Reproductive toxicity of dietary zinc to Daphnia magna

Aquatic Toxicology 70 (2004) 233–244

Reproductive toxicity of dietary zinc to Daphnia magna K.A.C. De Schamphelaerea,∗ , M. Canlia , V. Van Lierdeb , I. Forreza , F. Vanhaeckeb , C.R. Janssena a

Laboratory of Environmental Toxicology and Aquatic Ecology, Faculty of Applied Biological Sciences, Ghent University, Jozef Plateaustraat 22, B-9000 Gent, Belgium b Laboratory of Analytical Chemistry, Institute for Nuclear Sciences, Ghent University, Proeftuinstraat 86, B-9000 Gent, Belgium Received 9 August 2004; received in revised form 14 September 2004; accepted 20 September 2004

Abstract Regulatory assessments of metals in freshwaters are mostly based on dissolved metal concentrations, assuming that toxicity is caused by waterborne metal only. Little attention has been directed to the toxicity of dietary metals to freshwater invertebrates. In this study the chronic toxicity of dietary zinc to Daphnia magna was investigated. The green alga Pseudokirchneriella subcapitata was exposed for 64 h to a control and three dissolved zinc concentrations, i.e. 23, 28 and 61 ␮g L−1 , resulting in internal zinc burdens in the algae of 130, 200, 320 and 490 ␮g g−1 dry weight, respectively. These algae were used as a food source in chronic, 21-day bioassays with D. magna in a test medium to which no dissolved zinc was added. None of the treatments resulted in effects on feeding rates or somatic growth of D. magna. In contrast, a significant 40% decrease of total reproduction (number of juveniles per adult) was observed in the 28 and 61 ␮g L−1 treatments. Time to first brood was not affected, whereas the mean brood size and the fraction of reproducing parent daphnids were reduced from the second brood onwards and the magnitude of these reductions increased with each subsequent brood. The reduced reproduction was accompanied with an elevated zinc accumulation in the 61 ␮g L−1 treatment only, suggesting that total body burden is no good indicator of dietary zinc toxicity. Overall our data suggest that dietary zinc specifically targets reproduction in D. magna through accumulation in particular target sites, possibly cells or tissues where vitellogenin synthesis or processing occur. Further, our data illustrate that the potential importance of the dietary exposure route should be carefully considered and interpreted in regulatory assessments of zinc. © 2004 Elsevier B.V. All rights reserved. Keywords: Daphnia magna; Dietborne exposure; Zinc; Biotic ligand model

1. Introduction ∗

Corresponding author. Tel.: +32 9 2643764; fax: +32 9 2643766. E-mail address: [email protected] (K.A.C. De Schamphelaere). 0166-445X/$ – see front matter © 2004 Elsevier B.V. All rights reserved. doi:10.1016/j.aquatox.2004.09.008

Regulatory assessments of metal toxicity in freshwaters are mostly based on dissolved metal concentrations, assuming that toxicity is caused by waterborne metal only (Janssen et al., 2000). In this context the

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biotic ligand model is a useful tool to predict changes in bioavailability and toxicity of waterborne metal as a function of the prevalent physico-chemical water characteristics (e.g., Di Toro et al., 2001; Paquin et al., 2002; De Schamphelaere and Janssen, 2002; Heijerick et al., 2002a, 2004). However, this approach does not take into account the potential toxicity of dietborne metal. Until now dietary metal exposure studies have almost exclusively focused on uptake and assimilation (Wang and Fisher, 1999). Although some authors have investigated effects of dietary metal exposure in fish species (e.g. Clearwater et al., 2002), too little attention has been directed to effect studies with invertebrate species. Hook and Fisher (2001a,b, 2002) demonstrated reproductive toxicity in the marine copepods Acartia tonsa and Acartia hudsonica, fed with the diatom Thallasiosira pseudonana contaminated with Hg, Cd, Zn, Ag and Mn, and in the freshwater cladocerans Simocephalus sp. and Ceriodaphnia dubia, fed with the green alga Chlorella vulgaris contaminated with Ag. They attribute this effect to disturbance of vitellogenesis as evidenced by the reduction of ovary development and the decreased yolk protein content of the eggs. Conversely, De Schamphelaere and Janssen (2004) demonstrated an increased reproduction and growth in the freshwater cladoceran Daphnia magna fed with Cucontaminated algae Pseudokirchneriella subcapitata. The effect of dietborne Zn to freshwater invertebrates has to our knowledge not been investigated yet. D. magna has, however, been shown to take up dietary Zn from contaminated diets (Chlamydomonas reinhardtii and Scenedesmus obliquus), which may indicate potential toxic effects (Guan and Wang, 2004). In the present study the green alga P. subcapitata was exposed for 64 h to a control and 20, 30 and 60 ␮g L−1 dissolved Zn. These algae were subsequently used as food for D. magna in 21-day bioassays in which bioaccumulation, feeding rate, growth and some reproductive parameters were determined to investigate the chronic toxicity of dietary zinc. Special care was taken to accurately monitor zinc concentrations in algae, daphnids and exposure solutions in order to be able to interpret and assess the importance of potentially confounding factors such as changes in zinc burdens in algae during storage or leaching of zinc from algal food into the daphnid exposure solution, i.e. providing a waterborne exposure pathway next to the dietary one.

2. Materials and methods 2.1. Algae exposure P. subcapitata starter cultures were obtained from the Culture Collection of Algae and Protozoa (CCAP 278/4, Ambleside, UK). Stock cultures of algae were maintained in carbon-filtered aerated tap water (Gent, Belgium) supplemented with the modified Provasoli’s ES enrichment (Bold and Wynne, 1978) at 1/2 strength, and 1.4 mg L−1 FeSO4 ·7H2 O, 15 mg L−1 NaH2 PO4 ·2H2 O, 150 mg L−1 NaNO3 and 2.35 mg L−1 MnCl2 ·4H2 O. These stock cultures were kept at 20 ± 1 ◦ C under continuous light (240 ␮E m−2 s−1 ) and were continuously aerated. Only algae exhibiting exponential growth were used for starting the exposures. Nominal zinc concentrations tested were 0, 20, 30 and 60 ␮g L−1 . For each zinc concentration, 10 L of test medium was prepared, spiked with zinc and transferred into a 12 L polyethylene bag. The test medium used is that described by OECD (1984), with the modifications described below. The concentrations of Ca and Mg were adjusted to 2 and 0.5 mM, respectively, using CaCl2 and MgCl2 . pH was buffered using 750 mg L−1 MOPS (3-Nmorpholino-propane sulfonic acid, Sigma–Aldrich, Steinheim, Germany) and then adjusted to 7 by adding dilute NaOH. The use of MOPS in toxicity testing is discussed in De Schamphelaere et al. (2004a). The strong metal-chelator ethylene-diamine-tetra-acetic acid (EDTA) was replaced with 32 ␮g DOC L−1 (humic acid, Aldrich) for reasons mentioned by Heijerick et al. (2002b). The spiked test media were equilibrated for 48 h at 20 ◦ C before the exposure. At the start of the exposure, each bag was inoculated with 5 × 105 cells mL−1 . Exposures were carried out under continuous illumination (240 ␮E m−2 s−1 ). The temperature in the media during the exposure period was 20 ◦ C. The cell density was recorded after 64 h with the aid of an electronic particle counter (Beckman Coulter Z1, Beckman, Miami, FL, USA). After 64 h, algae were harvested by centrifuging cell suspensions using a continuous-flow centrifuge (IEC Chemical centrifuge, International Equipment Company, USA) with a volume of 300 mL at a flow rate of 2.5 mL s−1 and a g-force of 1000 × g. The supernatant was carefully pipetted off and the obtained pellets were suspended in a specified volume of supernatant (i.e. the same solu-

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tion chemistry and zinc content as they were exposed to) to yield a cell density of 8 × 107 cells mL−1 . Finally, suspensions of the cells were stored in darkness at 4 ◦ C throughout the daphnid testing period. Resuspension in the supernatant was done to minimize changes in zinc burdens of the algae during storage of the suspensions.

weighed again. Dry weight of algal cells was obtained by difference between weight before and after drying. Weighing was performed to the nearest 10 ␮g (Mettler Toledo balance, model AX105, Greifensee, Germany). Algal dry weight was determined on the first and 21st day of the daphnid exposure.

2.2. Concentration of zinc in algal food

2.4. Exposure of D. magna to dietary zinc

Internal and external zinc concentrations of the algae were determined on the first and the last day of the daphnid exposure. After determining cell density in the cell suspensions (as above), 2 mL of cell suspension was transferred into a polyethylene vial and centrifuged for 10 min at 1000 × g and the supernatant was discarded. For each zinc concentration, six replicates were analyzed. Three replicates were taken for determination of total zinc (=external + internal zinc) and three for determination of the internal zinc concentration. External zinc was calculated as the difference between total and internal. For the determination of total zinc concentration, the pellets were destructed in 300 ␮L of 14 N HNO3 (p.a. VWR international, Leuven, Belgium) with the aid of a microwave oven (2 min at 90 W, 2 min at 180 W, 2 min at 270 W, 2 min at 360 W, with cooling to room temperature between subsequent destruction steps). The internal zinc concentration was determined by first washing the pellets with 2 mL of a 5 mM Na2 EDTA solution to remove external (surface-bound) zinc. The suspensions were then manually shaken for 10 min and subsequently centrifuged at 1000 × g for 10 min. The supernatant was discarded and 300 ␮L of 14 N HNO3 was added to the pellets for destruction as above. After destruction of the pellets 2.7 mL of deionized water was added and the zinc content of the samples was determined to yield internal and total zinc concentrations of the algae (see below for more information about analytical techniques). Internal and total zinc concentrations were calculated as g Zn cell−1 .

Test organisms originated from a healthy D. magna clone which has been cultured in the laboratory under standardized conditions in M4 medium (Elendt and Bias, 1990) for several years. Tests were performed in a synthetic freshwater containing 2 mM CaCl2 , 0.5 mM MgSO4 , 0.75 mM NaHCO3 and 0.078 mM KCl. 3-N-Morpholino-propane sulfonic acid (MOPS; 750 mg L−1 ) was added to the test medium and the pH was adjusted to 7 using dilute NaOH. The exposure media for D. magna and P. subcapitata are the same with respect to pH and hardness, thus improving the ecological relevance of this study. Chronic bioassays were performed according to OECD guideline 211 (OECD, 1998). At the start of each test, 10 juvenile animals (<24 h old, 8 ␮g dry weight) per concentration were transferred individually to polyethylene cups containing 50 mL of the test medium (i.e. 10 replicates of one organism per concentration). Daphnids in the different dietary treatments were fed with the algal cells exposed to a control, 20, 30 and 60 ␮g Zn L−1 (see above). Every day the daphnids were fed with 8 × 106 cells per day from day 0 to day 6, 12 × 106 cells per day from day 7 to day 8 and 16 × 106 cells per day from day 9 to day 20. On days 2, 5, 7, 9, 12, 14, 16 and 19 the medium was renewed. Each day the number of produced juveniles was noted. After 21 days the parent daphnids were collected for determination of dry weight and the zinc body burden. This was accomplished by washing the daphnids in control test medium for 10 min (to remove adhered particles) and subsequently in a 5 mM Na2 EDTA solution for 20 min (to remove Zn bound to the exoskeleton). Daphnids were then dried individually for 48 h at 40 ◦ C and were weighed to the nearest 1 ␮g. Daphnids from each treatment were then pooled and digested in 300 ␮L of 14 N HNO3 . The 2.7 mL of double deionized water was added to the digest and the Zn content of the sample was analyzed using flame atomic absorption spectrometry (AAS, see below). The Zn content of the samples was used to cal-

2.3. Dry weight of algal cells Dry weight of algal cells was obtained by filtering a known amount of algal cells through a moistened, dried (48 h at 40 ◦ C) and pre-weighed 0.8 ␮m filter (Supor 800, Pall Corporation, Ann Arbor, MI, USA). Filters + algae were then dried for 48 h at 40 ◦ C and

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culate internal body burdens of the daphnids, expressed as ␮g Zn g−1 dry wt. 2.5. Analyses of zinc in solutions, algae, and daphnids Next to the analysis of the total and internal zinc concentrations of P. subcapitata and of the body burden of D. magna, intensive monitoring of zinc in exposure solutions was performed. Samples from test media, filtered through a 0.45 ␮m filter (Gelman Sciences, Ann Arbor, MI, USA) were taken were taken daily during the algal exposures (triplicate) and at every medium renewal during the daphnid exposures (a pooled sample from 10 replicate vessels). In the daphnid exposures this was performed to assess the potential leaching of zinc from the Zn-contaminated algal food, which might result in significant exposure of D. magna to zinc via the water. The leaching of metals from food in dietary toxicity studies is recognized as an important confounding factor (Clearwater et al., 2002). The water samples were immediately acidified with 0.14 N HNO3 and stored at 4 ◦ C until analysis. All samples (water samples, and digests of algae and daphnids) were first analyzed using flame AAS (SpectrAA100, Varian, Mulgrave, Australia). Calibration standards (Sigma–Aldrich, Steinheim, Germany) and a reagent blank were analyzed with every ten samples. The practical quantification limit was about 20 ␮g L−1 as determined with the method described in Clesceri et al. (1998). Two certified reference samples, TMDA-62 and TM-25.2 (National Water Research Institute, Burlington, Ont., Canada) with certified Zn concentrations (mean ± 95% confidence interval) of 110 ± 15.5 and 24 ± 4.6 ␮g L−1 , respectively, were analyzed at the beginning and end of each series of Zn measurements. Measured values were always within 10% of the certified value. With each determination of algal zinc concentration or daphnid zinc body burden, a reference plankton sample (BCR-414, Institute for Reference Materials and Measurements, Geel, Belgium) was analyzed and zinc levels were always within 10% of the certified value. When flame AAS yielded concentration readings below 20 ␮g L−1 , samples were analyzed with an inductively coupled plasma-sector field-mass spectrometer (ICP-SF-MS) (Giessmann and Greb, 1994), which

offers a lower detection limit, i.e. 0.5 ␮g L−1 than flame AAS. In practice this was only necessary for exposure solution samples. 64 Zn, 66 Zn and 59 Co were monitored, with Co (final concentration: 50 ␮g L−1 ) acting as the internal standard, correcting for matrix effects, signal drift and instrument instability (Vanhaecke et al., 1996). The ICP-SF-MS unit used (Finnigan MAT Element, Bremen, Germany) was operated in the medium resolution mode (mass resolution = 3000) to eliminate the spectral overlap of SO2 + and Zn+ signals and is fitted with a guard electrode, surrounding the ICP torch, to achieve the highest sensitivity possible (Appelblad et al., 2000). With this method too, measurements of reference samples were all within 10% of the reference values. 2.6. Data treatment and statistics All data are reported as mean ± 1 standard error (S.E.). Statistical comparisons between exposures and controls were carried out with the Mann–Whitney Utest (Siegel and Castellan, 1988) using Statistica 6 software (Statsoft, Tulsa, OK, USA).

3. Results and discussion 3.1. Algae exposure and algal zinc burdens Details on the zinc exposure of algae are summarized in Table 1. The dissolved zinc concentration in the control was 3.7 ± 0.9 ␮g L−1 . Dissolved concentrations in the zinc exposures decreased with time. At the start of the test, 23, 28 and 61 ␮g L−1 Zn was measured in the 20, 30 and 60 ␮g L−1 nominal exposures, respectively. After 64 h of exposure those concentrations had decreased to 9.9, 11 and 31 ␮g L−1 due to uptake and adsorption of zinc by the algae and to zinc adsorption to the exposure bags. Based on mass balance calculations (using data reported in Table 1), the latter process may have resulted in 10–15% loss of dissolved Zn during the 64 h exposure. Cell growth rates of all treatments were all within 3% of eachother, indicating no biologically important difference between the different exposures. The dry weight of Zn-exposed algae was in all cases within 10% of the control dry weight. All dry weights decreased with 18–34% during the 21-day storage due to dark respiration. The dry

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Table 1 Details of the 64 h exposure of P. subcapitata to a control and three zinc concentrations and of the measurements of dry weights and internal zinc concentrations during the storage of the exposed algae (all values are mean ± standard error of mean) Nominal Zn

Control

20 ␮g L−1

30 ␮g L−1

60 ␮g L−1

Measured dissolved Zn at start of exposure (␮g L−1 ) Measured dissolved Zn after 64 h of exposure (␮g L−1 ) Cell density after 64 h (×106 cells mL−1 ) Cell specific growth rate (day−1 )

3.7 ± 0.9a

23

28

61

11

31

3.7 ± 0.9a

9.9

2.23

2.14

2.23

2.23

0.56

0.55

0.56

0.55

Dry weight at start of storage (×10−11 g cell−1 ) Dry weight at end of storageb (×10−11 g cell−1 ) Internal Zn concentration at start of storagec,d (×10−15 g cell−1 )

2.02

1.84

1.93

1.80

1.46

1.50

1.27

1.37

1.8 ± 0.2 W (89 ␮g g−1 )

3.1 ± 0.2 X (170 ␮g g−1 )

4.3 ± 0.2 Y (220 ␮g g−1 )

8.7 ± 0.7 Z (480 ␮g g−1 )

Internal Zn concentration at end of storagec,d (×10−15 g cell−1 )

2.3 ± 0.2 W (160 ␮g g−1 )

3.4 ± 0.2 X (230 ␮g g−1 )

5.2 ± 0.1 Y (410 ␮g g−1 )

6.9 ± 0.2 Z (500 ␮g g−1 )

Mean internal Zn concentration during storaged,e (×10−15 g cell−1 )

2.1 ± 0.2 W (130 ␮g g−1 )

2.9 ± 0.2 X (200 ␮g g−1 )

4.9 ± 0.3 Y (320 ␮g g−1 )

7.2 ± 0.5 Z (490 ␮g g−1 )

a b c d e

Mean ± S.E. (n = 4) of daily samples. End of storage was 21 days after start of storage. Internal concentrations measured during the D. magna exposure (n = 3). Values within a row, followed by a different capital letter are significantly different from one another, two-sided Student’s t-test (P < 0.05). Average of all measurements at start and end of storage (n = 6).

weights of the Zn-exposed cells were within 15% of the control weight. The internal zinc concentrations in the algae varied up to 30% on per cell basis (i.e. as g Zn cell−1 ) during the storage of the concentrated algal suspension, but more, by approximately 100%, on dry weight basis (i.e. as g Zn g−1 dry wt.) (Table 1). However, hardly any overlap (comparing between start and end of the storage) of internal zinc concentrations was observed between the different treatments. Moreover, mean internal zinc in the algae in the different treatments varied by a factor of 3–4 and they were all significantly different from each other. Both findings indicate that the daphnids feeding on algae from different treatments were truly exposed to different levels of dietary zinc. Internal concentrations always accounted for 80–90% of the total zinc associated with the algae (data not shown). However, it is suggested that the externally bound fraction is of little importance with regard to potential dietary toxicity of zinc to D. magna as it has been shown that this fraction is

rapidly desorbed once the algae are introduced as food in the daphnids’ chronic test medium (Wolterbeek et al., 1995). Internal zinc concentrations, expressed as ␮g Zn g−1 dry wt. (Table 1), were in the same range as those reported for C. reinhardtii exposed to Zn concentrations ranging from 2 to 200 ␮g L−1 , i.e. between 27.2 and 280 ␮g g−1 dry wt. (Guan and Wang, 2004). However, at the same zinc exposure concentrations, internal concentrations in Scenedesmus subspicatus were much higher, i.e. up to 7460 ␮g g−1 dry wt. at 200 ␮g L−1 (Guan and Wang, 2004). The differences in Zn-accumulation by different algal species point to the need for future research on the effect of food type (i.e. species) on dietary metal uptake and/or toxicity. 3.2. Daphnia exposure and zinc body burdens As no zinc is added, the dissolved zinc concentrations in the D. magna test media should in principle be close to zero (apart from some background Zn content of deionized water, and slight contamination by precip-

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Table 2 Effects of a 21-day dietary zinc exposure on D. magna growth (as dry wt.), reproduction, algal ingestion rates, zinc body burdens, reproduction, time to first brood, and intrinsic rate of natural increase (rm ), after feeding with P. subcapitata exposed for 64 h to a control and three zinc concentrations (20, 30, and 60 ␮g L−1 ) Nominal Zn in algae exposure Dissolved zinc in D. magna test at renewals (␮g L−1 ) Body burden of parent daphnid after 21 days of exposure (␮g Zn g−1 dry wt.) Algal ingestion rate between day 19 and day 21 of the exposure (×107 cells day−1 ) Dry wt of parent daphnid after 21 days of exposure (␮g) Time to first brood (day) Total reproduction (number of juveniles per parent) rm (day−1 )

Control 4.0 ± 0.7 62.3 1.54 (225 ␮g day−1 )

20 ␮g L−1

30 ␮g L−1

60 ␮g L−1

4.1 ± 0.8

6.4 ± 1.1

5.6 ± 1.0

50.9 1.48 (222 ␮g day−1 )

55.6 1.86 (236 ␮g day−1 )

84.0 1.64 (225 ␮g day−1 )

401 ± 22

401 ± 26 NSa

390 ± 23 NS

410 ± 21 NS

10.1 ± 0.5 45.3 ± 2.8

10.2 ± 0.2 NS 43.8 ± 3.5 NS

10.4 ± 0.2 NS 18.0 ± 3.1***

10.4 ± 0.4 NS 17.4 ± 2.0***

0.266 ± 0.005

0.270 ± 0.005 NS

0.240 ± 0.011**

0.238 ± 0.010**

Error intervals represent standard errors (n = 10). Asterisk indicates significant differences with control (Mann–Whitney U-test). a Not significantly different. ∗∗ P < 0.01. ∗∗∗ P < 0.001.

itation of airborne Zn). However, as a result of adding Zn-contaminated food to the exposure solutions, zinc in the dissolved phase may increase during the period between two test media renewals. If such increases are high enough to cause toxic effects of zinc via the dissolved phase, they may confound the interpretation of the effects of the dietary exposure. Background zinc concentrations in the control test water (before daphnids or algae were introduced) was 1.8 ± 0.8 ␮g L−1 (mean ± standard error, n = 9). Mean (±standard error, n = 9) dissolved zinc concentrations at test media renewal were 4.0 ± 0.7, 4.1 ± 0.8, 6.4 ± 1.1 and 5.6 ± 1.0 ␮g L−1 in the control, 20, 30 and 60 ␮g L−1 exposures, respectively (Table 2). There seems to be a trend towards slightly higher zinc concentrations in the daphnid test media which received food with higher zinc contents. The increases of dissolved zinc in the test media may occur either via desorption or elimination of zinc from the algal cells (Wolterbeek et al., 1995) and/or from elimination by the daphnids that have incorporated dietary zinc in their tissues (Guan and Wang, 2004). However, at such low concentrations no toxic effects on D. magna of the dissolved zinc are expected. Indeed, in the same test medium we have previously observed no effects on daphnid reproduction up to a dissolved Zn concentration of 155 ␮g L−1 (De Schamphelaere et al., 2004).

When examining the effects of the dietary exposures, no daphnid mortality was observed in any of the treatments. Body burdens of 21-day-old daphnids exposed to dietary zinc are reported in Table 2. Control daphnids had a body burden of 62.3 ␮g g−1 dry wt. When daphnids were fed algae exposed to 20 and 30 ␮g L−1 , body burdens were 50.9 and 55.6 ␮g g−1 dry wt., respectively, i.e. within 20% of the control body burden. Perhaps, this indicates the existence of a regulation mechanism in D. magna upon dietary zinc exposure. Regulation of zinc burdens upon waterborne exposure in D. magna has previously been demonstrated by Muyssen and Janssen (2002). The regulation for dietary zinc occurs — most likely — at the uptake side. Indeed, using C. reinhardtii food containing similar Zn concentrations as in our study (i.e. between 27 and 280 ␮g g−1 dry wt.), Guan and Wang (2004) observed a trend of lower zinc assimilation efficiency by D. magna at higher algal zinc burdens but did not note an increase in elimination rate constant (i.e., ∼0.29 day−1 ). When fed algae exposed to 60 ␮g L−1 , the body burden of the daphnids used in our study increased up to 84 ␮g g−1 dry wt. We hypothesize that the down-regulation of assimilation efficiency in this treatment did no longer compensate the increased zinc burden of the food and that consequently increased body burdens were observed.

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3.3. Effects of dietary zinc on D. magna Table 2 also summarizes the effects of dietary zinc on growth and reproduction. Daphnid growth was not affected by any treatment, as evidenced by the dry weights of the 21-day-old daphnids which were not significantly different. Conversely, total reproduction was severely affected when daphnids were fed algae which were exposed to 30 and 60 ␮g L−1 , but not when fed algae grown in 20 ␮g L−1 . For both the 30 and 60 ␮g L−1 treatments reproduction was reduced to about 40% of the control value. This is in line with the observations of Hook and Fisher (2002), who exposed marine copepods to dietary Hg, Ag, Cd, Zn and Mn. They suggested that a 50% reduction of total reproduction appears to be a threshold effect, i.e. all metals reduced reproduction to about the same level of 50% of the control value reduction and once this level was reached no further reduction was observed at higher metal concentrations. Contrary to the total reproductive output, time to first brood was not significantly affected (Table 1). According to Kooijman (2000) both observations suggest a direct effect of dietary zinc on reproduction, i.e. dietary zinc adversely affects the conversion of energy reserves and resources into the number of offspring. Hook and Fisher (2002) also demonstrated that it is unlikely that the decrease in reproductive output following dietary metal exposure is due to a general stress response as in their study other sublethal endpoints such as respiration rate and behaviour were unaffected. Similarly, the fact that in our study ingestion rates of the daphnids — another general stress endpoint — were very similar in all treatments (Table 2) supports the hypothesis of a direct effect on reproduction. According to Kooijman (2000), a reduction of food uptake would be reflected in both reproduction and growth effects. Hence, it is concluded that dietary zinc specifically acts on the reproductive physiology of D. magna. The specific effect of dietary zinc on reproduction as observed in aquatic invertebrates, such as D. magna (this study) and in marine copepods (Hook and Fisher, 2002) has, to our knowledge, not been investigated in aquatic vertebrates, such as fish. For some fish species dietary zinc has been shown to reduce survival or growth (for a review, see Clearwater et al., 2002). Mortality and growth were not affected in D. magna or marine copepods upon exposure to dietary

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Zn. The mechanisms of dietary zinc toxicity in fish and invertebrates may thus be different. Given the suspected impact of dietary zinc on vitellogenesis in D. magna and copepods, further research towards reproductive toxicity of dietary zinc in fish species is warranted, as the latter also require vitellogenesis for egg production. The observed specific effect on reproduction most likely also excludes the possibility of a diet quality effect, i.e. that the effects would not be due to zinc but to a reduced nutritional value of the Zn-exposed algae. According to Clearwater et al. (2002) this possible confounding factor should not be neglected in dietary toxicity studies. Food quality in cladocerans is mainly determined by parameters such as the essential fatty acid content, the C:P ratio and the C:N ratio (Sundbom and Vrede, 1997; Park et al., 2002; Ferrao et al., 2003). Whenever effects of these food quality parameters on cladoceran reproduction have been observed, concurrent effects on somatic growth were observed, an observation not made in the present study. Moreover, McLarnon-Riches et al. (1998) demonstrated only very moderate shifts in fatty acid composition of P. subcapitata following exposure to a zinc concentration that caused a 50% inhibition of growth, whereas no growth inhibition of the algae was observed in our exposures. It is thus unlikely that food quality played a role in the daphnids’ reduced reproductive output following exposure to dietary zinc. At first sight, it may seem odd that reproduction is reduced at a dietary zinc level (i.e. when algal food had been exposed to 30 ␮g L−1 ) where the internal zinc burden of D. magna was not higher than that in the control. Moreover, the much higher zinc burden in the 60 ␮g L−1 treatment than in the 30 ␮g L−1 treatment resulted in a quantitatively similar adverse effect on reproduction. This suggests that direct effects on reproduction may not be related to total internal zinc burdens and that the toxic effect is most probably related to accumulation in specific target tissues or cells. Hook and Fisher (2001a,b, 2002) suggested that the reduced reproduction in marine copepods following dietary exposure to metals was due to a disturbance of vitellogenesis, a necessary step in the reproduction of oviparous organisms like copepods and cladocerans. This is supported by the fact that Cd uptake via the food resulted in a reduced accumulation of yolk protein (lipovitellin) in blue crabs Callinectes sapidus (Lee and Noone, 1995).

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Thus, the target sites where accumulation of dietary zinc would lead to reproductive toxicity are most likely sites where production of vitellogenin occurs or where vitellogenin is further processed to yolk protein. In parthenogenetically reproducing Daphnia vitellogenin synthesis is believed to occur in the so-called fat cells (Zaffagnini and Zeni, 1986; also termed ‘storage cells’ by Bodar et al., 1990), which are scattered along the digestive tract (Bodar et al., 1990). Vitellogenin is then transferred via the haemolymph into the oocytes where it is further processed to yolk protein. Potential sites of toxic accumulation of dietary zinc in D. magna are most probably the fat cells (inhibition of vitellogenin production) or the oocytes (inhibition of processing of vitellogenin to lipovitellin). The exact mechanism by which inhibition of vitellogenin production or processing by zinc takes place is unknown. Hook and Fisher (2002) suggested that inhibition of vitellogenesis is due to binding of metals to enzymes involved in vitellogenesis, a process which could occur in the vitellogenin production (fat cells) or processing (oocytes) stage of the vitellogenesis. An alternative hypothesis for the inhibition of vitellogenin synthesis can be found in Bodar et al. (1990). In a cytopathological study with D. magna these authors observed an ultra-structural alteration of mitochondria (accompanied by a reduced mitochondrial energy production) and a reduced glycogen content of the storage cells upon exposure to Cd and tributyltin chloride. They argued that storage cells in this state would be less able to properly carry out their vitellogenin synthesizing function. Indirect evidence for this was obtained in the tributyltin chloride experiments in which no neonates in the brood pouch were noted. To investigate the reproductive effect in more detail, we determined two additional reproductive traits of the daphnids, i.e. the brood size and the fraction of parent daphnids producing broods (Fig. 1), both of which determined the ultimate total number of juveniles. All organisms were able to produce at least one brood. In the control and in the 20 ␮g L−1 treatment 70 and 90% of the daphnids produced four broods, respectively. The remainder of the organisms in these exposures produced three broods. In the 30 ␮g L−1 treatment, 70, 30 and 10% managed to produce two, three and four broods, respectively. In the 60 ␮g L−1 treatment 90, 20 and 20% of the daphnids achieved two, three, and four broods, respectively. Hence, it is concluded that the fraction of reproducing daphnids

Fig. 1. Effects of a 21-day dietary zinc exposure on D. magna reproduction, depicted per brood, after feeding with P. subcapitata exposed for 64 h to a control and three zinc concentrations (20, 30, and 60 ␮g L−1 ), represented as the percentage of parent daphnids still reproducing (top), and the juveniles produced in each brood (bottom). Error bars denote standard errors. Asterisk denotes significant differences as compared to control treatment (P < 0.05, Mann–Whitney U-test).

progressively decreases with time of exposure, starting from the second brood. Considering brood sizes, a similar pattern is observed (Fig. 1). No reduction is observed in the ␮g L−1 treatment compared to the control. For the 30 and 60 ␮g L−1 treatments, a significant reduction of brood size compared to the control is observed in the third and fourth brood, but not in the first and second brood (although there seems to be a trend that the second brood in those two treatments is slightly smaller than in the control). Summarizing, the effects on brood size and fraction of daphnids producing broods, it seems that the adverse effect of dietary zinc exposure on brood size increases with exposure time, starting from the second brood which is released after only 13.7 days. The latter may have important consequences for the population ecology of D. magna exposed to dietary

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zinc. Population level effects can best be described in terms of the intrinsic rate of natural increase (rm ), which in this study was calculated according to Van Leeuwen et al. (1987). The rm takes into account all aspects of reproduction, including time to first brood, brood interval time, and age-specific brood size. The later effects on reproduction occur, the less impact it will have on rm . Unlike total reproduction which was inhibited by about 40% in the 30 and 60 ␮g L−1 treatment, the rm was only inhibited by about 9% in both treatments. Hence, the population level effect is expected to be rather limited. It should be stressed that here we investigated only one generation, whereas it is unclear what the effect of dietary zinc would be in a long-term multi-generation experiment. For example, Zn-induced disturbance of vitellogenesis in the parent generation may result in decreased fitness of the F1 generation and hence a further reduction in rm and population viability. Muyssen and Janssen (2001) have demonstrated that during a multi-generation exposure of D. magna to waterborne zinc rm values in the fifth generation were similar at optimal zinc concentrations but were significantly lower at lower (deficiency) and higher (toxic) concentrations (Muyssen and Janssen, 2001). It is noted that the addition of food in their exposures may have provided some uptake via the dietary zinc route. Mechanistically, provisioning of the oocytes that form the second brood occurs during the instar at the end of which the first brood is released (Bradley et al., 1991). Taking into account the time to first brood (10.3 days) and an estimated inter-moult period of about 3 days, provisioning of the second brood would have occurred between 7 and 10 days after initiating the exposure. This is the period during which biologically important effects on vitellogenesis should occur first. It remains unclear why we did not observe effects on the first brood (i.e. the brood that was provisioned between approximately 4 and 7 days after exposure initiation). Two hypotheses can be put forward. First, the accumulation of zinc in target cells might be rather slow, only reaching toxic levels for vitellogenesis after at least 7 days of exposure. Second, a cascade of events (e.g. via the disruption of storage cell functioning) may be needed before vitellogenesis is inhibited. Further research towards tissue and/or cell specific accumulation of dietary zinc in D. magna is needed, combined with determinations of cytopathology, other cell processes (e.g. mitochondrial activity) and vitellogenin synthesis

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and/or processing to test the above-mentioned hypotheses. 3.4. Implications for water quality criteria and bioavailability models The finding that exposure to dietary zinc has been shown to result in toxicity in different species (cladocerans, copepods and several fish species) may have some consequences for risk assessment and water quality criteria (WQC) setting of zinc in aquatic environments. Indeed, currently it is assumed that toxicity primarily occurs via the dissolved phase. According to Clearwater et al. (2002), the preferred approach to derive WQC would be to use only data from chronic toxicity studies which used live (i.e. natural) food. As such the complex mechanisms by which organisms deal with dietary metal stress are taken into account. For example, upon incorporation of the food, a metal may be sequestered into insoluble metal detoxification granules, resulting in a lower dietary availability of the metal (Bryan and Gibbs, 1983). To our knowledge, next to the present study, only two other have worked with live diets in investigating dietary zinc toxicity. Hook and Fisher (2002) fed marine copepods with Zn-contaminated diatoms, and Mount et al. (1994) fed rainbow trout with Zn-contaminated Artemia nauplii. More studies are needed to be able to take the dietary pathway into account in a scientifically defensible manner. The occurrence of dietary zinc toxicity may also have consequences for the use of biotic ligand models (BLM) in regulatory assessments. Biotic ligand models, which can be used to predict chronic zinc bioavailability and toxicity in the dissolved phase as a function of water chemistry (Heijerick et al., 2004; De Schamphelaere and Janssen, in press), may need to include the dietary toxicity pathway. It is noted that the experimental data based on which the chronic Zn-BLM for D. magna was developed, were obtained through chronic toxicity tests in which algal food was presented to the daphnids (Heijerick et al., 2004), thus providing the potential of dietary zinc uptake by D. magna after uptake of zinc from the daphnids’ test medium by the algae. Nevertheless, this BLM predicts chronic reproductive zinc bioavailability and toxicity in synthetic and field surface waters with a reasonable accuracy even without explicitly directly considering the dietary toxicity

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pathway (Heijerick et al., 2004; De Schamphelaere et al., 2004b). Three explanations are possible. First, in those tests, the uptake of dissolved zinc by the algal food under the conditions of a chronic D. magna test might be too slow to reach toxic algal zinc concentrations before the algae are ingested. Second, the presence of elevated waterborne zinc concentrations may induce an increased resistance against the dietary zinc in D. magna (e.g. via metallothionein induction). These two hypotheses would also explain why daphnids experience reduced reproduction when chronically fed with algae exposed to 30 ␮g L−1 and why they do not when chronically exposed to a waterborne concentration of 155 ␮g L−1 while being fed daily with control algae that accumulate zinc from the daphnid test solution (De Schamphelaere et al., 2004b). A third possible explanation is that the derived chronic Zn-BLM parameters (for the waterborne exposure) partly reflect the contribution of the dietary route in “waterborne” exposures. This may be especially true for the BLM parameter that accounts for the effect of pH on chronic zinc toxicity, i.e. the constant which is assumed to describe competition between H+ and Zn2+ for toxicity via the water and which thus predicts the higher chronic toxicity of waterborne zinc at increased pH (Heijerick et al., 2004). In reality, however, this constant may also describe the increased uptake of zinc by the algal food at higher pH levels. Indeed, it is well known that algae can accumulate more metal at increased pH levels (e.g. Xue and Sigg, 1990). When the algal food in the chronic “waterborne” exposure would be able to take up enough Zn to reach dietary toxic levels at higher pH levels, this hypothesis could be plausible. This hypothesis is supported by the fact that the chronic BLM tends to underestimate the toxicity of waterborne zinc mostly at pH levels above 8 (De Schamphelaere et al., 2004b). This would also explain the marked increase of chronic toxicity of “waterborne” zinc to D. magna with increased pH, whereas pH did not affect acute toxicity (Heijerick et al., 2002a, 2004). Further research is clearly needed to: (1) assess the relative importance of and the possible interaction between the waterborne and the dietary uptake and toxicity pathways of Zn in D. magna, (2) determine the implications for biotic ligand modeling of chronic zinc toxicity and (3) further improve regulatory assessments of zinc in aquatic environments.

4. Conclusions The exposure of D. magna to dietary Zn resulted in reduction of reproduction, but not of growth and feeding rates. Reproduction effects were not related to total internal zinc burdens. Overall, our results thus indicate that dietary zinc specifically targets reproduction in D. magna through accumulation in particular target sites, possibly cells or tissues where vitellogenin synthesis or processing occur. Our data also indicate that the dietary exposure route may need to be considered next to the waterborne route when assessing the bioavailability of zinc in freshwater, e.g. with biotic ligand models. Finally, our data illustrate that the potential importance of the dietary exposure route should be carefully considered and interpreted in regulatory assessments of zinc.

Acknowledgments Karel De Schamphelaere is supported by a postdoctoral research grant of the Fund for Scientific Research-Flanders (FWO). Frank Vanhaecke would like acknowledge the FWO for their financial support (research project G.0037.01). Veerle Van Lierde is a research assistant of the FWO. Additional funding was obtained from the International Lead Zinc Research Organization (ILZRO) and from the Ghent University Research Fund (BOF No. 01110501). The authors thank Dr. Patrick Sorgeloos and Kristof Dierckens for their assistance in interpreting food quality related issues and Emmy Pequeur, Jill Van Reybrouck, and Gis`ele Bockstael for their technical assistance.

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