Advertisement
Research Article

Tough Adults, Frail Babies: An Analysis of Stress Sensitivity across Early Life-History Stages of Widely Introduced Marine Invertebrates

  • M. Carmen Pineda,

    Affiliation: Department of Animal Biology (Invertebrates), University of Barcelona, Barcelona, Spain

    X
  • Christopher D. McQuaid,

    Affiliation: Department of Zoology and Entomology, Rhodes University, Grahamstown, South Africa

    X
  • Xavier Turon,

    Affiliation: Centre for Advanced Studies of Blanes (CEAB-CSIC), Blanes, Spain

    X
  • Susanna López-Legentil,

    Affiliation: Department of Animal Biology (Invertebrates), University of Barcelona, Barcelona, Spain

    X
  • Víctor Ordóñez,

    Affiliation: Department of Genetics, University of Barcelona, Barcelona, Spain

    X
  • Marc Rius mail

    mrius@ucdavis.edu

    Affiliation: Department of Evolution and Ecology, University of California Davis, Davis, California, United States of America

    X
  • Published: October 12, 2012
  • DOI: 10.1371/journal.pone.0046672

Abstract

All ontogenetic stages of a life cycle are exposed to environmental conditions so that population persistence depends on the performance of both adults and offspring. Most studies analysing the influence of abiotic conditions on species performance have focussed on adults, while studies covering early life-history stages remain rare. We investigated the responses of early stages of two widely introduced ascidians, Styela plicata and Microcosmus squamiger, to different abiotic conditions. Stressors mimicked conditions in the habitats where both species can be found in their distributional ranges and responses were related to the selection potential of their populations by analysing their genetic diversity. Four developmental stages (egg fertilisation, larval development, settlement, metamorphosis) were studied after exposure to high temperature (30°C), low salinities (26 and 22‰) and high copper concentrations (25, 50 and 100 µg/L). Although most stressors effectively led to failure of complete development (fertilisation through metamorphosis), fertilisation and larval development were the most sensitive stages. All the studied stressors affected the development of both species, though responses differed with stage and stressor. S. plicata was overall more resistant to copper, and some stages of M. squamiger to low salinities. No relationship was found between parental genetic composition and responses to stressors. We conclude that successful development can be prevented at several life-history stages, and therefore, it is essential to consider multiple stages when assessing species' abilities to tolerate stress. Moreover, we found that early development of these species cannot be completed under conditions prevailing where adults live. These populations must therefore recruit from elsewhere or reproduce during temporal windows of more benign conditions. Alternatively, novel strategies or behaviours that increase overall reproductive success might be responsible for ensuring population survival.

Introduction

Abiotic factors such as temperature, salinity and habitat characteristics have long been considered primary factors affecting survival, fitness and distribution of marine organisms [1]. More recently, anthropogenic changes to the environment have yielded new agents of selection, with resistance to pollution being one of the most important [2], [3]. Thus, the persistence of human-mediated stressors in the environment nowadays contribute to shaping the distribution of marine organisms, excluding some (e.g. [4]) and facilitating the establishment of others (e.g. [5]). Moreover, a species' long-term performance is modulated by abiotic factors across multiple life-history stages, including adulthood [6][9], and embryonic and larval development (e.g. [10][12]). Among these, embryogenesis, settlement and metamorphosis are critical life-history phases for many organisms (e.g. [13], [14]), especially when exposed to anthropogenic stressors [15][17]. For sessile marine organisms, where adults are unable to escape unfavourable abiotic conditions, the importance of successful early stages is even more striking as it determines the viability of local adult populations [18][20]. This in turn can have community-level consequences as many sessile species act as ecosystem engineers, sensu [21], providing habitat for multiple associated organisms while excluding competitors for space.

The arrival and establishment of non-indigenous species (NIS) via man-mediated transport is a major factor altering communities worldwide (e.g. [22], [23]). Shipping facilities such as harbours and marinas often act as entrance gates for NIS [24][28], and thus newcomers have to be able to cope with the stressful conditions (e.g. pollution, disturbance) that characterize these altered habitats. Establishment of NIS in such environments depends on physical and biological conditions being suitable not only for adults [29][31] but also for juvenile stages (e.g. [32]).

Genetic diversity is an important factor influencing the establishment of NIS [33][36] and it is generally assumed that the richer the genetic composition of a species' population, the wider its potential ability to adapt to stressful environmental situations [35], [37]. The heritability of traits under selection depends on stress-response variation within a population, and the potential for rapid evolution in new environments [17], [38]. For NIS, the latter can be problematic as introduced species often experience genetic bottlenecks that can reduce the genetic diversity needed for selection [35], [39], [40]. The study of genetic variability of introduced populations is essential to understanding NIS tolerance of environmental stresses and their potential to spread. To date, however, few studies have considered how different levels of parental genetic diversity in NIS influence offspring responses to multiple stressors.

Genotype-environment interactions are generally considered when differences in response between genotypes are not consistent from one environment to another, and have been investigated to assess, for instance, phenotypic stability [41] or genotypic responses to lethal and non-lethal stresses [42]. Most studies on genotype-environment interaction have analysed the influence of abiotic conditions during adulthood (e.g. [43]), while studies covering different, presumably more sensitive, early life-history stages remain rare. In line with this, genetic markers can be used to characterize different populations and to relate differences in biological response to genetic diversity and differentiation between and within populations.

Here we investigated the performance across multiple life-history stages of two widely introduced marine invertebrate species in locations where both species coexist. The solitary ascidians Styela plicata (Lesueur, 1823) and Microcosmus squamiger (Michaelsen, 1927) are sessile organisms that have been introduced worldwide [44], [45] and that often inhabit places with highly variable abiotic conditions [7], [46][48]. The success of introductions of S. plicata to new habitats has been linked to its high tolerance of polluted waters and changes in temperature and salinity [7], [10], [49], while M. squamiger is known to be resistant to low salinities as adults [46]. In addition, previous genetic studies of these widespread species based on a fragment of the mitochondrial gene Cytochrome Oxidase subunit I (COI) have revealed the existence of two highly divergent and widely distributed haplogroups for each species [44], [45]. No information is available, however, on the functional significance of this intraspecific genetic structure in terms of responses to stress. In this study, we targeted several early life-history stages (fertilization, larval development, settlement, and metamorphosis) and we genetically characterized the progenitors. We tested species performance under thermal and salinity stress, and with several concentrations of a heavy metal (Cu). We hypothesised that S. plicata and M. squamiger offspring would develop well under realistic environmental conditions found in sheltered habitats where adults occur, although different haplogroups might respond to stress differently.

Methods

Field sites and general methods

Adult individuals of Styela plicata and Microcosmus squamiger were collected during the austral spring of 2010 (October and November) when both species are known to reproduce [50][52]. Two sites along the South African coast, approximately 160 km apart, were sampled: Port Elizabeth (33°57′44″S, 25°38′8″E) and Knysna (34°02′32″S, 23°02′40″E). The selection of the sampling sites was based on the co-occurrence and abundance of the studied species and the availability of genetic data from previous studies [44], [45]. The sites feature slightly different characteristics in terms of abiotic parameters and ship traffic. Port Elizabeth harbour is an industrial port, with temperature oscillating around 16–18°C in winter and 22–22°C in summer, although there can be rapid fluctuations of up to about 10°C (range c. 14–24°C) due to periodic influx of warm water from the Agulhas Current and cooler upwelled water from nearby Cape Recife [53]. There is little freshwater input into the bay as a whole and salinities are consistently close to oceanic norms of 35.2‰ [54]. Copper concentrations in this harbour are between 0.5 µg/L and 11.3 µg/L [55]. In contrast, Knysna lagoon supports only recreational boating and can have a partially estuarine regime, with alternation of freshwater influx and tidal input. Salinities in the lagoon are generally similar to the open ocean [56], although freshwater conditions can reign in the area during unusually wet periods, displacing non-euryhaline organisms (M.R., pers. obs.). Temperatures generally oscillate around 15°C in winter and 22–24°C in summer. Copper levels in the waters of Knysna Lagoon range between <0.1 to 4.7 µg/L [57], [58].

After collection, individuals were transported to the laboratory (located less than 6 h away) in insulated containers containing water from the collection site. Since seawater from harbours usually contains high concentrations of pollutants [55], [57], [59], [60], we collected seawater from nearby clean sites far from any urban or industrial influence (33°58′47″S, 25°39′29″E for Port Elizabeth, 34°03′42″S, 23°22′38″E for Knysna). Animal storage and all laboratory experiments were conducted using this seawater, which we previously filtered using a vacuum filtration unit with 10 µm pore filters. Individuals were kept in the laboratory at constant temperature (20°C) and water aeration, for a minimum of 12 h and a maximum of four days for acclimatization. During the storage period in aquaria, temperature and salinity were monitored and modified to maintain the desired conditions. We used constant artificial illumination to prevent light-induced spawning [61]. All the samples were obtained according to current South African regulations. This species are not protected by any law and all sampling was conducted outside protected areas.

Experimental trials

We chose an array of abiotic factors (temperature, salinity and pollution) that are known to influence survival of marine invertebrates [10], [46], and analysed four early life-history processes: fertilisation, development of the larvae, settlement and metamorphosis. Temperatures were set to either 20°C (control) or 30°C (treatment) in a Constant Environment (CE) room. Seawater temperature of 30°C represents the higher values occasionally reached in summer within the distributional area of both species [62]. Distilled water was added to seawater to achieve reduced salinity values (26‰ and 22‰) similar to those that are known to affect ascidian development and survival and can be found in estuaries [48], [63]. For the pollution treatments, we used copper because it is known to be one of the most toxic heavy metals for marine invertebrates [64], especially during early life-history stages [65][67]. We added liquid copper (Spectrosol® ref.14139 1000 ppm copper standard solution) to filtered seawater to attain the desired concentrations: 25 µg/L, mean concentration in a polluted harbour [2]; 50 µg/L, common in highly polluted harbours or near boats recently painted with antifouling paint [68] and 100 µg/L, an extreme copper concentration often used in this type of studies [69].

Gamete extraction, fertilisation and experimentation

Gametes were extracted by dissecting the ripe gonads as described in Svane and Young [70] and Marshall et al. [71]. A mix of eggs and sperm was poured through a 100-µm filter with seawater into a small beaker to retain the eggs in the filter and gather the sperm and seawater in the beaker. For each fertilisation attempt (see Table 1 for details), around 10 individuals were dissected: 5 individuals for eggs and 5 for sperm (both species are simultaneous hermaphrodites). The oocytes obtained from the 5 female donors (around 12 to 18 ml per individual, ~500 eggs ml−1) were subsequently pooled together, and the same was done with the sperm obtained from the 5 male donors (~107 sperm ml−1).

thumbnail

Table 1. Artificial fertilisation runs for each species and sampled site.

doi:10.1371/journal.pone.0046672.t001

For the fertilization and larval development assays, 6 ml of the oocyte suspension, 12 ml of the corresponding treatment solution (filtered seawater for the temperature treatment, other treatments adjusted to obtain the desired final concentrations after mixing with gametes), and 2 ml of concentrated sperm mix were added to a 65 mm Petri dish. The cultures were then immediately taken to the appropriate CE room for fertilisation. After 1 hour, the eggs were washed with the treatment solution to remove excess sperm using a 100-µm filter and then distributed among five Petri dishes (~100–500 eggs per dish) containing 12 ml of the treatment solution at the appropriate concentrations, and closed with a lid to avoid evaporation during the experimental period. This first set of cultures was used to assess fertilisation and development rates.

To obtain enough larvae to conduct the settlement and metamorphosis assays, new individuals were obtained from each species (Table 1) and fertilized in an aerated beaker containing 500 ml filtered seawater and maintained in a CE room at 20°C to maximize development rates [72]. Post-hatching experiments consisted of 40 larvae carefully pipetted out and placed in a Petri dish with 12 ml of the corresponding treatment solution (5 replicates per treatment and location). Petri dishes were previously submerged in seawater for 24 h to develop a biofilm in order to facilitate larval settlement [73], [74]. All Petri dishes were then placed in CE rooms (30°C for the temperature treatment and 20°C for the rest of experimental conditions) and kept for 4 days.

Data collection and analyses

For both species, most of the larvae hatched within 14 hours of fertilisation at 20°C. Numbers of viable larvae, larvae with deformities (or immature larvae), undeveloped embryos and unfertilized eggs (Fig. 1) were then recorded using a stereomicroscope. Likewise, the numbers of settled, completely metamorphosed and unattached larvae were assessed every 24 h over 4 days (96 h) in the settlement and metamorphosis assays. The fertilisation rate (FR), development rate (DR), settlement rate (SR) and metamorphosis rate (MR) were calculated as follows:



We analyzed two types of variables, the proportion of success at each developmental stage (i.e. fertilization rate, development rate, settlement rate, and metamorphosis rate) for controls and treatments, and the relative success ratios (RS) obtained by dividing the value of each rate by the mean of the corresponding controls. The former was used to assess treatment effects against the controls. For site effects, as differences between sampled sites often occurred even in the controls, the RS were an appropriate assessment of the effect of interest (i.e., whether development was impaired differentially in one site with respect to the other, after eliminating the effect of differences in controls).

thumbnail

Figure 1. Picture of eggs and larvae under stereomicroscope.

a) S. plicata and b) M. squamiger (ue: unfertilized egg; um: undeveloped embryo; ul: unviable larvae; vl: viable larvae).

doi:10.1371/journal.pone.0046672.g001

For both types of variables, we performed separately two-way analyses of variance (ANOVA) per species with site and treatment as fixed factors. We used a logit transformation of the FR, DR, SR and MR data as it is known to stabilize the variances of proportional data better than other commonly used methods [75]. Our transformed data had homogeneity of variances in all datasets, although normality was only accomplished in a few cases. Nonetheless, we performed the ANOVA tests as they are robust to departures from normality when variances are homogeneous [76]. For the relative success rates (RS), the data complied in all cases with the homoscedasticity assumption, although they weren't normally distributed in some cases. As several transformations tried did not improve this, we proceeded with the raw data in the analyses.

For the proportion data, used to assess treatment effects, if the interaction between factors was significant, post-hoc analyses of treatments were performed at each site against the control with Dunnett's test. If the interaction was not significant, post-hoc tests on treatment levels were done combining both sites. For the RS variables, used to determine site differences, when interaction was significant, site effects were assessed within each level of treatment (using a post-hoc Student-Newman-Keuls test). If interaction was not significant, no test was necessary as site had only two levels. In all post-hoc analyses, the residual mean square obtained from the original two-way ANOVAs was used to calculate the standard errors of the means for the post-hoc comparisons [76], [77]. Statistical analyses were performed using the software STATISTICA v. 6.1 (©StatSoft, Inc. 1984–2004).

In order to obtain an overall estimate of success (from egg fertilisation to post-metamorphic formation), we also calculated the cumulative % success of the different stages, for each of the treatments. For this purpose, each of the different rates (FR, DR, SR, MR) was multiplied by the mean of the previous stage.

Screening of parental genotypes

A piece of muscular tissue from the mantle or the siphon of each individual used for fertilisation was dissected and immediately preserved in absolute ethanol (Table 2). After a few hours, the tainted ethanol was replaced by new absolute ethanol and samples were then stored at −20°C until extracted. Total DNA was extracted using the REDExtract-N-Amp Tissue PCR Kit (Sigma-Aldrich). The universal primers LCO1490 and HCO2198 described in Folmer et al. [78] were used to amplify a fragment of the COI gene (maternally inherited). Amplifications were performed in a final volume of 20 µL using 10 µL of REDExtract-N-amp PCR reaction mix (Sigma-Aldrich), 0.8 µL of each primer (10 µM), and 2 µL of template DNA. The PCR program consisted of an initial denaturing step at 94°C for 2 min, 30 amplification cycles (denaturing at 94°C for 45 seconds, annealing at 50°C for 45 seconds and extension at 72°C for 50 seconds), and a final extension at 72°C for 6 min, on a PCR System 9700 (Applied Biosystems). PCR products were sent for purification and sequencing to Macrogen Inc. (Seoul, Korea). Sequences were edited and aligned using BioEdit® v.7.0.5.3 [79]. Number of alleles (Nh), gene diversity (Hd), and nucleotide diversity (π) were computed with DnaSP v.5 [80]. Pairwise genetic distances (FST) using allele frequencies were calculated with Arlequin v.3.1 [81] and their significance was assessed by performing 10,000 permutations. Note that because each fertilization attempt involved a combination of gametes from ten different donors, offspring resulted from a random combination of these genotypes.

thumbnail

Table 2. Diversity measures and population differentiation values (FST) for the mtDNA sequences (COI gene).

doi:10.1371/journal.pone.0046672.t002

Results

Experimental trials

The results of the ANOVA on the different logit-transformed rates are presented in Tables 3 and 4. The results of post-hoc site comparisons using relative success rates are presented in Fig. 2, where these rates are depicted. All abiotic conditions analyzed: temperature at 30°C, salinity values of 22‰ (22S) and 26‰ (26S), and copper at a concentration of 25 µg/L (Cu25), 50 µg/L (Cu50) and 100 µg/L (Cu100), produced important effects on the relative success ratio of each developmental stage considered, with differences due to both species and the site of adult collection. There was no consistent trend of one of the sites having higher or lower success rates, although many outcomes differed significantly between sites (Tables 3,4; Fig. 2).

thumbnail

Figure 2. Relative success ratios of each developmental stage.

a) S. plicata and b) M. squamiger. Values are relative to the controls (i.e., values<1 indicate less success than in the corresponding controls). Treatments include: temperature (30°C), salinity (circles for 26‰ salinity, triangles for 22‰) and copper (circles for copper concentration of 25 µg/L, triangles for 50 µg/L and squares for 100 µg/L). Black symbols correspond to the population at Port Elizabeth, while white symbols correspond to the population at Knysna. Values are means ±standard errors. Asterisks indicate significant differences between locations.

doi:10.1371/journal.pone.0046672.g002
thumbnail

Table 3. ANOVA examining the effects of site and treatment at four developmental stages for S. plicata.

doi:10.1371/journal.pone.0046672.t003
thumbnail

Table 4. ANOVA examining the effects of site and treatment at four developmental stages for M. squamiger.

doi:10.1371/journal.pone.0046672.t004
S. plicata .

There were significant interactions of treatment and site for all dependent variables. S. plicata showed significantly reduced fertilisation rates (FR) in most treatments (Table 3). Knysna gametes seemed somewhat less affected by the treatments than Port Elizabeth (Fig. 2a). All treatments had significant effects in Port Elizabeth, while gametes from Knysna were unaffected by temperature and Cu25 (Table 3). Significant site differences were found for 26S and Cu25, where fertilization relative to controls was significantly higher in Knysna (Fig. 2a).

The development of viable larvae (DR) was probably the most sensitive stage in the early development of this species (Fig. 2a) and was significantly impaired by all treatments, except for Cu25 and Cu50 (Table 3). Notably, the presence of Cu25 increased DR relative to the controls (thus relative success rates were above 1), although the effect is significant only in Knysna. Significant inter-site differences in relative success rates were found only for Cu50, with embryos from Port Elizabeth being more resistant.

Settlement rate (SR) tended to show higher relative success values (Fig. 2a) than the previous variables, indicating that this stage is somewhat more tolerant. All treatments except 26S yielded significantly low values for Port Elizabeth larval settlement, while no significant effect was detected for Knysna (Table 3). Relative success values in Port Elizabeth were significantly lower for 22S, and significantly higher for Cu50. Although the effect of salinity (26S) on SR was not significant, low salinities did appear to accelerate settlement within 24 hours (Fig. 3a). On the other hand, Cu100 seemed to accelerate settlement for larvae from Port Elizabeth adults but to delay it for Knysna (Fig. 3a).

thumbnail

Figure 3. Percentage of total settled (above) and metamorphosed (below) individuals each 24-hour interval.

a) S. plicata and b) M. squamiger. Left bars for Port Elizabeth (PE); right bars for Knysna (KN). Asterisks indicate zero success.

doi:10.1371/journal.pone.0046672.g003

As for settlement, metamorphosis in S. plicata (MR) was also a relatively tolerant process under most treatments and for both sites. The strongest inhibition effect on MR occurred at 22S and Cu100 for both sites (Fig. 2a), and these treatments yielded significantly lower metamorphosis than the controls at both sites (Table 3). In addition, the metamorphosis of Knysna larvae was also impaired at 26S, Cu25 and Cu50 (Table 3). Site differences were significant in the three copper treatments, with relative success rates higher in Port Elizabeth. Increased temperature accelerated the metamorphosis of the settled individuals within 72 hours, although low salinities had the opposite effect, causing a delay in metamorphosis (Fig. 3a). Most of the larvae from the 22S and Cu100 treatments never achieved complete metamorphosis within 96 hours, and none did so in Port Elizabeth at Cu100 concentration (Fig. 3a).

M. squamiger .

All treatments except Cu25 significantly reduced the fertilisation rates (Fig. 2b, Table 4) at both sites combined (no significant interaction term), and the most drastic reduction was observed after exposure to 26S, 22S, Cu50 and Cu100 (Fig. 2b). For the relative success rates (RS), the interaction term was not significant, and there was an overall effect of site, with mean success rates higher in Port Elizabeth.

As for S. plicata, larval development was the most sensitive stage (Fig. 2b). The interaction was not significant and, combining sites, all treatments significantly reduced DR, especially high temperature, 26S, 22S, Cu50 and Cu100 (Table 4, Fig. 2b). When analysing relative success rates, the interaction proved significant, and this was due to the outcome of the Cu25 treatment, being significantly higher in Knysna.

Settlement was also less affected by temperature and salinity treatments than the previous processes (Fig. 2b). The three copper concentrations resulted in significantly lower SR than the controls in Knysna, while only Cu50 and Cu100 reduced settlement of larvae from Port Elizabeth (Table 4). High temperatures and low salinities increased the number of settlers relative to the controls (values above 1, Fig. 2b), with a significant positive effect for Knysna larvae kept at 22S and Port Elizabeth larvae at 26S (Table 4). Significant site differences in relative success rates were found for temperature and 22S (higher rates in Knysna), and Cu25 (higher rates in Port Elizabeth, Fig. 2b). Moreover, Settlement was accelerated at higher temperature (Fig. 3b), while Cu50 and Cu100 delayed settlement of larvae from Knysna but not from Port Elizabeth (Fig. 3b).

All treatments except 26S significantly decreased the MR from Port Elizabeth larvae, while only 22S, Cu50 and Cu100 impaired metamorphosis of Knysna larvae (Table 4). On the other hand, more larvae metamorphosed at high temperature and Cu25 than in the controls in Knysna (leading to relative rates higher than one, Fig. 2b), although this outcome was not significant. The relative success rates were significantly higher in Knysna for temperature and for Cu25, and in Port Elizabeth for 26S (Fig. 2b). Cu25 also accelerated the timing of metamorphosis at Knysna (Fig. 3b). No metamorphosis was observed for larvae from Port Elizabeth subjected to the 22S treatment, larvae from Knysna at Cu50, or larvae from either sampled site at Cu100 (Fig. 3b).

S. plicata and M. squamiger comparison.

When the whole developmental sequence was considered, from fertilisation of the egg to post-metamorphic juveniles, clear differences in cumulative success were found between the species, with S. plicata being overall more tolerant of harsh conditions than M. squamiger (Fig. 4). As previously stated, the development of larvae seems to be the most sensitive stage for both species, acting as a bottleneck that results in a sharp reduction in the number of viable larvae in most treatments (Fig. 4).

thumbnail

Figure 4. Cumulative success over fertilisation (F), development (D), settlement (S), and metamorphosis (M) for each treatment.

a) S. plicata and b) M. squamiger. Legend: T: temperature at 30°C; 22S: 22‰ salinity; 26S: 26‰ salinity, Cu25: copper concentration of 25 µg/L; Cu50: 50 µg/L; and Cu100: 100 µg/L.

doi:10.1371/journal.pone.0046672.g004

It is particularly relevant that the complete process of reproduction and recruitment only occurred in non-negligible numbers in the controls and the treatments with the lower copper concentrations assayed (Cu25, Cu50) in S. plicata (Fig. 4a), and only for the controls in the case of M. squamiger (Fig. 4b). In all other treatments, failure of one step or another (particularly development of larvae) prevented successful completion of the early life-history stages completely or almost so.

Genetic screening

All adults used for the fertilization experiments were sequenced (Tables 1, 2), except for two individuals of M. squamiger that failed to amplify. Three haplotypes were obtained for S. plicata, corresponding to haplotypes already described by Pineda et al. [45]. For M. squamiger, we found ten haplotypes. Six of these had previously been reported [44], while the sequences of the remaining four haplotypes (Hap 53–56) were new and were deposited in GenBank with accession numbers JQ815436-JQ815439 (Table 2). S. plicata showed two clear groups of haplotypes, with Knysna composed entirely of Lineage I (50% Hap 1 and 50% Hap 2) and Port Elizabeth mainly represented by Lineage II (95% Hap 5, 5% Hap 2), sensu [45]. Thus, although these three haplotypes are globally distributed [45], Port Elizabeth and Knysna were highly differentiated (FST = 0.728, P<0.001) (Table 2). Regarding M. squamiger, the two most frequent haplotypes were Haps 1 and 5 (Table 2) for both populations, and together represented ca. 60% of the genetic pool. Haplotypes corresponding to Lineage II, sensu [44], represented around 90% of each population, and the two populations did not differ significantly (FST = 0.048, P = 0.991) (Table 2).

Discussion

Increased temperature, decreased salinity and elevated copper concentrations affected several life-history stages of the introduced ascidians Styela plicata and Microcosmus squamiger at the two studied sites. Differences according to sensitivity to abiotic stressors and life-history stages were observed but overall, fertilisation and larval development were the most sensitive stages for both species. Thus, although later stages (settlement and metamorphosis) seemed in general more tolerant, the initial stages (fertilisation and development) must necessarily happen under more benign conditions.

A few of the stressors had apparent positive effects on some stages (resulting in the corresponding rates being greater than in the controls, or accelerating processes). It has been reported that moderate concentrations of pollutants can enhance some early life-history stages of marine invertebrates but eventually lead to detrimental effects (e.g. [82], [83]). Similarly, our combined rates show that, notwithstanding these rare positive effects, the overall effect through the developmental stages considered is negative in all cases. Therefore, considering a single stage independently can lead to misleading conclusions about the ability of a species to overcome stressful conditions during the early life-history stages.

In general, S. plicata was more resistant to copper pollution, and both species coped similarly with increased temperature. Decreased salinity prevented complete development in both cases; however, some stages of M. squamiger (e.g. fertilization, settlement) are less affected or enhanced by low salinities. These tolerances correlate well with the types of environments where these species are commonly found. S. plicata is often found in harbours, which are known to accumulate copper [15], [45], and M. squamiger in estuaries, which are characterized by frequent salinity changes [84]. In fact, Lowe [46] found that adults of M. squamiger could withstand reduced salinity levels (15–25‰) for extended periods of time, outcompeting native species such as Molgula manhattensis in southern California harbours. Similarly, estuarine sites along the southeast coast of South Africa (e.g. Port Alfred, Bushman's River Mouth and East London) are dominated by M. squamiger while S. plicata is consistently absent in estuarine conditions but found in nearby harbours (M.R., pers. obs.).

Sensitivity differences according to development stages and stressors have been observed across phyla for other marine invertebrates, including molluscs [1], [14], echinoderms [85] and ascidians [86][88]. Our results indicate that complete development, from fertilisation to metamorphosis, is impaired by all studied stressors, affecting several early life-history stages. In fact, we recorded completion of early stages only in S. plicata if copper concentrations were at/below 50 µg/L. Thus, the wide distribution of these species in environments where high temperature, low salinity or extreme pollutant concentrations are present cannot be inferred from laboratory or manipulative studies, but must be explained by novel strategies or behaviours in nature that increase overall reproductive success [89]. In this sense, Bellas et al. [65] suggested that the ascidian Ciona intestinalis could probably detect trace metals in the water with the adhesive papillae and delay or inhibit attachment. Although increasing the swimming period may decrease the probability of post-settlement survival due to the high metabolic cost required for the latter [10], [90][92], the successful settlement and survival of a few individuals could result in successful introductions to new habitats. Even if recruitment failures were a common outcome, the prolonged reproductive period observed for both species [50][52] would increase the chances of a propagule finding favourable temporal windows of tolerable conditions.

The sensitivity of S. plicata embryos and larvae to temperature and salinity changes was in accordance with Thiyagarajan and Qian [10], who studied S. plicata in Hong Kong and reported recruitment failure when seawater temperature reached values of 26–30°C and salinities of 22–30‰ in summer. In our study, these conditions prevented both S. plicata and M. squamiger from completing development, with the earlier stages (embryo fertilisation and larval development) being especially sensitive, while settlement was hardly affected. The lowest salinity tested (22‰), however, prevented most larvae of either species to complete metamorphosis even after successful settlement, as previously described for other ascidians [63], [93]. High temperatures and low salinities also tended to accelerate development, with most larvae of both species settling within 24 h, which would limit options for escape to more favourable sites. Thus, the current climate change predictions of increasing temperatures and decreasing salinities [94] suggest that these species, particularly S. plicata [48], would find it difficult to cope with these predicted stresses [95].

Copper has been shown to inhibit embryo development, reduce successful settlement and metamorphosis, and reduce growth in many marine invertebrates, including ascidians (e.g. [17], [96][98]). Elevated copper concentrations also negatively affected the early life-history stages of S. plicata and M. squamiger, with more dramatic effects on developmental stages of the latter species. Even at copper concentrations similar to those found in highly polluted harbours (25–50 µg/L), fertilisation success of S. plicata was still around 50% that of the controls for both populations, and development through metamorphosis was possible. At the highest concentration (100 µg/L), though, there was no development of the larvae and the metamorphosis of settled individuals was seriously impaired. In contrast, even the lowest concentration of copper assayed (25 µg/L) had detrimental effects on early development of M. squamiger. This suggests that S. plicata will continue to perform better in polluted habitats than M. squamiger and has important implications for understanding the distributions of the two species across overlapping ranges.

The genetic patterns found were clear-cut: genetic differentiation was high between populations of S. plicata, while it was negligible for M. squamiger. However, we could not detect a clear correlation of this pattern with differential responses to abiotic stress. In general, although some particular outcomes were significantly different, all populations responded similarly to the tested stressors. Genetic diversity within populations was lower for S. plicata than for M. squamiger, but again this has no clear connection with our results as, if any, the low genetic diversity species S. plicata was overall more tolerant to stress than M. squamiger. The only emerging pattern was found when comparing the responses to low salinity and high copper concentrations between populations of S. plicata. For instance, fertilization rates at low salinities (26‰) were considerably higher for the eggs from Knysna than for the eggs from Port Elizabeth. Adult samples from Knysna exclusively displayed haplotypes from Lineage I, which is the most widespread haplogroup in the world [45]. In contrast, adults from Port Elizabeth mainly belonged to Lineage II, which is also found in salt marsh habitats [48]. Thus, the slightly different response of these two populations of S. plicata may be related to differences in their genetic composition. Differential adaptation to environmental factors (e.g. temperature, salinity) of mitochondrial sequences within one species has been previously described in marine invertebrates [99][103]. Of course this adaptive capability need not be directly linked to the studied gene, but can be related to other genes that vary between lineages. In order to assess whether there is any genetic basis in the responses featured by both species, a more precise genetic characterization (for example, using microsatellites), together with controlled crossings and transplant experiments are necessary.

In conclusion, we found that several early life-history stages of the ascidians S. plicata and M. squamiger were seriously impaired by exposure to realistic scenarios of abiotic stressors, independent of the haplogroup tested. Moreover, abiotic factors do not affect animals in isolation but will normally combine as multiple stressors, often resulting in additive or synergistic effects. Thus, our results are likely to overestimate the resilience of the life-history processes studied here, a surprising fact given the abundance of these species in habitats such as harbours where such stressors are the norm. Behavioural strategies that can only be observed in the field (e.g. delay in spawning until suitable conditions are restored, strong propagule pressure with arrival of larvae from more benign environments, extended reproductive periods) seem plausible explanations for the presence of adults in these localities. Basic knowledge of reproduction, larval development and survival of these species in new habitats coupled with further information on their genetic variability is therefore essential to predict possible areas of establishment and spread worldwide.

Acknowledgments

The authors would like to thank all the staff at the Department of Zoology and Entomology at Rhodes University and especially, Tracy Lindsay, for their priceless assistance with the experimental settings and administrative paperwork. We are in debt to Dr. M. Martínez Azorín, A. Martínez Soler, Dr. M. Noyon, P. Smallhorn-West and E. Arias Galán for valuable assistance in the field and laboratory. Finally, we also thank Gretchen and Charles Lambert for their useful advice on artificial fertilisation in ascidians.

Author Contributions

Conceived and designed the experiments: MCP CDM XT SL-L MR. Performed the experiments: MCP VO MR. Analyzed the data: MCP XT SL-L MR. Contributed reagents/materials/analysis tools: CDM XT SL-L MR. Wrote the paper: MCP CDM XT SL-L VO MR.

References

  1. 1. Kinne O (1964) The effects of temperature and salinity on marine and brackish water animals. II. Salinity and temperature salinity combinations. Oceanogr Mar Biol 2: 281–339.
  2. 2. Hall LW, Scott MC, Killen WD (1998) Ecological risk assessment of copper and cadmium in surface waters of Chesapeake Bay watershed. Environ Toxicol Chem 17: 1172–1189. doi: 10.1002/etc.5620170626
  3. 3. Johnston EL, Keough MJ (2005) Reduction of pollution impacts through the control of toxicant release rate must be site- and season-specific. J Exp Mar Ecol 320: 9–33. doi: 10.1016/j.jembe.2004.12.024
  4. 4. Osovitz CJ, Hofmann GE (2005) Thermal history-dependent expression of the hsp70 gene in purple sea urchins: Biogeographic patterns and the effect of temperature acclimation. J Exp Mar Ecol 327: 134–143. doi: 10.1016/j.jembe.2005.06.011
  5. 5. Piola RF, Johnston EL (2006) Differential tolerance to metals among populations of the introduced bryozoan Bugula neritina. Mar Biol 148: 997–1010. doi: 10.1007/s00227-005-0156-5
  6. 6. Dunson WA, Travis J (1991) The role of abiotic factors in community organization. Am Nat 138: 1067–1091. doi: 10.1086/285270
  7. 7. Naranjo SA, Carballo JL, García-Gómez JC (1996) Effects of environmental stress on ascidian populations in Algeciras Bay (Southern Spain) Possible marine bioindicators? Mar Ecol-Prog Ser 144: 119–131. doi: 10.3354/meps144119
  8. 8. Byers JE (2002) Impact of non-indigenous species on natives enhanced by anthropogenic alteration of selection regimes. Oikos 97: 449–458. doi: 10.1034/j.1600-0706.2002.970316.x
  9. 9. Addison PF, Knott NA, Keough MJ (2008) Spatially variable effects of copper on sessile invertebrates across a marina. J Exp Mar Ecol 364: 19–23. doi: 10.1016/j.jembe.2008.06.020
  10. 10. Thiyagarajan V, Qian PY (2003) Effect of temperature, salinity and delayed attachment on development of the solitary ascidian Styela plicata (Lesueur). J Exp Mar Ecol 290: 133–146. doi: 10.1016/s0022-0981(03)00071-6
  11. 11. Przeslawski R, Davis AR, Benkendorff K (2005) Synergistic effects associated with climate change and the development of rocky shore molluscs. Global Change Biol 11: 515–522. doi: 10.1111/j.1365-2486.2005.00918.x
  12. 12. Marshall DJ, Cook CN, Emlet RB (2006) Offspring size effects mediate competitive interactions in a colonial marine invertebrate. Ecology 86: 214–225. doi: 10.1890/05-0350
  13. 13. Bayne BL, Thompson RJ, Widdows J (1976) Physiology I. In: Bayne BL (ed) Marine Mussels: Their Ecology and Physiology: Cambridge Scientific Press, UK, pp 121–206
  14. 14. Verween A, Vincx M, Degraer S (2007) The effect of temperature and salinity on the survival of Mytilopsis leucophaeata larvae (Mollusca, Bivalvia): The search for environmental limits. J Exp Mar Ecol 348: 111–120. doi: 10.1016/j.jembe.2007.04.011
  15. 15. Galletly BC, Blows MW, Marshall DJ (2007) Genetic mechanisms of pollution resistance in a marine invertebrate. Ecol Appl 17: 2290–2297. doi: 10.1890/06-2079.1
  16. 16. Polato NR, Voolstra CR, Schnetzer J, DeSalvo MK, Randall CJ, et al. (2010) Location-specific responses to thermal stress in larvae of the reef-building coral Montastraea faveolata. PLoS One 5: e11221. doi: 10.1371/journal.pone.0011221
  17. 17. McKenzie L, Brooks R, Johnston E (2011) Heritable pollution tolerance in a marine invader. Environ Res 111: 926–932. doi: 10.1016/j.envres.2010.12.007
  18. 18. Giangrande A, Geraci S, Belmonte G (1994) Life-cycle and life-history diversity in marine invertebrates and the implications in community dynamics. Oceanogr Mar Biol 32: 305–333.
  19. 19. Berkelmans R (2002) Time-integrated thermal bleaching thresholds of reefs and their variation on the Great Barrier Reef. Mar Ecol-Prog Ser 229: 73–82. doi: 10.3354/meps229073
  20. 20. Linares JC, Camarero JJ, Carreira JA (2010) Competition modulates the adaptation capacity of forests to climatic stress: insights from recent growth decline and death in relict stands of the Mediterranean fir Abies pinsapo. J Ecol 98: 592–603. doi: 10.1111/j.1365-2745.2010.01645.x
  21. 21. Jones CG, Lawton JH, Shachak M (1994) Organisms as ecosystem engineers. Oikos 69: 373–386. doi: 10.2307/3545850
  22. 22. Harris LG, Tyrrell MC (2001) Changing community states in the Gulf of Maine: Synergism between invaders, overfishing and climate change. Biol Invasions 3: 9–21.
  23. 23. Strayer DL, Eviner VT, Jeschke JM, Pace ML (2006) Understanding the long-term effects of species invasions. Trends Ecol Evol 21: 645–651. doi: 10.1016/j.tree.2006.07.007
  24. 24. Zibrowius H (1991) Ongoing modification of the Mediterranean marine fauna and flora by the establishment of exotic species. Mésogée 51: 83–107.
  25. 25. Glasby TM, Connell SD, Holloway MG, Hewitt CL (2007) Nonindigenous biota on artificial structures: could habitat creation facilitate biological invasions? Mar Biol 151: 887–895. doi: 10.1007/s00227-006-0552-5
  26. 26. Tyrrell MC, Byers JE (2007) Do artificial substrates favor nonindigenous fouling species over native species? J Exp Mar Ecol 342: 54–60. doi: 10.1016/j.jembe.2006.10.014
  27. 27. Dafforn KA, Johnston EL, Glasby TM (2009) Shallow moving structures promote marine invader dominance. Biofouling 25: 277–287. doi: 10.1080/08927010802710618
  28. 28. Bulleri F, Chapman MG (2010) The introduction of coastal infrastructure as a driver of change in marine environments. J Appl Ecol 47: 26–35. doi: 10.1111/j.1365-2664.2009.01751.x
  29. 29. Stachowicz JJ, Fried H, Osman RW, Whitlatch RB (2002) Biodiversity, invasion resistance, and marine ecosystem function: Reconciling pattern and process. Ecology 83: 2575–2590. doi: 10.1890/0012-9658(2002)083[2575:birame]2.0.co;2
  30. 30. Blackburn TM, Duncan RP (2001) Determinants of establishment success in introduced birds. Nature 414: 195–197. doi: 10.1038/35102557
  31. 31. Zerebecki RA, Sorte CJ (2011) Temperature tolerance and stress proteins as mechanisms of invasive species success. PLoS One 6: e14806. doi: 10.1371/journal.pone.0014806
  32. 32. Fowler AE, Gerner NV, Sewell MA (2011) Temperature and salinity tolerances of Stage 1 zoeae predict possible range expansion of an introduced portunid crab, Charybdis japonica, in New Zealand. Biol Invasions 13: 691–699. doi: 10.1007/s10530-010-9860-2
  33. 33. Holland BS (2000) Genetics of marine bioinvasions. Hydrobiologia 420: 63–71. doi: 10.1007/978-94-017-2184-4_7
  34. 34. Grosberg R, Cunningham CW (2001) Genetic structure in the sea. From populations to communities. In: Bertness MD, Gaines SD, Hay ME (eds) Marine Community Ecology Sinauer Associates, Inc., Sunderland, Massachusetts, pp 61–84
  35. 35. Sakai AK, Allendorf FW, Holt JS, Lodge DM, Molofsky J, et al. (2001) The population biology of invasive species. Annu Rev Ecol Syst 32: 305–332. doi: 10.1146/annurev.ecolsys.32.081501.114037
  36. 36. Geller JB, Darling JA, Carlton JT (2010) Genetic perspectives on marine biological invasions. Annu Rev Mar Sci 2: 367–393. doi: 10.1146/annurev.marine.010908.163745
  37. 37. Fisher RA (1930) The genetical theory of natural selection. Clarendon Press, Oxford, England
  38. 38. Reznick DN, Ghalambor CK (2001) The population ecology of contemporary adaptations: what empirical studies reveal about the conditions that promote adaptive evolution. Genetica 112: 183–198. doi: 10.1007/978-94-010-0585-2_12
  39. 39. Novak SJ, Mack RN (2005) Genetic bottleneck in alien plant species: influence of mating systems and introduction dynamics. In: Sax DF, Stachowicz JJ, Gaines SD (eds) Species invasions: insights into ecology, evolution, and biogeography Sinauer Associates Inc., Sunderland, Massachusetts, pp 95–122
  40. 40. Dupont L, Viard F, David P, Bishop JDD (2007) Combined effects of bottlenecks and selfing in populations of Corella eumyota, a recently introduced sea squirt in the English Channel. Divers Distrib 13: 808–817. doi: 10.1111/j.1472-4642.2007.00405.x
  41. 41. Pederson DG (1968) Environmental stress heterozygote advantage and genotype-environment interaction in Arabidopsis. Heredity 23: 127–138. doi: 10.1038/hdy.1968.11
  42. 42. Barata C, Baird DJ, Minarro A, Soares AMVM (2000) Do genotype responses always converge from lethal to nonlethal toxicant exposure levels? Hypothesis tested using clones of Daphnia magna Straus. Environ Toxicol Chem 19: 2314–2322. doi: 10.1002/etc.5620190922
  43. 43. Tomas F, Abbott J, Steinberg C, Balk M, Williams S, Stachowicz J (2011) Plant genotype and nitrogen loading influence seagrass productivity, biochemistry, and plant-herbivore interactions. Ecology 92: 1807–1817. doi: 10.1890/10-2095.1
  44. 44. Rius M, Pascual M, Turon X (2008) Phylogeography of the widespread marine invader Microcosmus squamiger (Ascidiacea) reveals high genetic diversity of introduced populations and non-independent colonizations. Divers Distrib 14: 818–828. doi: 10.1111/j.1472-4642.2008.00485.x
  45. 45. Pineda MC, López-Legentil S, Turon X (2011) The whereabouts of an ancient wanderer: Global phylogeography of the solitary ascidian Styela plicata. PLoS One 6: e25495. doi: 10.1371/journal.pone.0025495
  46. 46. Lowe AJ (2002) Microcosmus squamiger; a solitary ascidian introduced to southern California harbors and marinas: salinity tolerance and phylogenetic analysis. PhD thesis, California State University, Fullerton.
  47. 47. Rius M, Turon X, Marshall DJ (2009) Non-lethal effects of an invasive species in the marine environment: the importance of early life-history stages. Oecologia 159: 873–882. doi: 10.1007/s00442-008-1256-y
  48. 48. Pineda MC, López-Legentil S, Turon X (2012) Stress levels over time in the introduced ascidian Styela plicata: the effects of temperature and salinity variations on hsp70 gene expression. Cell Stress Chaperon 17: 435–444. doi: 10.1007/s12192-012-0321-y
  49. 49. Sims LL (1984) Osmoregulatory capabilities of 3 macrosympatric Stolidobranch ascidians, Styela clava Herdman, Styela plicata (Lesueur), and Styela montereyensis (Dall). J Exp Mar Ecol 82: 117–129. doi: 10.1016/0022-0981(84)90098-4
  50. 50. Yamaguchi M (1975) Growth and reproductive cycles of marine fouling Ascidians Ciona intestinalis, Styela plicata, Botrylloides violaceus, and Leptoclinum mitsukurii at Aburatsubo-Moroiso Inlet (Central Japan). Mar Biol 29: 253–259. doi: 10.1007/bf00391851
  51. 51. Rius M, Pineda MC, Turon X (2009) Population dynamics and life cycle of the introduced ascidian Microcosmus squamiger in the Mediterranean Sea. Biol Invasions 11: 2181–2194. doi: 10.1007/s10530-008-9375-2
  52. 52. Pineda MC (2012) A global wanderer: Biology, phylogeography and resilience of the introduced ascidian Styela plicata. PhD thesis, Universitat de Barcelona, Barcelona
  53. 53. Schumann EH, Churchill JRS, Zaayman HJ (2005) Oceanic variability in the western sector of Algoa Bay, South Africa,. African J Mar Sci 27: 65–80. doi: 10.2989/18142320509504069
  54. 54. Schumann EH (1998) The coastal ocean off southeast Africa, including Madagascar. The Sea 11: 557–582.
  55. 55. Fatoki OS, Mathabatha S (2001) An assessment of heavy metal pollution in the East London and Port Elizabeth harbours. Water SA 27: 233–240. doi: 10.4314/wsa.v27i2.4997
  56. 56. Largier JL, Attwood C, Harcourt-Baldwin J-L (2000) The hydrographic character of the Knysna Estuary. Trans Roy Soc S Afr 55: 107–122. doi: 10.1080/00359190009520437
  57. 57. Watling RJ, Watling HR (1982) Metal surveys in South African estuaries II. Kynsa River. Water SA 8: 26–44.
  58. 58. Allanson BRA (2000) The Knysna Basin Project reviewed—research findings andimplications for management. Trans Roy Soc S Afr 55: 97–100. doi: 10.1080/00359190009520435
  59. 59. Schiff K, Diehl D, Valkirs A (2004) Copper emissions from antifouling paint on recreational vessels. Mar Pollut Bull 48: 371–377. doi: 10.1016/j.marpolbul.2003.08.016
  60. 60. Valkirs AO, Seligman PF, Haslbeck E, Caso JS (2003) Measurement of copper release rates from antifouling paint under laboratory and in situ conditions: implications for loading estimation to marine water bodies. Mar Pollut Bull 46: 763–779. doi: 10.1016/s0025-326x(03)00044-4
  61. 61. West AB, Lambert CC (1976) Control of spawning in tunicate Styela plicata by variations in a natural light regime. J Exp Zool 195: 263–270. doi: 10.1002/jez.1401950211
  62. 62. Segar DA (1997) Introduction to Ocean Sciences. Brooks/Cole Pub Co
  63. 63. Vázquez E, Young CM (2000) Effects of low salinity on metamorphosis in estuarine colonial ascidians. Invertebr Biol 119: 433–444. doi: 10.1111/j.1744-7410.2000.tb00113.x
  64. 64. Piola RF, Johnston EL (2009) Comparing differential tolerance of native and non-indigenous marine species to metal pollution using novel assay techniques. Environ Pollut 157: 2853–2864. doi: 10.1016/j.envpol.2009.04.007
  65. 65. Bellas J, Beiras R, Vázquez E (2004) Sublethal effects of trace metals (Cd, Cr, Cu, Hg) on embryogenesis and larval settlement of the ascidian Ciona intestinalis. Arch Environ Con Tox 46: 61–66. doi: 10.1007/s00244-003-0238-7
  66. 66. Reichelt-Brushett AJ, Harrison PL (2005) The effect of selected trace metals on the fertilization success of several scleractinian coral species. Coral Reefs 24: 524–534. doi: 10.1007/s00338-005-0013-5
  67. 67. Xie ZC, Wong NC, Qian PY, Qiu JW (2005) Responses of polychaete Hydroides elegans life stages to copper stress. Mar Ecol-Prog Ser 285: 89–96. doi: 10.3354/meps285089
  68. 68. Haynes D, Loong D (2002) Antifoulant (butyltin and copper) concentrations in sediments from the Great Barrier Reef World Heritage Area, Australia. Environ Pollut 120: 391–396. doi: 10.1016/s0269-7491(02)00113-6
  69. 69. Marshall DJ (2008) Transgenerational plasticity in the sea: Context-dependent maternal effects across the life history. Ecology 89: 418–427. doi: 10.1890/07-0449.1
  70. 70. Svane I, Young CM (1991) Sensory structures in tadpole larvae of the ascidians Microcosmus exaspertus (Heller) and Herdmania momus (Savigny). Acta Zool 72: 129–135. doi: 10.1111/j.1463-6395.1991.tb00939.x
  71. 71. Marshall DJ, Styan CA, Keough MJ (2000) Intraspecific co-variation between egg and body size affects fertilisation kinetics of free-spawning marine invertebrates. Mar Ecol-Prog Ser 195: 305–309. doi: 10.3354/meps195305
  72. 72. Rius M, Branch GM, Griffiths CL, Turon X (2010) Larval settlement behaviour in six gregarious ascidians in relation to adult distribution. Mar Ecol-Prog Ser 418: 151–163. doi: 10.3354/meps08810
  73. 73. Keough MJ, Raimondi PT (1995) Responses of settling invertebrate larvae to bioorganic films - Effects of different types of films. J Exp Mar Ecol 185: 235–253. doi: 10.1016/0022-0981(94)00154-6
  74. 74. Wieczorek SK, Todd CD (1997) Inhibition and facilitation of bryozoan and ascidian settlement by natural multi-species biofilms: Effects of film age and the roles of active and passive larval attachment. Mar Biol 128: 463–473. doi: 10.1007/s002270050113
  75. 75. Warton DI, Hui FK (2011) The arcsine is asinine: the analysis of proportions in ecology. Ecology 92: 3–10. doi: 10.1890/10-0340.1
  76. 76. Underwood AJ (1997) Experiments in ecology: their logical design and interpretation using analysis of variance. Cambridge
  77. 77. Quinn GP, Keough MJ (2002) Experimental design and data analysis for biologists. Cambridge University Press, Cambridge
  78. 78. Folmer O, Black M, Hoeh W, Lutz R, Vrijenhoek R (1994) DNA primers for amplification of mitochondrial cytochrome c oxidase subunit I from diverse metazoan invertebrates. Mol Mar Biol Biotech 3: 294–299.
  79. 79. Hall TA (1999) BioEdit: a user-friendly biological sequence alignment editor and analysis program for Windows 95/98/NT. Nucl Acids Symp Ser 41: 95–98.
  80. 80. Librado P, Rozas J (2009) DnaSP v5: a software for comprehensive analysis of DNA polymorphism data. Bioinformatics 25: 1451–1452. doi: 10.1093/bioinformatics/btp187
  81. 81. Excoffier L, Laval G, Schneider S (2005) Arlequin (version 3.0): An integrated software package for population genetics data analysis. Evol Bioinform 1: 47–50.
  82. 82. Ng TYT, Keough MJ (2003) Delayed effects of larval exposure to Cu in the bryozoan Watersipora subtorquata. Mar Ecol-Prog Ser 257: 77–85. doi: 10.3354/meps257077
  83. 83. Cebrian E, Uriz MJ (2007) Do heavy metals play an active role in sponge cell behavior in the absence of calcium? Consequences in larval settlement. J Exp Mar Biol Ecol 346: 60–65. doi: 10.1016/j.jembe.2007.02.010
  84. 84. Mead A, Carlton J, Griffiths C, Rius M (2011) Introduced and cryptogenic marine and estuarine species of South Africa. J Nat Hist 45: 2463–2524. doi: 10.1080/00222933.2011.595836
  85. 85. Allen JD, Pechenik JA (2010) Understanding the effects of low salinity on fertilization success and early development in the sand dollar Echinarachnius parma. Biol Bull 218: 189–199.
  86. 86. Dybern BI (1967) The distribution and salinity tolerance of Ciona intestinalis (L.) f. typica with special reference to the waters around southern Scandinavia. Ophelia 4: 207–226. doi: 10.1080/00785326.1967.10409621
  87. 87. Vázquez E, Young CM (1996) Responses of compound ascidian larvae to haloclines. Mar Ecol-Prog Ser 133: 179–190. doi: 10.3354/meps133179
  88. 88. Pennati R, Groppelli S, Zega G, Biggiogero M, De Bernardi F, Sotgia C (2006) Toxic effects of two pesticides, Imazalil and Triadimefon, on the early development of the ascidian Phallusia mammillata (Chordata, Ascidiacea). Aquat Toxicol 79: 205–212. doi: 10.1016/j.aquatox.2006.05.012
  89. 89. Marshall DJ (2002) In situ measures of spawning synchrony and fertilization success in an intertidal, free-spawning invertebrate. Mar Ecol-Prog Ser 236: 113–119. doi: 10.3354/meps236113
  90. 90. Wendt DE (2000) Energetics of larval swimming and metamorphosis in four species of Bugula (Bryozoa). Biol Bull 198: 346–356. doi: 10.2307/1542690
  91. 91. Maldonado M, Young CM (1999) Effects of the duration of larval life on postlarval stages of the demosponge Sigmadocia caerulea. J Exp Mar Ecol 232: 9–21. doi: 10.1016/s0022-0981(98)00076-8
  92. 92. Bennett CE, Marshall DJ (2005) The relative energetic costs of the larval period, larval swimming and metamorphosis for the ascidian Diplosoma listerianum. Mar Freshw Behav Phy 38: 21–29. doi: 10.1080/10236240400029333
  93. 93. Svane I, Young CM (1989) The ecology and behaviour of ascidian larvae. Oceanogr Mar Biol Ann Rev 27: 45–90.
  94. 94. Drinkwater K, Mueter F, Friedland K, Taylor M, Hunt G, et al. (2009) Recent climate forcing and physical oceanographic changes in Northern Hemisphere regions: A review and comparison of four marine ecosystems. Prog Oceanogr 81: 10–28. doi: 10.1016/j.pocean.2009.04.003
  95. 95. Somero GN (2010) The physiology of climate change: how potentials for acclimatization and genetic adaptation will determine ‘winners’ and ‘losers’. J Exp Biol 213: 912–920. doi: 10.1242/jeb.037473
  96. 96. Bellas J, Vázquez E, Beiras R (2001) Toxicity of Hg, Cu, Cd, and Cr on early developmental stages of Ciona intestinalis (Chordata, Ascidiacea) with potential application in marine water quality assessment. Water Res 35: 2905–2912. doi: 10.1016/s0043-1354(01)00004-5
  97. 97. Cebrian E, Martí R, Uriz JM, Turon X (2003) Sublethal effects of contamination on the Mediterranean sponge Crambe crambe: metal accumulation and biological responses. Mar Pollut Bull 46: 1273–1284. doi: 10.1016/s0025-326x(03)00190-5
  98. 98. Agell G, Turon X, Caralt S, López-Legentil S, Uriz MJ (2004) Molecular and organism biomarkers of copper pollution in the ascidian Pseudodistoma crucigaster. Mar Pollut Bull 48: 759–767. doi: 10.1016/j.marpolbul.2003.11.001
  99. 99. Bastrop R, Jurss K, Sturmbauer C (1998) Cryptic species in a marine polychaete and their independent introduction from North America to Europe. Mol Biol Evol 15: 97–103. doi: 10.1093/oxfordjournals.molbev.a025919
  100. 100. Gerber AS, Loggins R, Kumar S, Dowling TE (2001) Does nonneutral evolution shape observed patterns of DNA variation in animal mitochondrial genomes? Annu Rev Genet 35: 539–566.
  101. 101. Schizas NV, Coull BC, Chandler GT, Quattro JM (2002) Sympatry of distinct mitochondrial DNA lineages in a copepod inhabiting estuarine creeks in the southeastern USA. Mar Biol 140: 585–594. doi: 10.1007/s00227-001-0728-y
  102. 102. Kelly DW, MacIsaac HJ, Heath DD (2006) Vicariance and dispersal effects on phylogeographic structure and speciation in a widespread estuarine invertebrate. Evolution 60: 257–267. doi: 10.1111/j.0014-3820.2006.tb01104.x
  103. 103. Roman J (2006) Diluting the founder effect: cryptic invasions expand a marine invader's range. P Roy Soc B-Biol Sci 273: 2453–2459. doi: 10.1098/rspb.2006.3597