In 2015, we collected more than 60 000 scavenging amphipod specimens
during two expeditions to the Clarion–Clipperton fracture zone (CCZ) in the
Northeast (NE) Pacific and to the DISturbance and re-COLonisation (DisCOL)
experimental area (DEA), a simulated mining impact disturbance proxy in the
Peru Basin in the Southeast (SE) Pacific. Here, we compare biodiversity patterns
of the larger specimens (
The abyssal deep sea (3000–6000 m) represents the largest ecosystem on
the planet, with the abyssal seafloor covering approximately 54 % of the
Earth's solid surface (Gage and Tyler, 1991; Rex et al., 1993). Since it is
one of the least investigated ecosystems, there are still extensive gaps in
our knowledge of deep-sea fauna (German et al., 2011). Marine research has
thus far focused on coastal areas, hydrothermal vents or chemosynthetic
habitats, whereas open-ocean abyssal plains have been less extensively
investigated (Ramirez-Llodra et al., 2010). This is unsurprising given the
challenges of sampling this remote environment, which is impeded by several
confounding factors. For example, deep-sea sampling is both financially
expensive and labour intensive and, furthermore, constrained by the
challenge of deploying equipment at low temperatures (0.01–4.0
In the traditional view of the deep sea, the abyss was considered to be homogeneous and many species were thought to have large biogeographical ranges, their dispersal aided by an apparent lack of barriers (Sanders, 1968). This hypothesis was challenged by the discovery of chemosynthetic habitats, e.g. hydrothermal vents (Lonsdale, 1977), cold seeps (Paull et al., 1984), seasonal fluctuations in primary productivity (Billett et al., 1983) and erratic whale falls (Smith et al., 1989). All of this research has demonstrated that the deep sea is a heterogeneous environment and is controlled by many factors, including particulate organic carbon (POC) flux, water depth, flow regime, current circulation, seafloor topography (Laver et al., 1985) and also historical factors, e.g. the opening of ocean basins (i.e. rifting), sea level rise and fall, and periods of deep-sea anoxia (Smith et al., 2006). All of these can result in a mosaic of different communities (Levin et al., 2001), many of which do not follow a latitudinal gradient (Brandt et al., 2007).
It has also been established that dispersal ability of species on the one hand, and their actual geographic and bathymetric distribution range on the other hand, are not always linked and are often dependent on habitat suitability, fragmentation and ecological flexibility (Lester et al., 2007; Liow, 2007). Therefore, although the deep seafloor includes some of the largest contiguous features on the planet, the populations of many deep-sea species are spatially fragmented and may become increasingly so with continued human disturbance (Hilário et al., 2015).
In the last decade, there has been a higher demand for exploitation of deep-sea resources, e.g. rare-earth element (REE) extraction (such as those concentrated in manganese nodule provinces) (Ramirez-Llodra et al., 2011). As a result, ecologists are increasingly asked to assess the ecological risks of these mining activities and to provide sustainable solutions for its mitigation, in order to prevent adverse changes to the deep-sea ecosystem (International Seabed Authority, 2017).
Glover et al. (2001) showed that abyssal sediments can contain high biodiversity, with more than 100 species of meiofaunal invertebrates (e.g. nematodes, copepods) and protists (e.g. foraminifers) found every square metre. Despite this, our knowledge of the deep-sea ecosystem structure and functioning is still limited, and there is a paucity of data about the distribution, drivers and origins of deep-sea communities at global scales. This is especially true for deep-sea invertebrates, including Amphipoda (Barnard, 1961; Thurston, 1990).
Although recent morphological and molecular studies have shed new light on
the distribution and habitat niches of certain bentho-pelagic amphipods
(e.g.
Here, we present distribution patterns of scavenging deep-sea amphipod communities, with the first comparisons of their biogeography and community structures in two oceanic basins. These two basins are the research areas for simulating and studying the anthropogenic impacts of deep-sea nodule mining.
We are investigating whether there are differences and similarities in the species compositions of the two basins (e.g. richness, abundances) and further exploit a disturbance experiment to compare the biodiversity of this mining impact proxy to the undisturbed reference areas. We discuss the possible implications of our findings, aiming to use them to formulate recommendations regarding the pending deep-sea mining of manganese nodule activities in the NE Pacific ecosystem.
We investigated the amphipod communities of two oceanic basins (Fig. 1):
(i) the Clarion–Clipperton fracture zone (CCZ,
Geographic locations of the two study areas, the
Clarion–Clipperton fracture zone (CCZ) (Northeast Pacific) and the
DISturbance and re-COLonisation (DisCOL) experimental area (DEA) (Peru
Basin, Southeast Pacific). There are nine Areas of Particular Ecological
Interest (APEIs) in the CCZ region, which are illustrated by
In 2015 (26 years after the first impact in the DEA in 1989), two research
expeditions with the RV
Amphipod samples were taken from the CCZ and DEA using a free-fall lander
(120 cm
The baited trap was deployed eight times across the CCZ at a depth range of 4116–4932 m (samples C1–C8) and five times in the DEA at a depth range of 4078–4307 m (samples D1–D5; Fig. 1, Table 1). In the CCZ, we sampled within four different contractor licence areas (Table 1) to obtain a pre-disturbance baseline and to then compare it with one of the nine protected APEIs around the CCZ. In contrast, in the DEA sampling was conducted once within the disturbed area (D1), twice 10 km away (D2, D3) and twice 40 km away (D4, D5) from D1 in four surrounding reference areas (see Fig. 1).
Column 1 provides the new the station codes used in this paper in Figs. 1, 3, 4, and 6 and Table 3. The original station codes from cruises SO239 and SO242-1 are in column 2. Depth refers to the trap depth at the time of deployment.
On recovery of the lander, all traps were disconnected and placed in
pre-cooled (4
Detailed sorting and identifications were performed using the morphological
species concept (Futuyma, 1998) and the keys of Schulenberger and Barnard (1976) and Barnard and Karaman (1991) to separate the samples into
taxonomic “morphotypes”. The larger fraction (
Of the 60 000 specimens, those with a size of less than 15 mm in length were excluded from the analysis because these were mostly juveniles and their morphological differences were not sufficiently pronounced to allow an accurate identification to the species or even genus level. Some pelagic amphipods were collected accidentally and omitted. Finally, genera containing multiple (and as yet) unidentified species have been summarized as “spp.”.
Our null hypothesis (
Secondly, to compare the beta biodiversity, we estimated the variability of the community compositions between sites. The Bray–Curtis dissimilarity metric (Bray and Curtis, 1957) was used to calculate differences between community compositions based on species densities, and the results were then visualized in 2-D using a non-metric dimensional scaling (NMDS) plot. The ANOSIM function in the vegan package of R (R Core Team, 2013; Taguchi and Oono, 2005) was used to test the statistical significance of the differences in species compositions between the two study areas.
In total, 6916 scavenging amphipods (
Histogram showing the species assemblage for the scavenging community in the Clarion–Clipperton fracture zone (CCZ) (black) and the DisCOL experimental area (DEA) (grey). The abundances of 17 morphotypes are shown.
Distribution and abundances of morphotypes across the Clarion–Clipperton fracture zone (CCZ) and DisCOL experimental area (DEA). For the numerical values the following format has been used: normal font is shared, italic font is DEA only and bold font is CCZ only.
There are eight morphotypes shared between the basins:
Two morphotypes were found only in the CCZ (
Due to differences in allocated ship times (the CCZ cruise being 52 d and the
DEA cruise being 29 d), the trap deployments were not identical, making
it necessary to check the effect of the different deployment times. The
resulting catch-per-unit-effort (CPUE) plot (Fig. 3) shows that there is
no statistically significant correlation between the length of time the trap
was at the seafloor and total number of amphipods caught (
Catch per unit effort (CPUE), illustrating the correlation between sampling time and number of individuals collected. Only the longer than 15 mm fraction was included here.
The rarefaction results (Fig. 4) show that the curves for nine stations
reach a plateau, indicating that sampling effort was sufficient to assess
diversity levels. These include all CCZ stations except C7. In contrast,
four of the five curves for the DEA (stations D1, D2, D4 and D5) are
unsaturated. A higher number of different species were collected at D1 and
D2; however, many of these were singletons or doubletons, with
Species rarefaction curves for each of the 13 trap stations across both areas, the Clarion–Clipperton fracture zone and the DisCOL experimental area. Only individuals longer than 15 mm were considered here.
Figure 5a and b show that the scavenging community in the CCZ is dominated
by three species,
Relative species abundances in the Clarion–Clipperton fracture zone and the DisCOL experimental area. These abundances represent the longer than 15 mm subsample of the scavenging amphipod community.
Comparison of biodiversity calculated using the Simpson
index (
The NMDS shows that the communities of the two basins are dissimilar
(ANOSIM:
NMDS plot showing the beta biodiversity (dissimilarities or similarities) for each of the 13 amphipod trap sampling stations associated with the two basins, Clarion–Clipperton fracture zone (CCZ) (black) and the DisCOL experimental area (DEA) (red). Data are supported by a low stress value of 0.01.
Although the most recent and comprehensive analysis of the animal diversity of the world's oceans estimates a total of less than a million species over all depths (Appeltans et al., 2012), it is not currently known how many species inhabit the deep sea. Over 7000 marine amphipod species have been found below 2000 m. These numbers are reduced to 173 known species, 87 genera and 37 families at depths below 3000 m, and 100 known species, 66 genera and 31 families are known to occur below 4000 m (Vader, 2005; Brandt et al., 2012).
The superfamily Lysianassoidea constitutes an important part of the abyssal amphipod fauna. Also, in our sampling, lysianassoid amphipods were collected in large numbers (99 % of the samples taken in both basins). As a superfamily, they comprise 23 % of all the species found below 2000 m, 35 % of the species found below 3000 m and 31 % of the species found below 4000 m (Brandt et al., 2012).
Many species in the Lysianassoidea occur in multiple abyssal basins and
some even have worldwide distributions (Thurston, 1990). Despite the Ocean
Biogeographic Information System (OBIS) database containing 615 650 records
of Amphipoda, many of these are shelf or pelagic species, with very few
records from the CCZ and DEA (OBIS, 2019). Here, we provide additional data
for the known bathymetric range of the seven amphipods, which we have
identified to species level:
While we only sampled
Despite the sampling campaign in the CCZ being twice as long as the DEA, the number of individuals and species collected does not correlate positively with deployment effort. We posit that this is rather an effect of abiotic and organic factors, such as the productivity-driven gradients in the CCZ, which decrease from east to west and from north to south (Hannides and Smith, 2003), and also the productivity differences between both basins.
Figure 5a and b clearly show that the DEA scavenging community has reduced
abundances of all species, including
The assemblages of the two basins have some overlap in their biodiversity (as is exemplified by the eight shared morphotypes). However, the sampling stations and the two basins as a whole are heterogeneous in their species compositions.
Thus, we can observe some negative influence (possibly attributed to the
disturbance in the DEA) on the scavenging amphipod community. This reduced
biodiversity is reflected in the higher Simpson index (
To explore whether this reduced diversity in the DEA was truly an artefact of
the simulated disturbance,
In the CCZ, the APEI (C8) shows a moderate level of biodiversity (
Within the DEA, the lowest biodiversities are calculated at the site of the disturbance (D1) and south of it (D2; Table 3), indicating that the reduced biodiversity in the DEA could indeed be caused by the simulated disturbance in 1989 (Thiel, 1992).
The highest abundances in the DEA were collected from station D5 (
Scavenging amphipods are resilient and dispersive, but most importantly they are highly mobile (Ingram and Hessler, 1983; Lörz et al., 2018). Often driven by their search for erratically deposited feeding opportunities (Smith et al., 1989), they are probably less constrained by local environmental abiotic conditions and seafloor topography.
Beta diversity can be regarded as the dissimilarities in species composition
between spatially different communities. As an indication for beta
biodiversity, the NMDS (Fig. 6) shows a significant separation in the
similarity index between the two basins (ANOSIM
In the CCZ, stations C1, C2, C3, C4 and C5 show a different Bray–Curtis index in comparison to stations C6, C7 and C8 (Fig. 6). The CCZ is a geomorphologically very heterogeneous region, with seamounts of 200 m altitude running from north to south. A barrier of this height would be expected to affect sedimentation rates, nodule presence and currents. Furthermore, the difference in depth from the eastern edge (3950 m) and the western edge (5150 m) is more than 1200 m. These combined factors very likely give rise to different trends in species composition (Glover et al., 2016). However, since it has been established that bentho-pelagic amphipods are less sensitive to such barriers (Havermans, 2016), at this stage other biotic (e.g. the productivity gradient) and abiotic factors causing this separation cannot be excluded as alternative explanations.
Whilst the NMDS (Fig. 6) illustrates a visual separation of the two basins, there is also some similarity in the amphipod fauna between the two areas (as is obvious by the eight shared species), indicating that the dispersal extent for these eight species might be up to at least 3000 km. However, this hypothesis will need to be confirmed with subsequent molecular analyses.
Abyssal amphipods have been shown to be able to travel actively at speeds of
almost 4 cm s
However, it is apparent that the dispersal of abyssal amphipods is not
always contingent on current direction but on passive dispersal.
Amphipods can also be carried passively over long distances by strong
currents (e.g. the circumpolar current of the Southern Ocean) (Laver et al.,
1985), but even weaker deep-sea currents have been suggested as a mechanism
for deep-sea dispersal of amphipods (e.g.
Recent research on
Unfortunately, in the absence of data on deep-sea currents in the study area, especially between the CCZ and DEA, it is not yet possible to fully explain the drivers and mechanisms of amphipod dispersal between these particular basins.
Higher abundances of scavenging amphipods were collected from the CCZ (3932 individuals) as opposed to the DEA (2984 individuals). Yet, we have identified more morphotypes in the DEA (15) than in the CCZ (10), indicating that the DEA is more speciose and thus more biodiverse.
However, although the DEA is more speciose, many of its morphotypes were collected in low abundances, with several of these being singletons or doubletons (collected from one or two sampling stations only). This is reflected in the rarefaction curves (Fig. 4), which indicate thorough sampling in the CCZ with all but station C7 reaching asymptotes. In contrast, four stations in the DEA (D1, D2, D4 and D5) are unsaturated. This pattern suggests first that the less abundant species that are present at fewer stations may not necessarily be rare species and second that there could be as of yet undetected biodiversity in the DEA.
Our preliminary (basin-scale) comparison of the scavenging communities of the two study areas shows that even if the DEA is a small-scale disturbance experiment, it is a very diverse area. Thus, the DEA is a well-chosen site for monitoring the impacts of disturbance and instrumental in its role as a proxy to assess impending mining activities in the CCZ.
At several stations in both basins, we collected amphipods in very high abundances (C1, C8, D3 and D5) (Table 2). Since biotic production is contingent on the sinking flux of particles from the euphotic zone (Sweetman et al., 2017), the biodiversity differences at each of the 13 stations could be driven by POC or erratic whale falls (Smith et al., 1989). However, not all feeding behaviour of scavenging amphipods is based on opportunistic or erratic availability of nutrients (Havermans and Smetacek, 2018). During future sampling campaigns, the POC of these amphipod sampling areas should be monitored, along with experiments on different types of food fall in addition to obtaining side-scan sonar and abiotic data. This will provide a more comprehensive view of the food types required for these species to thrive in the deep sea.
It is not clear from our results whether substrate type (i.e. nodule or non-nodule) has any effect on the amphipod communities (Smith and Demopoulos, 2003) since these kinds of data are only available for stations D3 and D4. To answer this question, resampling of the study areas in combination with an ocean floor observation system (OFOBS) (video or camera) is required.
Although our study only addresses the scavenging amphipod species longer than 15 mm, we have already found indications of a possible disturbance effect in the DEA. It is obvious that scavenging amphipods are only one of several bentho-pelagic impact-indicator groups. Other truly benthic groups such as sponges or less dispersive amphipods (e.g. collected by the EpiBenthic Sledge, EBS) may demonstrate an even more pronounced impact of mining activities and should be investigated in future studies.
With the application of molecular techniques to identify cryptic species (Delić et al., 2017), more realistic estimates of biodiversity can be obtained (Schön et al., 2012), improving our current knowledge of the biodiversity of this area. If these improved estimates of biodiversity also include cryptic species, it is possible that the biological impact of manganese nodule mining on amphipod and other deep-sea faunal communities may turn out to be even higher.
In summary, this study on the scavenging amphipod community of two abyssal oceanic basins has demonstrated that amphipods are present in high abundances across the CCZ and DEA, with eight shared species and some morphotypes possibly being unique to their respective basin.
Our results have indicated that the simulated mining experiment may have had
an impact on the biodiversity of these scavenging amphipods, as demonstrated
by the low alpha biodiversity of the DEA overall at the disturbance site
itself (D1) and the 60 % dominance of
Given the scarcity of sampling and industry experience of marine habitats at these depths, the formulation of effective regulations is challenging (International Seabed Authority, 2017). Nonetheless, our study provides the first results on possible effects of disturbance activities on the abyssal amphipod biodiversity of deep-sea basins.
Multibeam scan showing the location of the 78 track marks created by the plough harrow in the DisCOL experimental area to simulate manganese nodule extraction activity (D1).
Photograph showing the baited free-fall lander trap designed and deployed by RBINS. Equipped with an acoustic release transponder, flashlight, Novotech radio beacon and Posidonia positioning signal to monitor position at the sea floor and ascension through the water column.
Relief changes in the DisCOL Experimental Area. © GEOMAR-Helmholtz-Centre for Ocean Research Kiel, Germany. Acknowledgement: Anne Henke.
Simpson index was used for the calculation of alpha biodiversity as follows.
The data discussed in the
paper are available at
Biological samples pertaining to this paper are stored at the Royal Belgian Institute of Natural Sciences, Brussels, Belgium.
TP was responsible for the execution of fieldwork during expedition SO242/1, onboard and subsequent taxonomic work, data analysis, and writing of the manuscript. HR was responsible for the execution of fieldwork during expeditions SO239 and SO242/1 and onboard identifications of the amphipods. CD'UD'A was responsible for providing taxonomic expertise and the reading, editing and approval of the manuscript. KM, IDM, SD and IS were responsible for the design of this study and the reading, editing and approval of the manuscript.
The authors declare that they have no conflict of interest.
This article is part of the special issue “Assessing environmental impacts of deep-sea mining – revisiting decade-old benthic disturbances in Pacific nodule areas”. It is not associated with a conference.
The authors would like to acknowledge the Belgian Science and Policy Office
(BELSPO) and the German Federal Ministry of Research and Education for
their funding. We would also like to thank the crew of the Research Vessel
“
This research has been supported by the JPI-Oceans project “Mining Impact” (BELSPO grant no. BR/15/MA/JPI-DEEPSEA2).
This paper was edited by Ann Vanreusel and reviewed by Charlotte Havermans and Cene Fiser.