Infertility in male aquatic invertebrates: A review

Aquatic Toxicology 120–121 (2012) 79–89

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

Review

Infertility in male aquatic invertebrates: A review Ceri Lewis a,∗ , Alex T. Ford b a b

Biosciences, College of Life and Environmental Sciences, Geoffrey Pope Building, University of Exeter, Stocker Road, Exeter EX4 4QD, UK School of Biological Sciences, University of Portsmouth, King Henry Building, Portsmouth PO1 2DT, UK

a r t i c l e

i n f o

Article history: Received 31 January 2012 Received in revised form 15 April 2012 Accepted 2 May 2012 Keywords: Male fertility Endocrine disruption Spermatozoa Toxicity

a b s t r a c t As a result of endocrine disruptor studies, there are numerous examples of male related reproductive abnormalities observed in vertebrates. Contrastingly, within the invertebrates there have been considerably less examples both from laboratory and field investigations. This has in part been due to a focus of female related endpoints, inadequate biomarkers and the low number of studies. Whether contaminant induced male infertility is an issue within aquatic invertebrates and their wider communities therefore remains largely unknown and represents a key knowledge gap in our understanding of pollutant impacts in aquatic wildlife. This paper reviews the current knowledge regarding pollutants impacting male infertility across several aquatic invertebrate phyla; which biomarkers are currently being used and where the science needs to be expanded. The limited studies conducted so far have revealed reductions in sperm numbers, examples of poor fertilisation success, DNA damage to spermatozoa and inhibition of sperm motility that can be induced by a range of environmental contaminants. This limited data is mainly comprised from laboratory studies with only a few studies of sperm toxicity in natural populations. Clearly, there is a need for further studies in this area, to include both laboratory and field studies from clean and reference sites, with a focus on broadcast spawners and those with direct fertilisation. Biomarkers developed for measuring sperm quantity and quality in vertebrates are easily transferable to invertebrates but require optimisation for particular species. We discuss how sperm tracking and techniques for measuring DNA strand breaks and sperm viability have been successfully transferred from human infertility clinics to aquatic invertebrate ecotoxicology. Linking sperm toxicity and male infertility effects to higher level impacts on the reproductive biology and dynamics of populations requires a much greater understanding of fertilisation dynamics and sperm competition/limitation for invertebrate species and represents the next challenge in our understanding of male toxicity effects in natural populations. © 2012 Elsevier B.V. All rights reserved.

Contents 1. 2.

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Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.1. Sperm morphometrics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2. Viability and staining assays . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3. Resazurin activity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4. Comet assay . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.5. Motility assays . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Studies on aquatic invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1. Crustacea . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2. Polychaetes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3. Molluscs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4. Echinoderms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.5. Other aquatic invertebrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Future directions and unanswered questions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

∗ Corresponding author. Tel.: +44 1392 263782; fax: +44 1392 263434. E-mail addresses: [email protected] (C. Lewis), [email protected] (A.T. Ford). 0166-445X/$ – see front matter © 2012 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.aquatox.2012.05.002

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1. Introduction The sperm cell is often misconstrued as being a simple cell that’s merely a swimming parcel of paternal DNA, yet this ‘simple’ cell is the most diverse cell type in the animal kingdom. Numerous specialisations in sperm form have been described, including loss of flagellum, multiple flagellum, gigantism and polymorphism (Pitnick et al., 2009). Aquatic invertebrates adopt a vast array of reproductive strategies, with variations in sperm design depending on fertilisation strategy. These range from the primitive ect-aquasperm and ent-aquasperm of the broadcast spawning mating strategies of sessile marine invertebrates (Jamieson and Rouse, 1989) to the non-motile sperm of copulating crustaceans (Fig. 1). In some cases water-borne sperm are packaged as spermatophores or spermatozeugmata, but most species with this pattern of mating release dispersing, unpackaged sperm, which arrive at the female as individual gametes. Conditions range from strong sperm competition in copulating and brooding species such as the majority of crustaceans (Diesel, 1990) and colonial invertebrates (Pemberton et al., 2003) to sperm limitation in low density broadcast spawning populations (Levitan and Petersen, 1995) hence the selection pressures acting on sperm production, morphology and physiology vary accordingly.

It is often assumed that the large numbers of sperm produced by an aquatic invertebrate, particularly those that free spawn their sperm into the water column, means there is large scale redundancy associated with this spawning strategy. This may be true under certain circumstances; however, in situ fertilisation is rarely 100% for any species (Marshall, 2002; Miller and Mundy, 2005), and there are energetic costs associated with sperm production (Lewis and Wedell, 2007). That fact that sperm are the most diverse cell type known suggests strong evolutionary pressures acting on their production. This does not reconcile well with the idea of wastage. Contaminant induced impairment of sperm quality must therefore have the potential to impact upon population reproductive success. We can postulate that reduced sperm function might be overcome by greater sperm production in males living in contaminated environments, since there is field evidence that male reproductive success is related to sperm allocation (Yund and McCartney, 1994). However, as yet, there is very little information describing the phenotypic plasticity of sperm production in aquatic invertebrates. Successful fertilisation for any species depends on the production of high quality sperm, maintenance of DNA integrity and fertilisation capacity and appropriate motility responses to oocytes and spawning medium (in this case water). All of these aspects of fertilisation have the potential to be disrupted by environmental

Fig. 1. (A) Patterns of Molluscan sperm head structure in a number of bivalves; i: Anadara broughtonii, ii: Arca boucardi, iii: Pododesmus macrochisma, iv: Macoma tokyoensis, v: Crassostrea gigas, vi: Mya japonica, vii: Trapezium liratum, modified from Drozdov et al. (2009). (B) Crustacean sperm morphology showing typical sperm morphology for the following groups; i: Brachyuran (Decapod); ii: Mysidae; iii: Euphasia (modified from Tudge, 2009); iv: Palaemonid shrimps; v: Branchiopods; vi: Cephalocardia; vii: Anispidacea (modified from Jamieson, 1989); (C) variation among echinoids in sperm head dimension. From Eckelbarger et al., 1989): a: Phrissocystis multispina (paraspermatozoan), b: Phrissocystis multispina (euspermatozoan), c: Aspidodiadema jacobyi, d: Salenia goesiana, e: Coelopleurus floridanus, f: Conolampus sigsbei, g: Brissopsis atlantica, h: Paleopneustes tholoformis, i: Paleopneustes cristatys, j: Cidaris blakei, k: Linopnuestes longispinus, l: Calocidaris micans, m: Phormosoma placenta, n: Araeosoma fenestratum, o: Araeosoma belli, p: Hygrosoma sp. q: unidentified echinothurrid; (D) SEMs of sperm morphology for the polychaetes i: Marphysa elityeni and ii: Marphysa sanguinea. With permission from Lewis and Karageorgopoulos (2008).

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perturbances or exposure to xenobiotics. Whilst somatic cells and oocytes contain a variety of proteins, anti-oxidants and DNA repair enzymes that repair and protect against environmentally induced damage, sperm are generally considered to have little or no capacity for DNA repair or anti-oxidant defence (Aitken et al., 2004). Human health research has clearly demonstrated that the male reproductive system is a major target of environmental chemicals and that sperm DNA integrity can be adversely affected by exposure to ubiquitous pollutants such as polycyclic aromatic hydrocarbons (PAHs) and phthalates. Sperm are thought to be particularly susceptible to oxidative damage due to the abundance of polyunsaturated fatty acids acting as substrates for reactive oxygen species. In humans this oxidative damage to sperm DNA has been implicated in a range of patho-physiological conditions, however virtually nothing is known about the impact of oxidative DNA damage in the sperm of aquatic organisms. Despite this, male toxicity effects are rarely studied compared to the wealth of knowledge that exists for maternal effects. Aquatic invertebrate sperm can be exposed to any environmental contaminants present in the seawater both during the stages of spermatogenesis and in free spawning species during spawning and fertilisation. Spermatogenesis in aquatic invertebrates can be an extended process lasting anything from a few week to many months (for example gametogenesis can last up to 9 months in the lugworm Arenicola marina (Betteley et al., 2008). Invertebrates generally lack the blood-gonad barrier found in mammals, which means that gametes are exposed to environmental toxicants during their development and maturation as well as during the process of reproduction. For vertebrates there is a vast wealth of knowledge regarding sperm biology, physiology and biochemistry, ranging from the evolutionary aspects of sperm competition theory to the bourgeoning fields of sperm proteomics (du Plessis et al., 2011; MartinezHeredia et al., 2006) and epigenetics (Singh et al., 2011). The majority of current knowledge of sperm toxicity effects and infertility is derived from the need to understand and treat human male infertility issues. Techniques for the assessment of sperm quality of natural populations for environmental monitoring purposes have mostly been derived from human health and IVF clinics. For example human health studies are now looking at how environmental toxins/drugs may affect fertility via epigenetic modifications. Drugs such as the anticancer agent 5-aza-20-deoxycytidine have been shown to cause a decrease in global DNA methylation that leads to altered sperm morphology, decreased sperm motility, decreased fertilisation capacity, and decreased embryo survival (Oakes et al., 2007). Such studies demonstrate the complex nature of male infertility pathology and the strength of epigenetics as tool for its study. Similarly, endocrine disruptors, such as methoxychlor (an estrogenic pesticide) and vinclozolin (an anti-androgenic fungicide) have been found to affect epigenetic modifications that may cause spermatogenic defects in subsequent generations (Anway et al., 2005). Both of these chemicals have environmental applications but very little information is available on their environmental fates or male fertility effects in wildlife. For invertebrates, and particularly aquatic invertebrates, there is very little known regarding the occurrence or consequences of contaminant induced male infertility. A few authors (Kobayashi, 1980; Mahadevan, 1986) have investigated the relative merits of sperm, fertilisation and embryo development for generalised acute toxicity bioassays, mostly demonstrating that sperm were generally more sensitive to contaminant induced disruption. In particular sperm motility and oxygen consumption were found to be highly sensitive toxicity endpoints for a number of contaminants, including heavy metals. Recent data also suggests that the change in seawater pH associated with CO2 induced acidification of the world’s oceans can act to reduce sperm swimming speeds, and hence fertilisation success in marine invertebrates (Havenhand et al., 2008; Morita

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et al., 2010). The possibility of interactive effects of reduced pH and contaminant exposure has yet to be investigated, and might be critical to our understanding of how environmental change will affect reproduction in marine animals. 2. Techniques Ecotoxicology studies looking at the impacts of environmental contaminants on male reproductive health generally use some measure of ‘sperm quality’ as their end point. The term ‘sperm quality’ usually refers to some measure of the ability of sperm to bring about successful fertilisation, but this can relate to a range of sperm physiological and morphological parameters. 2.1. Sperm morphometrics Assessing morphological abnormalities in mature and developing sperm is probably the oldest, yet still the simplest and most inexpensive ways of assessing sperm quality and predicting sperm function (Trivedi et al., 2010; Wyrobek et al., 1975). Evaluation of sperm head morphology is still regularly used in mammalian studies for the screening of chemicals which might be teratogenic, mutagenic and carcinogenic in nature. Sperm are stained, examined under a microscope and sperm head morphology categorised as being normal, quasinormal and grossly abnormal using a set of standardised characteristics first described by Burruel et al. (1996). Sperm morphology has been shown to be negatively affected by a range of environmental pollutant in mammals (Apostoli et al., 1998; Bakare et al., 2005; Trivedi et al., 2010), fish (Hatef et al., 2011; McAllister and Kime, 2003) and amphibians (Feng et al., 2011). Contaminant-induced abnormalities in mature sperm does not appear to have been investigated for any marine or freshwater invertebrate species, although this has been shown for UV exposures in sea urchin sperm (Pruski et al., 2009). Yang et al. (2008) investigated the impacts of environmental contaminants on the earlier stages of spermatogenesis and gonad development has been investigated for a number of marine invertebrates and found testicular abnormalities and reduced sperm counts in amphipods from contaminated sites. Morphological abnormalities in the mature sperm were not determined in this case. 2.2. Viability and staining assays Viability assays generally use staining techniques to assess the proportion or live/dead spermatozoa in any sample or can look at more precise measures of sperm function according to the type of stain used. These viability biomarkers can then be assessed using simple microscopy techniques or using flow cytometry as a faster, high-throughput technique for sperm analysis (Graham et al., 1990). A suite of specific fluorescent marker dyes can be used to analyse a range different sperm functions, the simplest of which is the live/dead sperm viability assay. For example SYBR® -14 dye, a nucleic acid stain that labels live cells with intact membranes with fluoresces bright green, can be used in combination with propidium iodide (PI) dye which labels cells with damaged cell membranes a fluorescent red (Garner and Johnson, 1995). More specific dyes can be used to look at more detailed sperm functions, such as the mitochondrial stain MitoTracker-Red CMXRos (M-7512; Molecular Probes, Eugene, OR), a red fluorescent dye that stains mitochondria in live cells, with dye accumulation being contingent upon membrane potential. The acrosome reaction can also be analysed using fluorescein isothiocyanate (FITC)-conjugated Arachis hypogaea (peanut) lectin (FITC-PNA). The PNA lectin is specific for terminal ␤-galactose moieties and so will bind to the acrosome in acrosome-reacted sperm and fluoresce green (Ashizawa et al., 2006).

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Fig. 2. Comet assay analysis of sperm DNA damage following paternal exposures to increasing concentrations of the known genotoxin methyl-methane sulphonate (MMS) in the mussel Arenicola marina.

Sperm viability and functional biomarkers have been widely used in mammalian sperm toxicity tests, but have been tested in a just few aquatic invertebrate species. Favret and Lynn (2010) demonstrated that these biomarkers are effective at detecting changes in several sperm viability parameters in two echinoderms and a molluscs after paternal laboratory exposure to pesticides and therefore have great potential for the analysis of field specimens. Employing these types of analyses on sperm for ecotoxicity studies could provide a fast and cost effective means of investigating the damage to sperm from a toxicant exposure.

Resazurin is a non-toxic blue redox dye, itself non-fluorescent until it is reduced to the pink coloured and highly red fluorescent resorufin by dehydrogenase activity (Dart et al., 1994). Resazurin is effectively reduced in mitochondria, making it useful to assess mitochondrial metabolic activity, and has been applied as a general biomarker of sperm ‘quality’ in a number of human (Rasmussen, 1999), mammalian (Zrimsek et al., 2004) and fish (Hamoutene et al., 2000) studies. In particular it has been used as an indicator of semen quality in breeding programmes for bulls (Zrimsek et al., 2004) and has been found to correlate well with other sperm parameters such as motility and concentration (Venkata Rami Reddy et al., 1998). In marine invertebrates, resazurin activity has been measured in sperm samples of the green sea urchin Strongylocentrotus droebachensis and scallops (Hamoutene et al., 2000) where it was shown to be inhibited by TBT exposure. Whilst currently not widely used for environmental toxicity testing, this simple, rapid colorimetric assay has the potential for field assessments of sperm quality for natural aquatic invertebrate populations.

damage DNA either directly or indirectly via the production of free radicals or after metabolic activation. The body of evidence documenting these effects in sperm for humans and vertebrates is fairly large (Hauser et al., 2007; Xu et al., 2003), very little attention has been given to the potential for environmentally induced genotoxic damage in the sperm of marine or freshwater invertebrate species. A commonly used technique for assessing DNA damage in cells, including sperm, is the alkaline version of the single-cell electrophoresis technique, otherwise known as the Comet assay (example Comet images for the polychaete A. marina are shown in Fig. 2). A small number of studies have used the Comet assay to look at the induction of DNA damage in marine invertebrate sperm, but so far these have mainly been based on laboratory exposures, with very little existing data on environmentally induced DNA damage of sperm. A few studies have looked at DNA damage in the sperm of the polychaetes Nereis virens and A. marina (Lewis and Galloway, 2008, 2009; Caldwell et al., 2011) and the mussel Mytilus edulis (Lewis and Galloway, 2009). These used laboratory exposures to demonstrate that a range of genotoxic chemicals (benzo(a)pyrene; copper and the diatom aldehyde 2E,4E-decadienal) not only induce DNA damage in the somatic cells of exposed adults but also induce significant levels of DNA damage to both stored and freshly-spawned sperm (Lewis and Galloway, 2009). Whilst sperm DNA damage alone does not affect sperm motility or the fertilisation reaction, it does lead to severe developmental abnormalities of the resulting embryos and larvae (Lewis and Galloway, 2009). (Kadar et al., 2011) used the comet assay looked at DNA damage induced in the sperm of the mussel Mytilus galloprovincialis by exposures to zero-valent iron nanoparticles. Significant DNA damage was detected in sperm exposed to the highest exposure concentrations (10 mg L−1 ).

2.4. Comet assay

2.5. Motility assays

A significant proportion of the chemicals entering our marine and freshwater environments have the potential to induce DNA damage or interfere with the processes involved in cell division (Depledge, 1998; Livingstone et al., 2000). These include, amongst others, persistent organics such as PAHs and metals, which can

Sperm motility is regularly used as a proxy for ‘sperm quality’ and is generally measured using computer assisted sperm analysis (CASA). This normally comprises an optics system with a negative phase contrast microscope with ×20 objective and cooled stage with high speed camera and computer software programme.

2.3. Resazurin activity

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Fig. 3. Example screen grab from a CASA analysis showing tracking profiles of Arenicola marina (Polychaetea) sperm in clean seawater and after exposure to 5 ␮M copper sulphate.

Individual sperm can be tracked for various time intervals providing data on percent motility and a number of swimming speed and directional parameters (e.g. curvilinear velocity) which have been demonstrated to be positively correlated with capacity for fertilisation for a number of species (Au et al., 2002; Linhart et al., 2000). Fig. 3 is an example of a CASA screen grab showing sperm motility in the polychaete A. marina before and after exposure to the environmental contaminant copper. CASA analysis has been used for a few marine invertebrate toxicity studies to date and is starting to provide some insight into the mechanisms behind reduced fertilisation success observations for a number of environmental contaminants. Fitzpatrick et al. (2008) looked at the impacts of environmentally relevant copper concentrations on sperm motility in the blue mussel Mytilus trossulus and found a significant reduction in sperm motility with increasing copper concentrations. This reduction in motility was thought to be a result of physiological rather than morphological disruptions and correlated strongly to a significant reduction in subsequent fertilisation success. Caldwell et al. (2011) found a similar result for the polychaete N. virens. They applied CASA to demonstrate the proportion of motile sperm for any male show a decrease with exposure to increasing concentrations of copper sulphate or the diatom aldehyde decadienal with a correlated reduction in fertilisation success. Fabbrocini et al. (2010) investigated the suitability of CASA as a toxicity bioassay using sperm of the urchin Paracentrotus lividus and pore water as the contaminant. They also found sperm motility to be both sensitive and discriminatory in comparison to other standardised toxicity tests, with impacts of pore water on a number of sperm motility parameters. No data on field studies looking at contaminant effects on sperm motility in a natural population could be found.

3. Studies on aquatic invertebrates 3.1. Crustacea Crustacean sperm vary considerably in their external morphology and ultrastructure (see Fig. 1) so much so that for some species they are used for taxonomic purposes (Jamieson, 1990; Koch and Lambert, 1990; Subramoniam, 1993). Studies into the effects of environmental contaminants on spermatogenesis and male fertility are currently limited (Yang et al., 2008). However, despite this there is a considerable body of work that has looked into male fertility and sperm quality/quantity in relation to the crustacean aquaculture including impacts of disease (Primavera and Quinitio, 2000) food quality (Bray and Lawrence, 1998; Leung-Trujillo and

Lawrence, 1991), husbandry (Rodriguez et al., 2007) and cryopreservation techniques (Bart et al., 2006; Gwo, 2000). Yang et al. (2008) recently found that sperm counts in the amphipod Echinogammarus marinus were 20% lower in sites known to be contaminated with PCBs, metals and hydrocarbons than reference sites and asked the question as to whether we have been ignoring the effects on male invertebrates. Studies with other amphipods have shown reduced sperm counts can have dramatic impacts of fertilisation success. For example, Dunn et al. (2006) investigated the impact of sperm limitation in estuarine amphipods (Gammarus duebeni) following successive matings with receptive females. They observed that after 3 successive matings, the sperm count had significantly dropped by approximately 56% and the sizes of broods in females by approximately 55%. In addition, females that were paired with sperm limited males showed approximately 30% reduced embryo success. In a similar study, Lemaitre et al. (2009) found an over 50% reduction of sperm in male Gammarus pulex between paired (and ready to copulate) and recently copulated individuals which suggest a large proportion of sperm are utilised during each copulation event. Studies of the snow crab (Chionoecetes opilio) has also demonstrated clear differences in the spermathecal load (weight, mg) and fertility of clutches. A one order of magnitude drop in spermathecal load, from 31–50 mg down to 3–4 mg results in few or no fertilised eggs (Rondeau and SainteMarie, 2001). Results such as these suggest that in some crustaceans the impacts of contaminants on the number of sperm produced could result in subsequent effects at the population level. The population level effects of reduced sperm numbers were recently investigated by Ford et al. (2012) in several arthropods species. All species reviewed investigated demonstrated reduced fertility with reduced sperm with the relationship varying between species and their results suggested that relatively small reductions in sperm can have variable and possibly dramatic impacts on fertilisation success for the species examined. For example, based on the species reviewed, a 5% reduction in sperm is predicted to have between 0.4 and 9.1% reductions in female brood size dependant on the species. A 5% reduction in sperm in the estuarine amphipod G. duebeni was predicted with have a between 5 and 9% reduction in fertilisation success. Using an amphipod population model the effects of reduced fertilisation revealed that whilst 5% reductions in brood sizes allow the population to still exist after ten years (albeit at critically low densities), brood sizes reductions of 10% or greater result in population collapse in less than 6–7 years. Several studies have demonstrated that laboratory exposure to contaminants can result in abnormal testicular development which may ultimately result in reduced male fertility. Vandenbergh et al. (2003) exposed the amphipod Hyella azteca to the synthetic

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oestrogen 17␣-ethinylestradiol (EE) ranging from 0.1 to 10 ␮g/L. In all EE concentrations histological aberrations of the reproductive tract were observed including disturbed maturation of germ cells and spermatogenesis. Yamaguchi et al. (2008) exposed Japanese freshwater crabs (Geothelphusa dehaani) to arsenic and observed different proportions of sperm cells (spermatids, spermatozoa and spermatocytes) between groups suggesting delayed spermatogenesis was occurring. Wang et al. (2002) injected the freshwater crab (Sinopotamon yangtsekiense) with cadmium and found changes in the ultra-structure of sperm cells in all stages of development with damage caused to the structures such as the acrosome, mitochondria, endoplasmic reticulum and cell membrane. Testing a range of metals (silver, cadmium, copper and zinc), Zhang et al. (2010) calculated the EC50 for two stages changes in the acrosome reaction of the mud crab (Scylla serrata). Results indicated that the 2nd stage of the acrosome reaction (injection of acrosomal filament) was more sensitive with EC50 of 1.96, 0.20, 1.46 and 0.34 ␮g/L for Ag+ , Cd2+ , Cu2+ and Zn2+ , respectively. The authors also observed that sperm cells exposed to heavy metals showed an increased rate of swelling, shape irregularities and the acrosomal filament of some sperm cells were abnormally shaped. Cary et al. (2004) found 73% inhibition reproduction in the copepod Amphiascus tenuiremis when males were exposed to the insecticide Fipronil and mated with control females. In addition, the authors also observed a delayed time (3 days) to brood sac extrusion in exposed males mated with control females. Cary et al. (2004) suggested that Fibronil may reduce or inhibit sperm production, viability and/or mobility. Several recent studies have looked at genotoxicity in different tissues of the amphipod Gammarus fossarum (Lacaze et al., 2011a,b,c, 2010). Lacaze et al. (2010) explored the use of the comet assay in assessing the genotoxic affects of several compounds in testes, ovarian and haemocytes in G. fossarum. The authors found clear concentration–response effects for all tissues, low variability and high sensitivity concluding the comet assay as an effective biomarker of the genotoxicity in this species. Interestingly the authors found that spermatozoa were more sensitive to the comet assay than oocytes (Lacaze et al., 2011a, 2010). Further studies using the same model organism have demonstrated a clear correlation between DNA damage in spermatozoa and embryo abnormalities when pairing exposed males with unexposed females to the known genotoxicants methyl methanesulphonate (MMS) and potassium dichromate (K2 Cr2 O7 ; Lacaze et al., 2011a). The authors noted that the assay demonstrated a clear link between toxic effects at the molecular level with impacts at the individual level of the next generation (Lacaze et al., 2011c). Transferring this assay to the field, Lacaze et al. (2011b) found that when G. fossarum were caged below wastewater treatment plants there was a significant increase in DNA damage in spermatozoa compared to above measured using the comet assay, again confirming the application of this biomarker to the field. 3.2. Polychaetes Polychaetes exhibit great variety in both mode of reproduction and sperm morphology. In many polychaetes, both the sperm and eggs are spawned straight into the water and there is no interaction of any sort between the gametes and individuals (Wilson, 1991). Often, however, sperm, spermatozeugmata or spermatophores may be released freely into the water and are gathered by, or swim to, the female (Rouse, 1999). Wilson (1991) identified 17 different reproductive modes within the polychaetes, with the majority of polychaetes utilising some form of brooding of larvae. Sperm are often packaged for transfer into bundles of sperm that are surrounded by a sheath or capsule that isolates them from the surrounding environment. Whether this ‘packing’ provides any form of protection to the sperm from environmental contaminants once

released from the male is unknown. In polychaetes that reproduce by free spawning their sperm directly into the water column and where there is no contact between sexes, sperm usually conform to the morphology termed ‘primitive’. Such sperm are released directly into the water column where they are exposed to any contaminants present without any form of protection. Many polychaetes, particularly those that free spawn their sperm such as A. marina and the nereids, are ideal candidates for male ecotoxicological research due to lengthy, synchronised spermatogenesis periods and ease of sampling developing and mature sperm from the coelomic cavity (Lewis and Watson, 2012). Despite this only a few studies exist that have examined male toxicity effects in polychaetes. The research undertaken to date has shown that pollutants can disrupt sperm function in polychaetes by affecting sperm morphology during gametogenesis (Rossi and Anderson, 1978; Zheng et al., 2010); disrupting sperm swimming ability (Caldwell et al., 2004, 2011); and inducing sperm DNA damage (Lewis and Galloway, 2009). These have all been laboratory based exposure studies with no field data currently available for male toxicity effects in any polychaete population. Field data does exist for environmentally induced damage to adult somatic cells demonstrating high levels of DNA damage in polychaetes living in contaminated habitats (Lewis and Galloway, 2008) suggesting DNA damage in the sperm of males from these natural populations is also highly likely. In the lugworm A. marina, laboratory exposure to the water soluble fraction of crude oil (WAF) at concentrations equivalent to 3.8 mg L−1 PAHs severely reduced sperm motility with the equivalent effect of increasing sperm dilution by a factor of 104 (Lewis et al., 2008). This resulted in significantly reduced sperm: egg collision rates, which negatively impacted upon fertilisation kinetics and fertilisation success. In N. virens and A. marina, adult exposure to the genotoxic PAH benzo(a)pyrene causes significant damage to sperm DNA (Lewis and Galloway, 2009) but has no impact upon sperm swimming or fertilisation success. Sperm DNA damage does, however, cause teratogenic effects in the resulting embryos and larvae. Both mechanisms of sperm toxicity result in reduced numbers of offspring reaching settlement stage. Naturally occurring toxins (e.g. the diatom aldehyde DDE) have also been shown to induce DNA damage in the sperm of N. virens (Caldwell et al., 2004, 2011) and have the potential to cause reproductive failure in populations breeding periods were to coincide with algal blooms. The surpulid tube worm Hydroides elegans has been used in a number of toxicity bioassay studies (Gopalakrishnan et al., 2008; Thilagam et al., 2008) and is one of the few polychaetes in which sperm toxicity effects has been investigated. Gopalakrishnan et al. (2008) looked at the effects of the heavy metals mercury, cadmium, lead, nickel and zinc on ‘sperm viability’. What they were actually measuring, however, was fertilisation success in experiments where the sperm were exposed to the metal solutions for 20 min prior to being added to unexposed oocytes. Significant reductions in fertilisation success were observed in a dose dependant response. Whether this reduction in fertilisation success was due to reduced sperm viability alone or a combination of effects including reduced motility and ability to undergo the acrosome reaction is not clear from this experimental design. However it does clearly demonstrate sperm toxicity effects separately from egg toxicity effects. 3.3. Molluscs Of the thousands of recognised extant Molluscan species only a very small number, mostly broadcast spawning bivalves, have been examined for any form of male reproductive health impacts due to environmental contaminant exposure. The broadcast spawning bivalve species, such as the mussel M. edulis or the eastern oyster Crassostrea virginica, are the most commonly studied species for

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both male and female reproductive toxicity end points due to the ease of collection of their gametes. The common mussel M. edulis has been used widely for the traditional 48 h embryo development toxicity test, to generate the commonly used no observable effect concentration (NOEC), the lowest observable effect concentration (LOEC), and the EC50 values for various contaminants, however this standardised assay using embryos fertilisation prior to exposure and hence is not representative of any potential spermiotoxicity effects. There are a few studies that have investigated sperm toxicity effects of common environmental contaminants in this popular ecotoxicology species. Lewis and Galloway (2009) demonstrated, that as in polychaetes, paternal exposure to genotoxins in M. edulis resulted in significant DNA damage in sperm and subsequent teratogenic impacts on larval development. Nanoparticles are a major emerging class of contaminant and studies have recently been carried out to assess their effects on spermiotoxicity in the mussel M. galloprovincialis (Kadar et al., 2011). Stabilised (i.e., coated with organic polyacrylic stabiliser) and non-stabilised forms of zero-valent nanoiron (nZVI) on sperm viability and the development of M. galloprovincialis embryos following 2 h exposure of the sperm prior to in vitro fertilisation. Kadar et al. (2011) found both forms of nZVI caused 30% mortality among spermatozoa with a subsequent 20% decline in fertilisation success and significant DNA damage in sperm exposed to the highest exposure concentrations (10 mg L−1 ). Sperm motility in mussels can also be disrupted by contaminants, for example low level copper exposures have been clearly shown to reduce sperm motility in both the blue mussel M. trossulus (Fitzpatrick et al., 2008) and both cadmium and phenol reduced sperm swimming in the green lipped mussel Perna viridis (Au et al., 2000). The numbers of male Crassostrea gigas with motile sperm following exposure to nonylphenol were found to be significantly lower than non exposed males (Nice, 2005). Wintermyer and Cooper (2007) developed a gametogenesis protocol to serve as an aquatic bivalve model for evaluating the impacts of dioxin-like compounds and other xenobiotics on both oogenesis and spermatogenesis using the eastern oyster (C. virginica). They demonstrated that spermatogenesis was disrupted by laboratory exposures to the chemical 2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD), an environmental chemical of concern known to have reproductive toxicity effects in mammals and humans. Spermatogenesis became unsynchronised in males exposed to just 2 pg/g TCDD and spermatogenic cells were morphologically altered. The majority of gastropod molluscs and cephalopods reproduce via copulation, hence due to the inaccessibility of sperm there is very little toxicological information regarding their susceptibility to environmental disruption. Other than the studies looking at tributyltin induced imposex in marine gastropods, which generally focus on morphological abnormalities and use histological end points (Horiguchi, 2006; Matthiessen and Gibbs, 1998; Oehlmann et al., 1991), we found only two marine gastropod reproductive toxicity studies in the literature. One of these used egg production rates as its end point and did not examine any potential sperm toxicity effects at all (Ramasamy and Murugan, 2007). The second, a recent study into the abalone Haliotis diversicolor supertexta, demonstrated reduced ATP-ase activity and ultra-structural changes in the form of mitochondrial swelling and distraction of cristae in sperm after exposure to the endocrine disruptor dimethyl phthalate (DMP) (Zhou et al., 2011). This second study used environmentally measured concentrations of DMP, demonstrating just how sensitive aquatic invertebrate sperm may be to environmental levels of priority contaminants. In cephalopods mating generally starts with a courtship, followed by the transfer of a spermatophore (sperm packet) by a male to a female through her mantle opening. The spermatophore is transferred by the male using either a penis or a modified arm

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called a hectocotylus. Reproductive toxicity studies are generally lacking for this group of molluscs, with most toxicity studies on them concentrating on their accumulation of metals with regard to the safety of their use for human consumption. Embryotoxicity of metals has been demonstrated in Sepia officinalis (Lacoue-Labarthe et al., 2009) but is the only reproductive toxicity report found for the cephalopods. Even the basic sperm morphological studies are lacking for many cephalopod groups, with no reported studies on spermiotoxicity or male reproductive health found to date at all. The females of most gastropods and cephalopods have some kind of sperm storage organ that would allow for sperm to be collected from post-copulation females for use in toxicity studies, however to date no such studies appear to have been undertaken. Different reproductive modes and associated sperm morphologies between these different Molluscan classes are likely to lead to differing sensitivities to environmental pollutants, meaning this is a large knowledge gap in our current understanding of the potential impacts of environmental contaminants on male reproductive health within this large group of ecologically and economically important invertebrates. 3.4. Echinoderms Much of what we know of generalised sperm physiology, form and functioning was initially described in an echinoderm, with sea urchins in particular being used as a model species for developmental biology (Epel, 1975; Summers et al., 1975; Tilney et al., 1978). The sea urchin embryo test (SET) has been widely used and adopted internationally to evaluate the seawater quality for a wide range of pollutants (Allen, 1971; ASTM, 1995; Saco-Alvarez et al., 2010). Despite this huge wealth of data many of these studies do not record parameters relating to sperm quantity or quality and primarily focus fertilisation tests. Au et al. (2000) investigated the effects of cadmium and phenol on the motility (CASA) and ultrastructure of both sea urchin and mussel spermatozoa. The authors found that the urchins were more sensitive than the mussel spermatozoa and demonstrated a good concentration response relationship to cadmium. They also observed that both compounds altered the morphology of the spermatozoa (heads and tails), suggesting that this may alter the swimming ability and hence fertilisation (Fig. 4). Ultrastructual changes observed by Au et al. (2000) included disorganisation of the mitochondrial membranes which the authors suggested may also impact sperm movement (Fig. 4). More recently, several studies have been conducted using the sea urchins have looked at sperm quality parameters using CASA. These have found reduced sperm motility, velocity, percent sperm, and abnormal mulitnucleate sperm and abnormal sperm tails exposed to a range of pollutants including copper, cadmium, phenol and UV-B radiation (Au et al., 2002, 2000, 2001a,b, 2003). Many of these studies were also able to correlate these sperm abnormalities with fertilisation success in relation to exposure concentration and length of exposure (Au et al., 2001a). Hinkley and Wright (1986) observed reduced fertilisation success in sea urchin embryos including abnormal development at first cell division when exposed to halothane. The authors suggested polyspermy (multiple sperm entry) as the cause of the reduced fertilisation. Volpi Ghirardini et al. (2005) used two sea urchin toxicity bioassays (% fertilisation and embryo toxicity) to discriminate pollution at several field stations in the Lagoon of Venice. When comparing the two bioassays the authors confirmed previous work (Novelli et al., 2003) that the embryo toxicity test was more sensitive than the fertilisation test (referred to as sperm cell toxicity). However, the study did concluded that the combined use of the spermiotoxicity and embryotoxicity tests to be more effective in highlighting

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clear toxicity fingerprints due to the complementarities of the endpoints. These particular studies expose mature sperm to various toxicants for limited periods of time (e.g. 60 min) and assess the ability of the sperm to fertilise. Bearing in mind we currently know that longer term exposure can results in a variety of abnormalities in sperm morphology and ultra-structure (Au et al., 2003) the impacts of pollutants on male fertility over the early development, sperm maturation and the life time of echinoderms are not fully understood and thus maybe impacted within the field. 3.5. Other aquatic invertebrates

Fig. 4. (a) Spermatozoa from sea urchin exposed to 0.01 mg l−1 Cd2+ . Sperm with intact tail (right) and sperm with short, incomplete ‘broken’ tail (left). Bar scale = 5 ␮m. (b) Sperm from sea urchin exposed to 0.1 mg l−1 Cd2+ SEM micro-graph. Sperm with incomplete ‘broken’ tail (arrow-headed), sperm head with convoluted profile (arrowed). Bar scale = 2 ␮m. (c) Spermatozoon of sea urchin exposed to 0.01 mg l−1 Cd2+ . Annular mitochondrion showing swollen and deformed cristae (arrow-headed). (d) Spermatozoon of sea urchin exposed to 0.1 mg l−1 Cd2+ . Severe mitochondrial cristae dis-organization (arrow-headed). Bar scales = 0.5 ␮m. (e) Control spermatozoon. (f) Spermatozoon of urchin exposed to 0.1 mg l−1 Cd2+ . Acrosome vesicle remains intact except the plasma membrane is convoluted (arrowed). (g) Spermatozoon of urchin exposed to 0.1 mg l−1 Cd2+ . Convoluted sperm plasma membrane (arrowed). A, acrosome; M, midpiece; N, nucleus; T, tail. Bar scales = 0.5 ␮m.

Given the considerable current research effort into understanding and identifying anthropogenic impacts on reef building corals and reef habitats there is surprisingly little information on environmental reproductive toxicity for coral species. Corals show a range of reproductive modes from brooding to broadcast spawning with some hermaphroditic species that release buoyant bundles containing both eggs and sperm during precisely timed spawning events. There has been some work on into the impacts of various metals, pesticides and petroleum products on fertilisation success in broadcast spawning corals which have not looked at sperm toxicity impacts directly but use fertilisation success as their reproductive endpoint. Heyward (1988) was the first to study the effects of copper and zinc sulphates on fertilisation rates in the scleractinian corals Favites chinensis and Platygyra ryukyuensis providing the first published data on the effects of trace metals on fertilisation in corals. Subsequently several authors have observed similar dose response declines in fertilisation success with contaminant exposures amongst a range of scleractinian and soft coral species (Markey et al., 2007; Mercurio et al., 2004; Negri and Heyward, 2000, 2001; Reichelt-Brushett and Harrison, 2005). Sperm toxicity effects are likely to be associated with these reductions in fertilisation success but have not been quantified. Disruption of sperm motility under future ocean CO2 conditions has been demonstrated in coral sperm (Morita et al., 2010) but similar studies to look for potential at contaminant induced motility disruption have not been performed. Estrogens and other bio-regulatory molecules are thought to play a role in the control of reproduction in corals, possibly acting through novel or primitive mechanisms common to all cnidarians, raising the possibility of environmentally induced endocrine disruption of reproductive processes in this group. One study to date has examined this, finding that sperm egg bundle production in Montipora capitata was reduced by 29% after exposure to the oestrogen estradiol-17beta in laboratory exposures (Tarrant et al., 2004). The only other aquatic invertebrate phyla for which paternal toxicity data could be found is the Ascidians. The toxicity of mercury, copper, cadmium and chromium on sperm viability and fertilisation in Ciona intestinalis was examined by Bellas et al. (2001) who found no reduction in percentage viability or fertilisation success at the metal concentrations examined. This suggests that perhaps Ascidian spermatozoa are more robust to toxic insult than those of echinoderms or molluscs which showed response at similar metal concentrations. The reproductive ecology of the Bryozoans has been well studied with much known now about sperm morphology, transfer and competition (Manriquez et al., 2001; Pemberton et al., 2003), yet we could find no male toxicology data for this important group of biofouling organisms.

With permission from Au et al. (2001b).

4. Future directions and unanswered questions Despite the recent laboratory evidence that aquatic invertebrate sperm are highly susceptible to disruption from environmental contaminants there are many examples of populations surviving

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and reproducing successfully in highly polluted environments. For example the population of the harbour ragworm Nereis diversicolor living in the highly contaminated upper reaches of Restronguet Creek, Cornwall, must free spawn their sperm onto sediments containing normally fatal levels of copper and cadmium. The severe lack of field collected data on environmental disruption of sperm form and function for aquatic invertebrates means that at present we can only hypothesise as to the nature of the behavioural and biochemical adaptations that enable populations living in contaminated habitats to reproduce successfully. Biochemical adaptations might include up-regulation of anti-oxidant enzymes in the sperm of males living in contaminated habitats; however, our knowledge of the presence or inducible nature of defence enzymes in invertebrate sperm is severely lacking. Behavioural adaptations that might evolve to compensate for reduced sperm function include increased spawning aggregation and highly synchronous spawning in order to maintain sufficiently high sperm concentrations for successful fertilisation. Oocyte diameter is another important factor determining sperm: egg collision rates, and is also under strong selective pressures to maximise maternal fitness (Marshall and Keough, 2008). This may result in selection for increased egg size and/or fecundity in order to maximise fertilisation success when living in an environment where sperm function is reduced. Many of the points discussed here are largely ignored by current ecotoxicological research, yet would provide valuable insight into the mechanisms that allow aquatic invertebrates to survive and reproduce in contaminated and changing habitats, and put us better placed to predict the consequences of pollution and climate change on the reproductive ecology of this key group of aquatic organisms. Clearly, there is a need for further studies in this area, to include both laboratory and field studies from clean and reference sites, with a focus on broadcast spawners and those with direct fertilisation. We suggest the following priority research areas:

1. Field monitoring is required to assess whether male fertility is being impacted by industrial pollutants in natural populations living in contaminated environments. There are currently only a limited number of studies that have assessed sperm quantity (Yang et al., 2008) and none that we could fine that have examined sperm quality from natural populations in the field. 2. Research to assess the relationship between reduced spermatozoa (quantity) and fertilisation success needs to be expanded across many phyla. There are a number of fertilisation dynamics models for broadcast spawners that attempt to quantify this but very little actual field data. 3. Studies indicating DNA damage in spermatozoa need to be further linked to fertilisation success and subsequent development for a wider range of reproductive strategies and up-scaled to assess population level effects. 4. Studies are required to develop and adapt protocols for assessing sperm quality – so far studies have shown that these are relatively easily transferable from existing vertebrate studies. 5. Background levels and specific types of sperm abnormalities need to be quantified in natural populations. 6. Mechanistic studies into the biology of sperm–oocyte interactions are required with specific emphasis of disruption along these axes. 7. Assessments into the impacts of known industrial pollutants on male fertility both broadcast spawners and internal brooders. 8. Research into the reproductive biology and sperm physiology of populations successfully surviving and reproducing in highly contaminated habitats are required to enable any adaptations in sperm physiology or reproductive partitioning to be identified.

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