Environmental Research

Non-native marine invertebrates are more tolerant towards environmental stress than taxonomically related native species: Results from a globally replicated study ^{$, $$}

Mark Lenz

, Bernardo A.P. da Gama

, Nadine V. Gerner

, Judith Gobin

, Frederike Gr ¨oner

, Anil Harry

, Stuart R. Jenkins

, Patrik Kraufvelin

, Corinna Mummelthei

, J ¨org Sareyka

, Eduardo A. Xavier

, Martin Wahl

aMarine Ecology Department, Leibniz-Institut f¨ur Meereswissenschaften an der Christian-Albrechts Universit¨at Kiel (IFM-GEOMAR), D¨usternbrooker Weg 20, D-24105 Kiel, Germany

bDepartamento de Biologia Marinha, Universidade Federal Fluminense (UFF), Caixa Postal 100644, CEP 24001-970 Nitero´i, Rio de Janeiro, Brazil

cDepartment of Aquatic Ecotoxicology, Johann Wolfgang Goethe Universit¨at Frankfurt am Main, Siesmayerstrasse 70, D-60054 Frankfurt, Germany

dDepartment of Life Sciences, Faculty of Science and Agriculture, University of the West Indies, St. Augustine, Trinidad and Tobago

eInstitute of Zoophysiology, Westf¨alische Wilhelms-Universit¨at, Hindenburgplatz 55, D-48143 M¨unster, Germany

fSchool of Ocean Sciences, Bangor University, Menai Bridge, Anglesey LL59 5AB, UK

gARONIA, Coastal Zone Research Team, ˚Abo Akademi University and Novia University of Applied Sciences, Raseborgsv¨agen 9, FI-10600 Eken¨as, Finland

hDepartment of Evolutionary Biology and Animal Ecology, Albert-Ludwigs-Universit¨at Freiburg, Hauptstrasse 1, D-79104 Freiburg, Germany

iDepartment of Animal Ecology, Evolution and Biodiversity, Ruhr-Universit¨at Bochum, Universit¨atsstrasse 150, D-44801 Bochum, Germany

a r t i c l e i n f o

Keywords:

Stress tolerance Survival Respiration Native species Non-native species

a b s t r a c t

To predict the risk associated with future introductions, ecologists seek to identify traits that determine the invasiveness of species. Among numerous designated characteristics, tolerance towards environ- mental stress is one of the most favored. However, there is little empirical support for the assumption that non-native species generally cope better with temporarily unfavorable conditions than native species. To test this concept, we ran five pairwise comparisons between native and non-native marine invertebrates at temperate, subtropical, and tropical sites. We included (natives named first) six bivalves: Brachidontes exustus and Perna viridis, P. perna and Isognomon bicolor, Saccostrea glomerata and Crassostrea gigas, two ascidians: Diplosoma listerianum and Didemnum vexillum as well as two crustaceans: Gammarus zaddachi and G. tigrinus. We simulated acute fluctuations in salinity, oxygen concentration, and temperature, while we measured respiration and survival rates. Under stressful conditions, non-native species consistently showed less pronounced deviations from their normal respiratory performance than their native counterparts. We suggest that this indicates that they have a wider tolerance range. Furthermore, they also revealed higher survival rates under stress. Thus, stress tolerance seems to be a property of successful invaders and could therefore be a useful criterion for screening profiles and risk assessment protocols.

&2011 Elsevier Inc. All rights reserved.

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Environmental Research

0013-9351/$ - see front matter & 2011 Elsevier Inc. All rights reserved.

doi:10.1016/j.envres.2011.05.001

$Funding sources: This study was run as part of the international research and training program GAME (Global Approach by Modular Experiments) of IFM-GEOMAR which is generously funded by the Mercator Stiftung (Essen, Germany) and the foundation of the Christian-Albrechts Universit ¨at zu Kiel (Kiel, Germany). Both foundations did not exert any influence on the design of the study as well as on the analysis and interpretation of the data.

$$Ethics statement: Vertebrate animals were not part of this study. All organism handling and subsequent procedures were in accordance with European laws [European Communities Council Directive of November 24, 1986 (86/609/EEC)] and the national ethics regulations of the participating countries.

nCorresponding author. Fax: þ49 431 600 1671.

E-mail addresses: mlenz@ifm-geomar.de (M. Lenz), bapgama@pq.cnpq.br (B.A. da Gama), nadine.gerner@ufz.de (N.V. Gerner), Judith.Gobin@sta.uwi.edu (J. Gobin), frederike09@googlemail.com (F. Gr ¨oner), anilharry01@yahoo.com (A. Harry), S.Jenkins@bangor.ac.uk (S.R. Jenkins), pkraufve@abo.fi (P. Kraufvelin),

cmummelthei@yahoo.de (C. Mummelthei), J.Sareyka@gmx.de (J. Sareyka), xavier.eduardo@gmail.com (E.A. Xavier).

1Present address: Department System Ecotoxicology, Helmholtz Centre for Environmental Research, Permoserstrasse 15, D-04318 Leipzig, Germany.

2Present address: Leibniz Center for Tropical Marine Ecology (ZMT), Fahrenheitstr. 6, 28359 Bremen, Germany.

3Present address: ARONIA, Coastal Zone Research Team, ˚Abo Akademi University and Novia University of Applied Sciences, Raseborgsv ¨agen 9, FI-10600 Eken ¨as, Finland.

1. Introduction

Species traits such as dispersal characteristics, life-history strategies, growth rates, distributional ranges, association with humans, and tolerance towards environmental stressors have been suggested as determinants of invasiveness in animal and plant species (Williamson and Fitter, 1996; Lepp ¨akoski and Olenin, 2000; Bayne, 2002; Lockwood et al., 2005; Braby and Somero, 2006; Chapman et al., 2006; Jeschke and Strayer, 2006).

Although our understanding of invasiveness grew rapidly during the last decades (Williamson, 1999; Richardson and Pysek, 2008), the list of traits is not complete and it is not clear whether the named characteristics are relevant for all organisms and habitats.

However, knowing the traits and their relative importance would help to understand past invasions and assess the risk of future introductions (Ricciardi and Rasmussen, 1998; Kolar and Lodge, 2001).

To address this problem, ecologists used comparisons between species that differ in their invasion status, including: (a) native species that have not been introduced elsewhere, (b) non-native species that remained inconspicuous in their new range (non- invasive non-natives), and (c) non-native species that established and spread in their new habitats (invasive non-natives) (van Kleunen et al., 2010a). We will follow these definitions through- out this paper, but see Lockwood et al. (2007) for a detailed discussion about the terminology used in invasion ecology.

The majority of these comparisons have been conducted primarily on fish and birds (Jeschke and Strayer, 2006) and higher plants (van Kleunen et al., 2010b) of terrestrial and freshwater systems. So far, only a small number of studies have been done in the marine environment. They were mainly focusing on oysters and mussels (but see Piola and Johnston (2009) for a study on bryozoans), which are dominant mollusk components of tempe- rate rocky shores (Braby and Somero, 2006; Krassoi et al., 2008;

Schneider, 2008; Zardi et al., 2008). This is in contrast with the massive increase in the number of marine bioinvasions by a broad variety of taxa in recent decades (Ruiz et al., 2000; Lepp ¨akoski et al., 2002).

Braby and Somero (2006) compared heart rates in two mussels from the Californian coast after acute and acclimatory changes in temperature and salinity. For this they used the non-native mussel Mytilus galloprovincialis and the native M. trossulus. They confirmed differences between the two bivalves which they anticipated on the base of the biogeographic patterning in mytilids: M. galloprovincialis is more heat tolerant than its native congener that is adapted to the colder California (upwelling) waters.

A further study on these species, conducted in the same region, showed that not only subtidal but also intertidal populations of M. galloprovincialis show a high thermal tolerance (Schneider, 2008). Moreover, Zardi et al. (2008) investigated the performance of this non-native mytilid in the face of wave and sand stress in its introduced range along the west and south coast of South-Africa.

The authors compared the invader to the native brown mussel Perna perna and found M. galloprovincialis to be more resistant towards sand inundation, while P. perna was more tolerant towards hydrodynamic stress.

At the east coast of Australia, Krassoi et al. (2008) found that abiotic stress modifies the outcome of competitive interactions between two oyster species that differ in competitiveness. In most habitats the non-native Crassostrea gigas is dominant over the native Saccostrea glomerata (Bayne, 2002) except in the high intertidal, where environmental conditions fluctuate widely.

These few examples illustrate that the picture for the marine environment is incomplete, taxonomically biased, biogeographi- cally restricted, and, most frustrating for invasion ecologists,

ambiguous. It seems that the sign of a given difference in stress resistance between a native and a non-native species is deter- mined by the type of stressor investigated.

It is intuitive that invasion success in marine habitats is controlled by the interplay between species-specific adaptations and site-specific environmental characteristics (Facon et al., 2006). Fluctuating habitats, such as the intertidal, should more frequently bear successful invaders, since species that thrive in these environments are pre-adapted to variations in abiotic variables like temperature, salinity, light intensity, and oxygen availability. This should predestine them for tolerating adverse conditions during transport and after arrival in a new habitat.

However, the data we have from marine systems do not yet permit to even answer the question whether successful invaders do feature wider abiotic tolerance. To improve the knowledge basis, we ran interspecific comparisons between successful non- native and resident species in various marine regions across the globe. The latter we classified as either truly restricted to their home range or as residents which are non-native elsewhere (van Kleunen et al., 2010a). All comparisons were done in the intro- duced range of the non-natives. Our experiments covered three climate zones (temperate to tropical) in 5 different sea areas worldwide. In all comparisons, we strove to select combinations of species which are similar with regard to their ecological niches, their tolerance towards common environmental stressors, e.g. we did not compare eury- to stenohaline species, and their taxon- omy. Finally, to expand the range of comparisons beyond the group of bivalves, we included crustaceans and tunicates in our study.

2. Material and methods

2.1. Study sites and organisms

For this comparative study, we used three categories of species of which the first two were investigated in their home range and the third in their introduced range: (a) native species that have so far not been found outside their native range (Brachidontes exustus, S. glomerata, Gammarus zaddachi), (b) native species that are non-native in other habitats (P. perna, Diplosoma listerianum), and (c) non-native species (P. viridis, Isognomon bicolor, C. gigas, Didemnum vexillum, G. tigrinus). These marine invertebrates comprise three major taxonomic groups and were collected in five biogeographic regions located in tropical, warm-temperate, temperate and cold-temperate coastal systems (Fig. 1). For the pairwise comparisons we chose organisms that are taxonomically related and that occupy comparable ecological niches. We ran experiments in which the organisms were exposed to acute abiotic stress for short (d) and intermediate (wk) time intervals, using the laboratory facilities of institutions that participate in the international research and training program GAME (Global Approach by Modular Experiments, www.ifm-geomar.de/

game).

In the tropical southern Caribbean, we investigated the native mytilid B.

exustus and the non-native P. viridis. Both species were collected from intertidal rocks or man-made structures, such as sea walls, in the Gulf of Paria, southwest Trinidad (10117⁰N, 61128⁰W). The scorched mussel B. exustus is native to the Gulf of Mexico and the southern Caribbean (Barber et al., 2005), while the greenlip mussel P. viridis stems from the Indo-Pacific region and it first appeared at the coastline of Trinidad in the 1990s (Agard et al., 1992). Both species belong to the family Mytilidae, are similar in their lifestyle, and co-exist in the habitats we sampled.

In the tropical Southern Atlantic, we collected specimens of the brown mussel P. perna, which is commonly considered as being native to the Atlantic coasts of South-America and Africa as well as to India (Hicks et al. (2001)and references therein; but seeSouza et al. (2003)), and the non-native purse oyster I. bicolor from two rocky shores near Nitero´i, Brazil (22158⁰S, 43104⁰W). I. bicolor originates from the Caribbean and was firstly described for Brazilian waters in the 1980s (Domaneschi and Martins, 2002). Both species belong to the subclass of the Pteriomorpha and co-occur in intertidal habitats of the State of Rio de Janeiro.

In the warm-temperate environment of New Zealand’s Northern Island, we compared the native New Zealand rock oyster S. glomerata and the non-native Pacific oyster C. gigas. We collected adult individuals of both species from tidal flats and rocks in the Mahurangi estuary (36145⁰S, 174170⁰E). C. gigas was first observed in the harbor of Mahurangi in 1971 (Powell, 1979) and spread rapidly also to shores of the Southern Island in following years. C. gigas and S. glomerata

both belong to the family Ostreidae and co-exist in many habitats along the coast of New Zealand (Krassoi et al., 2008).

In the temperate climate of the Isle of Anglesey, Wales, (53119⁰N, 04138⁰W) we worked on two compound ascidians. D. listerianum is a cosmopolitan species whose origin has not unequivocally been identified. However, first reports of it from the British Isles date back to the 1840s (Hansson, 1989) and it is common at the west coast of the UK (Vance et al., 2008). D. vexillum presumably has its native range in the northwest Pacific and was first found at the coasts of Wales in 2008 (Griffith et al., 2009;Lambert, 2009). The two ascidians belong to the family Didemnidae and were both collected from one marina where they grow on artificial hard-substrata.

In the brackish and cold-temperate environment of the northern Baltic Sea, in the Gulf of Finland (59150⁰N, 23115⁰E), individuals of the two gammarid species G. zaddachi and G. tigrinus were collected from shallow shore habitats which are dominated by the brown alga Fucus vesiculosus. G. zaddachi is one of five gammarid species that are native to this sea area (Korpinen and Westerbom, 2010), while G. tigrinus is a recent invader that was first reported from the Gulf of Finland in 2003 (Pienim ¨aki et al., 2004;Paavola et al., 2008). Both crustaceans have similar food preferences, are of similar size, share the same habitat, and reproduce during the same period in summer (Sareyka et al., 2011).

2.2. Collection of study organisms

All organisms were collected in their natural habitats: mussels and oysters were carefully detached from rocks or artificial substrata using a knife; ascidians were gently scraped from pontoons with a scalpel, and gammarids were picked from submerged fronds of F. vesiculosus with hand nets and pipettes. All animals were transported to the laboratory within 2 h. During transport they were kept in cooler boxes in which they were submerged in natural seawater. In the laboratory, animals were maintained in aerated aquaria with water quality control that was achieved by: (a) a flow-through system (in New Zealand and Finland), (b) regular manual renewal (in Trinidad and Wales) or (c) filtering (in Brazil) until the experiments were started 1–2 weeks later (Table 1). During this period, bivalves were not fed beyond the plankton contained in the unfiltered seawater provided, ascidians received 2 ml of a Rhinomonas reticulata culture per day and container (0.9 L volume), while gammarids were fed with pieces of F. vesiculosus. After arrival in the laboratory, ascidian colonies were cut into fragments of 5  10 mm and were then induced to attach themselves to microscope glass slides (Epelbaum et al., 2009). Re-attachment of colony fragments was verified 2 d later by touching the colonies with a pipette: attached fragments remained in place, while the unattached fell off the slides (Bullard et al., 2007). Only well-attached colonies were later used in the experiments.

2.3. Abiotic stressors

Experimental conditions varied among countries due to the use of (a) different test organisms, (b) different infrastructures, and (c) different experimental

logistics. However, they were always comparable for the pairs of non-native and native species and consequently did not affect the relative difference in stress tolerance which was our response variable throughout. We screened the different experimental regions for environmental variables that fluctuate strongly in time and therefore constitute a relevant stress for the organisms tested (Table 1). Hypo- salinity is commonly induced by land runoff and/or heavy rainfall, while these events can affect intertidal or shallow water habitats at temperate (Braby and Somero, 2006) and tropical (Goodbody, 1961;Coles and Jokiel, 1992) coasts. Since the study sites in Trinidad, Brazil, Wales, and New Zealand are all located in humid regions and exhibit yearly average precipitation rates between 1000 and 2000 mm, we identified hypo-salinity as one of the most relevant stressors for local shallow water biota. Hypoxia in coastal habitats is mainly an indirect consequence of eutrophication that can be induced, for example, by decaying plant material (Jewett et al., 2005). Since drifting algal mats are a common phenomenon in the northern and northeastern parts of the semi-enclosed Baltic Sea (Norkko and Bonsdorff, 1996), we opted for hypoxia as a relevant stressor for gammarids from the Finnish coast. In the same system, summer heat waves can raise temperatures in small rock pools and very shallow (o10 cm), semi-enclosed bays beyond 32 1C (Ganning, 1971). This suggests that elevated temperatures are another relevant stressor for gammarids in this region.

In general, stress intensity and duration were adjusted to mimic common local regimes at all sites. In the following, we indicate the magnitude of an experimental deviation from normal conditions by a ‘‘delta-value’’. It comes in the unit of the environmental variable we manipulated (despite oxygen concentrations) and its sign indicates whether we increased or decreased it. In all experiments, the animals were exposed to stressful conditions abruptly to prevent acclimatization.

We measured bivalve respiration rates before and during exposure to stress (short- term experiments only) as well as mortality during stress. In all studies, we kept reference groups at local ambient conditions in parallel to the stress trials to monitor (a) natural fluctuations in respiration rates and (b) background mortality.

Hypo-osmosis was achieved by mixing seawater with unchlorinated tap water or with rainwater that we collected in a storage basin prior to the experiment. Salinity levels were carefully adjusted with a hand-held refractometer or conductivity meter. All experiments were conducted between January and October 2009.

When we manipulated salinity in New Zealand, we decoupled the experi- mental units from the flow-through system, but aerated them throughout the experimental period. The number of replicates, aquaria with a volume of o5 L, ranged between 10 and 12 in Trinidad, Brazil, Wales and Finland, while it was 6 in New Zealand (Table 1). The deviation from the ambient salinity at the different sites ranged from  7 units (Wales) to  30 units (Brazil), while the time of exposure was between 4 d (Trinidad, Brazil and New Zealand) and 14 d (Wales) (Table 1). In Trinidad, Brazil and New Zealand we additionally monitored the recovery from stress-induced metabolic depression during another 24 h with a temporal resolution of 12 h (New Zealand) and 24 h (Trinidad and Brazil). To initiate the recovery phase, we slowly replaced the water in all experimental units by seawater within the course of 1 h.

In Finland, we exposed adult gammarids to three different levels of thermal stress: 30 1C (þ 10 units), 32 1C ( þ12 units), and 35 1C (þ 15 units), while juveniles experienced hypoxia (1% oxygen saturation at 20–22 1C,  79 units). In the first Fig. 1. Interspecific comparisons between native and non-native species were conducted at five sites, Trinidad, Brazil, New Zealand, Wales, Finland, and covered three climate zones. Asterisks mark non-native species.

experiment, test groups had 10 replicates containing 10 equally sized gammarids each. Thermal stress was applied by warming the containers in a 5 l water bath over 1–2 h to target temperature using an aquarium heater. We started to record mortality after the respective target temperature was reached in the test contain- ers and stopped the experiment immediately when 50% mortality was exceeded.

For controlling background mortality, we had one reference unit (18 1C) with 10 gammarids per treatment replicate. For the second experiment on hypoxia tolerance, we used four to six-week-old gammarids. Juveniles were chosen because they frequently inhabit decaying mats of filamentous algae such as Cladophora glomerata, which they simultaneously use as a food source (Salovius and Kraufvelin, 2004;Kraufvelin and Salovius, 2004). These micro-habitats often exhibit low oxygen concentrations. The animals with an average length of 4.5 mm were put in sealable chambers (10  10  4 cm³, 10 individuals per chamber constituted one replicate, n¼ 10). The lid of the chamber had one opening to insert an oxygen probe and a second one to bring in a nitrogen pipe. We reduced the oxygen concentration in the chambers by N2-bubbling, monitored the oxygen tension every 20 min and stopped the trial when 50% mortality was reached.

A further 10 replicates served to assess background mortality at normoxia.

2.4. Response variables

We used three different response variables in this study: survival as a function of time (Brazil, Wales), time until 50% mortality was reached (Finland), and the change in respiration rate after the onset of stress (Brazil, New Zealand, Trinidad).

We used simple criteria to determine whether an individual died during the course of the experiment. In Brazil and New Zealand, we considered a gaping bivalve dead when it did not close its valves following mechanical stimulation. In Finland, we carefully observed limb movements of the gammarids. If they ceased completely for longer than two minutes, an individual was counted as dead.

Finally in Wales, we had a set of criteria to determine whether an ascidian colony died. For this we checked whether (a) colonies or single zooids detached from glass slides, (b) unicellular algae accumulated on the bottom of the tank, what indicated that the zooids were not feeding anymore, and (c) colony colour changed from yellow to brown-greyish, while small brown spots appeared on them. These were clumps of fecal pellets caught in the cloacal canals and cavities near the colony’s surface (Valentine et al., 2007). Colonies were considered dead whenever at least one of these criteria was met. At all sites mentioned, we never observed the recovery of an individual that was previously considered dead.

Monitoring respiration required a setup that allowed the detection of subtle changes in the oxygen content in the experimental units. For this we had small, seawater-filled and air-tight respiration chambers, which differed in shape and

volume between the stations. For measuring respiration rates, we transferred all individuals of one experimental unit from the aquarium into a chamber and back.

After a short recovery phase, we monitored the decline in oxygen concentration over time.

Trinidad was the only exception from this protocol. Here the experimental units themselves were sealed air-tight and aeration was switched off. Then we inserted an oxygen probe and measured the decline in oxygen concentration.

Afterwards, the seal was removed and aeration was switched on again.

Despite the differences in the setup, measurement procedures were similar at the different stations and will be exemplified by the case of Brazil. Here the chamber had a volume of 1 l and we measured the change in oxygen concentra- tion over the course of 1.5 h by quantifying the O2content in the beginning and in the end of this phase. We exchanged the water in the chamber after measuring one replicate, and re-adjusted its salinity according to the different treatment levels. Chambers were always completely filled to avoid the presence of air bubbles and were sealed with a lid that had an opening to allow the insertion of an oxygen probe. Types of hand-held oxymeters varied between sites: YSI Environ- mental DO 200 (Trinidad), Microx TX3 by PreSens (New Zealand), GMH 3630 by Greisinger GmbH (Brazil). All devices were carefully calibrated following the manufacturer’s information and therefore we do not assume that differences between them biased our results. To ensure the mixing of the water near the sensor, we used a magnetic stirrer that was placed underneath the container and that moved a magnet on the bottom of the chamber. Respiration rates were measured immediately before the stress was imposed and again two and/or four days after the onset of stress. Finally, we obtained the dry weight of all bivalve individuals and standardized respiration rates by biomass. Deviations from the Brazilian protocol at other sites are indicated inTable 1.

2.5. Data analyses

The choice of statistical test procedures for the different experiments depended on (a) type of response variable (interval vs. survival data), (b) data distribution (parametric vs. non-parametric tests) and (c) independency of samples (independent vs. dependent tests). Respiration rates before and during hypo-osmosis (within-species comparisons) in New Zealand were compared with the Wilcoxon matched-pairs test. We tested for normality of data with the Shapiro–Wilks-W test and confirmed homogeneity of variances with Levene’s test. Data were square root transformed to achieve normality for parametric test procedures if needed. To compare the effect of hypo-salinity on respiration rates between species, we calculated a ratio that quantifies the relative change in oxygen consumption after the onset of the stressor. For this, we divided the Table 1

Overview over species, stressors, stress levels, experimental conditions and the log effect ratios (respiration during stress/respiration before stress) measured at the five different sites included in this study.

Country Species (* ¼ non- native)

Exposure period

Absolute stress levels

Delta Acclimatiza-

tion to laboratory conditions

Ambient water temperature (1C)

Number of replicates

Number of test organisms per replicate

Log effect ratios

Trinidad &

Tobago

Brachidontes exustus

4 days 10 and

20 psu

25 and

15 units

1 day 23–25 10 1 0.62

Perna viridis* 0.04

Brazil Perna perna 4/12 days 5 and

20 psu

30 and

15 units

1 week 19 12 3 (P. perna)

5 (I. bicolor)

0.94

Isognomon bicolor*

1 0.43

New Zealand

Saccostrea glomerata

4 days 12 psu 23 units 1 week 26 6 1 3.00

Crassostrea gigas*

1.90

Wales Diplosoma

listerianum

14 days 20 and

27 psu

14 and

7 units

2 weeks 16 11 1 –

Didemnum vexillum*

–

Finland Gammarus

zaddacchi

5.5h 30, 32 and

35 1C

þ13, þ15 and þ18 units

2 weeks (thermal stress only)

17 10 10 –

Gammarus tigrinus*

–

Finland Gammarus

zaddacchi

3.5 h 1% oxygen

saturation

79 units – 17 10 10 –

Gammarus tigrinus*

–

respiration rate during stress/after stress by the oxygen consumption before stress (effect ratios), i.e. we made an intra-individual comparison. For graphical presentation and statistical analysis the ratios were log10-transformed. We then tested whether the obtained ratios were significantly different between species by comparing them with a Mann–Whitney-U test. In case of multiple testing, we adjusted the significance level with the Bonferroni correction to compensate for the increase in the type I error rate.

Survival rates of colonial ascidians in Wales were compared with the Log-rank test. We used a t-test or the Mann–Whitney-U test to detect significant differences in the time spans until 50% mortality was reached in the Finnish amphipod populations.

3. Results

All species tested showed significant deviations from their normal performance when we exposed them either to hypo- salinity, hypoxia, or heat stress. In all cases, these deviations were more pronounced in the native than in the non-native inverte- brates. In the following paragraphs, we will first provide informa- tion about the stress-induced change in animal performance and then we will document the difference between the respective native and non-native species. Experimental sites are listed according to their latitude (101, 221, 361, 531 and 591).

3.1. Trinidad: B. exustus (native) vs. P. viridis (non-native)

Mean respiration rates in the native B. exustus decreased by more than 50% after the onset of osmotic stress, regardless of the magnitude of the challenge (  15 or 25 units). Furthermore, they did not return to normal during the 72 h of stress exposure.

The reference group for B. exustus, which was simultaneously kept under oceanic salinity, exhibited no significant change in mean O

-uptake rates (data not shown). When exposed to identical stress regimes, the non-native P. viridis exhibited a much lower decline in mean respiration rates than the native. Furthermore, it maintained normal respiration rates at least for 48 h, while the native B. exustus exhibited metabolic depression already after 24 h (Fig. 2). The pronounced differences in median relative performances (consumption under stress in relation to consump- tion before stress) confirm that the native mussel reacted more

strongly to hypo-osmosis than the non-native (Fig. 2). In 20 psu, the difference between species was significant throughout the experiment (Mann–Whitney-U test: 24 h: U ¼8, pr0.006; 48 h:

U¼11, p r0.006; 72 h: U¼12, pr0.006; n¼7–10 for all compar- isons). In 10 psu, though the trend was the same, it was never significant due to the large spread in the B. exustus group.

Interestingly, some individuals of the latter species returned to their pre-stress performance within 24 h after salinity was re-adjusted to oceanic conditions, while none of the non-native P. viridis exhibited such a high resilience (Fig. 2). However, the large discrepancy in median relative performance between the native and the non-native bivalve in 20 psu persisted until the end of the experiment (Mann–Whitney-U test: 96 h: U¼13, p r0.006, n¼8; Fig. 2).

3.2. Brazil: P. perna (native) vs. I. bicolor (non-native)

Both species showed a similar and significant decline in their respiration rates during the first 48 h under hypo-osmosis ( 30 units). Metabolic depression persisted until the experiment was terminated after 96 h. During this time, some individuals even became inactive, closed their valves, and ceased oxygen uptake.

However, after the bivalves were returned to fully saline sea- water, both species reached their normal respiratory performance within 24 h (Fig. 3). Reference groups for P. perna (native) and I. bicolor (non-native), which we kept at oceanic salinity through- out the experiment, revealed no change in their oxygen con- sumption during the same time interval (data not shown).

Though both species showed a similar response to unfavorable osmotic conditions, the decline in respirations rates relative to pre-stress levels was significantly stronger in the native species after two but not after four days of exposure to hypo-osmosis (Mann–Whitney-U test: 2nd day: U ¼159.00, pr0.01; 4th day:

U¼211.00, p¼0.11; n ¼12 for all comparisons, Fig. 3).

In the long-term experiments (12 d exposure period), we observed no mortality in the reference groups that experienced oceanic salinity (data not shown). Mortality under both stress regimes was significantly higher in the native than in the non-native bivalve (5 psu: Log-rank test: w

¼23.8, p r0.001;

20 psu: Log-Rank-test: w

¼22.6, pr0.001; Fig. 4aþb, n ¼12 for all comparisons).

3.3. New Zealand: S. glomerata (native) vs. C. gigas (non-native)

Both species exhibited a significant reduction in their median respiration rates during the 96 h of osmotic stress (Wilcoxon matched-pairs test for both species: T¼0.00, p r0.05, n¼6, Fig. 5). Twelve hours after salinity was re-adjusted to normal conditions, both species exhibited respiration rates that clearly exceeded their pre-stress oxygen uptake (Fig. 6). The stress- induced effect on respiration, however, was blurred by the remarkable variability in the response of the non-native oyster, which became apparent in the stressed but also in the reference group and may reflect the natural variability inherent in the population sampled (Fig. 5). Even when osmotically challenged some specimens showed respiration rates close to their normal performance. Consequently, the spread in the data was signifi- cantly different between the oyster species (Levene’s test: F¼7, pr0.05, n¼6, Fig. 5). When we compared the stress-induced changes in oxygen consumption between species (median log effect ratio: C. gigas: 1.9, S. glomerata: 3.0), we did not find a significant difference (Mann–Whitney-U test: U¼ 12, p¼0.38, n¼ 6, Fig. 6), what is presumably due to the high variability in the data.

Fig. 2. Relative change in oxygen uptake during and after hypo-osmotic stress (10 and 20 psu) in the native Brachidontes exustus (plain boxes) and the non-native Perna viridis (hatched boxes) from Trinidad. For the log10effect ratios, oxygen consumption during stress (24 h, 48 h, 72 h) as well as after stress (96 h) was divided by oxygen consumption before stress. The vertical dashed line indicates when salinity in the experimental units was re-adjusted to oceanic conditions.

Asterisks mark significant differences between species (Mann–Whitney-U test:

pr0.006). N¼10. Boxplots show medians, interquartiles, and the non-outlier range.

3.4. Wales: D. listerianum (native) vs. D. vexillum (non-native)

Under hypo-osmotic stress of 7 units all colonies of the native tunicate were dead after 10 d, while 82% of the D. vexillum colonies were still alive after 14 d (Fig. 7a). This difference in mortality rates between species was significant (Log-Rank test:

w

¼14.6; pr0.001, n¼11). We observed the same trend when we exposed the tunicates to 20 psu (  14 units) (Fig. 7b).

However, under these more stressful conditions both species showed substantial mortality within 14 d: 100% in D. listerianum (native) and 62% in D. vexillum (non-native). Again the stress resistance of the non-native species was significantly higher (Log- Rank test; w

¼19.7; pr0.001, n¼11). At oceanic salinity the two tunicate species showed negligible mortality over 14 d (1 colony per species).

Fig. 3. Relative change in oxygen uptake during and after hypo-osmotic stress (5 psu) in the native Perna perna (white boxes) and the non-native Isognomon bicolor (gray boxes) from southern Brazil. For the log10effect ratios, oxygen consumption during stress (48 h, 96 h) as well as after stress (120 h) was divided by oxygen consumption before stress. The vertical dashed line indicates when salinity in the experimental units was re-adjusted to oceanic conditions. N¼24. Asterisks mark significant differences between species (Mann–Whitney-U test: pr0.01).

Boxplots show medians, interquartiles, and the non-outlier range.

Fig. 4. (a þb) Survival of the native Perna perna (dotted line) and the non-native Isognomon bicolor (solid line) in Brazil under (a) 5 psu and (b) 20 psu. N ¼12.

Fig. 5. Effect of osmotic stress (12 psu) on respiration rates in the native Saccostrea glomerata (plain boxes) and the non-native Crassostrea gigas (hatched boxes) in New Zealand. Measurements were taken immediately before manipula- tion and after 96 h of exposure to hypo-osmotic conditions. Boxes in gray show respiration rates in reference groups that were kept at 35 psu. N¼ 6. Boxplots show medians, interquartiles, and the non-outlier range.

Fig. 6. Relative change in oxygen uptake during and after hypo-osmotic stress (12 psu) in the native Saccostrea glomerata (plain boxes) and the non-native Crassostrea gigas (hatched boxes) from New Zealand. For the log10effect ratios, oxygen consumption during stress (96 h) as well as after stress (108 h) was divided by oxygen consumption before stress. The vertical dashed line indicates when salinity in the experimental units was re-adjusted to oceanic conditions.

N¼ 6. Boxplots show medians, interquartiles, and the non-outlier range.

3.5. Finland: G. zaddachi (native) vs. G. tigrinus (non-native)

The two gammarid species differed conspicuously in their thermal stress tolerance: all adult individuals of G. tigrinus (non-native) survived to the end of the experiment after 10 h both at 30 1C and 32 1C, while 50% of the G. zaddachi (native) individuals had died after an average of 5.6 h (SD ¼1.73 h) and 1 h (SD ¼0.22 h), respectively. Furthermore, while 50% of all G. tigrinus (non-native) individuals survived at 35 1C for an average time span of 5.6 h (SD ¼1.3 h), all individuals of G. zaddachi (native) died immediately.

Under hypoxic conditions juvenile specimens of the non-native G. tigrinus survived significantly longer than individuals of the native G. zaddachi (mean time interval until 50% mortality was reached: G.

tigrinus¼ 3.5 h, SD¼ 0.44 h; G. zaddachi¼3.1 h, SD¼0.43 h). Time spans until 50% mortality was reached were significantly different between species (t-test: t¼2.27, p¼0.018, n¼10).

4. Discussion

Despite the diversity of the experiments – 10 different species belonging to 3 different classes tested in 5 different biogeographic regions covering 3 climatic zones – the resulting pattern is

surprisingly clear: whenever stress resistance differed between functionally similar and phylogenetically related non-native and native species, the non-natives were more robust. Their higher tolerance was attested by (a) their sustained maintenance of normal respiration in the face of stress (Trinidad), (b) the fact that their metabolic change under unfavorable conditions was less distinct than in the native species (Trinidad, Brazil, New Zealand), and (c) their higher survival when environmental variables were extreme (Brazil, Wales, Finland). The clearness of our results is remarkable since the stress applied was one encountered naturally in the native species’ habitat. Thus, the pattern was robust across the different climate zones and regions. It was not superimposed by habitat- specific adaptations to, for instance, annual variability in water temperature (high in Finland and Wales, moderate in New Zealand, low in Trinidad and Brazil) or food availability (variable in Finland, Wales and New Zealand, constant in Trinidad and Brazil), which may differ between native and non-native species. Osovitz and Hofmann (2007) reviewed recent macrophysiological studies from the marine environment and found that spatial and temporal variation in environmental variables largely defines the physiologi- cal tolerance of species. The fact that the pattern we observed – non- native species were more stress tolerant than native species – did not change from variable temperate to more stable subtropical/

tropical habitats speaks for its robustness.

van Kleunen et al. (2010a) emphasized that in interspecific comparisons one need to distinguish between native species, which have never been reported from outside their home range, and residents, which are successful invaders elsewhere. Experi- ments in which the first or the second type of native species serves as a reference differ in their explanatory power. If a high physiological tolerance is an indispensable pre-adaptation for the establishment in new environments, we should in the second case not expect a difference between the non-native species and the species that is native at the study site, but invasive elsewhere- unless there are differences regarding stress tolerance among native and invasive populations of given species. Such a negative result could then lead to the (false) rejection of models about the relevance of stress tolerance for invasion success. The largest difficulty for an experimenter here is to find reliable information about the distributional range of species and whether they have successfully invaded other habitats. This information is unavail- able for many organisms, since invasions may have gone unrec- ognized or unreported. In our study, the majority of species we considered as native have not been encountered outside their home range. Exceptions are the brown mussel P. perna and the ascidian D. listerianium. The common assumption that P. perna is native to the coast of Brazil (e.g. Global Invasive Species Database, 2011) has been doubted by Souza et al. (2003), who, on the base of archeological records, suggested that P. perna was introduced during the 16th century with vessels that arrived from West Africa. However, to this date there are no genetic studies that could verify this hypothesis. Even if it is correct, it would mean that P. perna has been part of the Brazilian coastal fauna for approximately 500 years, while its non-native counterpart in our study, the purse oyster I. bicolor, arrived not more than 30 years ago. Even in case P. perna is non-native to Brazil in the sense that it has not evolved in this region, populations of the two species have had very different time spans to re-build genetic diversity after invasion and for adaptation to the Brazilian environmental conditions. Almost synchronously to the establishment of I. bicolor in Brazilian waters, P. perna expanded its distributional range northwards and invaded the Gulf of Mexico (Hicks et al., 2001). Hence, P. perna, independent of its actual status in Brazil, exhibited the potential to invade new habitats and is therefore different from the majority of native organisms we included in this study.

Fig. 7. (a þ b) Survival of the native Diplosoma listerianum (solid line) and the non-native Didemnum vexillum (dotted line) from Wales under (a) 27 psu and (b) 20 psu. N ¼ 11.

The ambiguous reports about the status of D. listerianum from the Eastern North-Atlantic suggest that it should be classified as cryptogenic for this part of the world ocean. Though reports from the British Isles date back to the 19th century (Hansson, 1989) and Berrill (1950) describes it also for Scandinavia and the Mediterra- nean, recent studies report it as non-native for The Netherlands (Gittenberger, 2007) and for a location at the east coast of the UK (Vance et al., 2008). This either hints at a discontinuous distribu- tion of this species in Europe or may indicate that it is still in the process of expansion after an introduction that occurred before the mid of the 19th century. However, the early reports from the British Isles and the fact that it is well established at the west coast of the UK (Vance et al., 2008) convinced us to regard it as native in the context of our study. There are no indications of invasions by the three other native invertebrates, i.e. B. exustus, S. glomerata and G. zaddachi, we included in this multi-site comparison. However, even in those cases in which the native species successfully colonized habitats outside their home range, we found them less robust to stress than their non-native counterparts. This may indicate that the notoriously stressful invasion process by selec- tive mortality enhances the robustness of non-native populations.

There are indications that populations of non-native species are less stress resistant in its region of origin than in its introduced range (Richards et al., 2006; Hammann and Weinberger, pers.

comm.). This disparity may be a consequence of another property of successful invaders: the ability to quickly adapt evolutionarily to new challenges (Lee, 2002).

The robust pattern that we found despite the heterogeneities (different biogeographic regions, species, stressors, and response variables) inherent in our experiments suggests that non-native species are indeed more stress resistant than natives. However, at the moment we may not decide in each case whether this difference resides at the species or at the population level.

The interpretation of survival rates is clear and unambiguous - higher survival rates under hypo-salinity (I. bicolor vs. P. perna, D. vexillum vs. D. listerianum), hypoxia (G. tigrinus vs. G. zaddachi) or thermal stress (G. trigrinus vs. G. zaddachi) indicate a higher resistance. In contrast, changes in oxygen consumption under hypo-osmotic stress need more careful consideration. Environ- mental stress can increase respiration rates when the organism compensates for stress impact in some active way. Regulatory mechanisms necessitate ATP and therefore lead to an elevated oxygen demand. Intertidal bivalves, for instance, control their cell volume in the face of fluctuating salinity by altering the concen- tration of nitrogenous osmolytes in cytosol (Yancey et al., 1982;

Sadok et al., 1997). However, we never observed an increase in oxygen consumption under osmotic stress in our experiments.

On the contrary, external fluctuations may also cause respira- tion to decrease when the organism goes into a metabolic depression until conditions improve (Ekau et al., 2010). This is well documented for intertidal bivalves that have the ability to isolate their internal milieu from the environment. In this group, a sudden change in physical parameters often leads to valve closure as a primary reaction (de Vooys, 1991; de Zwaan and Mathieu, 1991; Nicholson, 2002; Braby and Somero, 2006). This behavioral response is then followed by metabolic depression, e.g. a reduction in cardiac activity (Braby and Somero, 2006), which minimizes energy turnover when gas and food supply is reduced or has ceased. M. edulis and the infauna clam Spisula solidissima, for instance, reduce respiration and cardiac activity when they are exposed to brackish water (DeFur and Mangum, 1979; Stickle and Sabourin, 1979; Bakhmet et al., 2005).

The bivalve species we tested followed the second strategy.

Although not all individuals closed their valves entirely, their behavioral responses were similar enough to allow this classifica- tion. We suggest that the strength of the deviation from normal

respiration during stress indicates the capability of an organism to tolerate hypo-salinity. The less pronounced the metabolic depression, the longer an organism can stand adverse conditions and the smaller are its energetic debts (e.g. due to anaerobiosis).

Therefore we consider those species more tolerant which showed the smaller stress-induced drop in respiratory performance. This concept is supported by our observation that the non-native purse oyster I. bicolor, which exhibited higher respiration rates under osmotic stress than the native P. perna, also showed lower stress-induced mortality than the latter.

In contrast to many bivalve species, ascidians are generally viewed as stenohaline organisms that do not tolerate pronounced fluctuations in ambient salinity (Lambert, 2005). However, the two species we investigated have been reported from oceanic as well as brackish habitats (Brunetti et al., 1988; Marshall and Keough, 2005; Osman and Whitlatch, 2007; Lambert, 2009). This indicates that they have a wider tolerance range than many other tunicates. The two species exhibited significantly different abil- ities to cope with osmotic stress. The non-native D. vexillum showed substantially less mortality under moderate ( 7 units) and severe (14 units) hypo-salinity than the native D. listerianum.

While it is unknown which physiological adaptation mediates the tolerance in D. vexillum, it should constitute a competitive advan- tage for this recently introduced species if precipitation rates will increase in coming years as it is predicted for Wales (Farrar and Vaze, 2000).

The amphipod G. tigrinus, which is non-native in the Baltic Sea, occurs, in its native range, from the cold-temperate waters of Canada to the subtropical Florida (Bousfield, 1958). It is not known whether the population in the Eastern Baltic stems from warm- or cold-adapted populations of the Western Atlantic.

Additionally, multiple introductions from various source areas to the eastern Baltic could also have led to a mixed gene pool that harbors genetic lineages exhibiting very different pre-adapta- tions. Roman (2006) documented such a process for non-native populations of the shore crab Carcinus maenas along North-West Atlantic shores. The fact that the Finnish populations tolerate very low temperatures (during the Finnish winter) as well as severe heat stress (during our experiment) depicts the wide thermal tolerance of G. tigrinus. The native G. zaddachi , in contrast, died quickly when water temperature exceeded 30 1C. Similarly, under hypoxia, the non-native G. tigrinus displayed higher survivorship than the native G. zaddachi. Since both stressors, high water temperatures and low concentrations of dissolved oxygen, reg- ularly co-occur in summer, the ability to withstand both stresses is an important adaptation to many coastal habitats. This could be mediated by the production of chaperones that stabilize proteins in many organisms when environmental conditions become suboptimal and metabolism is reduced (Velazquez and Lindquist, 1984; David et al., 2005; Anestis et al., 2010). However, to date, the physiological background of the stress response in the two gammarids is unknown. Should heat waves become more frequent and summer maxima temperatures more extreme in the Baltic region, shallow water habitats along the Finnish coast may become devoid of the native G. zaddachi during summer months facilitating the further spread of the non-native G. tigrinus.

In our experiments, stress, i.e. acute deviation from ambient

conditions, consistently provoked a stronger reaction in native as

compared to non-native species. The latter reduced their respira-

tion less and suffered lower mortality than the first. Our findings

therefore support the hypothesis that stress tolerance is a

common trait in introduced marine invertebrates that may be a

pre-requisite for the survival of transport and the establishment

in a new habitat. We cannot decide whether the tolerance width

we observed for the non-native species was (i) a species-specific

pre-adaptation, which is typical for organisms that originate from

harsh and/or fluctuating environments (i.e. a species-specific trait), (ii) the result of selective adaptation during the process of invasion (i.e. a population-specific trait) (sensu Richards et al., 2006), (iii) due to a broad genetic diversity in the non-native population caused by repeated introductions (i.e. a population- specific trait) (Kelly et al., 2006; Roman, 2006) or a combination of these. Intra-specific comparisons among native and non-native populations with regard to tolerance width could provide invalu- able insights to this question. Furthermore, it should be noted here that if stress tolerance can be acquired during relatively short time scales (scenario ii), this would limit the usefulness of this criterion for risk assessment.

Despite these limitations, our results suggest that a wide stress tolerance of a species in its native range could be a promising candidate for a catalog of traits helping to profile potential invasive species (e.g. Ricciardi and Rasmussen, 1998). This could be expected for species inhabiting strongly fluctuating coastal habitats such as the shallow subtidal, the intertidal as well as estuaries. These systems can be influenced by a multitude of physical forces and associated stressors: wave impact, storms, tides, rainfall, sediment transport, frost, intense sunlight and UV-radiation. The often substantial, predictable and unpredict- able fluctuations in abiotic conditions on very short time scales would select for a wide tolerance range rather than the capacity to rapidly acclimatize. Some of the most notorious invasive species in their native range live in intertidal/ shallow subtidal (e.g. Mytilus spp., C. gigas, Enteromorpha spp., Fucus spp.) or in estuarine habitats (e.g. Dreissena polymorpha, Eriocheir sinensis, Mnemiopsis leidyi, Asterias amurensis) (IUCN, 2010).

Even if broad stress tolerance is a typical trait of successful invaders it is certainly not the only pre-requisite. A candidate species must possess a whole suit of traits such as a life stage which is suitable for long distance transport by one of the common vectors (currents, ballast water, aquaculture material, ship hull, rafting, etc.), the aforementioned abiotic tolerance, and the pre-adaptation or phenotypic plasticity (Richards et al., 2006) to cope with new biotic interactions (competitors, consumers, parasites).

Acknowledgments

This study was run as part of the international research and training program GAME (Global Approach by Modular Experi- ments) of IFM-GEOMAR, which is generously funded by the Mercator Stiftung (Essen, Germany) and the foundation of the Christian-Albrechts Universit ¨at zu Kiel (Kiel, Germany). We would like to thank two anonymous reviewers and four guest editors for their valuable suggestions. Their input improved the quality of the manuscript substantially.

Appendix A. Supplementary material

Supplementary data associated with this article can be found in the online version at doi:10.1016/j.envres.2011.05.001.

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