomsten van de hoofdstukken 2 & 3 opnieuw geanalyseerd werden. Het model beschreef de effecten van de in de tijd variërende blootstelling goed en gaf aan dat wanneer er geen sterfte optrad na blootstelling aan fenvaleraat, volledige herstel van de populaties mogelijk was. De resultaten wezen verder uit dat de data van beide studies beschreven konden worden met dezelfde no effect concentrations (NECs) voor de modelparameters overleving en assimilatie, en met dezelfde sterftesnelheid en tolerantieconcentratie. Er werd daarom geconcludeerd dat de voedselafhankelijke gevoeligheid voor fenvaleraat verklaard kan worden door veranderde toxicokinetiek en energieallocatie, en niet door veranderingen in de gevoeligheid van de watervlooien.

Hoofdstuk 5 verifiëerde de resultaten die verkregen waren in studies aan cohorten (hoofstukken 2-4) met waarnemingen aan echte Daphnia magna populaties die verschilden in leeftijdsopbouw. Hiertoe werden de reacties van Daphnia magna populaties onderzocht op een kortstondige fenvaleraat blootstelling gedurende een vroege en een late fase van populatieontwikkeling. De dynamiek van het daarop volgende herstel werd geanalyseerd. Een nieuw

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ontwikkeld digitaal observatiesysteem werd toegepast om de grootte en de aantallen dieren te monitoren. Hoge fenvaleraat concentraties veroorzaakten hoge sterfte in zowel de exponentiele groeifase als in de voedselgelimiteerde stationaire fase. Echter, het herstel van de populaties duurde aanzienlijk langer in de stationaire fase dan in de exponentieel groeiende populaties. De slechte conditie van de hongerende adulten werden geïdentificeerd als de belangrijkste oorzaak van het langzame herstel van deze populaties. Er werd beargumenteerd dat populaties in de stationaire fase, die het maximale draagvermogen van hun omgeving hebben bereikt, kwetsbaarder zijn voor toxicanten, in een mate die niet voorspelbaar is op basis van observaties aan exponentieel groeiende populaties die normaal gesproken gebruikt worden in ecotoxiciteitexperimenten.

De huidige studie benadrukt het gebruik van populatiegroeisnelheid (r) als een essentiële parameter die in ecotoxiciteittesten moet worden meegenomen. Dit proefschrift toonde aan dat door het ontbreken van een observatieperiode ná de blootstelling in gestandaardiseerde toxiciteittesten vertraagde- en cumulatieve effecten en herstel niet waargenomen worden. Deze studie suggereerde verder dat de effecten van een pesticidetoepassing in het veld afhangen van het trofische niveau van het water, en dat volledig herstel van de overlevenden mogelijk is nadat de concentraties pesticiden in het omringende milieu zijn gedaald. Deze studie toonde tevens voor de eerste keer aan dat de leeftijdsstructuur van watervlooienpopulaties bepalend is voor de reactie op pesticiden, en voor het herstel na de blootstelling. Geconcludeerd werd dat heterogene populaties die uit individuen bestaan met een verschillende ontwikkelingsfase belangrijk inzicht verschaffen in de reactie van populaties op blootstelling aan pesticiden. Ecotoxiciteittesten kunnen verder verbeterd kunnen worden door het gebruik van snelle niet-destructieve optische methoden voor het monitoren van populaties, zoals toegepast in dit onderzoek. Deze studie toonde aan dat het blootstellingsregime van een pesticide en de tijdsafhankelijke expressie van effecten essentieel zijn om het totale effect op een populatie te kunnen vaststellen. Tijdsexpliciete blootstelling en tijdsexpliciete effectbeoordeling zijn dus elementaire factoren voor een verbeterde risicobeoordeling van pesticiden.

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Acknowledgements The research described in this thesis was financially supported by the the Department of System Ecotoxicology, Centre for Environmental Research (UFZ). I am very grateful for the supervision by my promotores Wim Admiraal and Matthias Liess. Matthias, during the first years of my research you were very important and you have taught me in a short time what science is all about. Although we had different opinions about a lot of things, you have made me a more complete and robust person. Wim, you were quite important in the second part of my research. Every time I thought things couldn’t get any worse you showed me the light at the end of the tunnel. And of course I would like to show my deep appreciation to my co-promotor, Michiel Kraak, your relentless enthusiasm and ‘overdrive’ will always be an inspiration to me. Thanx! Finally, I would like to thank A.M. Breure, T.C.M. Brock, A.J. Hendriks, R.W.P.M. Laane, M.W. Sabelis, N.M. van Straalen and P. de Voogt for participating in the Ph.D. committee and for providing useful comments.

I thank Dick de Zwart for helping out with the cover of this dissertation and Harm van der Geest who contributed in numerous ways. I am especially greatful for the help of Tjalling de Jager with the DEBtox modeling.

I enjoyed the companionship of my UFZ colleagues: Sebastián Reynaldi (private guru and Shakira connoisseur), Peter von der Ohe (coffee breaks and pc-wizard), Matthias Grote (organizing parties), and Carola Schriever (good spirits). Thumbs up for Klaus Seyfarth and Ingrid Ränker for their practical help in the laboratory. The technical assistance of Uwe Schröter with chemical analyses should also not be forgotten, never let a biologist pretend to be a chemist… I thank Steffen Wahrendorf for his support with the image analysis system and the microcosms.

Being in Leipzig was a great experience for me, mainly due to some very special people. Living together with Claudia and Steffi in the über-WG of the Südvorstadt was incredible and my safe-house. Wir haben das Recht auf Faulheit und Rotkäppchen Sekt! I thank Stefan, Steffi and Ioanna for just being there. I could always talk about anything what was on my mind and you made me feel at home in Leipzig, I think we will always stay in contact. Unfortunately, I left some people behind in The Netherlands which was definitely not easy in the beginning. However, all of you were continuously on my mind. Odin, Marcel, Niels & Marc, Michel, and Erik, I enjoyed your company during my short visits in Haarlem.

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Fons and Sander, your brief but intense visits in Leipzig were also perfect to relax and go to pubs and concerts.

I greatly appreciate the unconditional love, patience and support from my parents, without you there would not have been a thesis. But most of all, I love you Eugénie. Meeting you, my soulmate, in Leipzig was probably one of the most important things which have happened to me. Although it was not easy to live together and do our Ph.D. projects simultaneously, we made it through. We will always be together!

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Publications list Bleeker, E.A.J., B.J. Pieters, S. Wiegman, and M.H.S. Kraak. 2002. Comparative

(photoenhanced) toxicity of homocyclic and heterocyclic PAHs. Polycycl. Aromat. Comp. 22: 601-610.

Liess, M., B.J. Pieters, and S. Duquesne. 2006. Long-term signal of population disturbance after pulse exposure to an insecticide: rapid recovery of abundance, persistent alteration of structure. Environ. Toxicol. Chem. 25: 1326-1331.

Pieters, B.J., T. Jager, M.H.S. Kraak, and W. Admiraal. 2006. Modeling responses of Daphnia magna to pesticide pulse exposure under varying food conditions: intrinsic versus apparent sensitivity. Ecotoxicology 15: 601-608.

Pieters, B.J., and M. Liess. 2006. Maternal nutritional state determines the sensitivity of Daphnia magna offspring to short-term fenvalerate exposure. Aquat. Toxicol. 76: 268-277.

Pieters, B.J., and M. Liess. 2006. Population developmental stage determines the recovery potential of Daphnia magna populations after fenvalerate application. Environ. Sci. Technol. 40: 6157-6162.

Pieters, B.J., A. Paschke, S. Reynaldi, M.H.S. Kraak, W. Admiraal, and M. Liess. 2005. Influence of food limitation on the effects of fenvalerate pulse exposure on the life history and population growth rate of Daphnia magna. Environ. Toxicol. Chem. 24: 2254-2259.

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Curriculum Vitae Barry Pieters was born on March 25, 1976, in Haarlem. After finishing high school (Schoter Scholengemeenschap in Haarlem), he started to study Geology and Physical Geography in 1995 at the Vrije Universiteit (VU) in Amsterdam, followed by the specialization Environmental Sciences in the second year at the VU. Next, he started to study Biology in 1997 at the University of Amsterdam (UvA). During the course of his study he conducted a M.Sc. research (Toxicity and phototoxicity of Polycyclic Aromatic Hydrocarbons and N-heterocyclic analogues) at the Department of Aquatic Ecology and Ecotoxicology (UvA) and a literature study (Addressing uncertainties concerning Species Sensitivity Distributions (SSDs) of herbicides) at Alterra Green World Research (Agricultural University of Wageningen). The second M.Sc. research was carried out in Leipzig, Germany, where he investigated the effects of pesticides on the Cladoceran Daphnia magna at the Centre for Environmental Research (UFZ). He graduated in Biology with the specialization aquatic ecotoxicology in 2002 at the University of Amsterdam. After graduation he continued to work at the Department of System Ecotoxicology (UFZ) where he investigated the acute and chronic effects of pesticides on Daphnia magna and examined the influence of environmental variables such as food availability and exposure regime, which constituted the framework of the present thesis. Currently, he is working at the National Institute for Public Health and the Environment (RIVM) in projects supported by the Ministry of Housing, Spatial Planning and the Environment (VROM) with the general aim of improving the implementation of the European Water Framework Directive in The Netherlands.

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