Factors affecting the toxicity of trace metals to fertilization success in broadcast spawning marine invertebrates: A review

Aquatic Toxicology 184 (2017) 1–13

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Aquatic Toxicology journal homepage: www.elsevier.com/locate/aquatox

Factors affecting the toxicity of trace metals to fertilization success in broadcast spawning marine invertebrates: A review M. Hudspith, Amanda Reichelt-Brushett ∗ , Peter L. Harrison Marine Ecology Research Centre, School of Environment, Science and Engineering, Southern Cross University, Lismore, New South Wales, Australia

a r t i c l e

i n f o

Article history: Received 19 August 2016 Received in revised form 19 December 2016 Accepted 22 December 2016 Available online 28 December 2016 Keywords: Pollution Reproduction Ecotoxicology Corals Toxicity mechanisms

a b s t r a c t Significant amounts of trace metals have been released into both nearshore and deep sea environments in recent years, resulting in increased concentrations that can be toxic to marine organisms. Trace metals can negatively affect external fertilization processes in marine broadcast spawners and may cause a reduction in fertilization success at elevated concentrations. Due to its sensitivity and ecological importance, fertilization success has been widely used as a toxicity endpoint in ecotoxicological testing, which is an important method of evaluating the toxicity of contaminants for management planning. Ecotoxicological data regarding fertilization success are available across the major marine phyla, but there remain uncertainties that impair our ability to confidently interpret and analyse these data. At present, the cellular and biochemical events underlying trace metal toxicity in external fertilization are not known. Metal behavior and speciation play an important role in bioavailability and toxicity but are often overlooked, and disparities in experimental designs between studies limit the degree to which results can be synthesised and compared to those of other relevant species. We reviewed all available literature covering cellular toxicity mechanisms, metal toxicities and speciation, and differences in methodologies between studies. We conclude that the concept of metal toxicity should be approached in a more holistic manner that involves elucidating toxicity mechanisms, improving the understanding of metal behavior and speciation on bioavailability and toxicity, and standardizing the fertilization assay methods among different groups of organisms. We identify opportunities to improve the fertilization assay that will allow robust critical and comparative analysis between species and their sensitivities to trace metals during external fertilization, and enable data to be more readily extrapolated to field conditions. © 2017 Elsevier B.V. All rights reserved.

Contents 1. 2.

3. 4.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2 Fertilization as an ecotoxicology endpoint . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2 2.1. Biology of fertilization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2 2.2. Possible mechanisms of trace metal toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4 2.3. The importance of understanding metal toxicity mechanisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6 Trace metal toxicity and speciation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7 3.1. Sediment water interactions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8 The effect of experimental design on species sensitivities to trace metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8 4.1. Variables of the experimental design can impact trace metal toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8 4.1.1. Test seawater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8 4.1.2. Sperm and/or egg exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9 4.1.3. Duration of exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9 4.2. Other factors determining sensitivity to trace metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10 4.3. Future development of fertilization assays . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10

∗ Corresponding author. E-mail addresses: [email protected] (M. Hudspith), [email protected] (A. Reichelt-Brushett), [email protected] (P.L. Harrison). http://dx.doi.org/10.1016/j.aquatox.2016.12.019 0166-445X/© 2017 Elsevier B.V. All rights reserved.

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5.

Conclusions and direction for future research . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11 Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11 Appendix A. Supplementary data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11

1. Introduction Trace metals occur naturally in the marine environment but significant amounts have been discharged into both nearshore and deep sea environments in recent years, resulting in elevated levels that can be toxic to marine organisms (Hassan, 2006; Mason and Jenkins, 1995). Metal inputs are permanent additions to the marine environment. They do not get broken down by bacterial action and rendered inert. Metal contamination occurs primarily as a result of land-based activities, with rivers and estuaries acting as conduits for metals into the coastal and deep sea environment (van den Hove and Moreau, 2007). They are most notably associated with the mining and metal processing industry, and also enter the marine environment via industrial, sewage and stormwater discharges, anti-fouling paints and urban/agricultural run-off (Hart and Lake, 1987). Ecotoxicological testing is an important tool used to evaluate the potential toxicity of trace metals and other environmental contaminants to inform ecosystem management (van Dam et al., 2008). Early-life stages of marine organisms are used extensively in ecotoxicological assays because it is generally agreed that they are more sensitive to chemical contaminants than their adult counterparts (reviewed by His et al., 1999). Pelagic stages are exposed in the water column, and thus multiple environmental factors can impact upon their success and development. Fertilization success is an important endpoint in ecotoxicological assays because it is sensitive, exhibiting a measurable dose-response relationship at metal concentrations similar to those found in polluted environments. It is also ecologically relevant because it has a direct bearing on the dynamics of recruitment and natural populations (Shea, 2011). Both USEPA (1995, 2002) and Environment Canada (2011) have developed standard aquatic fertilization assays using sea urchin and sand dollar gametes, which are used to evaluate the toxicity of effluents to coastal marine waters and estuaries. Many studies have documented the negative effect of metals on external fertilization in marine invertebrates. Ecotoxicological data exist for all the major marine phyla and are used in the derivation of water quality guidelines (ANZECC/ARMCANZ, 2000; USEPA, 2002). However, there remain some uncertainties that impair our ability to confidently interpret and extrapolate these data. There is little information on precisely how metals affect fertilization (see Fitzpatrick et al., 2008) and this diminishes the extent to which we can reliably estimate the probability and extent of a toxicant’s harmful effects (Shanker, 2008). The toxicity of a given trace metal is intimately linked to chemical speciation, which is governed by the physico-chemical nature of seawater and can be variable in dynamic coastal waters (Elder, 1988). However, metal speciation is not always considered during ecotoxicological testing and data analyses. Indeed, of the ecotoxicological data examined in this review (see Table 1), only one study has attempted any predictive or measured speciation study (Ward et al., 2006). There are also differences in experimental designs between studies, which may influence the variation in species’ sensitivities across a range of invertebrate taxa with the same mode of fertilization (external). This limits our ability to draw direct comparisons between species. This review aims to resolve some of the uncertainty surrounding fertilization success as a toxicity endpoint by considering the

process of fertilization, metal toxicity and speciation, and the influence of experimental design, to enable us to accurately interpret data by adopting a more holistic view of metal toxicity. By exploring the fine-scale process of fertilization and the individual gametes themselves, possible mechanisms for metal toxicity will be highlighted. In considering the theoretical basis of metal behavior and speciation in marine waters, we can aid in the interpretation and prediction of bioavailability and toxicity. Finally, by scrutinizing differences in methodologies between fertilization assays, we can improve our ability to discern between genuine trends and artefacts of the experimental design. 2. Fertilization as an ecotoxicology endpoint 2.1. Biology of fertilization Fertilization, in its simplest form, is the fusion of two specialized gametes to form a single viable cell − the zygote (Rosati, 1995). Many marine organisms have external fertilization, whereby eggs and sperm are released into the environment, and this is thought to be an ancestral reproductive strategy (Lotterhos and Levitan, 2010). Behaviors such as spawning aggregations and synchronous gamete release (Babcock and Mundy, 1992; Harrison et al., 1984) greatly increase the likelihood of gamete interaction, and species-specific sperm chemotaxis also increases gamete encounters and reduces hybridization (Riffel et al., 2004). Once the gametes have come in contact with one another, the sperm binds to the egg and makes passage through the egg membrane. Entry of the sperm is facilitated by the acrosome reaction in several marine invertebrates (Niijima, 1963; Talbot and Chanmanon, 1980). When the acrosome of the sperm comes in contact with the extracellular matrix of the egg, the acrosome reaction (AR) is triggered. This causes the contents of the acrosome to be released and digest a passage through the egg envelope, allowing the spermatozoa to reach the egg surface (Baccetti, 1985; Levine et al., 1978). The second stage of the AR involves the extension of the acrosomal process/filament, which makes contact and fuses with the egg plasma membrane and facilitates the entry of the sperm (Franklin, 1970). However, not all marine invertebrates, including most cnidarians, have a specific acrosome (Harrison and Jamieson, 1999). Entry of the sperm into the egg triggers a variety of metabolic changes referred to as egg activation (Gilbert, 2000). The primary response of the egg upon sperm penetration is a rapid, transient depolarization of the membrane potential that effectively prevents another sperm from entering the egg, preventing polyspermy (Jaffe, 1976; Rothschild and Swann, 1952). This brief depolarization is supplemented by a second and permanent block to polyspermy (Schatten and Chakrabarti, 2000), where the cortical granules of the egg are exocytosed and the vitelline coat becomes elevated and hardened to form the fertilization membrane (Gould and Stephano, 2003), rendering the egg impermeable to spermatozoa (Schuel et al., 1973). Cnidaria do not have a vitelline coat like higher invertebrates, but a cortical reaction occurs in some sea anemones when exposed to sperm (see Harrison and Jamieson, 1999). Once inside the egg, the sperm undergoes several changes and becomes the pronucleus; the entry of the sperm also initiates the second meiotic division of the egg, resulting in a haploid egg nucleus known as the female pronucleus (Gilbert, 2000). The male pronucleus then

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Table 1 Median effective concentration (EC50 ) values for the effect of trace metals on fertilization success across a range of marine invertebrates (␮g L−1 ). Species

Metal

Salta

Wb

EC50

TEc

Exposure characteristicsd

Author

Cnidaria Acropora cytherea A. longicyathus A. millepora A. pulchra A. surculosa A. tenuis Goniastrea aspera G. aspera G. retiformis Montipora capitata Platygyra daedalea P. daedalea P. acuta Lobophytum compactum

Cu Cu Cu Cu Cu Cu Cu Cu Cu Cu Cu Cu Cu Cu

Cu(HNO3 )2 CuCl2 CuCl2 Cu(HNO3 )2 CuSO4 CuCl2 CuCl2 CuCl2 CuCl2 Cu(HNO3 )2 CuCl2 CuCl2 CuCl2 CuCl2

F SF F F F SF SF SF SF F SF SF F SF

69.4 15.2 17.4 75.4 45.2 39.7 14.5 18.5 24.7 21.9 33 73 145 261

C C C C C C C C C C C C C C

Gametes 4.5 h (Gametes 30 m) + 5 h Gametes 4 h Gametes 4.5 h Gametes 5 h (Gametes 30 m) + 5 h (Gametes 30 m) + 5 h (Gametes 30 m) + 5 h (Gametes 30 m) + 5 h Gametes 3 h (Gametes 30 m) + 5 h (Gametes 40 m) + 5 h (Gametes 30 m) + 5 h (Gametes 30 m) + 10 h

Puisay et al. (2015) Reichelt-Brushett and Harrison (2005) Negri and Heyward (2001) Puisay et al. (2015) Victor and Richmond (2005) Reichelt-Brushett and Harrison (2005) Reichelt-Brushett and Harrison (1999) Reichelt-Brushett and Harrison (2005) Reichelt-Brushett and Harrison (2005) Hédouin and Gates (2013) Reichelt-Brushett and Hudspith (2016) This study Kwok et al. (2016) Reichelt-Brushett and Michalek-Wagner (2005)

Cu Cu Cu Cu Cu Cu Cu Cu Cu Cu

CuCl2 CuSO4 CuSO4 CuCl2 CuCl2 NS CuSO4 NS Cu(NO3 ) 2 CuCl2

A N F F F F F F A F

12 29.9 18 200 26 70 17 14 57 59

M M M M M M M M NS M

(Sperm 1 h) + 20m (Sperm 1 h) + 20m (Sperm 1 h) + 10–20m (Sperm 20 m) + 1 h (Sperm 1 h) + 20 m (Sperm 1 h) + 20 m (Sperm 10 m) + 10 m (Sperm 1 h) + 20 m (Sperm 1 h) + 20 m (Sperm 1 h) + 20 m

Nacci et al. (1986) USEPA (1991) Larrain et al. (1999) Lee et al. (2004) Dinnel et al. (1989) Ramachandran et al. (1997) Thongra-ar (1997) Ringwood (1992) Novelli et al. (2003) Dinnel et al. (1989)

Cu Cu

CuCl2 CuCl2

F F

25 1.9

M M

(Sperm 1 h) + 20 m (Sperm 1 h) + 20 m

Dinnel et al. (1989) Dinnel et al. (1989)

Polychaeta Hydroides elegans H. elegans Nereis virens

Cu Cu Cu

CuCl2 CuSO4 CuSO4

F A F

10030 47 351.1

C C M

NS Gametes (cleavage) Gametes 4.5 h

Gopalakrishnan et al. (2007) Xie et al. (2005) Caldwell et al. (2011)

Mollusca Crassostrea gigas Isognomon californicum

Cu Cu

NS NS

NS F

12 55

NS C

NS (Sperm 1 h) + 1 h

Nacci et al. (1986) Ringwood (1992)

Crustacea Scylla serrata

Cu

NS

CF

13.69

AR

Sperm (duration NS)

Zhang et al. (2010)

Urochordata Ciona intestinalis

Cu

CuCl2

A

36.6

HL

Gametes 20 h

Bellas et al. (2004)

Cnidaria Oxypora lacera

Cd

Cd(NO3 ) 2

SF

>1000

C

(Gametes 30 m) + 5 h

Reichelt-Brushett and Harrison (1999)

Echinodermata A. crassispina A. punctulata A. spatuligera A. amurensis D. excentricus D. setosum D. setosum E. mathaei P. lividus S. droebachiensis S. franciscanus S. purpuratus

Cd Cd Cd Cd Cd Cd Cd Cd Cd Cd Cd Cd

CdCl2 CdCl2 Cd SO4 CdCl2 CdCl2 NS CdCl2 NS Cd(NO3 ) 2 CdCl2 CdCl2 CdCl2

F A F F F F F F A F F F

1700 38000 140900 154000 8000 950 6280 >100 8400 26000 12000 18000

M M M M M M M M NS M M M

Sperm 30 m (Sperm 1 h) + 20m (Sperm 1 h) + 10–20m (Sperm 20 m) + 1 h (Sperm 1 h) + 20m (Sperm 1 h) + 20m (Sperm 10 m) + 10 m (Sperm 1 h) + 20m (Sperm 1 h) + 20m (Sperm 1 h) + 20m (Sperm 1 h) + 20m (Sperm 1 h) + 20m

Vaschenko et al. (1999) Nacci et al. (1986) Larrain et al. (1999) Lee et al. (2004) Dinnel et al. (1989) Ramachandran et al. (1997) Thongra-ar (1997) Ringwood (1992) Novelli et al. (2003) Dinnel et al. (1989) Dinnel et al. (1989) Dinnel et al. (1989)

Polychaeta H. elegans H. elegans

Cd Cd

CdCl2 CdCl2

F F

94.3 227.6

C C

(Sperm 20m) + 1 h (Eggs 20 m) +1 h

Gopalakrishnan et al. (2008) Gopalakrishnan et al. (2008)

Mollusca C. gigas

Cd

NS

NS

11900

NS

NK

Nacci et al. (1986)

Crustacea S. serrata

Cd

NS

CF

2.14

AR

Sperm (duration NS)

Zhang et al. (2010)

Urochordata C. intestinalis

Cd

CdCl2

A

721

HL

Gametes 20 h

Bellas et al. (2004)

Cnidaria G. aspera

Zn

ZnSO4

SF

>500

C

(Gametes 30 m) + 5 h

Reichelt-Brushett and Harrison (1999)

Echinodermata Arbacia punctulata A. punctulata A. spatuligera Asterias amurensis Dendraster excentricus Diadema setosum D. setosum Echinometra mathaei Paracentrotus lividus Strongylocentrotus droebachiensis S. purpuratus S. franciscanus

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Table 1 (Continued) Species

Metal

Salta

Wb

EC50

TEc

Exposure characteristicsd

Author

Echinodermata A. amurensis A. punctulata A. spatuligera D. excentricus D. setosum P. lividus S. droebachiensis S. franciscanus S. purpuratus

Zn Zn Zn Zn Zn Zn Zn Zn Zn

ZnCl2 ZnCl2 ZnCl2 ZnCl2 ZnSO4 Zn(NO3 ) 2 ZnCl2 ZnCl2 ZnCl2

F A F F F A F F F

550 121 116 28 380 210 383 313 262

M M M M M NS M M M

(Sperm 20 m) + 1 h (Sperm 1 h) + 20 m (Sperm 1 h) + 10–20 m (Sperm 1 h) + 20 m (Sperm 10 m) + 10 m (Sperm 1 h) + 20 m (Sperm 1 h) + 20 m (Sperm 1 h) + 20 m (Sperm 1 h) + 20 m

Lee et al. (2004) Nacci et al. (1986) Larrain et al. (1999) Dinnel et al. (1989) Thongra-ar (1997) Novelli et al. (2003) Dinnel et al. (1989) Dinnel et al. (1989) Dinnel et al. (1989)

Polychaeta H. elegans H. elegans H. elegans

Zn Zn Zn

ZnCl2 ZnCl2 ZnCl2

F F F

44220 945.3 2025.6

C C C

NS (Sperm 20 m) + 1 h (Eggs 20 m) + 1 h

Gopalakrishnan et al. (2007) Gopalakrishnan et al. (2008) Gopalakrishnan et al. (2008)

Mollusca C. gigas

Zn

NS

NS

444

NS

NS

Nacci et al. (1986)

Crustacea S. serrata

Zn

NS

CF

2.21

AR

Sperm (duration NS)

Zhang et al. (2010)

Cnidaria A. tenuis A. longicyathus G. aspera

Pb Pb Pb

Pb(NO3 ) 2 Pb(NO3 ) 2 Pb(NO3 ) 2

SF SF SF

1801 1453 2467

C C C

(Gametes 30 m) + 5 h (Gametes 30 m) + 5 h (Gametes 30 m) + 5 h

Reichelt-Brushett and Harrison (2005) Reichelt-Brushett and Harrison (2005) Reichelt-Brushett and Harrison (2005)

Echinodermata A. punctulata D. excentricus P. lividus S. droebachiensis S. franciscanus S. purpuratus

Pb Pb Pb Pb Pb Pb

PbCl2 PbCl2 Pb(NO3 ) 2 PbCl2 PbCl2 PbCl2

A F A F F F

5400 13000 16210 19000 1300 8200

M M NS M M M

(Sperm 1 h) + 20 m (Sperm 1 h) + 20 m (Sperm 1 h) + 20 m (Sperm 1 h) + 20 m (Sperm 1 h) + 20 m (Sperm 1 h) + 20 m

Nacci et al. (1986) Dinnel et al. (1989) Novelli et al. (2003) Dinnel et al. (1989) Dinnel et al. (1989) Dinnel et al. (1989)

Polychaeta H. elegans H. elegans H. elegans

Pb Pb Pb

PbCl2 PbCl2 PbCl2

F F F

30370 380.8 691.7

C C C

NS (Sperm 20 m) + 1 h (Eggs 20 m) + 1 h

Gopalakrishnan et al. (2007) Gopalakrishnan et al. (2008) Gopalakrishnan et al. (2008)

Mollusca C. gigas

Pb

NS

NS

5500

NS

NS

Nacci et al. (1986)

Echinodermata A. amurensis A. punctulata A. punctulata D. excentricus S. droebachiensis S. purpuratus S. franciscanus

Ag Ag Ag Ag Ag Ag Ag

AgNO3 AgNO3 AgNO3 AgNO3 AgNO3 AgNO3 AgNO3

F F A F F F F

430 14 51 54 86 115 112

M NS M M M M M

(Sperm 20 m) + 1 h (Sperm 1 h) + 2 h (Sperm 1 h) + 20 m (Sperm 1 h) + 20 m (Sperm 1 h) + 20 m (Sperm 1 h) + 20 m (Sperm 1 h) + 20 m

Lee et al. (2004) Ward et al. (2006) Nacci et al. (1986) Dinnel et al. (1989) Dinnel et al. (1989) Dinnel et al. (1989) Dinnel et al. (1989)

Mollusca C. gigas

Ag

NS

NS

29

NS

NK

Nacci et al. (1986)

Crustacea S. serrata

Ag

NS

CF

10.02

AR

Sperm (duration NS)

Zhang et al. (2010)

Cnidaria P. daedalea

Ni

NiCl2

SF

1420

C

(Gametes 30 m) + 5 h

Reichelt-Brushett and Hudspith (2016)

Echinodermata P. lividus

Ni

Ni(NO3 )2

A

5130

NS

(Sperm 1 h) + 20 m

Novelli et al. (2003)

Polychaeta H. elegans H. elegans H. elegans

Ni Ni Ni

NiCl2 NiCl2 NiCl2

F F F

38300 773.46 1178.4

C C C

NS (Sperm 20 m) + 1 h (Eggs 20 m) + 1h

Gopalakrishnan et al. (2007) Gopalakrishnan et al. (2008) Gopalakrishnan et al. (2008)

a

NS: not specified. Experimental seawater used; SF: sperm free; F: filtered; A: artificial; N: natural; CW: calcium free. c Toxicity endpoint; C: the onset of cleavage; M: elevation of the fertilization membrane; AR: in the process/completion of the acrosome reaction; HL: percentage of normal hatched larvae. d Exposure time; (Gametes 30m) + 5 h: gametes dosed separately for 30 min then combined for a further 5 h; Gametes 4 h: gametes dosed together for 4 h; (Sperm 1 h) + 20: sperm dosed for 1 h then eggs added for a further 20 min; Gametes (cleavage): gametes dosed together until cleavage; Sperm (duration NS): sperm dosed, duration not specified. b

migrates toward the centrally located female pronucleus, where they fuse to form the diploid zygote nucleus (Elder and Dale, 2000). Fertilization is now complete and the zygote undergoes mitosis, known as the first cleavage.

2.2. Possible mechanisms of trace metal toxicity The cellular and biochemical events underlying trace metal toxicity in external fertilization are not clearly understood (Victor and

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Richmond, 2005). However, some studies have examined how the gametes themselves and key events in fertilization are affected by trace metals in marine broadcast spawners, bringing us closer to elucidating the mechanism(s) of toxicity. The effect of selected trace metals on the function and morphology of gametes has been studied in some marine invertebrate species, with results indicating that sperm are particularly sensitive to metal exposure. Au et al. (2000) found sperm of the tropical urchin Anthocidaris crassispina suffered impaired sperm motility on exposure to cadmium. A decline in sperm velocities was concomitant with an enlarged sperm midpiece (consisting of a single, large mitochondrion) and disorganized mitochondrial membranes and cristae, which may affect swimming movement and respiratory processes (and hence ATP supply), respectively. Ultrastructural damage was also evident in sperm of the mud crab Scylla serrata after exposure to the trace metals Ag+ , Cd2+ , Cu2+ , Zn2+ .1 Prominent damage occurred in acrosomal regions and the sperm cells became irregular, swollen, and less electron dense (Zhang et al., 2010). Sperm of the blue mussel Mytilus trossulus swam more slowly and exhibited a reduced fertilization capacity when exposed to 100 ␮g L−1 copper, which may have been caused by copper induced interference of mitochondrial activity (Fitzpatrick et al., 2008). Sperm activation and subsequent motility are controlled by reactive sulfhydryl (-SH) groups (Barron et al., 1948) and trace metals have a strong affinity for these groups. For example, when taken up into a cell, copper is reduced to the cuprous ion (Cu+ ) which readily binds with the uncoupled −SH groups (Viarengo et al., 1996); this ultimately inhibits ATP production necessary for flagella motility by affecting key enzymes involved in respiration (Mohri, 1956; Wimalasena et al., 2007). Some studies have found that unfertilized eggs are relatively unaffected by trace metal exposure compared to spermatozoa. Hollows et al. (2007) found that fertilization success in the polychaete Galeolaria caespitosa was not affected by exposing eggs to copper pre-fertilization, and likewise the viability of blue mussel eggs was not influenced by increasing concentrations of copper (Fitzpatrick et al., 2008). Eggs of the polychaete Hydroides elegans were less sensitive to trace metal exposure (Hg, Cd, Pb, Ni, Zn) than sperm, with egg toxicity median effective concentrations (EC50 s) being consistently higher than sperm toxicity EC50 s by a factor of at least two (Gopalakrishnan et al., 2008). In a recent study, we endeavoured to investigate the effect of increasing total copper concentrations on individual coral gametes (that is, exposure of egg and sperm separately), allowing us to ascertain the effects of copper on gamete functionality, specifically egg viability and sperm fertilization capacity. Egg viability was not affected by increasing copper exposures in the reef coral Platygyra daedalea, however, sperm exhibited a reduced fertilization capacity after copper exposure, with the highest copper concentration of 66 ␮g L−1 resulting in a lower average fertilization success rate of 26.7% (±6.2% S.D.) relative to the control (67.2%, ± 11.7%) (Fig. 1). Exposing sperm to low concentrations of copper (10 and 21 ␮g L−1 ) resulted in very high levels of fertilization success (>97%). Fertilization success was more sensitive to copper when only sperm were exposed (EC50 43 ␮g L−1 Cu) compared to when both egg and sperm were exposed (EC50 73 ␮g L−1 Cu), using the standard coral fertilization assay (for methods see Supplementary material). Other work on this species showed a more sensitive response to copper exposure (EC50 33 ␮g L−1 Cu) (Reichelt-Brushett and Hudspith, 2016), suggesting that there are factors beyond species differences and experimental protocol that can influence variability, including natural variation both temporally and between individuals within a

1 Unless otherwise stated, metals and their symbols (when referring to ecotoxicological studies) represent total metals.

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species (Chornesky and Peters, 1987). With a relatively long gametogenesis period of up to 9 months (Babcock et al., 1986; Harrison and Wallace 1990; see Prasetia et al., 2016), inter-annual variability in environmental conditions such as coral bleaching events and other stressors can influence the reproductive cycles and viability of coral gametes (Harrison and Ward 2001; Ward et al., 2002) and hence their susceptibility to contaminants. Initiation and activation of sperm motility and sperm chemotaxis have been well described in marine invertebrates. The phenomenon of sperm chemotaxis towards an egg was first described by Dan (1950) and is important for increasing gamete interactions and reducing hybridization (Riffel et al., 2004). The chemical nature of sperm chemo-attractants has since been identified in cnidarians, molluscs and echinoderms (see Morisawa and Yoshida, 2005). However, the effect of trace metals on sperm activation or chemotaxis/chemo-attractants has not been examined to date and is a potential pathway for metal toxicity. This is clearly an area that merits further scientific study. Other assessment methods related to the effects of stressors on reproductive success, such as fecundity (e.g. Harrison and Wallace 1990; Ward and Harrison, 2000), could be applied to metal studies. Research on how metals might affect an organism’s ability to detect environmental cues used to synchronise spawning activities are also little understood but worthy of further investigation. Trace metals can affect the viability of individual gametes prefertilization but they are also detrimental to many of the key events involved in fertilization. The acrosome reaction (AR) is necessary for fertilization success in many marine invertebrates (Dan, 1967; Tosti and Ménézo, 2016). Although assessment of the effect of trace metals on the AR is virtually absent from the literature, a study by Zhang et al. (2010) found that the AR in the mud crab was more sensitive to trace metal exposure than the widely used sea urchin sperm cell assay, where the elevation of the fertilization membrane is the typical toxicity endpoint. Elevated levels of Ag+ , Cd2+ , Cu2+ , and Zn2+ , significantly reduced the number of completed acrosome reactions, and it was suggested that these metals may block the active sites of the acrosomal filaments and potentially inhibit the AR by disrupting calcium ion (Ca2+ ) flow. Opening of the calcium channels in the sperm membrane and the subsequent influx of calcium is an essential initial phase of the AR, and it is thought that metal ions may prevent the AR by competitively inhibiting calcium influx by blocking calcium channels (Liévano et al., 1990). Once the egg has been successfully penetrated by the sperm, it undergoes significant changes in membrane ion permeability, so beginning the process of egg activation. Electrical modifications are facilitated by the activation of ion channels located on the plasma membrane, including calcium and sodium channels (Tosti, 2006; Tosti and Boni, 2004). Trace metals can bind competitively or adventitiously to numerous biological ligands (Cowan, 1997), and can readily interact with ion channels and disrupt the permeability of cells. Thus, while eggs may be relatively insensitive to trace metal exposure pre-fertilization, the initiation of egg activation and opening of specific ion channels present an opportunity for trace metals to interfere with gamete conductivity (Fitzpatrick et al., 2008). Schäfer et al. (2009) found that copper exposure affected calcium homeostasis in sea urchin eggs during the initial stages of egg activation, with a simultaneous reduction in fertilization success. Mercury was found to inhibit the functioning of ionic channels involved in the initial stages of the electrical response of the egg to fertilization in the ascidian Phallusia mammillata (Franchet et al., 1997). Exposure to highly toxic mercuric ions suppressed the electrical response of the egg by inhibiting the calcium and sodium currents, resulting in an inefficient electrical block to polyspermy and thus a higher incidence of polyspermy. Mercuric ions also inhibited the transformation of the penetrating male nucleus into the male pronucleus − an essential step in fertilization.

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Fig. 1. The effect of copper on fertilization success in the reef coral Platygyra daedalea following spawning in 2014. Eggs and sperm were pre-exposed to copper prior to combination as per the standard approach (a), only sperm were exposed for 40 min prior to crossing (b), only eggs were exposed for 40 min prior to crossing (c), and only eggs were exposed for 2 h prior to crossing (d). Error bars are standard deviation. Metal concentrations are measured.

Whilst the literature on the mechanisms of trace metal toxicity during external fertilization is limited, the above studies give insight into the underlying cellular processes of trace metal toxicity. They reveal the complex nature of toxicity, with different metals capable of affecting multiple aspects of fertilization to varying degrees. The strong link between the toxicity of a given metal and its affinity for the sulfhydryl groups of enzymes and metabolites, and their ability to interfere with ion transport across specific channels, highlight their potential as major pathways of trace metal toxicity (e.g. Quig, 1998; Valko et al., 2005). There are opportunities to expand upon this research area because while there are few studies that have explored cellular mechanisms of trace metal toxicity, very limited information is available on the mechanisms of toxicity at the molecular level. 2.3. The importance of understanding metal toxicity mechanisms Few authors have explored the mechanisms of trace metal toxicity in marine broadcast spawners and fewer have designed experiments to ascertain the potential pathways or targets of toxicity. However, the aspect of fertilization that trace metals affect is not irrelevant and has important ecological ramifications (Marshall, 2006). Hollows et al. (2007) noted that although the manifestation of the effect of trace metals on fertilization may be identical, i.e. a reduction in fertilization success, the way in which fertilization is affected (e.g. impaired sperm viability, gamete interaction, polyspermy block) has distinct environmental consequences. Reproductive success is strongly affected by population density in marine broadcast spawners (Levitan, 2005; Oliver and Babcock 1992), and the detrimental effect of a toxicant on fertilization can vary depending on its mode of action and the density of spawners to which it is released. Population densities of broadcast

spawners vary spatially and temporally and can occur in response to significant ecosystem impacts. Echinoderms, for example, are renowned for large scale population density variations which are often related to anthropogenic influences (Uthicke et al., 2009). In a low density population, a toxicant will have a greater effect on fertilization success if it affects sperm viability or sperm-egg interactions as opposed to a toxicant that affects polyspermy, because these populations are already sperm-limited and gamete interactions are low. This could prove detrimental to coral reefs in particular because anthropogenic climate change and other human impacts have resulted in considerable reductions in coral abundance in many reef regions around the world (Burke et al., 2011). In such degraded, low-density populations, sperm limitation already presents a problem for fertilization success (Nozawa et al., 2015), and could potentially be exacerbated by exposure to trace metals that reduce or impair gamete interactions. Conversely, a metal such as mercury could have a stronger effect on fertilization success in high density populations relative to low density populations, because in suppressing the electrical response of the egg during fertilization, mercury reduces the efficacy of the polyspermy block (Franchet et al., 1997). Therefore, in high density populations where the incidence of polyspermy is already high, mercury exposure will increase lethal polyspermy and further reduce fertilization success. These examples highlight the importance of describing metal toxicity mechanisms in marine invertebrate broadcast spawners. This is a major missing element in the study of metal effects on fertilization success. Studies that investigate the molecular, cellular and physiological pathways of trace metal toxicity on external fertilization would greatly enhance our understanding of toxicity mechanisms, and allow researchers to predict metal sensitivity using biological processes, such as sodium turnover rate (Grosell et al., 2002). A few previous studies have differentially exposed

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gametes to metals (e.g. Fitzpatrick et al., 2008; Gopalakrishnan et al., 2008) and while this method may not reveal the underlying toxicity mechanisms, it is still valuable because it highlights vulnerable targets and can therefore direct future mechanistic research. It is also a relatively straightforward addition to the fertilization assay and can therefore be readily incorporated into experimental approaches.

3. Trace metal toxicity and speciation One of the keys to understanding metal toxicity is understanding the effects of the physico-chemical characteristics of water on metal speciation and bioavailability, as not all forms of metals are equally toxic (Libes, 1992). The concentration of total or dissolved metal can vary greatly depending on the water chemistry. Marine waters tend to have a more stable chemistry than freshwaters and relatively little variability exists in the chemistry of various parts of the open ocean. In contrast, coastal areas can be markedly influenced by surrounding landmasses, which results in more variable physico-chemistry through rainfall and run off, desiccation, and disturbances such as dredging and land reclamation. Metal loads are also higher in coastal marine waters as a result of their proximity to the land and associated human activities (Furness and Rainbow, 1990). These systems are both temporally and spatially dynamic. Copper has a strong complexing capacity for organic material including humic substances and fulvic acids. In water types that are characteristically low in these substances, such as oligotrophic tropical marine waters (e.g. Ramachandran et al., 1997), copper has a relatively high bioavailability compared to freshwater. The organic matter in seawater may be reduced by up to 90% in comparison with freshwater (Furness and Rainbow, 1990). Unlike other metals, the dominant inorganic ligands in marine waters, including Cl− , F− , SO4 2− , HCO3 − , CO3 2− and HPO3 , do not complex strongly with copper. The availability of such ligands is influenced by pH, alkalinity, and salinity. Metal speciation can be predicted using computer models such as PHREEQC and these predictions have been used to support ecotoxicological studies (e.g. Hughes et al., 2005). Some studies have also directly measured specific effects of variations in a single physico-chemical parameter combined with metal exposure (e.g. temperature and copper effects on early life stages of coral (Kwok et al., 2016); humic substances and copper on sea urchin larvae (Lorenzo et al., 2002), and fulvic acids and copper on polychaete larvae (Qiu et al., 2007). Very few fertilization studies include speciation analysis, in part due to time and cost constraints. In addition to traditional extraction and electrochemical techniques for measuring actual concentrations of different metal species, advancing technologies in High Performance Liquid Chromatography − Inductively Coupled Plasma Mass Spectrometry (HPLC-ICPMS) are also enabling direct measurements of metal species in samples (Montes-Bayón et al., 2003). However, theoretical understanding of metal behavior and speciation in marine waters is important to aid in the interpretation and prediction of bioavailability and toxicity. Several theories have been proposed to explain the behavior of metals in different ligand (salinity) environments (Florence and Batley, 1988; Huheey, 1983; Morrison et al., 1989; Tessier and Turner, 1995). The hard-soft acid base theory describes the way in which hard acids (Class ‘A’ metals) and hard bases (e.g. F− , CO3 2− , NO3 − , PO4 3− , Cl− , NH3 ) tend to be small and slightly polarizable while soft acids (Class ‘B’ metals) and soft bases (CO, H− ,CN− , I, C2 H4 ) tend to be larger and more polarizable. As a ‘rule of thumb’, the hard acids form more stable complexes with hard bases and the soft acids form more stable complexes with soft bases (Huheey, 1983). The terms hard and soft are relative, with no sharp divid-

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ing lines between them, and some metals are better categorized as ‘borderline’ metals. The metals have been placed in these different categories based on their complexing preference for ligands that are either electron withdrawing or electron donating (Tessier and Turner, 1995). Interestingly, borderline metals show ligand-binding characteristics that are intermediate between class A and B metals. Typically they have between 1 and 9 outer electrons in their d shell, and include biologically essential elements required in trace quantities by cells; they are incorporated into a variety of different macromolecules and metalloenzymes (e.g. McCall et al., 2000). Turner et al. (1981) classified the stability of borderline metals depending on their binding strengths to fluoride and chloride ions as follows: Fe3+  Fe2+ ≈ Cu2+ < Mn2+ < Zn2+ < Pb2+ < Co2+ ≈ Ni2+ < Cd2+

Metals to the left of the sequence have class A character and form stronger complexes with unpolarizable ligands (fluoro and oxo), whereas those to the right have a stronger class B character and form more stable complexes with more polarizable ligands (chloro and nitro). Trace metals vary in their toxicity to fertilization success of marine invertebrates. The order of toxicity may vary depending on the species but in general terms the order is similar to the binding strength of borderline metals (many of which are trace metals of interest) to fluoride and chloride ions, as described above. This order is linked strongly to the complexing capacity of seawater and for comparison the complexes in freshwater are formed predominantly by oxygen (oxo) containing ligands (nitrates, phosphates, sulfates, and organic acids) whereas most metal complexes in seawater are chloro- and carbonate or bicarbonate complexes (Kester, 1986). Cadmium and copper exhibit the most notable differences in their toxicities between fresh and salt water. Cadmium is considered to be an extremely toxic metal (Hawker, 1990; Sadiq, 1992) and this is true in freshwater environments, where the chloride concentration is low and cadmium forms complexes with oxygen containing ligands. These cadmium oxo-complexes are more labile, and therefore the toxic free metal ion is more available. With the abundance of chloride in seawater, more thermodynamically stable cadmium complexes are formed, thus reducing the availability of the free metal ion and its overall toxicity (Wright, 1995). Conversely, copper in seawater forms more labile chloroand carbonate complexes, therefore the toxic free metal ion is more readily available (Steemann Nielsen and Wium-Andersen, 1970). According to Kester (1986) about 90% of copper in seawater forms carbonate complexes. However, Hawker (1990) contradicts this, suggesting that 90% of copper in seawater is in the form of copper hydroxide. In freshwater, chelation with complex organic acids and particulates effectively binds the copper forming non labile species, thus reducing the availability of the free metal ion (Hawker, 1990). Some other class B metals such as mercury, the borderline metal cobalt, and metalloids such as arsenic, can form bonds with carbon that are stable in water (Stumm and Morgan, 1996). Other mercury species such as mercuric chloride are highly soluble in marine waters but the duration of their existence in the water column depends on the type of species and environmental conditions (Sadiq, 1992). Mercury is transformed to methylmercury in marine sediments primarily by anaerobic bacteria. Once formed, methylmercury compounds are very stable and are lipophilic in nature, resulting in their tendency to biomagnify though the food chain and accumulate in higher organisms (Stumm and Morgan, 1996).

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3.1. Sediment water interactions Adsorption on to the surfaces of suspended particles plays an important role in the removal of trace metals from marine water. The capacity for trace metals to bind to these surfaces will depend on the size, composition and abundance of the particles, concentration of other ions in solution, the charge of the metal ion, and pH of the solution. Trace metal adsorption onto suspended particles is a significant mechanism controlling the solubility and dispersion of pollutants (Sadiq, 1992). Such sediment water interactions can be important factors influencing trace metal concentrations in the overlying water. Iron, for example, may transform from the soluble Fe2+ to the comparatively insoluble Fe3+ in marine waters. Additionally, freshly precipitated colloidal iron is particularly effective in scavenging trace metals compared to aged colloidal material (e.g. Förstner, 1987). Further to this, the formation of insoluble trace metal sulfides, under slightlyalkaline pH and anoxic conditions, limits the bioavailability of most metals in reduced sediments. Disturbances from natural or anthropogenic activities influence the redox conditions by enhancing oxidization and can cause sulfide to oxidize to sulfate. As a consequence of this it is possible that some trace metals will release from the sediment to the water column (e.g. Reichelt and Jones, 1994). Therefore, this process of metal mobilization from sediment particles needs to be considered in combination with the fact that the now soluble iron and manganese hydroxides and oxides may then bind to these newly released metals.

4. The effect of experimental design on species sensitivities to trace metals The standardization and development of the fertilization or sperm cell toxicity assay arose from the need for short-term sensitive tests for biomonitoring and hazard assessment in coastal marine and estuarine environments (Dinnel et al., 1989). These tests utilize the aforementioned sensitivity of gametes during external fertilization and provide an assay that is inexpensive, rapid, globally available, and sensitive to a wide range of toxicants (Dinnel et al., 1989; Larrain et al., 1999). This assay has also proved to be easily reproducible, reliable, and with good toxicant discriminatory abilities − all of which meet the criteria for assays used for routine monitoring purposes (Ghirardini et al., 2005; His et al., 1999). However, procedures for fertilization assays vary between studies: many expose only the sperm to the trace metal/s in question (known as the ‘sperm cell’ or ‘spermiotoxicity’ assay) whilst some expose both eggs and sperm to the toxicant. The precise endpoint often changes subtly between studies but fertilization success is normally defined as the elevation of the fertilization membrane or the onset of cleavage. Sperm cell concentrations are often adjusted by dilution prior to experimentation to provide suboptimal conditions for fertilization (∼80% fertilization success) (e.g. Willis et al., 1997). This prevents the effects of the toxicants being masked by excessive fertilization and allows hormesis effects to become apparent. Fertilization toxicity data exist across a range of marine broadcast spawners with some phyla, such as the echinoderms, being particularly well-represented. While the gametes of marine broadcast spawners might be expected to be similarly sensitive to trace metals, given that all gametes are equally exposed to the environment and must undergo the same process of external fertilization, including gamete interaction, entry of the sperm into the egg and fusion of the genetic material, the process of fertilization in these organisms shows a marked variation in its sensitivity towards trace metals (Table 1). The general trends in sensitivities between different metals maintained across phyla can be explained by the hierarchy of metal

toxicities − some metals are inherently more toxic than others (e.g. copper, mercury, silver) because of their biological mode of action and the chemistry, speciation, and hence bioavailability of the metal in seawater (Luoma et al., 1995). However, when comparing toxicity data for a specific trace metal, there are differences in species sensitivities of several orders of magnitude, even for species of the same class or genus (Table 1). It is difficult to ascertain whether these differences are genuine, for example as a result of morphological and ultrastructural differences in gametes seen across aquatic invertebrates (e.g. Harrison and Jamieson 1999; Lewis and Ford, 2012), or rather a result of disparities between experimental designs. Some jurisdictions have standardized the fertilization assay: USEPA and Environment Canada (EC) each have a standard fertilization assay protocol for specific species of sea urchins and sand dollars (EC, 2011; USEPA 1995, 2002). These tests form part of several core aquatic toxicity tests, which together measure toxic effects using organisms representing several taxonomic groups and trophic levels (Sergy, 1987). These protocols, which will be discussed later, are not included in this appraisal of the effects of experimental design on toxicity testing. They apply only to five species of echinoids, and the data from these tests are used for monitoring and regulatory purposes, and therefore are not within the framework of exploratory ecotoxicology. 4.1. Variables of the experimental design can impact trace metal toxicity The toxicity of various metals, expressed as EC50 s, to fertilization success across a range of marine taxa, is shown in Table 1 and also highlights the methodological variations between studies. Variations exist in almost all aspects of the experimental design: the duration of toxicant exposure, gamete exposure protocol, metal salt used, experimental media, and ecotoxicological endpoint all vary between studies to different degrees. When experimental methods vary between studies, authors become limited in their ability to interpret and compare their toxicity results with those of other relevant species and studies (e.g. Lee et al., 2004; Novelli et al., 2003). Issues can also arise when these methodological variations are not accounted for or considered during data analysis, which may lead to a misinterpretation of species and endpoint sensitivity. Variations in experimental designs not only limit our analytical powers but can also influence the calculation of water quality guideline values. Toxicity data are used to derive toxicant guideline values (GV), which are environmental concentrations or thresholds that, if exceeded, would indicate a potential environmental problem and elicit a management response. In Australia and New Zealand, the toxicant GV derivation method is being revised to incorporate scientific advances since the release of the 2000 Guidelines (ANZECC/ARMCANZ, 2000; Batley et al., 2014). While these guidelines have rules for toxicity data inclusion and stipulate data quality requirements (only those data sets of ‘high’ or ‘acceptable’ quality are deemed suitable for use in GV derivation), they do not differentiate between data that use different methodologies. However, the selection of the most sensitive data during the derivation process acts to minimize any effects associated with these variations in methodologies. 4.1.1. Test seawater Many factors of the experimental design have the potential to influence the test result and hence apparent sensitivity of an organism. For example, the nature of the assay media used during testing can affect trace metal speciation and ultimately the toxicity of the trace metal. Some studies use artificial seawater (e.g. Bellas et al., 2004; Xie et al., 2005), which contains little to no organic matter. While this might ensure the repeatability of lab-

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oratory results, it is undesirable because natural organic matter is present in field conditions and can potentially significantly affect trace metal toxicity. Organic matter provides ligands for binding trace metals and thus reduces metal bioavailability. For example, Donat et al. (1994) characterized the speciation of dissolved copper in South San Francisco Bay and found that it exists predominantly as organic complexes (80–92% of total dissolved copper). Trace metals also partition onto particulate matter in seawater which includes organic matter, clays and carbonates (Cauwet, 1978). Particulate matter is generally considered to be the fraction that is retained on a 0.45 ␮m filter and therefore studies that use filtered seawater, of which there are many, effectively exclude all or some of the particulate fraction, depending on the pore size used (typically between 0.2 ␮m − 0.7 ␮m). This ultimately affects metal speciation by limiting complexation with particulates in normal seawater. Although no published studies to date have compared the effect of using filtered and unfiltered seawater on the sensitivity of trace metal fertilization assays, larvae of the coral Acropora tenuis were found to be similarly sensitive to copper when filtered seawater (0.45 ␮m) was used as the assay media compared to unfiltered seawater (48 h LC50 for larval survival: 45 ␮g L−1 and 39 ␮g L−1 , respectively) (Reichelt-Brushett, unpublished data). While the sensitivity of coral larvae to copper exposure was not markedly affected by the nature of the assay media, it is not known whether its effect on fertilization success is similarly negligible. Therefore, the effect of using filtered and unfiltered seawater on the sensitivity of fertilization assays needs to be determined, which will allow for future assessment of the suitability of using filtered seawater during toxicity testing. In addition, dissolved and particulate matter loadings and other important environmental parameters vary between coastal and offshore waters, hence further studies examining fertilization responses of various taxa across coastal-offshore water quality gradients are needed to provide a more comprehensive and meaningful assessment of trace metal toxicities in future. Whilst the use of natural seawater in fertilization assays is favourable, there are difficulties associated with its use. Natural seawater may not be accessible for some laboratories, and the receiving water of choice may be contaminated with pollutants that are not easily characterized, and which could impact upon toxicity findings. Also, the use of artificial seawater may be appropriate and mandatory for some environmental monitoring programmes (see USEPA, 2002). However, there are uncertainties associated with using artificial seawater; Bielmeyer et al. (2004) measured the water quality characteristics of artificial seawater made from seven different synthetic sea salt brands, and used these characteristics to predict the effect of synthetic salt brand on copper toxicity to the saltwater mussel Mytilus edulis, using the Biotic Ligand Model (BLM). Predicted dissolved copper EC50 s varied by a factor of 10.7 among brands, therefore Bielmeyer et al. (2004) emphasized the need for caution when conducting toxicity tests using artificial seawater. Ultimately, the use of artificial and filtered seawater in toxicity testing comes at the expense of truly understanding the toxicity of a given contaminant in the receiving environment, and hence diminishes the ecological relevance of the findings. The use of uncontaminated receiving water as test dilution water is permitted by both EC (2011) and USEPA (1995, 2002) echinoid protocols, provided that the sample of seawater is taken beyond the influence of the contaminant or effluent. 4.1.2. Sperm and/or egg exposure The manner in which gametes are exposed to the toxicant can affect their sensitivity and this is exemplified by experiments conducted by Gopalakrishnan et al. (2007, 2008). Gametes of the polychaete H. elegans varied remarkably in their sensitivity to trace metals depending on whether they were exposed during or prefertilization. When the gametes were mixed with the toxicants Pb,

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Ni, and Zn, EC50 values for fertilization success were 30.37 mg L−1 , 38.30 mg L−1 and 44.22 mg L−1 , respectively. However, when the eggs were exposed to these metals for 20 min prior to the addition of untreated sperm, EC50 values fell to 0.692 mg L−1 , 1.178 mg L−1 and 2.025 mg L−1 , respectively. In this case, the variation in experimental protocols between studies may have been beneficial as it inadvertently revealed something of the nature of toxicity during fertilization. The apparent lower sensitivity of gametes when they were mixed simultaneously with the metal as opposed to separate pre-exposure led the authors to conclude that the formation of the fertilization membrane may preclude the entry of toxicants into the oocyte by acting as a physical block (Gopalakrishnan et al., 2008). However, without due consideration for the difference in experimental designs, these data may have been misleading, as fertilization can occur rapidly and may mask the effects of toxicants when mixed simultaneously. Gamete exposure protocols vary between published coral fertilization assays. Whilst some investigators dose the eggs and sperm separately before combination (‘pre-exposure’) (e.g. Kwok et al., 2016; Reichelt-Brushett and Harrison, 1999, 2005), others mix both gametes simultaneously with the toxicant (e.g. Puisay et al., 2015; Victor and Richmond, 2005). These methodological differences can affect the resulting sensitivities. Those studies that pre-expose the eggs and sperm to trace metals separately (Table 1–‘exposure characteristics’) usually do so for 30 min before combining them and allowing fertilization to take place with the toxicant for a further 5 h. Other studies mix the contaminant, egg and sperm simultaneously and let them develop for a set period of time (3–5 h) to allow fertilization and early cleavage to reach an endpoint. Given that trace metals have been shown to detrimentally affect gametes, and particularly sperm, prior to fertilization, we may deduce that assays which pre-expose gametes to toxicants may result in more sensitive ecotoxicological data than those that mix gametes and toxicants simultaneously. This is because in the preexposure assays, trace metals have a longer period of time to exert their toxicological effects solely on the gametes before fertilization and cleavage occurs. With no pre-exposure, immediate fertilization may take place before the toxicant can exert its influence and mask the effects of toxicants. In the field, gametes may be exposed to the elements during broadcast spawning events for some time before egg-sperm contact (e.g. Nozawa et al., 2015; Oliver and Babcock, 1992), and thus pre-exposure protocols may be more reflective of natural environmental conditions.

4.1.3. Duration of exposure Exposure time is an important factor determining trace metal toxicity and is arguably as important as dose concentration (Rozman, 1998). Both median effective and lethal effective concentration values (EC50 and LC50 ) are related to the period of exposure and both decrease with increasing exposure time (Walker et al., 2012). Trace metal exposure time can vary between ecotoxicology studies for similar species and of all the experimental design variables this may have the greatest impact on toxicity. Lee et al. (2004) were limited in their ability to compare the results of their fertilization assay for the Asteroidea echinoderm Asterias amurensis with data from another class of echinoderms, the Echinoidea, because the authors’ trace metal exposure time (sperm 20 min) was shorter in duration than for many other studies (sperm 60 min). Their estimated EC50 value of 200 ␮g L−1 Cu was one order of magnitude greater than the echinoid EC50 values − whether this was genuine or, more likely, as a result of different exposure times, was not discernible. Similarly, Thongra-ar (1997) could not meaningfully compare the sensitivity of fertilization in the sea urchin Diadema setosum to trace metals with other echinoids because their

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Table 2 Recommendations for standardizing fertilization assays in marine broadcast-spawning invertebrates to reflect environmental conditions and assist in future comparative assessment of organism sensitivity. Component of Protocol

Current practices

Recommendation

Nature of assay media used Gamete exposure

Artificial seawater, or filtered seawater. Simultaneous mixing of gametes with toxicant, no pre-exposure. Only sperm pre-exposed. Varying duration of pre-exposure times between studies. Varying between assays for the same groups (e.g. scleractinians: between 3–5.5 h).

Natural unfiltered seawater, where possible. Separate pre-exposure of gametes prior to crossing for all broadcast spawners. Duration of pre-exposure time uniform within groups of organisms. Standardization of exposure times post-crossing. Times should reflect the length of gamete viability in the field and gamete contact time necessary for optimal fertilization rates. Speciation and metal solubility studies in seawater need to be considered for some metals. Sperm concentration should be adjusted to provide sub-optimal conditions for fertilization (∼80% fertilization success). This may require a series of sperm dilution assays to determine optimal sperm concentrations in previously untested species.

Length of exposure

Metal speciation Sperm concentration

Effect of metal speciation not considered during most fertilization assay studies. Not always diluted to sub-optimal fertilization conditions.

experimental exposure duration of 20 min was markedly shorter than for the standard sea urchin sperm assay (total 80 min). Coral fertilization assays also vary in duration, with exposure times ranging from 3 to 5.5 h. These methodological differences again hinder our ability to make clear assertions about the relative sensitivity of a species to a given trace metal (Hédouin and Gates, 2013). Given the ecological importance of corals as foundation species on coral reefs (Harrison and Booth, 2007) and the paucity of toxicological data for regionally-relevant tropical species, it is unfortunate that our understanding of trace metals toxicities are diminished because of inconsistent experimental designs. 4.2. Other factors determining sensitivity to trace metals Some authors have studied the effect of trace metals on fertilization success across a range of species using the same protocol (e.g. Nacci et al., 1986; Ringwood, 1992), thus presenting an opportunity to distinguish the effects of experimental approach from species differences and other sources of variability. Dinnel et al. (1989) compared the effects of five trace metals on fertilization success in four species of echinoderms. Sensitivities to trace metals were of a similar magnitude across all species for each metal, although there were a few notable differences. The red sea urchin (Strongylocentrotus fransicanus) had a sensitivity to copper which was an order of magnitude less than for the other echinoderms (EC50 1.9 ␮g L−1 Cu, compared to EC50 s 25–59 ␮g L−1 Cu), and the sand dollar (Dendraster excentricus) was more sensitive to zinc by an order of magnitude compared to the other species (EC50 28 ␮g L−1 Zn, compared to EC50 s 262–383 ␮g L−1 Zn). These differences in sensitivities demonstrate that there are many factors that determine sensitivity. Species differences can cause variability because metals can exert their toxicity in a species-specific manner (Chan and Chiu, 2015); this could be a result of differences in gamete ultrastructure and morphology, biochemical processes, and underlying tolerances to stressors between species. The physiological condition of the adult can also influence the sensitivity of fertilization to trace metals, as stress has been shown to affect gamete viability and reproductive success in marine invertebrates (Au et al., 2001; Bayne et al., 1978). 4.3. Future development of fertilization assays Whilst most authors mention differences in experimental methodologies when comparing fertilization assay data and recognize the limiting effect it has on data comparison and interpretation, with some foresight these issues can be overcome. This clearly depends on the objectives of the study: if the aim is to further scientific knowledge on a specific aspect of trace metal tox-

icity mechanisms during external fertilization, then consideration of methodological variations may not be necessary. However, if the purpose of a study is to provide data for environmental management (e.g. water quality guidelines) or to assess the sensitivity of an organism in relation to others, then attention should be given to the applicability of the results to the marine environment and the experimental methodologies used in previous relevant studies so that data can be directly compared. The experimental design should, where feasible, reflect the conditions that the gametes would encounter in the field. It is important to remember that the ultimate goal of ecotoxicological testing is to predict the effect of a toxicant in the environment, at various organizational levels, and that this should underpin the rationale of the experimental design. Ensuring that assays are applicable to the marine environment includes consideration of the bioavailable fraction of the trace metal available to the organism (which is related to the type of assay media used), the likely duration of exposure before the zygote is formed and embryogenesis ensues, and whether the gametes will be separately exposed to the toxicant prior to fertilization. Some of these factors will vary depending on the biology of the organism in question and seasonality of spawning. Whilst the fertilization assay is relatively uniform within some groups of marine organisms (e.g. scleractinians, echinoderms), small variations in protocols are evident in some newer publications (e.g. Kwok et al., 2016; Victor and Richmond, 2005). Standardization of the fertilization assay is important for a number of reasons. It facilitates direct comparisons of fertilization assay data between studies, and also allows for more robust derivation of water quality guideline values by ensuring that differences in sensitivities (as effective concentration values) are genuine, and thus the calculated guideline values can be based on more accurate information. While the differences in some guideline values may be relatively small if these recommendations are adopted, it is still advisable that we strive for the most robust data and derivation method possible. If future fertilization assays deviate from the standardized methodologies, the variations should be minimised and justified. A summary of recommendations to improve and standardize fertilization assays in broadcast spawning marine invertebrates is provided in Table 2. There are standard fertilization assays in the U.S. and Canada that have been primarily designed for use in monitoring and regulatory programmes (EC, 2011; USEPA 1995, 2002). These protocols are not applicable to exploratory ecotoxicology, where novel species and taxonomic groups from a variety of habitats are tested, because they are species specific and logistically difficult for some laboratories, especially those located in developing tropical nations with limited resources. They can, however, aid in the development

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of standard protocols for other organisms. In comparison with our recommendations for standardizing the fertilization assay, USEPA and EC guidelines also have standard exposure duration periods, although these do differ. USEPA guidelines have only one exposure duration (1 h sperm exposure, eggs then added for a further 20 min), and EC guidelines have a choice of 3 exposure durations, depending on research objectives. The assay media (“test/dilution water”) can be artificial seawater, unpolluted receiving seawater, or unpolluted laboratory seawater, depending on research goals, whereas we recommend the use of natural seawater. Filtration of dilution/control seawater is not required or recommended for both North American guidelines, and similarly we advocate the use of unfiltered seawater. Both protocols stipulate that only the sperm is pre-exposed to the toxicant before eggs are added (‘spermiotoxicity’ assay), but we recommend exposing both sets of gametes to the toxicant before crossing as this is more reflective of environmental conditions. 5. Conclusions and direction for future research The continued generation of ecotoxicological data to assess the effects of trace metals on sensitive life-history stages such as fertilization is important if we are to reliably predict the effects of pollution and manage discharges in marine ecosystems. However, future research could expand upon the idea of metal toxicity by approaching it in a more holistic manner. This review has identified opportunities for improving the environmental relevance of ecotoxicological data concerning the effects of trace metals on external fertilization. This can be achieved by highlighting the cellular and molecular mechanisms that underpin trace metal toxicity in external fertilization processes, understanding the theory of metal behavior and speciation in marine waters and their interaction with organisms, and standardizing the fertilization assay among groups of organisms to allow critical and comparative analyses between species and their sensitivities to trace metal exposure. A more inclusive and considered approach to studying the effect of trace metals on external fertilization in marine invertebrates is essential in future because chronic low-level and intensifying pollution in most coastal environments is occurring in combination with other anthropogenically-induced environmental changes, and the consequences of these global shifts in climate and oceanographic conditions on trace metal toxicity is not known. Acknowledgments Sincere thanks to Barbara Harrison for help during experimental studies presented in this paper and to Abbey Bromley for completing the post-experiment fertilization assessment and some counts for the species Platygyra daedalea. Thanks to the Environmental Analytical Laboratory at Southern Cross University for metal analyses. Constructive comments on the manuscript by anonymous reviewers are gratefully acknowledged. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.aquatox.2016.12. 019. References ANZECC/ARMCANZ, 2000. Australian and New Zealand Guidelines for Fresh and Marine Water Quality. National Water Quality Management Strategy Paper No 4. Australian and New Zealand Environment and Conservation Council and Agriculture and Resource Management Council of Australia and New Zealand, Canberra, Australia.

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