Acidification alters seawater chemical speciation and biogeochemical cycles of many elements and compounds. One well-known effect is the lowering of calcium carbonate saturation states, which impacts shell-forming marine organisms Decreases in the availability of carbonate building and maintaining their shells or skeletons. n shell formation may leave less energy for other biological processes like growing, reproducing or responding to other stresses. Many shell-forming and carbonate ion concentrations; conditions ed for the coming decades may prove very stressful to these calcifying organisms. Corals, bivalves (such as oysters, clams, and mussels), pteropods (free-swimming snails) and certain The biological impacts of oceanacidification will vary, ferent groups of marine organisms have a wide range of sensitivities to changing seawater Impacts from oceanacidification at any life stage can reduce the ability of a population to m losses due to disturbance or stress. Therefore oceanacidification will also impact various economic sectors (eg: fisheries, aquaculture, tourism and coastal communities) and may also have heavy indirect effects on much broader segment of the world economy and production. Oceanacidification represents yet another stress on marine environments that may endanger the flow The atmospheric CO 2 can be controlled by
Oceanacidification has a range of impacts on biological systems. It will change the competition between marine plankton species in favour of those that rely less on calcium (Orr et al. 2005;Riebesell et al. 2000), it will negatively affect shellfish (Gazeau et al. 2007;Spicer et al. 2007), it will impact on fish (Ishimatsu et al. 2004), it may benefit highly invasive non- native algal species (Hall-Spencer et al. 2008), and it will reduce coral calcification (Hoegh- Goldberg et al. 2007). However, while the initial impact of oceanacidification is relatively clear, the eventual impact depends on the complex interaction of many species. The estimation of resulting changes in economic values, which generally derive from the higher trophic levels (e.g., top predator fish, marine mammals, sea birds), is therefore also pervaded by uncertainty. Coral reefs are an exception in that the impact of oceanacidification is relatively well understood and they have a range of direct and indirect use values for humans (e.g., coastal protection, fisheries, recreation, amenity). It is for these reasons that this paper is limited to assessing the economic impact of oceanacidification on coral reefs.
Oceanacidification will have complex effects on marine organisms. One predicted effect is a reduction in the calcification rate of some species. Coastal marine estuary ecosystems are highly biologically productive yet are more sensitive to changes in pH than is the open ocean due to their shallower depth, lower salinity, and lower alkalinity (Miller et al., 2009). Mollusks like mussels and oysters, which support valuable marine fisheries, may be especially vulnerable to acidification, resulting in economic losses and social disruptions in societies that depend on marine calcifiers (Cooley & Doney, 2009). Changes in fishery industries due to acidification are likely to disproportionately affect developing nations that often rely more on marine-related economic activities than developed nations (Cooley et al., 2009). Projected revenue losses from mollusk harvest declines due to acidification range from $0.6-2.6 billion through 2060 (Cooley & Doney, 2009). In addition, projections of lost coral reef area in Year 2100 due to acidification range from 16% to 27% and translate to economic losses of as much as $870 billion in Year 2100 (Brander et al., 2009).
Furthermore, our emphasis for the remaining sections of this monograph is mainly on the global coastal ocean, which was motivated when we started this work by the fact that little attention had been given to the role of the carbon cycle and associated elements of the coastal ocean in the context of global environ- mental change, despite the fact that this region is highly important at the global scale (Gattuso et al ., 1998; Ver et al ., 1999; Chen et al ., 2003, 2008; Mackenzie et al ., 2005) . At present time, the global coastal ocean (i.e. bays, estuaries, lagoons, banks, and continental shelves) constitutes only 7% to 8% of global ocean surface area, but is the location where approximately 10–30% of the world’s marine primary production occurs (Jahnke, 2010) . Its further importance lies in the fact that 80% of fluvial suspended sediment inputs, containing reactive nutrients, from the land to the sea are deposited in coastal environments . In addition, 85% of organic C and 45% of inorganic C are buried in the sediments within this region (Gattuso et al ., 1998; Wollast, 1998; Chen et al ., 2003; Mackenzie et al ., 2005) . Furthermore, the coastal ocean provides a myriad of critical services to a large proportion of the human population in terms of food, nutrition, and income from the fisheries and tourism economies . In addition, the global coastal ocean supports nearly 50% of all species on the planet and helps sustain that life through providing 20% of animal protein and five percent of the total protein in the human diet . Of that resource, coastal ocean waters out to 200 nautical miles supply 99% of the worldwide annual commercial fish catch .
Meanwhile, the estimated economic costs amount only to a very small fraction of world GDP or the total expected economic damage of climate change. The share of the mollusk loss to the world GDP in 2100 is 0.018% based on van Vuuren et al.’s GDP projections and 0.027% based on Gaffin et al.’s GDP projections. These figures correspond to 1.0% and 1.5% of the total expected damage of climate change (which corresponds to 1.8% of world GDP excluding the impacts of oceanacidification) based on the equation 15 from Tol’s (2009) meta-study on the economic impact of climate change impact combined with by the estimated increase of global surface temperature by the end of the 21 st century under A1B scenario (2.8°C). Estimates of the social cost of carbon would increase more that 1.8% if the effect on mollusks is included, because the ocean acidifies faster than the atmosphere warms. Nonetheless, it would be fair to argue that the recognition of negative effects of oceanacidification on mollusks would not have significant bearings on the discussions of global CO 2 emission policy. However, it is of course the case that the mollusk fisheries constitute
Ocean warming and acidification are predicted to have far- reaching impacts on marine biodiversity and fisheries before the end of this century (Doney et al., 2012; Hoegh-Guldberg et al., 2014; Pörtner et al., 2014; Phillips and Pérez-Ramírez, 2017). Indeed, changes in geographical distributions and phenology that are consistent with climate change predictions have already been observed in many marine species (Simpson et al., 2011; Poloczanska et al., 2013). Marine heatwaves are also increasing in frequency and duration, with devastating effects in some ecosystems (Hughes et al., 2018; Richardson et al., 2018; Stuart-Smith et al., 2018). Significant biological effects of contemporary oceanacidification are more difficult to ascertain, but there is evidence that anthropogenic carbon dioxide is already exacerbating the impacts of corrosive seawater incursions that occur during periodic upwelling events in some coastal ecosystems (Bednaršek et al., 2014; Ekstrom et al., 2015; Chan et al., 2017). Predictions about the impacts of future ocean warming and oceanacidification on marine organisms are primarily derived from laboratory experiments that rear animals at projected future temperatures, oceanacidification conditions, or both, and assess the performance of the animals in these conditions compared with a current-day control (Kroeker et al., 2013; Boyd et al., 2018). Such experiments usually last for several weeks, or maybe months at the most. Thus, while these experiments are informative about the potential effects of climate change conditions on the performance of marine organisms, they do not account for adaptation that might occur over the timeframe that the environmental changes are actually occurring (Riebesell and Gattuso, 2015). Therefore, to provide robust predictions about the likely impacts of ocean warming and acidification, it is also necessary to assess the adaptive potential of marine species to these environmental changes (Munday et al., 2013; Sunday et al., 2014).
Food webs are complex systems, often connecting a large number of species through the transfer of energy from the bottom to the top. Food webs are rarely linear and predators usually rely on a number of prey species. Key species form a central part of food webs. In the Southern Ocean food web Antarctic krill, Euphausia superba, is a key species. Krill forms a vital part in the diet of many animals in the Southern Ocean such as seals, seabirds, penguins and whales (Atkinson et al. 2004). Apart from sustaining higher trophic levels, and being a target for fisheries itself, krill products are also of commercial interest for humans and in animal nutrition (Virtue et al. 1993; Martin et al. 2006; Kawaguchi and Nicol 2007; Simmons and Isaac 2007).
For several reasons, oceanacidification has seri- ous implications for the type of policy interventions required to control climate change. First, since oceanacidification is exclusively driven by CO 2 , as opposed to climate change which is also caused by other greenhouse gases, the additional cost associ- ated with CO 2 emissions due to oceanacidification changes the trade-offs between the reductions of greenhouse gases. Second, the absorption of CO 2 by the oceans and the impact of oceanacidification occur over a short time scale, whereas the warming of the atmosphere substantially lags behind the build-up of greenhouse gases in the atmosphere. This changes the dynamics of optimal emission control. Third, the consideration of oceanacidification also has implica- tions for the choice of policy instrument for the control of climate change. Climate change may be countered by geo-engineering, but oceanacidification would continue unabated and may even accelerate if sul- phur particles are used to cool the planet. Therefore, valuing the impact of oceanacidification will not only increase the estimates of the Pigouvian tax required to achieve efficient greenhouse gas emissions abate- ment 22 , but it will affect other trade-offs and policies
both, in a 3-day closed system experiment measuring chemical bioerosion via the alkalinity anomaly technique (Wisshak et al. 2013), and in a 10 days experiment with flow-through conditions measuring total bioerosion rates by recording changes in buoyant weight (Wisshak et al. 2012). Most recently the relationship of C. orientalis bioerosion rates with global climate change was again confirmed by experiments carried out by Fang et al. (2013) in longer exposures of 8 weeks in a series of combined acidification and warming scenarios. Our experimentally obtained rates of chemical bioerosion of C. orientalis and C. celata clearly reveal the similarity in response (Fig. 4). Thereby, rates in the tropical C. orientalis are roughly five times higher than those determined herein for the cold- temperate C. celata, but the degree to which pCO 2
Fishes in early life history stages are presumably more vulnerable to changing oceanic conditions. Higher metabolic demand throughout metamorphosis coupled with increases in diffusion surface area with development of gill lamellae at juvenile stages could account for observed sensitivity to alterations in ocean chemistry (Kikkawa et al. 2003; Ishimatsu et al. 2004; Melzner et al. 2009); however, whether sensory disruptions observed in larval fishes are present in adult stages has yet to be determined. Accurate interpretation of environmental sensory cues is essential for critical adult behaviours including location of foraging (Chave 1978; Heyman et al. 2001) and breeding sites (Tesch 1967; Dittman and Quinn 1996). Therefore, it is important to test the tolerance of adults to elevated CO 2 in order to predict the dynamics of future marine fish
meso-predators and the outcome of interactions between P. fuscus and their prey. Further investigations are required to determine and test the effects of changed behaviour on predator success in the field and the impact this has on prey populations. However, the extreme attraction to predator odour by larval fishes under acidified conditions, compared to the relatively small avoidance of prey cues by meso-predators detected in this study suggest that it is unlikely that negative effects on predators will fully compensate for the increase in mortality rates of larval fish returning to the reef in an acidified ocean.
On the 29 th of January 2013, ten “Kiel Off-Shore Mesocosms for Future Ocean Simulations” (KOSMOS, M1-M10; ) were moored by research vessel Alkor in Gullmar Fjord on the Swedish west coast (58˚ 15’ 50” N, 11˚ 28’ 46” E). Study site, key events, deployment, and mesocosm manipulation procedures are described in detail in the abovementioned overview paper . In brief, each mesocosm was composed of an 8 m tall floatation frame and an 18.7 m long cylindrical polyurethane bag with a diameter of 2 m. The bags were folded and installed in the floatation frame before mesocosm deployment by Alkor. After deployment, bags were unfolded and lowered underwater to allow water exchange with the fjord. Water inside the bags was isolated from the fjord water by attaching 2 m long conical sediment traps to the bot- tom and pulling the upper end of the bag about 1.5 m above the surface [13,14]. The mesocosm bags with the attached sediment traps reached ~19 m deep after the closing procedure.
Experimental design. Experiments were conducted in two sealed respiration incubation chambers with a volume of 27 litres (Submersible Photosynthesis-Respiration System, Aquation Pty Ltd, Umina Beach, Australia), deployed on the seabed over patches of MPB. The chambers were made of UV-transparent acrylic with 25 cm diameter (chamber diameter) stainless steel sleeves on the base for insertion into the sediment. The chambers pumped acidified or control water from a larger long-term CO 2 enrichment experiment (Antarctic Free Ocean
environmental changes, including seawater acidification (for examples see Zondervan et al., 2002; Schippers et al., 2004). Limitations of these small enclosures, however, are unnatural physico-chemical and surface attachment effects. Significant changes to community composition and to rates of metabolism, independent of the treatment being tested, may occur within 24 hours, in what is known as the “bottle effect” (ZoBell and Anderson, 1936). Large mesocosms, especially located in situ rather than in the laboratory, can help reduce the bottle effect, allowing a pseudo-environmental investigation in which chosen parameters can be controlled. Bacterioplankton ecological studies have made use of mesocosms to assess: bacterial abundance and activity (Lebaron et al., 2001), microbial community dynamics and changes in genetic diversity (Schäfer et al., 2001), microbial organic matter production (Engel et al., 2004); to follow bacteria-virus interaction in a bacterial community during the course of a viral infection (Hewson et al., 2003), and effects of increased pCO 2 on bacterial
This is the first study to investigate the potential effects of near-future oceanacidification on skeletal development in marine fishes. Skeletal tissue is not expected to be susceptible to reduced ocean pH, because it is primarily formed of calcium phosphate minerals (hydroxylapatite) and complex cartilaginous tissues (Witten et al. 2010), rather than calcium carbon- ate. However, it is still possible that changes to extra- cellular concentrations of bicarbonate and non- bicarbonate ions due to acid –base regulation in hypercapnia-exposed fish could affect skeletal devel- opment. As expected, we observed no consistent effects of exposure to elevated CO 2 on the size of rep-
A third consideration emerges from the differential influence of acidification on shell area and body mass. Desiccation during low tide joins predation as one of two dominant agents of mortality in newly settled juveniles (Gosselin and Qian, 1997; Bownes and McQuaid, 2009). Although drying rates of emergent organisms depend on multiple factors tied to evaporative water loss and heat gain, they are strongly dependent on the ratio of surface area to mass. Projected areas determined in the present study are not precisely coincident with surface areas, but provide a relevant index for scaling rates of thermal input. Likewise, dry tissue mass correlates strongly with wet tissue mass, and thus provides a measure of how much dehydration could take place before an individual dries out during low tide. Area/mass ratios in day-8 larvae therefore yield a first-order estimate of susceptibility to desiccation at the larval–juvenile transition. The dramatic 40% increase in this ratio for late-stage larvae reared in elevated-CO 2 conditions suggests the
We are grateful to D. Luquet, L. Gilletta and J.-Y. Carval for sea urchin sampling in the field. We thank T. Lepage for providing P. lividus transcript sequences and C. M. Moore for advice regarding statistics related to molecular data. The IAEA is grateful for the support provided to its Marine Environment Laboratories by the Government of the Principality of Monaco. S.D. is funded through the Linnaeus Centre for Marine Evolutionary Biology at the University of Gothenburg and supported by a Linnaeus grant from the Swedish Research Councils VR and Formas. This work is a contribution to the ‘European Project on OceanAcidification’ (EPOCA), which received funding from the European Community’s Seventh Framework Programme (FP7/2007-2013) under grant agreement no. 211384.