Sublethal mechanisms of Pb and Zn toxicity to the purple sea urchin (Strongylocentrotus purpuratus) during early development

Accepted Manuscript Title: SUBLETHAL MECHANISMS of Pb and Zn TOXICITY TO The PURPLE SEA URCHIN (stroNgylocentrotuS purpuratus) DURING EARLY DEVELOPMENT Author: Margaret S. Tellis Mariana M. Lauer Sunita Nadella Adalto Bianchini Chris.M. Wood PII: DOI: Reference:

S0166-445X(13)00310-X http://dx.doi.org/doi:10.1016/j.aquatox.2013.11.004 AQTOX 3673

To appear in:

Aquatic Toxicology

Received date: Revised date: Accepted date:

9-7-2013 1-11-2013 5-11-2013

Please cite this article as: Tellis, M.S., Lauer, M.M., Nadella, S., Bianchini, A., Wood, Cs.M.,SUBLETHAL MECHANISMS of Pb and Zn TOXICITY TO The PURPLE SEA URCHIN (stroNgylocentrotuS purpuratus) DURING EARLY DEVELOPMENT, Aquatic Toxicology (2013), http://dx.doi.org/10.1016/j.aquatox.2013.11.004 This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

Highlights

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! ! Pb and Zn exposure resulted in disruption of Ca homeostasis. ! ! Ca uptake rates and whole body Ca accumulation were severely inhibited by both metals. Ca2+ ATPase activity was significantly inhibited by Zn but not by Pb. ! ! Pb exhibited marked bioaccumulation whereas Zn was well–regulated. ! ! Both metals had little effect on growth, measured as larval dry weight, or on Na, K, or Mg concentrations. ! ! Embryos exposed to Pb showed some capacity for recovery.

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SUBLETHAL MECHANISMS OF PB AND ZN TOXICITY TO THE PURPLE

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SEA URCHIN (STRONGYLOCENTROTUS PURPURATUS) DURING EARLY

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DEVELOPMENT

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*Margaret S. Tellis1a,b, Mariana M. Lauerb,c, Sunita Nadellaa,b, Adalto Bianchinib,c,

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Chris.M. Wooda,b

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900 Rio Grande, Rio Grande do Sul, Brazil

McMaster University, Department of Biology, Hamilton, Ontario, Canada L8S 4K1

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Bamfield Marine Sciences Centre, Bamfield, British Columbia, Canada V0R 1B0

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Instituto de Ciências Biológicas, Universidade Federal do Rio Grande (FURG), 96203-

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Corresponding author: Tel (+1) 9059628469 [email protected], 1280 Main St. W., McMaster University, Dept. of Biology, Hamilton Ontario, Canada L8S4K1

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Abstract

27 In order to understand sublethal mechanisms of lead (Pb) and zinc (Zn) toxicity,

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developing sea urchins were exposed continuously from 3 h post-fertilization (eggs) to 96

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h (pluteus larvae) to 55 (± 2.4) µg Pb/L or 117 (± 11) µg Zn/L, representing ~70% of the

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EC50 for normal 72 h development. Growth, unidirectional Ca uptake rates, whole body

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ion concentrations (Na, K, Ca, Mg), Ca2+ ATPase activity, and metal bioaccumulation

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were monitored every 12 h over this period. Pb exhibited marked bioaccumulation

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whereas Zn was well–regulated, and both metals had little effect on growth, measured as

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larval dry weight, or on Na, K, or Mg concentrations. Unidirectional Ca uptake rates

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(measured by 45Ca incorporation) were severely inhibited by both metals, resulting in

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lower levels of whole body Ca accumulation. The greatest disruption occurred at

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gastrulation. Ca2+ ATPase activity was also significantly inhibited by Zn but not by Pb.

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Interestingly, embryos exposed to Pb showed some capacity for recovery, as Ca2+ATPase

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activities increased, Ca uptake rates returned to normal intermittently, and whole body

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Ca levels were restored to control values by 72-96 h of development. This did not occur

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with Zn exposure. Both Pb and Zn rendered their toxic effects through disruption of Ca

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homeostasis, though likely through different proximate mechanisms. We recommend

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studying the toxicity of these contaminants periodically throughout development as an

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effective way to detect sublethal effects, which may not be displayed at the traditional

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toxicity test endpoint of 72 h.

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Keywords: echinoderm, lead, zinc, embryonic development, toxicity, spicule,

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calcification

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Highlights

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! ! Pb and Zn exposure resulted in disruption of Ca homeostasis. ! ! Ca uptake rates and whole body Ca accumulation were severely inhibited by both metals. Ca2+ ATPase activity was significantly inhibited by Zn but not by Pb. ! ! Pb exhibited marked bioaccumulation whereas Zn was well–regulated. ! ! Both metals had little effect on growth, measured as larval dry weight, or on Na, K, or Mg concentrations. ! ! Embryos exposed to Pb showed some capacity for recovery.

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1. Introduction

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Metals enter marine ecosystems through various anthropogenic and natural modes. At high concentrations, metals, including Pb and Zn, are known to be extremely

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toxic to aquatic organisms, reducing their general survival as well as hampering

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embryonic development. While extensive research exists on the effects of Pb and Zn in

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freshwater environments (e.g. European Union, 2007 and 2009; Wang, 1987; Naimo,

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1995), there is only limited research on the effects of these metals in marine and estuarine

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environments (e.g. Supanopas et al., 2005; França et al., 2005; Bielmyer et al., 2012), and

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most of our understanding of mechanisms of toxicity comes from research on fish.

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Pb is classified as a priority contaminant in European Union regulations on water

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policy (European Commission, 2012). As a xenobiotic, Pb has no known role in the

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biological processes of organisms and prolonged exposure results in purely toxic effects.

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In general, Pb toxicity is thought to occur by replacing divalent ions such as Zn and Fe

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and by calcium mimicry (Ballatori, 2002); not surprisingly, Pb is found in the calcified

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skeleton of fish (reviewed by Mager, 2012). In contrast, Zn is an essential ion as well as a

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toxicant, and therefore has considerable importance in the aquatic environment. Zinc is a

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vital component of over 300 enzymes and other proteins (Vallee and Falchuk, 1993).

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However at high concentrations, Zn has detrimental effects on the development and

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survival of many aquatic organisms (Eisler, 1993). Similar to Pb, Zn can also mimic Ca

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(Ballatori, 2002). Zn disrupts Ca homeostasis in freshwater fish through the induction of

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hypocalcaemia and also causes a disturbance of acid-base balance, but in contrast to Pb,

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Zn exhibits minimal tendency for bioaccumulation (reviewed by Hogstrand, 2012).

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Sea urchin embryo bioassays have been historically utilized as a means of

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monitoring marine water quality, because this life stage is extremely sensitive to a variety

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of contaminants (Kobayashi, 1971). A number of previous studies have assayed the

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toxicity of Pb and Zn to sea urchin embryos (e.g. Warnau and Pagano, 1994; Phillips et

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al., 1998; Nacci et al., 2000; Radenac et al., 2001; Novelli et al., 2003; Kobayashi and

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Okamura, 2004; Sanchez-Marin et al., 2010; Nadella et al., 2013), but this research has

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been directed at quantifying toxicity, rather than understanding mechanisms of toxicity.

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Understanding mechanisms behind sublethal effects is important, because it may shed

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light on more sensitive endpoints, which can be detected earlier and at lower

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concentrations than traditional endpoints such as mortality and inhibited growth. Indeed,

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more research is needed in order to reduce the uncertainty surrounding marine water

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quality guidelines for these two metals. This is of particular importance in Canada where

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national authorities have yet to establish marine water quality criteria for Pb and Zn (with

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the exception of Pb in British Columbia) and in the U.S.A, where current chronic criteria

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for Zn and Pb in sea water were derived from data available in the 1980s (Hogstrand,

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2012; Mager, 2012).

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The objective of the current research was to elucidate the mechanisms of Pb and

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Zn toxicity over the first 96 h of development of the purple sea urchin

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(Strongylocentrotus purpuratus). Recently, we have characterized the changes in ionic

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status occurring over this period during normal development (Tellis et al., 2013). This

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publication provides the control data set for the current study. Particularly remarkable

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were 15-fold increases in whole body Ca concentration, accompanied by up to 7-fold

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variations in unidirectional Ca uptake from the external sea water (measured by 45Ca

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incorporation) and 2-fold variations in Ca2+ ATPase activity, a key enzyme of Ca

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metabolism. These are very likely involved in the rapid development and calcification of

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the internal skeleton or spicule at this time, which is essential for normal development

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(Wilt, 1999, 2002). Therefore, our research focused specifically on Ca homeostasis,

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especially as this has been shown to be a major target of both Pb and Zn toxicity in fish

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studies, as discussed earlier.

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Our working hypothesis was that both Pb and Zn, at continuous exposure

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concentrations below those causing acute mortality, would have marked deleterious

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effects on various aspects of Ca metabolism. To test this hypothesis, a variety of

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endpoints of Ca homeostasis were analyzed at 12-h intervals including unidirectional Ca

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uptake rate (measured by 45Ca incorporation), whole body Ca (as well as Na, K, and Mg)

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concentrations, metal accumulation, Ca2+ ATPase activity, and dry weight growth. The

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nominal exposure concentrations, 52 µg/L (Pb) and 106 µg/L (Zn), were chosen to

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represent 70% of the EC50s for normal development through 72 h (Pb EC50 = 74 µg/L

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Zn EC50 = 151 µg/L), as determined in a companion study (Nadella et al., 2013). These

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concentrations were environmentally realistic for polluted sites with Zn being found at

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levels of 75-882 µg/L and Pb being found at levels of 6-2000 µg/L in mining disturbed

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areas (Eisler, 1993; Mager, 2012). By way of reference, recent surveys (Hogstrand, 2012;

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Mager, 2012) of the very few jurisdictions that have chronic ambient marine water

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quality guidelines for these metals reported a range of regulatory guidelines (allowable

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upper limits) for Pb from 4.4 µg/L (Australia/New Zealand) to 8.1 µg/L (U.S.A), and for

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Zn from 15 µg/L (Australia/New Zealand) to 86 µg/L (U.S.A).

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2. Materials and methods

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2.1 Experimental Organisms Methods followed those described by Tellis et al. (2013). In brief, reproductively

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ripe adult sea urchins (Strongylocentrotus purpuratus) were obtained in June by divers

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from the natural benthic populations of Barkley Sound, B.C., Canada (48°50'30" N,

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125°08'00" W). At Bamfield Marine Sciences Centre, they were held with minimal

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handling in aerated tanks supplied with flowing sea water (32 ppt) at 15 ˚C. After

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spawning, they were returned to the wild. Spawning was induced by an injection of 1 ml

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of a 0.5 M KCl solution into the haemocoel as described by Hinegardner (1975). Eggs

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were collected into filtered (0.2 µm SteritopTM filter– Millipore, Billerica, MA, USA)

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seawater (32 ppt). The eggs from different females were then pooled, and filtered

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through a mesh to remove debris from the egg solution. Sperm from spawning males was

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diluted in 50 ml of filtered (0.2 µm) seawater, then 1 ml of this diluted sperm was added

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to the pooled sea urchin eggs to initiate fertilization, which was normally achieved in

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under 0.5 h. Once fertilization of 80% of the eggs was verified, the stock was diluted to a

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final concentration of 60,000 eggs/L for each replicate exposure in 800 ml plastic

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NalgeneTM beakers. Full strength sea water (32 ppt) was used in all exposures. The

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embryos were then allowed to develop in an incubator at 15 ! C with 16 h light: 8 h dark

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light cycle for the desired time of each test.

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Three experimental series were performed as outlined below. The same

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fertilization batch was used for series 1 and 2 and a second fertilization batch was used

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for series 3. Series 1 and 2 were run simultaneously and series 3 was run two weeks later.

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In series 1 and 3 every 12 h over the first 96 h of development, 5-6 replicates were harvested (800 ml each, entire volume sampled) for the various endpoints. In series 2 at

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each time point 5-6 replicate tubs were sampled in each treatment. The tub was stirred

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gently and a small volume of test water containing embryos was removed. Therefore the

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volume of the exposure would decrease but the density of the embryos in the exposure

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would remain the same. Metal exposures and accompanying control treatments were set

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up to ensure that 5-6 replicates of each treatment (control, Pb and Zn) could be sampled

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every 12 h over 96 h of development.

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2.2 Metal Exposure Solutions

Pb and Zn metal stock solutions were made by diluting their respective inorganic

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salts (Pb(NO3)2 and ZnSO4), (Sigma-Aldrich; St. Louis, MO, USA - trace metal grade)

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in deionized water. These stock solutions were stored in NalgeneTM bottles (NalgeneTM

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Rochester, NY, USA) under refrigeration. Metal exposure solutions were made by adding

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the required volumes of metal stock solutions to filtered (0.2 μm) seawater (32 ppt) to

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obtain the desired metal concentrations. Exposure solutions were prepared 24 h in

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advance of the tests to allow time for equilibration of the metal salts with seawater in the

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test containers. Analytical samples were taken immediately before addition of the

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organisms.

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2.3 Series 1. Whole body metal accumulation, ion content and embryonic weights over

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development

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Developing embryos were continuously exposed to either control conditions or 52 µg/L (Pb) or 106 µg/L (Zn) (nominal concenrations) for 96 h. Every 12 h, 5 replicates of

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each exposure treatment were sampled by filtration through pre-weighed filters

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(Whatman Nucleopore Track-Etch Membrane PC MB 47 MM 8.0 μm), via a vacuum

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pump. The density of embryos in the original suspension was counted under a light

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microscope using a Sedgewick-Rafter slide. The collected embryos on the filter were

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rinsed with nanopure water and the filter was then placed in an open EppendorfTM tube

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and left to dry at room temperature. To determine embryonic weights, the dried filter with

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collected embryos on it was weighed. This weight minus the filter weight divided by the

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number of embryos collected was determined to be the mean dry weight of the

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developing embryo.

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The dried filter paper and embryos were then digested in full strength HNO3 at 65

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˚C for 48 h (samples were vortexed at 24 h to aid in the digestion process). Analysis of

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filter paper blanks revealed negligible concentrations of Na, K, Ca, Mg, and metals in the

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filters.

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2.4 Series 2. Whole body unidirectional Ca uptake rates during 96 h exposures

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Unidirectional uptake rates of Ca from the water, as determined by 45Ca

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incorporation, were measured in embryos at 12-h intervals over the first 96 h of

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development, using the same exposure treatments as in Series 1. Every 12 h, each of the 6

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replicates was gently stirred to ensure homogeneity of embryos in suspension and a small

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volume was removed for flux rate determination. The extracted volume was decreased to

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a few ml by using a SteritopTM filter (0.2 µm – Millipore, Billerica, MA, USA) to filter

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out some of the sea water by gravity. The resulting concentrated embryo solution was re-

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suspended with new sea water and reduced in volume again. This was repeated 3 times to

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wash the embryos. The embryos were finally resuspended in fresh sea water containing

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the appropriate metal level to achieve a nominal target density of 2500 embryos/ml for

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the 45Ca uptake rate measurements.

Unidirectional Ca influx rates were measured by incubating 0.5 ml of the sea

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urchin embryo suspension with 0.5 ml of radioactive 45Ca (0.17 µCi/ml as CaCl2,

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PerkinElmer, Woodbridge, ON, Canada) in sea water in 2-ml EppendorfTM tubes for 20

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min. The radioactive 45Ca solution also contained the appropriate concentration of metal.

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The flux period was based on preliminary time series experiments in which 20 min was

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determined to be optimal for 45Ca uptake analysis, representing the longest period of

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linear uptake before significant radioisotopic backflux occurred. At the end of the 20-

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min flux period, the embryos were removed from the EppendorfTM tube via a 1-ml

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syringe. The syringe contents were then ejected through a 0.45-µm syringe tip filter

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(NalgeneTM Rochester, NY, USA), leaving the embryos on the filter. Then 10 ml of clean

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sea water was immediately passed through the filter to wash the embryos. The filter was

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then reversed and 3 x 1 ml (i.e. each ml separately) of clean sea water was passed through

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the filter and collected in a scintillation vial to recover the embryos. Scintillation fluid (5

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ml; Aqueous Counting Scintillant, Amersham, Little Chalfont, UK) was added to each

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vial and the 45Ca radioactivity in the sample was measured by scintillation counting (Tm

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Analytic, Beckman Instruments, Fullerton, CA, USA). Tests showed that quench was

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constant. A dummy run (without 45Ca) was performed at each time point and the embryos

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recovered from the filter were counted under the microscope to determine embryo

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concentrations used in the flux measurements. Unidirectional Ca uptake rates per larva were calculated from the counts per minute of each replicate (CPM), mean measured specific activity (SA) of the incubation

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solution, number of embryos in each replicate and exact time (t), and expressed as pmol

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Ca/embryo/h:

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Specific activity was calculated by dividing the measured Ca concentration (pmol/ml) in

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the incubation solution by 45Ca radioactivity (CPM/ml).

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2.5 Series 3. Ca2+ ATPase enzyme activities during 72 h exposures

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At each 12-h time point until 72 h, 5 replicates of each exposure (control, Pb and Zn)

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were sampled. Samples were not collected for the 84 h and 96 h time points due to

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insufficient time on the research trip. The volume of each replicate was reduced to 20 ml

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by filtering the solutions by gravity through a 0.45 µm NalgeneTM filter. The concentrated

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embryos were resuspended in clean sea water and concentrated again. This was

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performed 3 times to wash the embryos. The washed, concentrated embryos were then

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centrifuged at 12000 g for 5 min, the supernatant was decanted and the resulting pellet

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was transferred to an EppendorfTM tube. Sampled embryos were immediately frozen

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and kept at -80 ˚C to preserve them for Ca2+ATPase analysis, which was performed a

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few months after the research trip.

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2.6 Analytical techniques Cation levels (Ca2+, Mg2+, Na+, K+) as well as Zn in whole body digests and total Ca2+ levels in sea water were measured using flame atomic absorption spectroscopy

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(220FS; Varian Instruments, Palo Alto, CA, USA) against commercial standards (Fisher

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Scientific, Toronto, ON, Canada). Pb in digests was measured using graphite furnace

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atomic absorption spectroscopy (220, Varian Instruments, Palo Alto, CA, USA). Water

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samples were acidified to 1% with HNO3. Metal concentrations in seawater exposure

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solutions were determined using a protocol developed by Toyota et. al. (1982). This

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method entails precipitating the metal from solution using lanthanum oxide and Na2CO3

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in order for it to be analyzed without interference from the many electrolytes in saltwater.

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Full details are provided by Nadella et al. (2013). Reference standards used for Zn and Pb

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were TM24 and TM25 (Environment Canada certified reference material, recovery 90 %

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- 95 %).

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Ca2+ATPase activity was measured in embryos that had been immediately frozen and kept at -80 ˚C to preserve them for Ca2+ATPase analysis a few months after the

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research trip. Ca2+ATPase activity was determined in the supernatant fraction of

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homogenized embryos using a method which measured the liberation of inorganic

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phosphate by the ATPase enzyme (Vijayavel et al., 2007). Embryos were homogenized

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using a buffer containing Tris HCl at 100mM, 2mM EDTA and 5mM of MgCl2 adjusted

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to a pH of 7.75. Homogenized samples were centrifuged at 10,000 g for 20 min at 4°C.

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Ca2+ATPase activity was determined as the amount of inorganic phosphate per mg of

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protein per hour released in the presence/absence of 0.5 mM CaCl2 in a medium

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containing 80 mM NaCl, 5 mM MgCl2, 3 mM ATP, 20 mM Tris–HCl (pH 7.4 and 1 mM

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ouabain. Activities were normalized to the protein content of the homogenate, as

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determined using BSA standards (Sigma-Aldrich) and Bradford’s reagent (Sigma-

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Aldrich; Bradford, 1976).

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Statistical analysis was performed using the software SigmaPlot 10.1. Two-Way

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ANOVAs (time, metal) were used to detect variation among multiple treatment groups

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and where the F value indicated significance, Fisher’s LSD post hoc tests were used to

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identify specific significant differences over time within individual treatment groups.

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Prior to the test, all data were checked for homogeneity of variances and normality of

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distribution, and where necessary were transformed using natural logarithm or square

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root functions. Student’s t-tests (two-tailed) were used to determine differences between

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control and metal-exposed embryos at the same times. All data are presented as means !

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SEM (N, number of replicates) on non-transformed data. Differences were considered

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significant at p < 0.05.

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3. Results

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3.1 Exposure concentrations and seawater Ca levels

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Nominal target concentrations for the exposures of Series 1, 2, and 3 were set to

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70% of the EC50 values reported in parallel studies by Nadella et al. (2013). Actual

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measured values during the tests were 55 ± 2.4 µg/L (Pb) and 117 ± 11 µg/L (Zn). The

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levels of Ca in the seawater used were 7.87-8.98 mM.

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3.2 Series 1. Whole body ion concentrations Whole body Ca concentrations demonstrated a significant interaction between Pb and time by two-way ANOVA. In controls, Ca levels increased progressively over 96 h.

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Embryos continuously exposed to Pb suffered lower levels of Ca accumulation during the

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initial stages of development, significant at 24, 36 and 60 h of development. However Ca

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levels in Pb exposed embryos returned to control amounts by 72 h onwards (Figure 1a).

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There was also a significant interaction of time and Pb exposure for Na, but not

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for K and Mg; the influence of time was significant for the latter. K and Na

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concentrations tended to decrease until 48 h, then increase thereafter, but continuous Pb

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exposure did not affect K and Na levels over 96 h of development (Figures 1b and c).

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Mg concentrations tended to increase with time, but exposure to Pb had no effect on Mg

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over most of the 96 h of development, apart from an increase at 12 h (Figure 1d).

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ANOVA indicated significant interactions of time and Zn exposure for all ions measured.

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Whole body Ca levels were markedly lower than controls at every time point after 24 h,

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with the exception of the 84 h time point (Figure 2a). The reductions ranged from 55-

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70%. K and Na levels were generally not affected by continuous Zn exposure apart from

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at 24 h where they were significantly lower for K with respect to controls (Figures 2b and

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c). Mg concentrations in Zn exposed embryos were significantly higher than controls at

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12h, 36 h, and 84h (Figure 2d).

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3.4 Series 1. Whole body metal accumulation Embryos continuously exposed to Pb showed significant Pb accumulation at

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every time point after 12 h through 96 h of development, with a peak at 84 h (Figure 3).

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Two –way ANOVA indicated a significant interaction of time and Pb exposure. In

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contrast, whole body Zn levels were well regulated with a tendency to decline by 96 h.

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Indeed, by the end of the exposure period, there was no significant Zn accumulation in

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Zn-exposed embryos relative to controls. However, Zn exposed embryos exhibited

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significantly higher Zn contents than controls at the 36 h and 72 h time points (Figure 4).

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Two –way ANOVA indicated no significant main effects or interaction effects.

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3.5 Series 1. Embryonic weights

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Two-way ANOVA revealed no significant effects of time or metal exposure, though there was an interaction effect with Pb. Nevertheless, mean dry weight had

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increased significantly in all three treatment groups by 96h. Both Pb and Zn exposed

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embryos were significantly heavier than controls at 60 h, but there were no other

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treatment-related effects (Table 1).

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3.6 Series 2. Whole body unidirectional Ca uptake rates A parallel series focusing specifically on unidirectional Ca uptake rates was

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performed because of the marked effects on Ca regulation in Series 1. Two –way

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ANOVA indicated significant interaction of time and Pb exposure.

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varied greatly over time in controls, with peaks at the blastula stage (24 h), gastrulation

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(48 h), and late pluteus larva stage (96 h). In embryos continuously exposed to Pb,

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inhibition of Ca uptake was recorded at these time points when Ca uptake was highest in

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control embryos, i.e. at the blastula stage through gastrulation (24h, 36h, and 48 h), and

Ca uptake rates

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during the pluteus larval stage (84 h and 96 h) (Figure 5). Two –way ANOVA indicated

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significant interaction of time and Zn exposure. Similar to Pb, continuous exposure to Zn

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resulted in inhibition of Ca uptake during the blastula stage (24 h) as well as at

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gastrulation (48 h). Significant inhibition of Ca uptake by Zn was also observed during

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the early pluteus larval stage (72 and 84 h) (Figure 6).

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3.7 Series 3. Whole body Ca2+ ATPase activities

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As Series 2 and 3 highlighted effects on Ca regulation, we followed up with an examination of a key enzyme of Ca metabolism, Ca2+ ATPase . Two –way ANOVA

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indicated significant interactions of time and metal exposures. In controls, Ca2+ ATPase

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activity increased until gastrulation at 48 h, then declined thereafter. Pb exposure caused

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significant increases in whole body Ca2+ ATPase activity at 12h, as well as during the

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gastrulation stage at 48 h and at 60 h (Figure 7). In Zn exposed embryos whole body Ca2+

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ATPase activity was significantly lower at 24, 36 and 48 h of development, but was

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higher at 72 h during the pluteus larval stage (Figure 8).

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4. Discussion

Classic toxicity testing by Nadella et al. (2013) showed S. purpuratus embryos to

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be highly sensitive to both Zn and Pb. In general, EC50’s for Zn were very similar to

372

previous values reported for developing sea urchin embryos, whereas those for Pb were

373

towards the lower end of a disparate range in the literature surveyed by Nadella et al.

374

(2013). For both metals, the levels tested in the present investigation (~ 70% of EC50

375

values) were environmentally relevant, and reasonably close to levels of regulatory

17 Page 17 of 44

significance (see Introduction). In support of our original hypothesis, the current study

377

demonstrated that both Pb and Zn toxicities are associated with disruption of Ca

378

homeostasis, as evidenced by analysis of various biomarkers periodically over

379

development. Notably, these disturbances caused only small effects on dry weight

380

growth (Table 1), because during this early development, the embryos are mainly re-

381

organizing their structure rather than increasing in mass. Had morphometric criteria been

382

used to define growth, inhibitory effects would likely have been seen (e.g. Warnau and

383

Pagano, 1994). The stages of development were confirmed in the controls visually by

384

observation through a microscope. In general, the metal exposures did not result in a

385

similar but delayed pattern of ionoregulatory changes relative to the controls, as would be

386

expected if the effect was simply a retardation of development, but rather significant

387

inhibitions of Ca metabolism at specific times over development.

390

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4.1 Whole body ion content measured periodically over 96 h of development In Series 1, Ca, and to a lesser extent Mg, were the only ions that displayed

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391

marked differences in embryos continuously exposed to Pb and Zn over development.

392

Both ions are integral constituents of the spicule (internal skeleton) which is forming

393

during this period (Wilt, 2002; Raz et al., 2003); MgCO3 makes up about 5% of the

394

mineral phase, the remainder being CaCO3 (Decker and Lennarz, 1988).

395

In both metal exposures, embryo Mg levels were higher than controls during the

396

initial stages of development (Figures 1d and 2d). Pb exposed embryos exhibited

397

increased levels of Mg at 12 h of development and Zn exposed embryos displayed higher

398

levels of Mg at 12 h, 36 h, and 84 h. The early elevations could represent a compensatory

18 Page 18 of 44

399

mechanism employed by the embryos to counter metal-induced disruption of Ca

400

incorporation. Notably by 72 h onwards, whole body Ca concentration had recovered in

401

Pb-exposed embryos (Fig. 1a), but not in Zn-exposed embryos (Fig. 2a). Ca was the only ion out of the four ions measured that increased steadily in a

ip t

402

consistent fashion, reaching a plateau by 72-96 h (Figures 1a and 2a). The approximate

404

15-fold increase over this period undoubtedly reflects the deposition of CaCO3 into the

405

developing spicule (Orström and Orström,1942; Yasumasu, 1959; Beniash et al., 1997;

406

Decker and Lennarz, 1988; Wilt, 1999, 2002; Raz et al., 2003). Under control conditions

407

the spicules contain approximately 70% of the whole body Ca content in developing

408

embryos of another sea urchin, Paracentrotus lividus (Pinsino et al., 2011). Both Pb and

409

Zn exposure clearly impeded this process (Figures 1a, 2a), and with Zn, the continuing

410

inhibition amounted to 55-70% of the whole body Ca content, comparable to the amount

411

thought to be in the spicules. Overall, these results point to inhibition of Ca accumulation

412

as a key mechanism of toxicity, though not necessarily by the same precise mode for each

413

metal. Notably, whole body Ca levels recovered later in the Pb exposure (Figure 1a) but

414

not in the Zn exposure (Figure 2a), and the continuing Zn effect occurred in the absence

415

of marked metal bioaccumulation relative to control organisms (Figure 4), in contrast to

416

the Pb effect which was accompanied by dramatic metal bioaccumulation (Figure 3).

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These results suggest that both Pb and Zn inhibit the Ca uptake process, which at

418

least in the first few hours post-fertilization is thought to occur via voltage-gated Ca2+-

419

channels (De Araújo Leite and Marques-Santos, 2011). However, in our earlier study

420

(Tellis et al., 2013), we calculated that unidirectional Ca uptake exceeded net Ca

421

accumulation by at least 10-fold, so most of the Ca taken is lost to efflux, rather than

19 Page 19 of 44

incorporated into the developing skeleton. Therefore, an alternate explanation is that

423

these metals inhibit calcification of the spicule rather than the uptake process itself, so

424

that even more of the influxed Ca is lost to efflux. This underscores the importance of

425

directly measuring the effects of continuous Pb and Zn exposure on unidirectional Ca

426

uptake itself in Series 2.

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429

4.2 Pb and Zn accumulation measured periodically over 96 h of development

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In Series 1, Pb exposure resulted in significant accumulation of the metal at all time points over the first 96 h of development (Figure 3). Accumulation of this non-

431

essential metal followed a somewhat similar pattern to that of Ca reaching a plateau at 72

432

h – 96 h, in accord with the concept of molecular mimicry (Ballatori, 2002). Using

433

radiolabelled 210Pb, Nash et al. (1981) reported a similar pattern of accumulation over

434

time up to 70 h of development in another sea urchin, Lytechinus pictus. Notably

435

however, in the present experiments, the accumulation of Pb in molar terms was only

436

about 1/1000 of the accumulation of Ca over the same period (Figure 3 versus Figure 1a).

437

Concentration-dependent accumulation of Pb was also seen in experiments performed on

438

this same species by Nadella et al. (2013).

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In marked contrast there was negligible bioaccumulation of the essential metal Zn

440

relative to control organisms (Figure 4) at an equi-toxic exposure concentration, yet an

441

even more profound inhibition of Ca accumulation (Figure 2a). Similarly, there was no

442

concentration-dependent bioaccumulation of Zn in the range finder experiments

443

performed by Nadella et al. (2013). Notably, Nadella et al. (2013) reported that the Zn

444

concentrations used here were lower than those at which Zn starts to accumulate in the

20 Page 20 of 44

embryos. Clearly, there is strong homeostatic regulation of Zn levels at these sublethal

446

exposure concentrations. These contrasting patterns for Pb and Zn bioaccumulation were

447

very similar to those reported earlier by Radenac et al. (2001) for similar experiments

448

with embryos of Paracentrotus lividus.

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4.3 Unidirectional Ca uptake over 96 h of development

Under control conditions, unidirectional Ca uptake by the developing sea urchin

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450

showed a variable pattern over 96 h of development (Figures 5 and 6). Earlier we

453

interpreted this as reflecting the differing Ca requirements of each developmental stage

454

(Tellis et al., 2013). In embryos continuously exposed to Pb and Zn, severe inhibition

455

was observed mainly at those time points when Ca uptake rates were highest.

456

Specifically, this inhibition occurred during the blastula (24 h), gastrulation (48 h) and

457

pluteus (84 h and 96 h) stages for Pb exposed embryos (Figure 5) and during the blastula

458

(24 h), gastrulation (48 h) and pluteus (72 h and 84 h) stages for Zn exposed embryos

459

(Figure 6). Thus the inhibition of whole body Ca accumulation (Figures 1a, 2a) is

460

associated, at least in part, with inhibited unidirectional Ca influx. Presumably either the

461

Ca channels (De Araújo Leite and Marques-Santos, 2011) or the internal Ca-handling

462

pathways Gunaratne and Vacquier, 2007; Pinsino et al., 2011), as discussed subsequently,

463

were compromised by Pb and Zn, and were therefore unable to meet the increased Ca

464

demands during these stages.

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465

In these continuous exposures, both Pb and Zn caused greatest inhibition of Ca

466

uptake at the gastrulation stage (48 h; Figures 5 and 6). This stage of development is

467

known to be an especially critical and vulnerable landmark in development. In early 45Ca

21 Page 21 of 44

uptake experiments on Pseudocentrotus depressus, gastrulation was identified as the

469

stage at which Ca uptake and incorporation increased dramatically (Nakano et al.,1963).

470

On completion of this developmental stage, three germ layers (ectoderm, mesoderm and

471

endoderm) and a rudimentary gut are formed and skeletogenesis is initiated.

472

Understandably, abnormalities at this phase often result in complications in later

473

development of the skeleton (Yaroslavetseva and Sergeeva, 2002).

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One reason why the gastrulation stage might display such sensitivity to metal

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474

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468

toxicity is that the maternal reserves of metallothioneins are depleted by this time

476

(Warnau et al., 1996). Metallothioneins are a group of low molecular weight proteins,

477

which protect organisms by virtue of their high affinity for metals, which is a result of

478

their cysteine-rich content. A decrease in available metallothioneins by gastrulation might

479

leave the embryos vulnerable until an increase in de novo synthesis of these proteins is

480

initiated (Warnau et al., 1996). Newly synthesized metallothioneins may also be more

481

effective in their protective role against metals, as they have not been previously exposed

482

to metals as the maternal metallothioneins might have been. This argument appears to be

483

more applicable to the Pb exposure results where metal bioaccumulation occurred.

484

Evidence for this is seen in a return to normalcy of Ca levels in Pb exposed embryos in

485

stages after gastrulation (Figure 1a) whereas this did not occur in Zn exposed embryos

486

(Figure 2a).

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475

487

Ca uptake rates in control embryos over time corresponded with Ca accumulation

488

over time. However, there was a latent period between the increase in Ca uptake rate and

489

the increase in whole body Ca content as the ion presumably required time to accumulate.

490

Similarly, there was a latent period between inhibition of Ca uptake by metal exposure

22 Page 22 of 44

and lower levels of Ca accumulated in the metal exposed embryos. In Zn exposed

492

embryos, inhibition of Ca uptake at 24 h (Figure 6) resulted in lower levels of Ca at 36 h

493

onwards (Figure 2a). Continuous Zn exposure had a major effect on Ca homeostasis with

494

a greater decrease in Ca levels than seen in Pb exposed embryos (Figure 2a versus Figure

495

1a).

In Pb exposed embryos, an inhibition of Ca uptake at 24 h (Figure 5) coincided

cr

496

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491

with lower concentrations of Ca at the same time point (Figure 1a). This could be because

498

inhibition of Ca uptake is occurring before 24 h (when uptake measurements are

499

performed) and by the 24 h mark had a significant effect on Ca concentrations. A return

500

to control uptake levels at 60 and 72 h (Figure 5) resulted in a return of Ca concentrations

501

to control levels for the remainder of the 96 h (Figure 1a). Perhaps the inhibition of Ca

502

uptake seen at 84 and 96 h (Figure 5) would have resulted in lower Ca accumulation later

503

in development beyond the last time point of the experiment (Figure 1a). Regardless, as

504

noted earlier, metal effects on net Ca accumulation may be due to effects on the efflux as

505

well as the influx components of Ca exchange in view of the rapid turnover of Ca in these

506

developing embryos (Tellis et al., 2013).

508 509

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4.4 Ca2+ ATPase activity over 72 h of development In developing sea urchin embryos, Ca2+-ATPase appears to be involved in the

510

internal handling and compartmentalization of Ca (Gunaratne and Vacquier, 2007) rather

511

than in its direct uptake which is instead mediated by voltage-gated Ca2+-channels (De

512

Araújo Leite and Marques-Santos, 2011). There is circumstantial evidence, summarized

513

by Pinsino et al. (2011), that Ca2+ ATPase is involved in Ca-trafficking into the

23 Page 23 of 44

skeletogenic cells. At least for Zn, substantial inhibition of Ca2+ ATPase activity, which

515

occurred at 24 – 48 h of development, the blastula and gastrulation stages respectively

516

(Figure 8), may have been at least part of the toxic mechanism contributing to depressed

517

whole body Ca levels (Figure 2a). Thus Zn appears to interfere both with Ca uptake from

518

the environment and with calcification of the developing spicules. Inactivation of

519

enzymes by metals has been widely documented (Viarengo et al., 1996). Specifically,

520

inhibition of ATPase enzymes has been associated with metals binding with the

521

sulfhydral groups of these enzymes (Pivovarova and Lagerspetz, 1996). Inactivation of

522

these enzymes has also been attributed to damage caused by reactive oxygen species

523

(Stark, 2005).

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In contrast, Pb exposure did not inhibit whole body Ca2+ATPase activity, but

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524

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514

rather caused significant increases in activity during the gastrulation stage at 48 h and at

526

60 h (Figure 7). This maintenance and later increase in Ca2+ATPase activity, together

527

with the return to normal Ca uptake levels at 60 and 72 h (Figure 5) may have contributed

528

to the return of whole body Ca concentrations to control levels by 72 – 96 h in Pb

529

exposed embryos (Figure 1a). Clearly embryos possess some capacity to recover during

530

ongoing Pb exposure, and this does not appear to occur during continuous Zn exposure,

531

reinforcing the conclusion that the two metals have different proximate mechanisms of

532

action in causing disruption of Ca homeostasis. However, an important caveat to this

533

conclusion is that in the closed incubation systems used, endogenously produced

534

dissolved organic carbon (DOC) may have built up over time, affecting the

535

bioavailability of the metals. Indeed Nadella et al. (2013) showed that exogenously added

536

DOC protected against Pb toxicity, but not against Zn toxicity to early life stages of two

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24 Page 24 of 44

marine mussel species, so this could explain the apparent recovery during continued Pb

538

exposure, but not during continued Zn exposure. On the other hand, exogenous DOC did

539

not protect S. purpuratus larvae against either metal. Furthermore, the apparent recovery

540

from the effects of Pb exposure in the present study does not exclude the possibility that

541

irreversible damage to development could still be occurring during early key

542

developmental stages, which might not be apparent through measuring Ca levels only up

543

to 96 h of development.

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544

546

5. Conclusions and Perspectives

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Our research suggests that the toxicity of Pb and Zn during early development stems primarily from a disruption of Ca homeostasis, with small or no effects on Na, K,

548

and Mg homeostasis. However the two metals may disrupt Ca metabolism by different

549

proximate mechanisms of action. Interestingly, Pinsino and colleagues (2010, 2011) have

550

recently demonstrated in Paracentrotus lividus embryos that exposure to manganese can

551

completely block formation of the spicule and greatly reduce whole body Ca

552

accumulation. This raises the possibility that many different metals may interfere with Ca

553

homeostasis in developing sea urchin embryos. In the studies of Pinsino et al. (2010,

554

2011), the threshold for this Mn effect appeared to be about 0.14 mmol/L, and complete

555

inhibition occurred at 1.14 mmol/L. In the present study, marked effects of Pb and Zn on

556

Ca metabolism were seen at only 0.3 µmol/L and 1.8 µmol/L respectively, emphasizing

557

the much greater sensitivity of the organisms to these two these metals.

558 559

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Overall, Pb and Zn had the most pronounced effects on the gastrulation stage of development, as Ca2+ uptake and accumulation and Ca2+ ATPase levels were the most

25 Page 25 of 44

impacted at this stage. This is the point at which spicule formation accelerates, so likely

561

calcification of the spicule was impacted. Although Pb was greatly bioaccumulated and

562

Zn was not relative to levels in control organisms, the embryos displayed some capacity

563

for recovery during continued Pb exposure, but not during continued Zn exposure when

564

both metals were present at ~70% of the EC50 concentrations for normal development.

565

This was evident from the return to normalcy of whole body Ca concentration, increases

566

in whole body Ca2+ ATPase activity, and intermittent recovery of unidirectional Ca

567

uptake rates, suggesting that the occurrence of damage-repair during continued Pb

568

exposure. Toxic effects of metal exposure seen at earlier time points during development

569

may not be apparent at just one time point. We propose measuring endpoints of toxicity

570

periodically over early development as a more effective way of studying the toxic stress

571

of contaminants, rather than using just one time point (e.g. 72 h) as in traditional toxicity

572

testing.

573

575

Acknowledgements

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This research was supported by an award from the International Development

576

Research Centre (IDRC, Ottawa, Canada) and the Canada Research Chair Program to AB

577

and CMW. CMW is supported by the Canada Research Chair Program. A. Bianchini is a

578

Research Fellow from the Brazilian ‘Conselho Nacional de Desenvolvimento Científico e

579

Tecnológico’ (CNPq, Proc. #304430/2009-9, Brasília, DF, Brazil), and is supported by

580

the International Research Chair Program from IDRC. Financial support was also

581

provided by CAPES (Coordenação de Aperfeiçoamento de Pessoal do Ensino Superior,

582

Brasilia, DF, Brazil) grant (AB, P.I.) in the scope of the ‘Programa Ciências do Mar’ and

26 Page 26 of 44

by two NSERC CRD grants (CMW and Scott Smith, P.I.’s) with co-funding from the

584

International Zinc Association (IZA), the International Lead Zinc Research Organization

585

(ILZRO), the Nickel Producers Environmental Research Association (NiPERA), the

586

International Copper Association (ICA), the Copper Development Association (CDA),

587

Teck Resources, and Vale Inco. Many thanks to Jasim Chowdhury and Eric Van

588

Genderen for useful comments on the mansuscript.

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Beniash, E., Aizenberg, J., Addaddi, L., Weiner, S., 1997. Amorphous calcium carbonate transforms into calcite during sea urchin larval spicule growth. Proc. Roy. Soc. Lond. B. 264, 461-465.

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Bielmyer, G.K., T. Jarvis, B.T. Harper, B. Butler, L. Rice, S. Ryan. 2012. Metal accumulation from dietary exposure in the sea urchin, Strongylocentrotus droebachiensis. Arch. Environ. Contam. Toxicol. 63, 86–94.

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Bradford, M.M., 1976. A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein-dye binding. Anal. Biochem. 72, 248-254. De Araújo Leite, J.C., Marques-Santos, L.F., 2011. Extracellular Ca2+ influx is crucial for the early embryonic development of the sea urchin Echinometra lucunter. J. Exp. Zool. B. Mol. Dev. Evol. 318,123-33. Decker, G.L., Lennarz, W.J., 1988. Skeletogenesis in the sea urchin embryo. Development 103, 231-247. Eisler, R., 1993. Zinc hazards to fish, wildlife and invertebrates: a synoptic review, Contaminant Hazard Reviews, Report 26. U.S. Department on the Interior Fish and Wildlife Service, Patuxent Wildlife Research Center, Laurel. 1–126. European Commission (EC), 2012. Directive of the European parliament and the council amending Directives 2000/60/EC and 2008/105/EC as regards priority substances in the field of water policy. http://ec.europa.eu/environment/water/waterdangersub/pdf/com_2011_875.pdf 27 Page 27 of 44

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Kobayashi, N., 1980. Comparative sensitivity of various developmental stages of sea urchins to some chemicals. Mar. Biol. 58, 163-171. Kobayashi, N., Okamura, H., 2004. Effects of heavy metals on sea urchin embryo development. 1. Tracing the cause by the effects. Chemosphere. 55, 1403-12. Mager, E., 2012. Lead. In Homeostasis and Toxicology of Non-essential Metals. Fish Physiology, Vol. 31B (eds. C.M. Wood, A.P. Farrell, and C.J. Brauner) 185-236. Academic Press. New York. Meyer, J.S., Santore, R.C., Bobbitt, J.P., DeBrey, L.D., Boese, C.J., Paquin, P.R., Allen, H.E., Bergman, H.L., and DiToro, D.M., 1999. Binding of nickel and copper to 28 Page 28 of 44

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Nash, W.W., Poor, B.W., Jenkins, K.D. 1981. The uptake and subcellular distribution of lead in developing sea urchin embryos. Comp. Biochem. Physiol. 69C, 205-2011.

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Orström, A., Orström, M., 1942. Uber die Bundig von Kalzium in Ei und larve von Paracentrotus lividus. Protoplasma 36, 475-490.

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Pinsino, A., Matranga, V. , Trinchella, F., and Roccheri, M.C. 2010. Sea urchin embryos as an in vivo model for the assessment of manganese toxicity: developmental and stress response effects. Ecotoxicology 19, 555-562. Pinsino, A., Roccheri, M.C., Costa, C., Matranga, V. 2011. Manganese interferes with calcium, perturbs ERK signaling, and produces embryos with no skeleton. Toxicol. Sci. 123, 217-230. Pivovarova, N.B., Lagerspetz, K.Y.H., 1996. Effect of cadmium on the ATPase activity in gills of Anodonta cygnea at different assay temperatures. J. Therm. Biol. 21, 77-84. Radenac, G., Fichet, D., Miramand, P., 2001. Bioaccumulation and toxicity of four dissolved metals in Paracentrotus lividus sea-urchin embryo. Mar. Environ. Res. 51, 151-166.

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Raz, S., Hamilton, P.C., Wilt, F.H., Weiner, S., Addadi, L., 2003. The transient phase of amorphous calcium carbonate in sea urchin larval spicules: the involvement of proteins and magnesium ions in its formation and stabilization. Adv. Funct. Mater. 13, 480-486.

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Reddy, M.P., Philip, H.G., 1992. Changes in the levels of respiration and ions in the tissues of freshwater fish, under fenvalerate stress. Chemosphere. 25, 843-852.

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Rogers, J.T., Wood, C.M., 2004. Characterization of branchial lead-calcium interaction in the freshwater rainbow trout Oncorhynchus mykiss. J. Exp. Biol. 207, 813-825.

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Sancho, E., Fernandez-Vega, C., Ferrando, M. D., Andreu-Moliner, E., 2003. Eel ATPase activity as biomarker of thiobencarb exposure. Ecotoxicol. Environ. Safety. 56, 434-441.

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Sanchez-Marin, P., Santos-Echeandía, J., Nieto-Cid, M., Álvarez-Salgado, X.A., Beiras, R.,2010. Effect of dissolved organic matter (DOM) of contrasting origins on Cu and Pb speciation and toxicity to Paracentrotus lividus larvae. Aquat. Toxicol. 96, 90-102.

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Spry, D.J., Wood, C.M., 1985. Ion flux rates, acid–base status, and blood gases in rainbow trout, Salmo gairdneri , exposed to toxic zinc in natural soft water. Canadian Journal of Fisheries and Aquatic Sciences. 42, 1332-1341.

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Stark, G., 2005. Functional consequences of oxidative membrane damage. J. Membr. Biol. 205, 1-16. Supanopas, P., Sretarugsa, P., Kruatrachue, M., Pokethitiyook, P., Upatham, E.S., 2005. Acute and subchronic toxicity of lead to the spotted babylon, Babylonia areolata (Neogastropoda, Bussinidae). J. Shellfish Res. 24, 91-98.

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Tellis, S.M., Lauer, M.M., Nadella, S., Wood, C.M. 2013. Ionic status, calcium uptake, and Ca2 +-ATPase activity during early development in the purple sea urchin (Strongylocentrotus purpuratus). Comp. Biochem. Physiol., Part A: Mol. Integr. Physiol. 166, 272-277. Toyota, Y., Okabe, S., Kanamori, S., Kitano, Y., 1982. The determination of Mn, Fe, Ni, Cu and Zn in seawater by atomic absorption spectrometry after coprecipitation with lanthanum hydroxide. J. Oceanogr. Soc. Jap. 38, 357-361. Vallee, B.L., Falchuck, K.H., 1993. The biochemical basis of zinc physiology. Physiol. Rev. 73, 79-118. Viarengo, A., Pertica, M., Mancinelli, G., Burlando, B., Canesi, L., 1996. In vivo effects of copper on the calcium homeostasis mechanisms of mussel gill cell plasma membranes. Comp. Biochem. Physiol. C Pharmacol. Toxicol. Endocrinol.113, 30 Page 30 of 44

421-425.

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Vijayavel, K., Gopalakrishnan, S., Balasubramanian, M.P.. 2007. Sublethal effect of silver and chromium in the green mussel Perna viridis with reference to alterations in oxygen uptake, filtration rate and membrane bound ATPase system as biomarkers. Chemosphere. 69, 979-86. Wang, W., 1987. Factors affecting metal toxicity to (and accumulation by) aquatic organisms – overview. Environ. Int. 13, 437-457.

cr

Warnau, M.,Pagano, G. (1994). Developmental toxicity of PbCl2 in the Echinoid Paracentrotus lividus. Bull,. Environ. Contam. Toxicol. 53, 434-441.

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Warnau, M., Iaccarino, M., De Biase, A., Temara, A., Jangoux, M., Dubois, P. Pagano, G., 1996. Spermiotoxicity and embryotoxicity of heavy metals in the echinoid Paracentrotus lividus. Environ. Toxicol. Chem. 15, 1931-1936. Wilt, F.H., 1999. Matrix and mineral in the sea urchin larval skeleton. J. Struct. Biol. 126, 216-226.

M

Wilt, F.H., 2002. Biomineralization of the spicules of sea urchin embryos. Zool. Sci. 19, 253-261.

te

d

Yaroslavetseva, L. M., Sergeeva, E.P. 2002. Sensitivity of embryos and larvae of the sea urchin Strongylocentrotus intermedius to effects of copper ions. Russ. J. Mar. Biol. 28, 393-397. Yasumasu, I., 1959. Spicules of the sea urchin larvae. Zool. Mag. 68,42 (in Japanese).

Ac ce p

757 758 759 760 761 762 763 764 765 766 767 768 769 770 771 772 773 774 775 776 777 778 779 780 781 782 783 784 785 786 787 788 789 790 791 792 793 794 795 796 797 798 799 800 801 802

31 Page 31 of 44

ip t

Time (h)

24

0.177AB

0.018

0.175 a

36 0.184ABC

0.019

0.151AB

Zn exposed (µg) 0.150 x

SEM 0.010

d

SEM 0.017

0.016

0.161 x

0.014

te

12

Control (µg) 0.136A

Pb exposed (µg) 0.167 a

an

us

cr

Table 1 Embryonic dry weights measured every 12 h over the first 96 h of embryonic development in embryos exposed to Zn (117 µg/L - measured) or Pb (55 µg/L measured). Control data from Tellis et al. (2013). An asterisk (*) indicates a significant difference from control levels at the same time point as determined with a Student’s t-test (P<0.05). Values with different letters are significantly different as determined by a Fisher LSD post hoc test. Letters of different cases indicate comparisons within treatments over time; upper case letters represent comparison within controls and lower case letters represent comparison within Pb or Zn treatments. Values are means ± SEM (N = 5).

0.177 a

0.017

0.136 x

0.028

0.048

0.165 a

0.022

0.137 x

0.019

60 0.186ABC

0.006

0.281 c *

0.017

0.256 y*

0.025

72

0.209BC

0.021

0.220 ab

0.020

0.187 x

0.019

84

0.247C

0.009

0.257 bc

0.032

0.220 xy

0.019

96

0.356D

0.050

0.257 bc

0.016

0.338 z

0.028

Ac ce p

48

M

803 804 805 806 807 808 809 810 811 812 813 814 815 816 817 818 819 820 821

SEM 0.010

822

32 Page 32 of 44

Figure Legends

823

Figure 1 Whole body ion concentrations measured every 12 h in embryos exposed to Pb

824

(55 µg/L) over 96 h of development a) Calcium b) Potassium c) Sodium d) Magnesium.

825

Control data from Tellis et al. (2013). An asterisk (*) indicates a significant difference

826

from control levels at the same time point as determined with a Student’s t-test (P<0.05).

827

Values with different letters are significantly different as determined by a Fisher LSD

828

post hoc test. Letters of different cases indicate comparisons within treatments over time;

829

upper case letters represent comparison within controls and lower case letters represent

830

comparison within the Pb treatment. Values are means ! SEM (N = 5).

an

us

cr

ip t

822

M

831

Figure 2 Whole body ion concentrations measured every 12 h in embryos exposed to Zn

833

(117 µg/L) over 96 h of development a) Calcium b) Potassium c) Sodium d) Magnesium.

834

Control data from Tellis et al. (2013). An asterisk (*) indicates a significant difference

835

from control levels at the same time point as determined with a Student’s t-test (P<0.05).

836

Values with different letters are significantly different as determined by a Fisher LSD

837

post hoc test. Letters of different cases indicate comparisons within treatments over time;

838

upper case letters represent comparison within controls and lower case letters represent

839

comparison within the Zn treatment. Values are means ! SEM (N = 5).

te

Ac ce p

840

d

832

841

Figure 3 Whole body Pb accumulation (mg/kg dw) measured every 12 h over 96 h of

842

development in embryos exposed to Pb (55 µg/L). An asterisk (*) indicates a significant

843

difference from control levels at the same time point as determined with a Student’s t-test

844

(P<0.05). Values with different letters are significantly different as determined by a 33 Page 33 of 44

Fisher LSD post hoc test. Letters of different cases indicate comparisons within

846

treatments over time; upper case letters represent comparison within controls and lower

847

case letters represent comparison within the Pb treatment. Values are means ! SEM (N =

848

5).

ip t

845

cr

849

Figure 4 Whole body Zn accumulation (mg/kg dw) measured every 12 h over 96 h of

851

development in embryos exposed to Zn (117 µg/L). Values with different letters are

852

significantly different as determined by a Fisher LSD post hoc test. Letters of different

853

cases indicate comparisons within treatments over time; upper case letters represent

854

comparison within controls and lower case letters represent comparison within the Zn

855

treatment. Values are means ! SEM (N = 5).

M

an

us

850

d

856

Figure 5 Unidirectional Ca uptake rates measured every 12 h over 96 h of development

858

in embryos exposed to Pb (55 µg/L). Control data from Tellis et al. (2013). An asterisk

859

(*) indicates a significant difference from control levels as determined with a Student’s t-

860

test (P<0.05). Values with different letters are significantly different as determined by a

861

Fisher LSD post hoc test. Letters of different cases indicate comparisons within

862

treatments over time; upper case letters represent comparison within controls and lower

863

case letters represent comparison within the Pb treatment. Values are means ! SEM (N =

864

6).

Ac ce p

te

857

865 866

Figure 6 Unidirectional Ca uptake rates measured every 12 h over 96 h of development

867

in embryos exposed to Zn (117 µg/L). Control data from Tellis et al. (2013). An asterisk 34 Page 34 of 44

(*) indicates a significant difference from control levels as determined with a Student’s t-

869

test (P<0.05). Values with different letters are significantly different as determined by a

870

Fisher LSD post hoc test. Letters of different cases indicate comparisons within

871

treatments over time; upper case letters represent comparison within controls and lower

872

case letters represent comparison within the Zn treatment. Values are means ! SEM (N =

873

5)

cr

ip t

868

us

874

Figure 7 Ca2+ ATPase activity measured every 12 h over 72 h of development in

876

embryos exposed to Pb (55 µg/L). Control data from Tellis et al. (2013). An asterisk (*)

877

indicates a significant difference from control levels at the same time point as determined

878

with a Student’s t-test (P<0.05). Values with different letters are significantly different as

879

determined by a Fisher LSD post hoc test. Letters of different cases indicate comparisons

880

within treatments over time; upper case letters represent comparison within controls and

881

lower case letters represent comparison within the Pb treatment. Values are means !

882

SEM (N = 5).

M

d

te

Ac ce p

883

an

875

884

Figure 8 Ca2+ ATPase activity measured every 12 h over 72 h of development in

885

embryos exposed to Zn (117 µg/L). Control data from Tellis et al. (2013). An asterisk (*)

886

indicates a significant difference from control levels at the same time point as determined

887

with a Student’s t-test (P<0.05).Values with different letters are significantly different as

888

determined by a Fisher LSD post hoc test. Letters of different cases indicate comparisons

889

within treatments over time; upper case letters represent comparison within controls and

890

lower case letters represent comparison within the Zn treatment. Values are means ! 35 Page 35 of 44

891

SEM (N = 5).

Ac ce p

te

d

M

an

us

cr

ip t

892 893 894 895 896 897 898 899

36 Page 36 of 44

899 Figure 1 12

700

36

48

pluteus 60

84

96

300

12

24

A

100

ab

A

*

a

cd AC

150 bc

36

48

B

60

72

84

12

96

24

36

48

pluteus

60

72

84

96

350

d A a

AB

Mg mmol/kg

te bc

a

AB

Ac ce p

Na mmol/kg

CD ab BC

C

* a

36

48

60

72

84

96

d

gastrulation

blastula 12

24

36

48

pluteus 60

84 d

96 cd

bcd BC

abc

* ab ABC

200 150

72 B

250

c

ab AB

B

300

D

4000

3000

a

an gastrulation

blastula

M

c

24

ab

Control Pb exposed

Control Pb exposed

12

d

hours

hours

c

A

a

0 24

d

ab

50

b

0 12

A

100

*

96

AC

AC

bc

*

84

cr

AB c

A

72

C

200

CD

BC

200

2000

60

250

d

300

5000

48

e D

400

902 903 904 905

36

pluteus

bb

D e

500

Ca mmol/kg

72

aa

gastrulation

blastula

us

600

24

gastrulation

ip t

blastula

K mmol/kg

900 901

ABC

ab AC

a

a A

A

36

48

A

100

a

1000

50

0

12

24

36

48

60

72

0 84

96

12

24

hours

906 907 908 909 910 911 912 913 914 915 916 917

Control Pb exposed

60

72

84

96

hours Control Pb exposed

37 Page 37 of 44

Figure 2 gastrulation

blastula 12

600

24

36

48

60

72

84

96

12

300

D

a

gastrulation

blastula

pluteus

24

36

48

pluteus 60

CD

200

BC b

*

200

AB

A

* a

* * A

a

A a

12

24

* b

bc AC

150 A 100

48

60

72

84

96

12

12

4000

24

gastrulation 36

48

pluteus 60

72

84

96

D

c bc CD

3000 C

BC

bc

b

te

b

ab

b

2000

d

c

AB

AB

AB

Ac ce p

A a

1000

24

36

48

60

72

blastula

12

400

24

60

72

84

96

hours

gastrulation

36

48

pluteus 60

72

84

96

* d

d B

* cd

300 bc

bc

abc

* ABC bc

AC

200 A

BC

ABC

ab

A

A a

0 84

96

12

24

hours

Control Zn exposed

48

100

0

12

M

blastula

36

Control Zn exposed

Control Zn exposed

922 923 924

24

an

hours

Mg mmol/kg

36

ab

B

0

0

bc

a

B

a

a 50

a

ab

*

b

a

A

us

300

AC c

AC

cr

K mmol/kg

Ca mmol/kg

400

Na mmol/kg

96 C

D

925 926 927 928 929 930 931 932 933 934 935 936

84

b 250

500

100

72

ip t

918 919 920 921

36

48

60

72

84

96

hours Control Zn exposed

38 Page 38 of 44

gastrulation

blastula 12

24

36

48

60

84

96

* e

an

50

30

* de

* de

M

40 Pb mg/Kg

72

us

60

pluteus

ip t

Figure 3

cr

937 938 939 940 941 942 943 944 945 946

*

cd

d

20

*

*

*

10

te

bcd

bc

b a

A

A

Ac ce p A

A

A

A

A

A

0

12

947 948 949 950 951 952 953 954 955 956 957

24

36

48

60

72

84

96

hours

Control Pb exposed

39 Page 39 of 44

350

12

pluteus

gastrulation

24

36

48

60

A b

* b

250 200 ab

150

A

M

Zn mg/Kg

A

A

84

96

* b

A b

an

300

72

cr

blastula

ip t

Figure 4

us

958 959 960 961 962 963 964 965 966

A

A ab

a A

d

a

te

100

Ac ce p

50 0

12

967 968 969 970 971 972 973 974 975 976 977

24

36

48

60

72

84

96

hours

Control Zn exposed

40 Page 40 of 44

Figure 5

12

24

36

48

60

72

E

us

200

DE 150

CD

96

CD

an

d

BC

100

50

A

a

0

24

Ac ce p

12

B

c

36

48

B

* ab

d

* b

* c

M

* c

te

Ca uptake pmol/larva/h

84

cr

250

pluteus

gastrulation

blastula

ip t

978 979 980 981 982 983

60

72

84

* ab

96

hours

Control Lead

984 985 986 987 988 989 990 991 992 993 994 995 996

41 Page 41 of 44

12

24

36

48

60

E

an

DE CD d

M

* cd

100

CD

B

*

bc

d

Ac ce p

A a

96

BC

c 50

84

B *

*

cd

bc

ab

te

Ca uptake pmol/larva/h

200

150

72

us

250

pluteus

gastrulation

blastula

ip t

Figure 6

cr

997 998 999 1000 1001 1002 1003

0

12

1004 1005 1006 1007 1008 1009 1010 1011 1012 1013 1014

24

36

48

60

72

84

96

hours

Control Zinc

42 Page 42 of 44

blastula 12

gastrulation

24

36

pluteus

48

60

2.0

a

us AB

M

1.0

A

d

A

0.5

Ac ce p

0.0

12

1023 1024 1025 1026 1027 1028 1029 1030 1031 1032 1033 1034

B

B

an

* a

B

* b

te

Ca ATPase activity Pi mmol/mg/h

ab

72

ab

* b 1.5

ip t

Figure 7

cr

1015 1016 1017 1018 1019 1020 1021 1022

24

36

48

60

72

hours

Control Pb exposed

43 Page 43 of 44

blastula 1.5

12

gastrulation

24

36

pluteus

48

60

us AB

* a * a

* a

a

A

M

* a

an

1.0

d

0.5

Ac ce p

0.0 12

1042 1043

B

A a

72

te

Ca ATPase activity Pi mmol/mg/h

B B

ip t

Figure 8

cr

1035 1036 1037 1038 1039 1040 1041

24

36

48

60

72

hours

Control Zn exposed

44 Page 44 of 44