Development of the sea urchin Heliocidaris crassispina from Hong Kong is robust to ocean acidification and copper contamination

Accepted Manuscript Title: Development of the sea urchin Heliocidaris crasssispina from Hong Kong is robust to ocean acidification and copper contamination Authors: Narimane Dorey, Elizaldy Maboloc, Kit Yu Karen Chan PII: DOI: Reference:

S0166-445X(18)30605-2 https://doi.org/10.1016/j.aquatox.2018.09.006 AQTOX 5022

To appear in:

Aquatic Toxicology

Received date: Revised date: Accepted date:

6-7-2018 7-9-2018 10-9-2018

Please cite this article as: Dorey N, Maboloc E, Karen Chan KY, Development of the sea urchin Heliocidaris crasssispina from Hong Kong is robust to ocean acidification and copper contamination, Aquatic Toxicology (2018), https://doi.org/10.1016/j.aquatox.2018.09.006 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.

Development of the sea urchin Heliocidaris crasssispina from Hong Kong is robust to ocean acidification and copper contamination

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Narimane Dorey1†, Elizaldy Maboloc2 and Kit Yu Karen Chan1 Division of Life Science, Hong Kong University of Science and Technology, Clear Water Bay,

Kowloon, Hong Kong 2

School of Science, Hong Kong University of Science and Technology, Clear Water Bay,

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Kowloon, Hong Kong

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† Correspondence to: Narimane Dorey,

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Institute of Marine Research, Nordnesgaten 50, Bergen, 5020 Norway

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E-mail: [email protected]

Highlights

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Ocean acidification will increase the fraction of the most toxic form of copper, increasing its bioavailability to marine organisms We tested the hypothesis that copper contaminated waters are more toxic to sea urchin larvae under future pH conditions in three laboratory experiments Larvae are robust to the pH and the copper levels we tested (little/no mortality) However, significant sub-lethal effects, could have indirect consequences on survival

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ABSTRACT

Metallic pollution is of particular concern in coastal cities. In the Asian megacity of Hong Kong, despite water qualities have improved over the past decade, some local zones are still particularly affected and could represent sinks for remobilization of labile toxic species such as copper. Ocean

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acidification is expected to increase the fraction of the most toxic form of copper (Cu2+) by 2.3folds by 2100 (pH ≈7.7), increasing its bioavailability to marine organisms. Multiple stressors are likely to exert concomitant effects (additive, synergic or antagonist) on marine organisms.

Here, we tested the hypothesis that copper contaminated waters are more toxic to sea urchin larvae under future pH conditions. We exposed sea urchin embryos and larvae to two low-pH and

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two copper treatments (0.1 and 1.0 μM) in three separate experiments. Over the short time typically

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used for toxicity tests (up to 4-arm plutei, i.e. 3 days), larvae of the sea urchin Heliocidaris

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crassispina were robust and survived the copper levels present in Hong Kong waters today (≤0.19

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μM) as well as the average pH projected for 2100. We, however, observed significant mortality

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with lowering pH in the longer, single-stressor experiment (Expt A: 8-arm plutei, i.e. 9 days).

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Abnormality and arm asymmetry were significantly increased by pH or/and by copper presence (depending on the experiment and copper level). Body size (d3; but not body growth rates in Expt

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A) was significantly reduced by both lowered pH and added copper. Larval respiration (Expt A)

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was doubled by a decrease at pHT from 8.0 to 7.3 on d6. In Expt B1.0 and B0.1, larval morphology (relative arm lengths and stomach volume) were affected by at least one of the two investigated

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factors.

Although the larvae appeared robust, these sub-lethal effects may have indirect

consequences on feeding, swimming and ultimately survival. The complex relationship between pH and metal speciation/uptake is not well-characterized and further investigations are urgently

needed to detangle the mechanisms involved and to identify possible caveats in routinely used toxicity tests. Keywords: Ocean acidification; Metallic pollution; Developmental biology; Invertebrates;

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Larvae; Bioaccumulation

1. INTRODUCTION

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Coastal marine organisms are exposed to a number of anthropogenic stressors. In particular for

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coastal cities, the influence of chronic seawater pollution, including metallic pollutants, is not

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negligible. With a human population of seven million, metal pollution has been a historical concern

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for the sub-tropical megacity Hong Kong (see reviews by Blackmore, 1998; Morton & Blackmore, 2001). Extremely high levels of water pollution have been recorded, e.g., seawater copper

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concentration reached approx. 0.19 μM (12 μg L-1) in 1983-89 (Blackmore, 1998) and levels up to

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4300 μg g-1 dry sediment were measured in 2015 (EPD, 2016). Although seawater quality has

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generally improved over the last 15 years (annual reports from Hong Kong Environmental Protection Department are available since 1986: e.g. EPD, 2016), some local "hotspots", in

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particular enclosed areas (e.g. typhoon shelters, Victoria and Tolo Harbours; EPD, 2016) and areas influenced by river discharges (Deep Bay and the North Western waters), still contain sediments

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with high levels of metal pollutants. These sediments may serve as reservoirs and sources of rapid re-mobilization (e.g. sediment re-suspension during dredging for land reclamation, strong summer storms; Morton and Blackmore, 2001), making toxic metals readily bioavailable to marine organisms.

Global change is taking place under such background contamination, exposing marine organisms simultaneously to multiple stressors and their possible cumulative impacts. While the interaction between temperature and metal toxicity has been studied (see review by Sokolova and

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Lannig, 2008), fewer studies have focused on the effects of ocean acidification (OA) on metal bioaccumulation (Belivermis et al., 2016; Dorey et al., 2018; Houlbrèque et al., 2011; Ivanina et al., 2013; Lacoue-Labarthe et al., 2012, 2011, 2009) and toxicity (Benedetti et al., 2016; Campbell et al., 2014; Fitzer et al., 2013; Götze et al., 2014; Ivanina et al., 2015, 2014, Ivanina and Sokolova, 2015, 2013, Lewis et al., 2016, 2013; Li et al., 2017; Nardi et al., 2017; Shi et al., 2016). The

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average ocean pH drop expected for 2100 (-0.3 to -0.4 pH units) is likely alter metal bioavailability

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by 1) metal re-mobilization from the sediments (see e.g. Riba et al., 2003 at pH 7.5 and 6.5); 2)

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changing metal speciation: e.g. the concentration of the most toxic form of copper, Cu2+, is

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expected to increase by 115% by 2100 (Millero et al., 2009; Stockdale et al., 2016), and 3) interacting with ionoregulation activities, thus changing the organisms’ coping abilities toward

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toxicants. Studies investigating the combined effects of pH on copper toxicity tend to show an

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increase of metal toxicity associated with lowering pH (Campbell et al., 2014; Fitzer et al., 2013;

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Ivanina et al., 2013; Lewis et al., 2016, 2013). Additionally, OA alone already compromise organisms’ performance. For instance, lowered pH affects the physiology of the early

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developmental stages of sea urchins, e.g. disrupted acid-base regulation (e.g. Hu et al., 2018; Stumpp et al., 2011b), feeding (e.g. Hu et al., 2017; Stumpp et al., 2013), increased metabolism

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(e.g. Dorey et al., 2013; Pan et al., 2015) and delayed growth (Chan et al., 2015a; Dorey et al., 2013; Foo et al., 2016; García et al., 2015a; Martin et al., 2011). OA also modulated expression in genes related to immunity (Runcie et al., 2017), metabolism (Evans et al., 2017) or biomineralization (Kurihara et al., 2012; Padilla-Gamiño et al., 2013).

In this study, we performed three experiments to investigate the effects of pH alone (pHT 8.0, -0.4 and -0.7 pH units, Expt A) and the effects of pH combined with environmentally-relevant (Expt B0.1: 0.1 μM) and highly toxic (Expt B1.0: 1.0 μM) levels of copper on the early

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development of sea urchins (embryos to 4 or 8-arm plutei). We studied the larvae of the sea urchin Heliocidaris crassispina, an edible and ecologically-important species in Hong Kong (Wai and Williams, 2006). More broadly, sea urchins are commonly-used in toxicity assays as their fertilization success and developmental endpoints are simple and reliable indicators (Bougis, 1965; Carballeira et al., 2012; His et al., 1999). Copper, an essential element, is actively accumulated in

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the body of larval sea urchin (Radenac et al., 2001) but excess copper can have sub-lethal and

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lethal effects (EC50 for larvae of sea urchins ranged from 0.17 to 1.8 μM depending on the species:

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Bielmyer et al., 2005; Kobayashi and Okamura, 2004; Warnau et al., 1996). Here, we hypothesized

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that copper combined to lowered pH will result in additive negative effects, such that both lethal (mortality rates) and sub-lethal (abnormalities, morphology, growth and respiration) effects are

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larger when larval sea urchins are exposed to high level of copper and low pH.

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2. MATERIALS AND METHODS

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2.1. Experimental design

In the first experiment (Expt A), we focused on the effects of pH alone on the development of

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Heliocidaris crassispina (A. Agassiz, 1864) until the 8-arm pluteus stage (9 days postfertilization). We reared larvae for eight days at three different pH: control pH (~8.0), -0.4 pH units and -0.7 pH units (Table 1). At the collection site of these urchins, Sai Kung, Hong Kong, pH is on average 8.19 ± 0.22, but extreme values have been recorded from 7.2 to 8.6 pH units (Pecquet et al., 2017).

In the second set of experiments (Expt B1.0 and B0.1), we investigated the combined effects of pH and copper until the 4-arm plutei stage: larvae were reared at the aforementioned pH conditions (Table 1) and two copper treatments (i.e. six conditions per experiment). Copper was

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added in half of the jars (7.5 μl of a CuSO4 dilution, Sigma-Aldrich, for a final Cu2+ concentration of 1.0 μM; i.e. 63.50 μg L-1 for Expt B1.0; and 7.5 μl of a 1/10th dilution for a final concentration of 0.1 μM; i.e. 6.35 μg L-1 for Expt B0.1) before fertilization and on d1, d2 and d3. While the copper concentration chosen for Expt B0.1 was ecologically relevant, the concentration used in Expt B1.0 was five to twenty times higher than those reported from nearby contaminated areas

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(e.g. in Baijiao, Jiulong River Estuary, China average concentration was 0.15 μM and ranged from

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0.04 to 0.20 μΜ: Weng and Wang, 2014). We chose the 1.0 μM concentration as a sub-lethal

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challenge since earlier studies reported sub-lethal effects but no lethal effects in Paracentrotus

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lividus larvae at 0.8 μM (Radenac et al., 2001: 100% mortality at 4.0 μM; EC50 at 1.8 μM in His et al., 1999). Each treatment had three culture replicates (i.e. 9 jars in Expt A and 18 jars in Expt

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B1.0 and Expt B0.1).

2.2. Adult collection, spawning and larval cultures

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Adults of the sea urchin Heliocidaris crassispina were procured from a local urchin farm located

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in Leung Sheun Wan, Sai Kung (Hong Kong) in Spring 2015 and transported in seawater within the hour to the Coastal Marine Laboratory (Hong Kong University of Science and Technology,

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Hong Kong). Animals were acclimated for a minimum of three months in a continuously aerated flow of natural seawater (average pHT≈7.95), following natural fluctuations of temperature (mean±SD: 25.8±1.2°C, N=18) and salinity (32.9±1.8, N=24). Adults were fed three times a week on commercially available dried Kombu algae (Saccharina japonica) rehydrated.

Spawning was induced by KCl injection (1 mL at 0.5M in 0.2μm-filtered seawater: FSW) across the peristomial membrane at room temperature. Sperm (Expt A and Expt B1.0: two males and Expt B0.1: one male) was collected dry and kept on ice until use. Sperm motility was checked

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at the microscope. Eggs (Expt A and Expt B1.0: three females, Expt B0.1: four females) were collected in FSW, rinsed and then pooled in clean FSW and sperm was added at a final density of ≈1000 sperm mL-1. Fertilization took place for 15 minutes in FSW at pHT≈7.95. Fertilization success was confirmed by the presence of fertilization envelop. Seawater was changed twice in the first hour after confirmation of >95% fertilization success and embryos were left to develop at

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≈26°C and pHT≈7.95 until the two-cell stage (≈1h30 post-fertilization).

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In each experiment, embryos were added to acid-washed glass jars containing 1.5 l of

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aerated FSW previously stabilized to the appropriate three or six experimental conditions at a

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density of ca. 3.5-4.5 embryos mL-1 (see Experimental design). Embryos and subsequent larvae

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were reared at a salinity of 32.0-33.0 (pre-adjusted with MiliQ water) in a thermo-constant bath

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(26 °C, Table 1) for nine (Expt A) or three (Expts B) days. In Expt A, water was changed 5 days post-fertilization with new FSW pre-equilibrated to the right pH.

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From day 2 post-fertilization (48h, d2), larvae were fed daily with the microalgae

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Rhodomonas sp. cultured in the laboratory (F2 medium: Guillard & Ryther 1962; T=22°C; S=32; light/dark cycle: 12h/12h). To prevent changes in food availability due to the experimental

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conditions, algal densities were checked daily using a Coulter counter (Z™ Series Coulter Counter Cell and Particle Counter, Beckman Coulters, Brea, USA) and adjusted in every larval culture (3000 cells mL-1 on d2, 4000 cells mL-1 on d3, and 5000 cells mL-1 afterwards). In order to discern indirect effects of copper and pH levels on the algal food, we exposed Rhodomonas sp to similar treatments for 24-h (initial density: 5000 cells mL-1). After 24-h exposure, each jar was sampled

twice and counted at the Coulter counter. At this density and exposure period, copper had no effect on algal growth but pH did (Fig. S1): while the algae grew in the control pH treatment (46±31% for pH=8.01 ±0.02 pHT units, N=6), algal density decreased in both lowered pH treatments (-

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23±8% for both pH=7.64 ±0.02 and pH=7.38 ±0.06 pHT units, N=12). 2.3. Seawater carbonate and copper chemistry

Each culture jar was continuously mixed and aerated through air bubbling. The pH was controlled by constant addition of a mix of compressed air and pure CO2, controlled by a thermal mass flow

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controller (GFC 17 from Aalborg, Orangeburg, USA; ± 1% FS accuracy) – to the exception of

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three low pH cultures in Expt A. Due to logistic constraints, the pH in these three cultures was

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adjusted by discontinuous bubbling of pure CO2 into the seawater controlled by pH-stat computer

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systems (R-WP017 CO2 Regulator, Easy-Aqua, China). pH, alkalinity and temperature were monitored every day following the recommendations of Dickson et al. (2007). pH was measured

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on the total scale (pHT) after calibration using TRIS (Tris/HCl) buffer solution with a salinity of

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33.0 (provided by the Dickson Lab at Scripps Oceanographic Institute) and pHT was used for adjustment of the pH-stat systems settings. Total alkalinity (TA) was assessed potentiometrically

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on filtered samples (0.2 μm) using a computer-driven titration system (905 Titrando mounted with

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a glass electrode: Unitrode with Pt 1000, Metrohm, Switzerland, and calibrated on the National Bureau of Standards scale). In Expt A, samples for TA were taken in all the jars before (duplicates

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of N=9) water replacement on d5 and at the end of the experiment (duplicates of N=9). In both Expts B, samples for TA were taken on d0 (duplicate samples of stock FSW) and in each jar on d3 (N=18 samples). TA was calculated using a Gran function as described by Dickson et al. (2007). The carbonate system parameters (pCO2, Ωa and Ωc) were calculated from these measurements

using the R package seacarb (Lavigne and Gattuso, 2011) with the dissociation constants from Mehrbach et al. (1973) as refitted by Dickson and Millero (1987). To follow copper concentrations, seawater samples from jars containing copper (Expts

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B1.0 and B0.1: three random jars each) and from jars without copper (Expt B1.0: two random jars) were taken before the experiment. Water was also sampled during the subsequent days of each experiments (d1, d2 and d3: N=9 jars with copper and d1: N=1 jar without copper for Expt B1.0; d2: N=3 and d3: N=7 random jars with copper for Expt B0.1). Samples were stored at room temperature until copper concentration analysis. Before the analysis, nitric acid (Merek, ACS,Reag.

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Ph Eur 69%) was added to a final concentration of 0.04% in each sample (5.88 μl of 69% HNO3 in

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10-mL samples; Merck Suprapur). The metal concentration was measured in two 1-mL

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subsamples per seawater sample with inductively coupled plasma-atomic emission spectrometry

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(ICP-OES; Perkin-Elmer 7000 DV; instrument detection limit of one part per billion i.e. ≈0.0157

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μM range for copper), as described in Cao and Wang (2017). A calibration curve (R2 > 0.9999)

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was established before the analysis with a range of dilutions of standard Cu solution (Perkin-Elmer,

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Wellesley, MA, USA).

2.4. Biological measurements

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Mortality and abnormality

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Over the course of the three experiments, each larval culture was sampled daily (10 mL x 2) to assess larval density. Samples were immediately fixed with a drop of formalin solution (4% in FSW, buffered at pH 8.3), counted and stored in 4% formalin solution at 4°C until further measurements.

In Expt A, survival (S, %) was calculated for each culture (N=9) on each day (9 days) as the proportion of larval density divided by the maximum number of larvae ever counted in the corresponding culture. Mortality rates (MR) were then calculated from the regression coefficient

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of the significant linear regression of S with i) time (% larvae hr-1) and ii) larval theoretical body length (% larvae μm-1, see following section for the calculation).

Each day, the proportion of abnormal larvae was recorded after His et al. (1999) definition. Abnormality included wrong or incomplete early cellular divisions, wrong arm orientation and deformities (e.g. arm completely in the opposite direction), and very strong arm asymmetry (e.g.

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one atrophied arm).

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Morphometric analyses

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Photos of larvae were taken every day for the first three days of the three experiments as well as on d5, d6 and d8 for Expt A (Nikon Eclipse e400 mounted with an Olympus BX41). Larvae were

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measured with the software ImageJ (U. S. National Institutes of Health, Maryland, USA; 4 to 20

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larvae per sampling point and culture, depending on larval mortality; N=514 larvae measured for

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Expt A, N=739 larvae for Expt B1.0 and N=606 larvae for Expt B0.1).

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In Expt A, larval body length (BL, μm) and the length of both post-oral arms (POL) was measured. In Expts B, in addition to BL and POL, we measured the length of antero-lateral (AL)

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arms, post-oral (POG) and antero-lateral (AG) arms’ gap as well as the volume of the stomach (Svol, μm3; spherical volume calculated from the average diameter; see Dorey et al., 2013). Arm symmetry (Expt A: POLsym; Expts B: POLsym and ALsym; %) was assessed as the ratio between the pairs of arms (i.e. perfect symmetry when ratio is equal to 100%).

POL, AL and Svol were each significantly correlated with BL (Fig. S6: Pearson correlation, test for correlation between paired samples: correlation >0.83). Because the relationships were linear over the timeline considered, we normalized the length of the longest arm

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and the stomach volume, each by the corresponding BL. The arm gaps (POG and AG) were strongly positively correlated with their respective arm lengths (POL and AL respectively; Fig. S6; e.g. Pearson correlation, test for correlation between paired samples: POG-POL: cor=0.89, p<0.001 and AG-AL: cor=0.73, p<0.001 for Expt B1.0). These results on POG and AG are not presented in full for clarity.

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In Expt A, body length growth rates (GR) were calculated for each culture as the regression

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coefficient of the significant logarithmic relationship between measured BL and developmental

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time (t; hr): 𝐵𝐿 = 𝐺𝑅 × log (𝑡) + 𝐼𝑛𝑡𝑒𝑟𝑐𝑒𝑝𝑡. In order to use the BL for the mortality rates (%

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larvae μm-1; see paragraph above), this equation was used to calculate the theoretical BL on every

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𝐺𝑅𝑃𝑂𝐿 × 𝐵𝐿 + 𝐼𝑛𝑡𝑒𝑟𝑐𝑒𝑝𝑡.

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day of counting. In Expt A, POL growth rates (GRPOL) were calculated relative to BL: 𝑃𝑂𝐿 =

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Respiration

In Expt A, larval respiration was measured on d3 and d6 with modified protocols after Marsh and

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Manahan (1999) from all cultures available at the time of the measurement. Larvae were handpicked under the microscope, thoroughly rinsed with corresponding-pH FSW (0.2 μm, air bubbled

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at pH≈7.95, additional pure CO2 bubbling for lowered pH) and incubated into 200 μl respiration micro-wells filled with FSW at the corresponding pH (N= 10 larvae per well on d3 and N= 4 to 10 on d6, depending on size and availability) at 25°C. We duplicated measurements for each culture (i.e. 9 wells randomly filled with larvae in two simultaneous 24 micro-wells plates). Oxygen

concentration within the wells were measured every 15 to 30 sec. with fluorescence micro-plate readers (complete respiration system by Loligo PreSens, Netherlands). Oxygen concentration was monitored for 6 to 17 hours. Respiration due to potential bacterial contamination was evaluated in

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wells devoid of larvae (N= 5 to 6 bacterial controls: two wells per pH level and per micro-plate) following the same procedure and never exceeded 10% of the initial oxygen content. Respiration results were calculated using linear regressions and corrected by the number of larvae per well, the average measured size of the larvae in each culture (O2 consumption: nmol O2 hr-1 larva-1 μm-1) and the respiration in the bacterial controls. The pO2 inside the wells never dropped below 80%

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air saturation.

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Larval copper content

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On d3 of Expt B0.1, all jars were completely filtered (61 μm mesh) and larvae were thoroughly rinsed with FSW at the corresponding pH before being gently transferred to Eppendorf tubes. The

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larvae were left to settle for 5 minutes and the supernatant water was removed by pipetting. The

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larvae were immediately frozen in liquid nitrogen and conserved at -80°C until further analyzes. The average wet weight of each larval sample was determined (balance 262 SMA-FR, Dietikon,

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Precisa, Switzerland) with 0.01 mg accuracy. Larvae were then freeze-dried (ilShin Biobase

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FDA8518, Netherlands) at -80°C for 48 hours. The dried larvae were weighed (3.4 to 5.4 mg) and then digested with 0.1 mL 70% nitric acid in 80°C for two hours and then at 110°C overnight.

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Before copper analysis, a calibration curve (R2 > 0.9999) was established with a range of standard sample dilutions. Copper concentrations in diluted (35x), digested larval samples were quantified using inductively coupled plasma-mass spectrometry (ICP-MS: PerkinElmer NexlON 300X, MA, USA) with EPA Method 6020A (0.1 μg g-1 dry weight method detection limit for Cu).

2.5. Data analyses All the statistical analyses and graphic preparations were carried out with the software R (R Development Core Team 2008), with significance levels of α=5%. The relationships between

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parameters were tested using logarithmic or linear regressions. To discriminate between an indirect effect (due to delayed development) and a direct effect of pH, we normalized mortality and respiration rates to theoretical and measured BL respectively (Dorey et al., 2013; Stumpp et al., 2011a). All data were checked for normality (Shapiro-Wilk test) and homoscedasticity (Bartlett or Flinger tests). ANOVAs (one, two and three factors) were used to test differences between

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treatments, and time. When the distributions were skewed, we transformed the data to normal

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distribution before the ANOVAs, in those cases they are indicated in Table S2. One-factor

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ANOVAs were replaced by non-parametric tests when necessary (KW: Kruskal-Wallis). Details

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of the statistical analyses and results are presented in Table S2.

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3. RESULTS

3.1. Culture conditions (copper and pH)

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Temperature was uniform in all the jars (26.1 ± 0.3°C, N=207; Table 1, details in Table S1). The

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control conditions did not significantly differ across experiments (pHT≈7.97±0.03 units, pCO2≈465±48 ppm, Ωar= 2.79±0.19, N=69; KW: χ2=5.59, df=2, p=0.061). The medium and low

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pH conditions of Expt B0.1 were significantly different from the remaining two experiments (pHT=7.88 vs. pHT≈7.63-7.65, and pHT=7.21 vs. pHT≈7.27-7.31, p<0.0001) due to a miscalibration of the GF unit at the start of the Expt B0.1. In the low pH treatments, seawater was always under-saturated with respect to aragonite (Ωar= 0.66±0.18) and calcite. Alkalinity (TA)

was specific to each experiment (N=36, 19 and 20 for Expts A, B1.0 and B0.1 respectively), ranging from 1.98 to 2.37 mmol kg-1 but did not vary by more than 8% within each experiment. The measured average copper concentration in Expt B1.0 was of 0.54 μM i.e. 34.31±8.13

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μg L-1 (N=33) and of 0.07μM i.e. 4.71±1.38 μg L-1 (N=11) in Expt B0.1. Each sample was measured twice (average variation of 5.7% between two measurements). For the Cu-free control jars tested (N=3 for ExptB1.0 and ExptB0.1), the copper concentration measured was always under the detection limit.

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3.2. Expt A: Effect of pH on larval development to the 8-arm plutei stage

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Mortality, abnormality and arm symmetry

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By d8, two of the three cultures at pHT 7.3 had lost more than 95% of their larvae (recorded

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densities <0.1 larvae.mL-1) while the cultures at pHT 8.0 had at least 56% larvae remaining (48%

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at pHT 7.6; Table S2). Mortality rates did not significantly differ between the pH treatments (Fig.

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1.a by time: p=0.061 or Fig. 1.b by size: p=0.59). However, there was a significant negative relationship between average jar pHT and mortality rates (computed against time; slope= 0.24 %

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larvae hr-1 pHT-1, p=0.021, R2=55%), indicating that mortality increased when the jar pHT was lowered (0.45±0.03 % larvae hr-1 at pHT 7.3 vs. 0.28±0.05 and 0.32±0.05 at pHT 8.0 and 7.6

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respectively).

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The frequency of abnormalities increased over time (12±32% on d1 to reach 37±27% on

d7) and with lowering pH (ANOVAII: Time: p<0.0001, pH: p=0.0016 and pH x Time: p=0.02). At pHT 7.3 (Fig 1.c: 32.6±29.8%) abnormalities were 1.6 to 2.4 times more frequent than at 7.6 (20.3±22.3%) and 8.0 (13.6±14.9%). POL symmetry index remained unchanged (≈86.3±12.2%;

Fig. S2.a) with respect to time and pH treatment (ANOVAII: Time: p=0.93; pH: p=0.83; pH x Time: p=0.53). Growth rates: BL and POL

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Larval body growth rates relative to developmental time (≈87.5±8.6 μmBL log(hr)-1) were not significantly affected by the pH treatment (Fig. 1.d; ANOVAI: p=0.39), nor by the mean pHT of each jar (linear regression: p=0.35; pHT 8.0: 93.2±4.5 vs. 7.6: 83.2±5.1 and 7.3: 86.1±12.9 μmBL log(hr)-1).

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Post-oral arm (POL) growth rates relative to body length (≈1.7±0.3 μm μmBL-1) were not

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significantly affected by the pH treatment (Fig. S2.b; ANOVAI: p=0.14), nor by the mean pHT of

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each jar (linear regression: p=0.073; pHT 8.0: 2.00±0.09 μm μmBL-1 vs. 7.6: 1.62±0.23 μm μmBL-1

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and 7.3: 1.52±0.38 μm μmBL-1).

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Respiration on d3 and d6

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Larval respiration rates were normalized by body size (pmolO2 hr-1 μmBL-1). On d3, average respiration rate nearly doubled at the lowered pH (Fig. 1.e: pHT 7.3: 0.58±0.29 and 7.6: 0.50±0.22

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pmolO2 hr-1 μmBL-1) compared to that of the control (8.0: 0.30±0.17 pmolO2 hr-1 μmBL-1), although

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the values were not statistically significantly different (ANOVAI: p=0.058; linear regression with average pHT: p=0.053). On d6, the doubling in average respiration (Fig. 1.f: pHT 8.0: 0.71±0.13,

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7.6: 0.89±0.13 and 7.3: 1.43±0.03 pmolO2 hr-1 μmBL-1) was significant (ANOVAI: p=0.0001; linear regression with average pHT: slope=0.95 pmolO2 hr-1 μmBL-1 pHT unit-1; p=0.00016; R2=74%), even though we lacked statistical power due to high mortality in pHT 7.3.

3.3. Expts B1.0 and B1.0: Effect of pH and copper (1.0 and 0.1 μM) on 4 arm-plutei Mortality In Expt B1.0, regressions on the first three days did not show any significant correlations between

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larval density (≈ 5.0±0.9 larvae.mL-1) and time (the ANOVAIII of pH x Copper x Time did not highligth any significant relationships either). It was neither the case in Expt B0.1, albeit a decreasing trend (Fig S3.b; from d1: 3.9±0.7 to d3: 2.8±1.3 larvae.mL-1) with one exception (one replicate at 7.9 pHT without copper treatment that experienced 100% mortality after 3 days).

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Therefore, we only analyzed the data on d3 for the effect of pH or/and copper on mortality.

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Larval density (d3) was not affected by pH or/and copper in Expt B1.0 (Fig. S3.a;

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ANOVAII: pH: p=0.36, Copper: p=0.52; pH x Copper: p=0.58) nor in Expt B0.1 (ANOVAII: pH:

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p=0.18, Copper: p=0.16; pH x Copper: p=0.23). This was still true in Expt B0.1 even when the 8.0 and 7.9 pH treatment were pooled (ANOVAII: pH: p=0.08, Copper: p=0.17; pH x Copper:

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p=0.70).

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Abnormality and arm symmetry

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In Expt B1.0, the presence of copper significantly increased frequency of abnormalities (Fig. 2.a,

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d3: 24.8±32.1% with copper vs. 5.1±4.7% without; ANOVAII: Copper: p=0.019). While the pH treatment did not significantly affect abnormality (p=0.055), the effect of copper was mostly driven

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by an increase of the abnormality in two of the three copper replicates at control pH (Fig. 2.a: 69 and 90% abnormality; interaction pH x Copper: p=0.013). These high scores were associated with increased propotions of under-developed larvae and flat prisms or plutei. Symmetry in arms can be used as a second indicator for normal development. The pH treatment significantly affected POL symmetry index (Fig. S5.a, d3; 7.6: 94.6±5.7% vs. 8.0: 90.9±9.5% and 7.3: 89.7±9.0%;

N=233) but not copper (ANOVAII: pH: p=0.0001; Copper: p=0.85). The interaction of pH and copper significantly affected POL symmetry (pH x Copper: p=0.026), indicating a small decrease in asymmetry in the 7.3 pH treatment in the presence of copper (91.6±8.1% vs. 87.2±3.6%

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symmetry without copper). In contrast, copper significantly but only marginally increased AL symmetry index (Fig. S5.c, d3; 79.7±13.3% vs. 75.7±13.8% without copper; N=230) while it was not affected by the pH treatment (ANOVAII: Copper: p=0.020; pH: p =0.57; pH x Copper: p =0.26).

In Expt B0.1, only pH significantly influenced abnormality, with the lowest scores found

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at pHT 7.9 (Fig. 2.b, d3: 3.3±3.9% vs. 8.0: 9.0±5.9% and 7.2: 9.9±4.7%; ANOVAII: pH: p=0.006;

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Copper: p=0.10; pH x Copper p=0.22). POL symmetry index was slightly but significantly reduced

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by lowering the pH (Fig. S5.b, d3: 8.0: 95.4±4.8% and 7.9: 95.6±5.3% vs. 7.2: 93.1±6.4%; N=182;

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ANOVAII: pH: p=0.004) and significantly decreased by ~4% due to the presence of copper

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(Copper: p<0.00001; pH x Copper: p=0.65). AL symmetry index was significantly affected by

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copper and by the interaction between the pH and copper (Fig. S5.d, d3; ANOVAII on arc-sin square-root transformed data: Copper: p<0.00001; pH x Copper: p=0.041; pH: p=0.48; N=182).

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The symmetry index was differently affected by the presence of copper depending on the pH treatment, with a decrease by -7.6% (82.8-75.2%) and -2.2% (80.7-78.5%) at pH 8.0 and 7.9

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respectively and an increase by +4.6% (74.1-78.7%) at pHT 7.2.

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Body length

Larval body size (BL, μm, d1 to d3; Fig. 2.c-d) significantly increased with time by 57-48 μm between d1 and d2 and by 108-48 μm between d2 and d3 (Expt B1.0-B0.1respectively ANOVAIII: Time: p<0.0001, NB1.0=606, NB0.1=739).

Lowering pH significantly decreased body size in both experiments (pH: p<0.0001), depending on the day. On average, from pHT 8.0 to pHT 7.3 in Expt B1.0, body size changed by 6.9% on d1, -5.8% on d2 and -8.9% on d3 (pH x Time: p=0.00036). From pHT 8.0 to pHT 7.2 in

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Expt B0.1, body size changed by -9.9% on d1, -12.9% on d2 and -10.5% on d3 (p<0.0001). Copper significantly affected larval size in both experiments (p<0.0001). In Expt B1.0, this effect was highly dependent on the pH treatment (pH x Copper: p<0.0001), with stronger negative effect of copper on size at control pH. For instance, on d3, copper presence induced a 13.7% decrease of average size at pHT 8.0 but only induced a 3% decrease at pHT 7.3. We even observed

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small positive effects of copper addition at pHT 7.6 (+5.1%). This strong interaction between pH

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and copper was also visible on d2 (pH 8.0: -12.9%, 7.6: -5.5%, 7.3: -5.5%) and to a smaller degree,

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on d1 (pH 8.0: -5.4%, 7.6: +0.9%, 7.3: -3.0%). All the other interactions between the three factors

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were also statistically significant (Copper x Time: p=0.0037; pH x Copper x Time: p<0.0001).

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In Expt B0.1, while the presence of copper significantly decreased larval size by 8.9% on

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average, its effect was dependent on time (d1: -9.9%, d2: -2.2%, d3: -6.3%; Copper x Time: p<0.0001). Copper and pH did not have a significant interactive effect on the body size (p=0.53)

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but the interaction between copper, pH and time did (p<0.0001).

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Arm length and stomach volume

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In Expt B1.0, the relative length of both the post-oral (POL/BL) and anterolateral arms (AL/BL, d3) were significantly affected by pH alone (ANOVAII: Fig 3.a: pH: p<0.0001, N=435; Fig 3.c pH: p<0.0001, N=420, respectively). Arm lengths were longer in the pHT 7.6 treatment (+24% and +44% for POL/BL and AL/BL respectively). POL/BL was also relatively longer in pHT 7.3 (+8%) but AL/BL marginally decreased in the same treatment (-0.2%). Copper alone had overall negative

effect on relative POL (-3%, p=0.007) and AL (-13%, p<0.0001). This effect was mostly visible at control pH for both POL (-12%, pH x Copper: p=0.037), and AL (-28%), although the interaction between copper and pH was not significant for the latter (p=0.52).

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In Expt B0.1, POL/BL was significantly affected by pH and copper but not by their interaction (Fig 3.b, ANOVAII: pH: p<0.0001; Copper: p=0.0004; pH x Copper: p=0.11; N=403). These significant effects however, were subtle: the pHT 7.9 treatment slightly increased the relative POL (+8%) while the pHT 7.2 decreased it (-7%). The presence of copper had a subtle overall positive effect on relative POL (+5%). Meanwhile, AL/BL was only significantly affected by pH

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(Fig 3.d, d3; ANOVAII: pH: p<0.0001; Copper: p=0.33; pH x Copper: p=0.06; N=399). The

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relative AL decreased by a third at pHT 7.2 compared to pHT 8.0 while it marginally increased by

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4% at pHT 7.9.

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In both experiments, the volume of the larval stomach relative to the BL (Svol/BL, d3, μm3 μmBL-1) was significantly affected by pH (Fig. 3.e-f; ANOVAII: p<0.0001; NB1.0=438, NB0.1=379).

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The stomach volume (Svol, μm3) was 33% (Expt B1.0) and 22% (Expt B0.1) bigger in the control

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pH treatment than in the two lowered pH treatments. Stomach volume was additionally decreased

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by 7.5% due to copper in Expt B0.1 (Copper: p=0.047; pH x Copper: p=0.11; vs. no effect in Expt B1.0: Copper: p=0.20; pH x Copper: p=0.26).

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Larval copper content

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In Expt B0.1, larval copper concentrations were 1.5 to 1.8 times higher at pHT 7.2 compared to pHT 7.9 and 8.0 (Fig 4: 14.3 ± 5.6 vs. 7.9 ± 0.5 and 9.1 ± 4.8 μg g-1 respectively) although this difference was not statistically significant (KW: χ2=3.29, df=2, p=0.19; linear regression to average jar pHT: slope= -7.67 μg g-1 pHT units-1, F=3.75, R2=25%, df=7, p=0.094). It should be noted that the regression became significant when the outlier of the control treatment was removed

(linear regression: slope= -9.93 μg g-1 pHT units-1, F=8.94, R2=53%, df=6, p=0.024; KW: χ2=6.25, df=2, p=0.0439).

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4. DISCUSSION Coastal marine invertebrates experience multiple stressors that exert concomitant effects (additive, synergic or antagonist). In this study, we tested the hypothesis that seawater copper contamination is more toxic to sea urchin larvae under lowered pH. We exposed sea urchin embryos and larvae to two low-pH and two copper treatments (0.1 and 1.0 μM) in three separate experiments. Over

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the short time typically used for toxicity tests (4-arm plutei, 3 days), larval mortality was not

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significantly affected by the treatments. However, on longer exposure (9 days), mortality increased

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with decreasing pH. Various sub-lethal effects were recorded: frequencies of abnormality and

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degree of arm asymmetry were significantly increased by pH or/and by copper presence (see results summarized in Table S2). Body size was significantly reduced by both pH and copper,

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with strong interactions visible between pH and copper. Larval respiration was doubled by a

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decrease at pHT from 8.0 to 7.3. Larval morphology (relative arm lengths and stomach volume)

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were decreased by at least one of the two investigated factors. Uptake kinetics of copper under future ocean conditions, and the resulting effect on the organisms, appear complex and require

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further experimentations across realistic environmental conditions.

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4.1. Hong Kong sea urchin larvae are robust to a lowering of seawater pH and the copper levels tested Consistently with previous results, lowering pH to the average future pH scenario (≥7.6) did not increase mortality of sea urchin larvae (e.g. Chan et al., 2011; Clark et al., 2009; García et al., 2015a; Gonzalez-Bernat et al., 2013; Stumpp et al., 2011) but it did when pH was lowered further

(Dorey et al., 2013; García et al., 2015b; Martin et al., 2011). The pH levels used in our experiment are possibly experienced yearly by the adults at a nearby site and could therefore be well within H. crassispina’s tolerance range (Dorey et al., 2013; Ginger et al., 2013; Pansch et al., 2014). The

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spawning period for H. crassispina is long (8 months, Urriago et al., 2016) and - within this period - pH ranges from 7.1 to 8.7 (Trio Beach, Sai Kung: average of pHNBS 8.12 ± 0.22 units). On that period, 11% of the monthly records of pHNBS are lower or equal to 7.8 (Pecquet et al., 2017). In the same area, larvae of two other invertebrates are also quite resilient to pH changes (Crepidula onyx, -0.7 pH units: Maboloc et al., 2017 and Bugula neritina, -1.4 pH units: Pecquet et al., 2017).

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Parental experience of a large range of pH conditions could help confer resilience to the larvae

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(see e.g. with adult oysters pre-exposure to lowered pH in Parker et al., 2015, 2012) and may

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explain the resilience we observed in H. crassispina larvae.

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A similar process could also have enabled larvae of H. crassispina to withstand the copper

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levels found in Hong Kong area. We used copper levels (0.1 μM and 1.0 μM, for final

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concentrations of 0.07 and 0.54 μM) that ranged from ecologically-relevant to high levels (environmental levels: 0.0008-0.19 μM in Hong Kong waters in Blackmore, 1998). We did not

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observe significant lethal effects due to copper, even at the highest copper concentration, suggesting that the urchins from Hong Kong areas are adapted to the levels they are likely to

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encounter year-round or during exceptional re-suspension events in this area. In comparison,

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Paracentrotus lividus larvae had a Lower Observed Effect Concentration (LOEC) of 0.5 μM (EC50 between 0.75 and 1 μM; Warnau et al., 1996) and Diadema antillarum larvae displayed more than 80% mortality when exposed to copper concentration of 0.31 μM (Bielmyer et al., 2005). Finally, the combination of pH and copper did not reveal any additive/synergistic effects on mortality. H. crassispina appears robust to the levels of pH and copper tested in this study.

4.2. Significant sub-lethal effects could have indirect consequences on survival The lowered pH and the presence of copper had little effect on the abnormality of the larvae. The frequency of abnormality significantly increased with lowering pH only in the 9-days experiment

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(Expt A: +19% at pHT 7.3 vs 8.0) and a small lowering by 0.1 pHT units from 8.0 even seemed beneficial (3 days, Expt B0.1: -5%). Abnormal larvae were 20% more frequent when copper was present (1.0 μM only), mostly driven by a change at pHT 8.0. Radenac et al. (2001) and Manzo et al. (2007) report similar abnormality frequencies for P. lividus (+30-50% abnormality at copper concentrations of 0.79 μM, but see the more sensitive species H. erythrogramma in Kobayashi,

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1980).

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While the definition of abnormality did not always include small deviations from perfect

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arm symmetry, such changes could impact ecological functions. Echinoplutei overall morphology - and specifically arm length - is tightly linked to feeding (e.g. S. droebachiensis: Hart, 1991) and

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swimming performances (Grünbaum and Strathmann, 2003), and is shaped by functional trade-

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offs (Strathmann and Grünbaum, 2006). In our study, arm symmetry was significantly affected by

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pH for POL, in 2/3 experiments (Expt B0.1: -2% in 7.2 vs. 8.0 but Expt B1.0: +4% in 7.6 vs. 8.0). Copper presence significantly decreased the symmetry of the post-oral arms, but only in the 0.1

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μM experiment (-4%). Larval morphology was significantly affected by lowering pH (e.g. Expt B0.1: relative AL was reduced by a third at pHT 7.2) and, to a smaller extent, by the presence of

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copper (e.g. Expt B1.0: relative POL was reduced by 3%). These parameters – symmetry (Fig. S5) and length (Fig. 3) of the arms - generally followed a bell shape, showing little positive effects at intermediate effects and negative (or return to control values) effects at the lowest pH. This response may hint at the intermediate pH (7.6-7.9) being in the hormetic zone (i.e. favorable biological response to small doses of toxins/stressors) for arm growth. Altered morphology could

have negative consequences on feeding and swimming. Despite an overall shape change, Chan et al. (2015a) did not observe an effect of lowered pH (-0.4 pH units) on swimming speeds of the larvae of S. droebachiensis (see also Chan et al., 2011 for sand dollar larvae).

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Stomach volume was also affected by the treatments (e.g. relative stomach volume was reduced by 25% with lowering pHT from 8.0 to 7.3 in Expt B1.0 and slightly but significantly reduced in the 0.1 μM copper treatment). This result (also seen by Chan et al., 2011) adds to the over-looked impact of acidification on the digestion of these larvae. A decrease in stomach volume – and hence absorption surface – could incur additional loss of feeding efficiency. Maintaining

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alkaline pH in sea urchin larvae at lower pH (Hu et al., 2017; Stumpp et al., 2013), could also have

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large additional metabolic costs.

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Here, we observed that lowering pH increased metabolic rates (Expt A: doubling of respiration normalized by size from pHT 8.0 to pHT of 7.3; see similar results for S. dorebachiensis

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and S. purpuratus in Dorey et al., 2013; Stumpp et al., 2011b respectively). Lowering pH also

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significantly decreased larval size in 2/3 experiments (d3: Expt B1.0: -9% at pHT 7.3 vs. 8.0 and Expt B0.1: -12% at 7.2 vs. 8.0). This is a known delay for the genus Strongylocentrotus and is

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attributed to the energy re-allocation caused by increased energy demand (Stumpp et al., 2012,

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2011b; Stumpp and Hu, 2017). In our study, food abundance (Rhodomonas sp., added on d2) was negatively affected by lowering pH (but not copper, Fig. S1) and could have played a marginal

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role in the observed reduction in size on d3. Larval body growth rates were also significantly reduced by copper presence in a pH-dependent manner: while the presence of copper (1.0 μM) had a strong negative effect at pHT 8.0 (decreased body size by 14% on d3), it little affected larvae at pHT 7.3 (-3%) and increased body size of the larvae at pHT 7.6 (+5%). Fernández and Beiras (2001) noted a decrease in mean larval size of P. lividus larvae (d2) by 2.5%, 13.7% and 45.1% for copper

concentrations of 0.25, 0.50 and 1.0 µM. In comparison with other species, H. crassispina appears to sustain high levels of copper contamination without being affected (mortality, body growth rates), which may reflect an adaptation to Hong Kong’s long history of pollution. Regardless of

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the causes, Dupont et al. (2010) demonstrated how slow larval growth rates can increase mortality - due to their longer exposure to pelagic predation pressure. For instance, a reduction of the growth rates by 10% could lead to 25-50% less larval abundance at the time when larvae reach competency to settle.

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4.3. Interactive effects of pH and copper

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Metals and ocean acidification are likely to exert interactive effects on the physiology of marine

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organisms (see Ivanina and Sokolova, 2015 for a review), not only because copper bioavailable

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forms are pH-sensitive (e.g. +115% of Cu2+ by pH 7.6) but also because of the effects of pH on the organisms’ acid-base regulation mechanisms. In particular, sea urchin larvae use ionic pumps

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such as ATP-ase pumps (Stumpp et al., 2012) to maintain intracellular medium under elevated

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CO2 which in turn create electrochemical gradients used by secondary active transporters (Stumpp and Hu, 2017). Yet the latter ion channels represent a major pathway for the entrance of metallic

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positively-charged ions (e.g. fish: Webb and Wood, 2000) and therefore, organisms possessing

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acid-base regulation systems could see their metallic bioaccumulation increased at lowered pH. Copper uptake/accumulation within the tissues may thus be expected to rise at low pH due

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to increased bioavaliability of the free form and/or increased activity of ionic pumps but also due to an increase in active uptake. Here, we observed that copper was twofold more concentrated in larval tissues at pHT 7.2 than at pHT 7.9-8.0 in parallel with a twofold increase in respiration. Fitzer et al. (2013) also reported an increase in the copper content of gravid females of copepods (+1 μg

L-1 per female) when pH was decreased by 0.11-0.24 units. Since the uptake of copper – an essential element – is likely to be active in sea urchin larvae (Radenac et al., 2001), an increase in metabolic rate could increase copper accumulation in the larval body.

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Regardless of the reasons, lower pH has been shown to increase copper toxicity for a few marine invertebrates (Campbell et al., 2014; Fitzer et al., 2013; Ivanina et al., 2015; Lewis et al., 2016, 2013). For instance, Campbell et al. (2014) reported that survival of a polychaete larvae was decreased by an additional 24% when both stressors were combined (-0.6 to -1.0 pHNBS units and 0.2 to 20µM copper). Here, the toxicity of copper in the sea urchin larvae H. crassispina was

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increased at lowered pH, but this was only visible at the sub-lethal level: five out of eight

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parameters analyzed were significantly affected by the two stressors combination in the highest

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copper experiment (1.0 μM; 2/8 at 0.1μM; Table S2). While we expected increased negative

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effects of copper at low pH, this was not the type of interactive effects we observed. Copper

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addition (1.0 µM) appeared to be most toxic at pHT 8.0 with e.g. abnormality on average ~59 %

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(vs. 3-12% in all other conditions). In the 1.0 µM experiment, larval sizes were also most reduced by copper at pHT 8.0 (Fig. 2.c). Ivanina and Sokolova (2013) propose that moderate acidosis could

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protect mitochondria against metal toxic effects in intertidal clams and oysters. The authors

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theorize that low pH (i.e. increase in protons) may reduce copper toxicity due to competition of the positively charged metallic ions with protons for protein binding sites. This mechanism

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requires further investigation in sea urchin larvae but has been observed in unicellular freshwater alga (Wilde et al., 2006) and could explain the higher toxicity of copper at pHT 8.0 in this study. 5. CONCLUSION

Larvae of the sea urchin H. crassispina are robust and survived the copper levels present in Hong Kong waters today (≤0.19 μM) as well as the pH projected for 2100 (shift by -0.4 units of the present pH range of 7.1-8.8 units). This species seems particularly tolerant to the strong variations

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of pH and the pulses of metal contamination they already experience in Hong Kong. Sea urchin larvae however exhibit sub-lethal sensitivities to the combination of both stressors that may have indirect consequences on feeding, swimming and ultimately survival. The observed interactive effects highlight the complex relationship between pH and metal speciation/uptake and our current lack of understanding in the mechanisms involved. Understanding copper behavior (uptake,

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accumulation kinetics) in marine invertebrates and across a broad range of pH is essential when

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these species are routinely used in toxicity tests or as biomonitoring species. Further knowledge

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on metals and pH interactions can help guiding management/mitigation strategies in implementing

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safe pollutant levels in the face of global change.

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ACKNOWLEDGEMENTS

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We thank Abby Lo, Jenny Ngo, Esther Wong Antoine Pecquet and Dr. Tam Yi Ki for their kind assistance in the lab. We are grateful to Prof. W. Wang and Dr. Jian Wang from HKUST for

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running the copper analyses. We thank Sam Dupont for valuable feedback on the manuscript and

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Ceri Lewis for interesting discussion on the data. ND is partially supported by VPRG, HKUST. Funding for the project comes from the Research Grant Council [Project Number: 26102515] to

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KC. KC is additionally supported by the Croucher Foundation, Hong Kong.

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impacts sea urchin larvae pH regulatory systems relevant for calcification. Proc. Natl. Acad. Sci. U. S. A. 109, 18192–18197. doi:10.1073/pnas.1209174109 Stumpp, M., Wren, J., Melzner, F., Thorndyke, M., Dupont, S., 2011b. CO2 induced seawater acidification impacts sea urchin larval development I: Elevated metabolic rates decrease scope for growth and induce developmental delay. Comp. Biochem. Physiol. - A Mol. Integr. Physiol. 160, 331–340. doi:10.1016/j.cbpa.2011.06.022 U. S. National Institutes of Health, Bethesda, Maryland, U., n.d. ImageJ, Rasband, W.S., http://imagej.nih.gov/ij/. Urriago, J.D., Wong, J.C.C.Y., Dumont, C.P., Qiu, J., 2016. Reproduction of the short-spined sea urchin Heliocidaris crassispina (Echinodermata: Echinoidea) in Hong Kong with a subtropical climate. Reg. Stud. Mar. Sci. 8, 445–453. doi:10.1016/j.rsma.2016.03.005 Wai, T.C., Williams, G.A., 2006. Effect of grazing on coralline algae in seasonal, tropical, lowshore rock pools: Spatio-temporal variation in settlement and persistence. Mar. Ecol. Prog. Ser. 326, 99–113. doi:10.3354/meps326099 Warnau, M., Temara, A., Jangoux, M., Dubois, P., Iaccarino, M., De Biase, A., Pagano, G., 1996. Spermiotoxicity and embryotoxicity of heavy metals in the echinoid Paracentrotus lividus. Environ. Toxicol. Chem. 15, 1931–1936. doi:10.1002/etc.5620151111 Webb, N.A., Wood, C.M., 2000. Bioaccumulation and distribution of silver in four marine teleosts and two marine elasmobranchs: influence of exposure duration, concentration, and salinity. Aquat. Toxicol. 49, 111–129. doi:10.1016/S0166-445X(99)00063-6 Weng, N., Wang, W.X., 2014. Variations of trace metals in two estuarine environments with contrasting pollution histories. Sci. Total Environ. 485–486, 604–614. doi:10.1016/j.scitotenv.2014.03.110 Wilde, K.L., Stauber, J.L., Markich, S.J., Franklin, N., Brown, P., 2006. The effect of pH on the uptake and toxicity of copper and zinc in a tropical freshwater alga (Chlorella sp.). Arch. Environ. Contam. Toxicol. 51, 174. doi:https://doi.org/10.1007/s00244-004-0256-0

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TABLE 1: Seawater carbonate chemistry parameters during the three experiments (Expt) presented by pH treatment (N= 9 daily measurements in 9 jars in Expt A and 4 daily measurements in 8 jars

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in Expt B1.0 and B0.1). Measured seawater total scale pH (pHT), temperature (T; °C), mean total alkalinity (TA; mmol kg-1) and average salinity (S) were used to calculate the CO2 partial pressure

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(pCO2; μatm) and aragonite and calcite saturation states (respectively Ωar and Ωca) using the package seacarb for R. All the values are expressed as mean ± SD. Detailed results per jar can be

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found in Table S1. Control conditions did not significantly differ across experiments (N=69; KW: χ2=5.59,

df=2, p=0.06) but medium (KW: χ2=46.76, df=2, p<0.0001) and low pH conditions of Expt B0.1 differed from Expt A and B0.1 (KW: χ2=21.18, df=2, p<0.0001). Expt

pH treatment

pHT

T

TA (S)

pCO2

Ωar

Ωca

B1.0

7.96 ± 0.05 7.63 ± 0.09 7.31 ± 0.15 7.98 ± 0.02 7.65 ± 0.03 7.27 ± 0.06 7.98 ± 0.01 7.88 ± 0.03 7.21 ± 0.03

26.28 ± 0.40 26.23 ± 0.47 26.24 ± 0.49 25.92 ± 0.05 25.92 ± 0.04 25.95 ± 0.05 26.00 ± 0.02 26.00 ± 0.00 26.00 ± 0.03

2.176 ± 0.051 (S=32.0) 2.064 ± 0.054 (S=31.5) 2.190 ± 0.052 (S=33.0)

487 ± 67 1190 ± 343 2743 ± 1151 436 ± 21 1065 ± 82 2715 ± 354 463 ± 19 613 ± 49 3291 ± 273

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control -0.4 -0.7 control -0.4 -0.7 control -0.4 -0.7

2.79 ± 0.25 1.46 ± 0.24 0.76 ± 0.24 2.68 ± 0.08 1.38 ± 0.09 0.62 ± 0.09 2.89 ± 0.08 2.38 ± 0.11 0.59 ± 0.05

4.24 ± 0.38 2.22 ± 0.38 1.15 ± 0.37 4.08 ± 0.12 2.10 ± 0.14 0.94 ± 0.13 4.38 ± 0.12 3.61 ± 0.17 0.89 ± 0.07

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FIGURE LEGENDS: FIGURE 1: Expt A: Larval mortality rates, calculated relative to a. time (% larvae hr-1) and b. body length (BL; % larvae μmBL-1) by pHT. Each dot represents the regression coefficient extracted from significant linear relationship for each culture (a. N=8 and b. N=9). c. Larval abnormality (%, N=9

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days x 3 pH x 3 jars=81) and d. growth rates (μmBL log(hr)-1) by pHT and size-normalized respiration rates (pmolO2 hr-1 μmBL-1) on d3 (e.) and d6 (f.) by pH treatment (“N=” indicates the number of valid replicates, when not indicated: N=6: three replicates distributed in two plates).

FIGURE 2: Expt B1.0 ( left panels) and: Expt B0.1 (right panels) - a. and b. Larval abnormality

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index (% larvae) on d3 and c. and d. Larval size (BL; μm) on d3, presented by treatments (pH treatment: color boxes; left three boxes without copper: No Cu vs. right three boxes with copper:

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+Cu).

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FIGURE 3: Expt B1.0 ( left panels) and: Expt B0.1 (right panels) - Morphological parameters on d3: a. and b. relative post-oral arm size (POL/BL, μm μmBL-1), c. and d. relative anterolateral arm

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size (AL/BL, μm μmBL-1), e. and f. relative stomach volume (Svol/BL, μm3 μmBL-1).

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FIGURE 4: Expt B0.1: Larval copper concentration (μg g-1) on d3 depending on the average pHT

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FIGURE 4