Keywords

11.1 Introduction

Marine ecosystems throughout the world ocean have been damaged by human activities, and coral reefs have suffered especially severe impacts (Halpern et al. 2008). The challenge for coral-reef scientists is to determine the strongest causal pathways to degradation. Measuring the relative contributions of proximate and ultimate candidate-causes is not merely an academic exercise; the prescriptions for mitigating and reversing reef degradation differ depending on the scales, hierarchical levels, and identities of those causes. With limited resources available to conserve coral reefs, it is imperative that time, labor, and funds be devoted to corrective measures that will yield the maximum benefits .

Because reefs are geological as well as ecological entities, the physical sciences have been integral to their study from the start. Our understanding of how coral reefs operate strongly emphasizes physical drivers (e.g., Roberts et al. 1992; Hubbard 1997; Montaggioni and Braithwaite 2009). A few basic examples highlight the physical control of biological processes: (1) reef development is limited to latitudes warmer than the 18 °C winter-minimum isotherm (Dana 1843; Johannes et al. 1983; Kleypas et al. 2001); (2) upwelling driven by oceanic gyres restricts reef development off the west coasts of continents (Birkeland 1997; Hubbard 1997), and inimical waters suppress reef development on smaller scales (Neumann and Macintyre 1985; Hallock and Schlager 1986; Ginsburg and Shinn 1994); (3) antecedent topography and fluctuating sea level determine the growth and form of reefs and their scope for vertical accretion (Darwin 1842; Neumann and Macintyre 1985); (4) light and wave exposure combine with topography to create the biological zonation of reefs (Adey and Burke 1977; Geister 1977; Woodley et al. 1981; Hallock and Schlager 1986; Hubbard 1988; Acevedo et al. 1989; Graus and Macintyre 1989; Murdoch 2007); and (5) climatic fluctuations set the tempo and mode of reef development (Precht and Aronson 2004; Precht and Miller 2007; Toth et al. 2012, 2015). Today geology and paleobiology are helping us to distinguish natural from anthropogenic perturbations of coral reefs and to understand the scales at which those perturbations occur (Aronson 2007). An emerging theme—and the subject of this review—is the overriding influence of physical forcing in the recent, worldwide degradation of reefs.

What is signal to an ecologist is largely noise to a paleontologist. Fossil deposits from shallow, soft-bottom facies are often temporally and spatially averaged, obscuring the record of short-term variability on which the science of ecology still nourishes itself, for better or worse (but see Kidwell 2001, 2007). Time averaging and transport are generally less problematic in interpreting coral-reef deposits than biotas buried in soft sediments, because for coral reefs the benthic assemblages themselves construct their sedimentary fabrics . Coral colonies have longer lifespans than most soft-sediment invertebrates, and their skeleton s are to some extent resistant to taphonomic degradation. Furthermore, because coral skeletons are made of calcium carbonate and, in most cases, the skeletal framework and entombing sediments are cemented, corals are often buried and preserved in place and in sequence , or at most subject only to minor transport. Even where reefs are uncemented, the coral assemblages are often autochthonous and sequential (Aronson and Precht 1997; Aronson et al. 2002, 2004, 2005; Wapnick et al. 2004; Greer et al. 2009). Fossil and subfossil reef deposits, therefore, provide an excellent record from which to understand the ecology of coral reefs in times past and thereby discern the time frame and effects of natural and anthropogenic perturbations on modern reefs (Pandolfi 1996; Greenstein et al. 1998; Pandolfi and Jackson 2001, 2006; Precht and Aronson 2006; Greenstein and Pandolfi 2008; Lescinsky et al. 2012; Toth et al. 2012; and many others).

In this chapter we use evidence from fossil and modern coral reefs to review and critically evaluate three related propositions that have attained enormous popularity among coral-reef ecologists: (1) localized human activity, specifically overfishing , has been the primary cause of the decline of coral populations; (2) the cascading, top-down effects of overfishing are currently limiting the recovery of coral assemblages; and, therefore, (3) local management actions are capable of promoting the resilience of reefs to climate change. We contrast these notions with the idea that physical drivers are the primary determinants of reef dynamics at scales not much larger than the scale of the reef or reef system and should be important considerations in management and conservation. We focus on the reefs of Florida , the Bahamas , and the Caribbean (henceforth collectively termed ‘the Caribbean’), which are significantly altered (Gardner et al. 2003; Schutte et al. 2010) and for which the historical, ecological, and paleobiological data are complete enough to draw reasonably firm conclusions.

11.2 Causal Connections in the Degradation of Caribbean Reefs

Aronson and Precht (2001a, 2001b, 2006) suggested that larger-scale factors, specifically climate change and coral disease (which is related to climate change ), were the primary causes of reef degradation throughout the Caribbean region over the preceding three decades. The elkhorn coral Acropora palmata had dominated the reef-crest and shallow fore-reef habitats at 0–5 m depth on windward -facing Caribbean reefs, whereas the staghorn coral A. cervicornis had dominated intermediate, fore-reef depths of 5–25 m and some back-reef and lagoon al habitats. A regional outbreak of white-band disease (WBD : Fig. 11.1) was the primary cause of the Caribbean-wide mass mortality of these congeneric corals from the late 1970s to the early 1990s. WBD is an infectious, bacterial syndrome that appears only to affect the acroporids (Gil-Agudelo et al. 2006; Weil et al. 2006; Vollmer and Kline 2008; Kline and Vollmer 2011; Gignoux-Wolfsohn et al. 2012; Sweet et al. 2014). Because in many locations A. cervicornis and A. palmata were the dominant occupants of reef substratum and the dominant constructors of framework , Aronson and Precht (2001a, 2001b, 2006) concluded that WBD had been the most important cause of coral mortality in the Caribbean in recent decades. Hurricanes , coral bleaching from anomalously high sea temperatures, and additional factors such as corallivory had played subsidiary roles in killing the Caribbean acroporids (ABRT 2005; Gardner et al. 2005). Bleaching and other diseases later killed massive corals, including the formerly abundant, framework-building Orbicella annularis species complex (McWilliams et al. 2005; Aronson and Precht 2006; Eakin et al. 2010; Toth et al. 2014; see also Rogers 2008; Miller et al. 2009; Rogers and Miller 2013).

Fig. 11.1
figure 1

White-band disease on otherwise healthy branches of Acropora palmata. Note the characteristic progression of the disease from the bases to the tips of the branches. The photograph was taken at Carysfort Reef in the Florida Keys during the summer of 2004 (Photo credit: William F. Precht)

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Meta-analyses subsequent to Aronson and Precht (2001a, 2001b) have supported their conclusions (Côté et al. 2005; Alvarez-Filip et al. 2009; Schutte et al. 2010). Furthermore, paleoecological studies of reefs in several locations around the Caribbean have demonstrated that the recent mass mortality of acroporids was a novel event in at least the last three millennia (Aronson and Precht 1997; Greenstein et al. 1998; Aronson et al. 2002, 2005; Wapnick et al. 2004; Lescinsky 2012). Epidemiological work has suggested, albeit obliquely, that rising sea temperatures were responsible for the devastating outbreak of WBD in the Caribbean (Kline and Vollmer 2011), providing a link to physical processes. More recently, Randall and van Woesik (2015) linked outbreaks of WBD to increased thermal stress associated with climate change. Outbreaks of some other coral diseases have also been tied to rising temperatures (Rosenberg and Ben-Haim 2002; Selig et al. 2006; Bruno et al. 2007; but see Lafferty et al. 2004).

Jackson et al. (2001), in contrast, asserted in a highly publicized review that the disruption of trophic cascades by overfishing was the most important cause of ecological degradation in shallow-marine environments worldwide. For coral reefs, the scenario was that overfishing reduced herbivory, releasing macroalgae , or seaweeds, to overgrow and otherwise outcompete corals for space (see also Pandolfi et al. 2003). The review by Jackson et al. (2001) was immediately welcomed by conservation groups and the popular media, who touted it as a visionary breakthrough in our understanding of human threats to marine life. With equal rapidity Jackson et al. (2001) drew fire from scientists who pointed out that overfishing was neither the only human assault on marine ecosystems nor necessarily the most significant one. Jackson and colleagues responded that they had never intended to imply a negligible role for other drivers of ecosystem degradation (Peterson et al. 2001). In reality, Jackson et al. (2001) had acknowledged the existence of other factors but had downplayed them.

This group of authors later moderated their stance on overfishing to a more pluralistic view of causality by including sedimentation and nutrient loading from terrestrial sources as another major threat to coral reefs (Bellwood et al. 2004; Kuntz et al. 2005; Jackson 2008). Some of them have recently regressed to their initial stance that overfishing of herbivores , specifically parrotfish , is far and away the primary cause (Jackson et al. 2014).

Terrigenous input has certainly been an important cause of degradation in some situations (Rogers 1990; Cortés 1994; Aronson et al. 2004, 2014; De’ath and Fabricius 2010). On the other hand, one top-down scenario of the impacts of overfishing on coral reefs included speculation of a strong, cascading trophic connection between shark-fishing and a high prevalence of infectious diseases in coral populations (Sandin et al. 2008). That causal chain has not been demonstrated. More convincing are data showing that populations of the corallivorous seastar Acanthaster planci are reduced on Pacific reefs that support more intact stocks of predatory fish (Dulvy et al. 2004; McCook et al. 2010).

Claims about the primacy of overfishing are a step backward from Hughes (1994), who argued that overfishing, the regional mass mortality in 1983–1984 of the herbivorous echinoid Diadema antillarum from an infectious disease , and direct coral mortality from a hurricane had combined to drive a phase shift from coral to macroalgal dominance on Jamaican reefs. The loss of herbivores has been particularly egregious in Jamaica (Aronson 1990; Hughes et al. 1999; Aronson and Precht 2000), so overfishing was considered an important ingredient in the transition to dominance by macroalgae . Whether or not the construct for Jamaica can be generalized to the rest of the Caribbean is an important question (Côté et al. 2013). In fact, mass coral mortalities permitted macroalgae to rise to dominance opportunistically even in some locations with reasonably intact fish assemblages. It also turns out, surprisingly in hindsight, that only a minority of reefs in the Caribbean actually became dominated by macroalgae (Aronson and Precht 2006; Precht and Aronson 2006; Bruno et al. 2009; Dudgeon et al. 2010; Schutte et al. 2010; Bruno et al. 2014).

There is no doubt that much of the world is overfished and that in some situations overfishing can have drastic, cascading impacts on marine ecosystems. The model of Jackson et al. (2001), that the loss of top predators to overfishing fundamentally alters marine food webs and is the primary impact of human activity, works well for ecosystems with strong top-down trophic connections, such as kelp forests (Estes and Duggins 1995; Shears and Babcock 2003; Estes et al. 2011), but its applicability is far from universal. The overfishing hypothesis has been questioned or refuted for seagrass beds, oyster reefs, pelagic ecosystems, and some kelp forests, as well as coral reefs. In these cases, top-down trophic connections play a minor role in community structure , are too weak to respond substantially to the restoration of fisheries, or are complicated by bottom-up effects and other causes (Boesch et al. 2001; MacKenzie 2007; Waycott et al. 2009; Foster and Schiel 2010; Condon et al. 2012). Like the issue of terrigenous input, the question is not whether overfishing can be important on some coral reefs under some circumstances—because clearly it can—but what is its relative contribution to the overall decline of coral populations and coral reefs, and on what spatio-temporal scales?

Because the two Caribbean species of Acropora are now rare to the point of being threatened or endangered (Precht et al. 2004; ABRT 2005; Hogarth 2006; Carpenter et al. 2008; Aronson et al. 2009a, 2009b), accurate knowledge of the timing and causes of their decline is critical to their effective management and conservation (National Marine Fisheries Service 2015). Hughes et al. (2010) presented a history of the causes of decline of the acroporid corals in the Caribbean, which included the following language:

[T]wo meta-analyses of the loss of structural complexity of Caribbean reefs between 1969 and 2008 [Alvarez-Filip et al. 2009] and of coral cover from 1971 to 2006 [Schutte et al. 2010] have proposed that an unreported epidemic of white band disease [emphasis ours] killed off most branching staghorn and elkhorn corals across the region in the1970s. In reality, the loss of coral cover has been highly asynchronous, and disease is only one of many causes of the decline . For instance, cold water killed >90 % of staghorn corals in the Dry Tortugas, Florida in the winter of 1976–77 [Davis 1982]. The collapse of branching acroporids in Jamaica was overwhelmingly because of Hurricane Allen in 1980 [Woodley et al. 1981]. There is only one report of a significant outbreak of white band disease in the Caribbean before 1980, a localized die-off affecting 5 hectares of shallow reef in St. Croix , US Virgin Islands in 1976–1979 [Gladfelter 1982]. In contrast, hurricanes and coral disease were dismissed as causes of the steep decline in coral cover in the Dutch Antilles from 1973 to 1992 [Bak and Nieuwland 1995].

This passage raises several issues. First, Hughes et al. (2010) may be correct that a cold-water event was responsible for the mortality of vast fields of A. cervicornis in the Dry Tortugas in 1977. Porter et al. (1982) used photographs of permanent quadrats taken 6 months before and 6 months after the event as evidence. The photographs, unfortunately, do not establish causality and WBD cannot be ruled out as the cause of mortality of A. cervicornis. In fact, the before-and-after photographs look suspiciously as though they are displaying mortality from WBD (see especially Fig. 11.2 in the paper) and not bleaching from cold-exposure. Most of the losses of A. palmata and A. cervicornis throughout the Florida reef tract were from WBD , especially after 1978 (Precht and Miller 2007; references therein).

Fig. 11.2
figure 2

Approximate ranges of annual sea temperatures within shallow-water habitats from polar to tropical latitudes. Narrow ranges have driven the evolution of stenothermy in ectotherms at the poles and in the tropics, whereas temperate ectotherms tend to be eurythermal. The polar range is for Antarctica (Data are from Clarke and Crame 2010 and other sources. Modified from Aronson et al. 2011)

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Second, Hughes et al. (2010) pointed out that mortality from Hurricane Allen in 1980 was the principal cause of the collapse of Acropora populations at Discovery Bay and elsewhere along the north coast of Jamaica . Knowlton et al. (1981), however, also noted, “Unusual amounts of tissue exfoliation, resembling that termed ‘white band disease ’ were observed in some colonies of A. cervicornis before the hurricane . This exfoliation continued after the storm …,” and within 5 months there was a 100-fold decrease in the abundance of living colonies of A. cervicornis compared to the population immediately after the storm. Hurricane Allen was clearly a catastrophic disturbance on Jamaican reefs, but it is equally apparent that WBD was critical to the decline of A. cervicornis on these reefs. In fact, lagoonal populations of A. cervicornis at Discovery Bay were killed outright by WBD , not by Hurricane Allen (Wapnick et al. 2004).

Third, Gladfelter (1982) recognized the devastating effects of WBD on acroporids in St. Croix , spanning the years 1976–1979, prior to Hurricane Allen. Hughes et al. (2010) minimized Gladfelter’s work by asserting that the outbreak of WBD was localized to a small area in St. Croix. In contrast, Gladfelter (1982) stated the following:

Throughout much of its range, A. palmata is subject to a necrosis which can cause extensive local mortality of the coral. The author has observed this necrosis (= “white band disease ”) in the northeastern Caribbean Sea (Virgin Islands , St. Marten, Antigua), Curaçao, [and] Nicaragua (Miskito Cays), and it has been observed in Panama (P. Glynn, pers. comm.) and south Florida (A. Antonius, pers. comm.).

The outbreak of WBD in St. Croix was clearly part of an epidemic that was well underway throughout the Caribbean in the late 1970s, and it was recognized by Gladfelter as a regional phenomenon at the time.

Finally, citing Bak and Nieuwland (1995), Hughes et al. (2010) stated that coral disease was not responsible for the observed mortality of corals in the Netherlands Antilles. Although Bak and Nieuwland (1995) noted that factors such as diseases were unlikely to have been important in structuring the reefs in question, their study was confined to water depths of 10–40 m and their quadrats contained no Acropora species. They were, however, careful to note, “White-band disease is practically limited to the Acropora species and these are only common at depths shallower than 10 m along these coasts (Bak and Criens 1981; van Duyl 1985).” The two latter references were the same ones Aronson and Precht (2001a, 2001b) cited to describe the strong impact of WBD in the Netherlands Antilles.

Using data from a paleoecological study, Cramer et al. (2012) attempted to dispute the conclusions of Aronson and Precht (2001a, 2001b, 2006): (1) that the most significant losses of acroporid corals in the Caribbean occurred beginning in the late 1970s; and, related, (2) that WBD was the primary cause of the decline of acroporids . Aronson and Precht drew these conclusions based on their compilation of direct observations by a large number of scientists and other informed observers from 31 areas distributed among 16 countries or territories throughout the tropical and subtropical western Atlantic . Cramer et al. (2012) suggested instead that A. cervicornis had begun to decline in the Caribbean because of anthropogenic pressure decades before the outbreak of WBD and the frequent occurrence of coral-bleaching events. They based their alternative interpretation on the stratigraphic distribution of subfossil A. cervicornis in 18 circular trenches, each 60 cm in diameter and 60–80 cm deep, which they excavated at six stations in one area in one country: Bocas del Toro, Panamá .

Cramer dug three trenches at each of the six stations at Bocas del Toro. Half the sites (nine trenches) were dug in a lagoonal environment and the other half (another nine trenches) were situated ‘offshore ,’ in a near-coastal environment. At the lagoonal stations, which were located in Bahía Almirante and the adjacent Laguna de Chiriquí, three of the nine trenches exhibited a decline of A. cervicornis before the regional WBD outbreak (based on radiocarbon dates of Porites furcata—not A. cervicornis—derived from accelerator mass spectrometry, or AMS), a pattern consistent with the preferred scenario of Cramer et al. (2012). Five lagoonal trenches had little or no A. cervicornis at any level, ranging in proportional weight from 0 to 8 % of the total coral material. Radiocarbon dates from the ninth lagoonal trench showed reversals that indicated significant stratigraphic mixing. In the offshore environment, none of the trenches showed evidence of an early decline of A. cervicornis. One station, consisting of three trenches, had negligible quantities of A. cervicornis. At each of the other two stations, two trenches showed declines of A. cervicornis in the 1970s or later (again based on AMS dates from P. furcata rather than the focal species), whereas there was no clear pattern in the third trench at any of the stations due to mixing. Not only are the temporal patterns of coral dominance inconsistent, but AMS is inaccurate at the young ages of the corals that were dated, calling into question the age models on which the conclusions were based (Aronson et al. 2014). The claim that the decline of A. cervicornis at Bocas del Toro began prior to the outbreak of WBD and subsequent bleaching events is not supported by the data in Cramer et al. (2012), nor by more accurate chronologies from the same area based on 210Pb dating (Aronson et al. 2014).

There can be little doubt that the largest decline in coral cover on Caribbean reefs that occurred in recent decades resulted from the regional mass mortality of Acropora spp. from WBD (Gladfelter 1982; Bythell and Sheppard 1993; Aronson and Precht 2001a, 2001b; Schutte et al. 2010), subsequent mass mortalities of massive corals notwithstanding. The disease -induced mass mortality of acroporids that occurred from the late 1970s through the early 1990s was not demonstrably connected to overfishing or changes in land use, and it was more than likely related to warming sea temperatures (Kline and Vollmer 2011; Randall and van Woesik 2015). Why is this important? Cramer et al. (2012) stated, “[Our] results, coupled with increasing evidence that protection from local disturbances may increase reef resilience to climate change (Hughes et al. 2007; Knowlton and Jackson 2008), highlight the importance of managing local impacts such as fishing and land clearing to stem the tide of reef decline .” In fact, the primary causes of the decline of Acropora, and other Caribbean corals for that matter, operated—and still operate—on larger spatial scales, highlighting the importance of confronting regional and global impacts if we are to save coral reefs.

Below, we evaluate whether local management can promote resilience under current conditions. We emphatically agree that local problems should be addressed, for reasons spelled out at the end of this chapter. Contrary to the accusation of Knowlton and Jackson (2008) that we have been monolithic in our view of the importance of WBD , we have always explicitly subscribed to the pluralistic view of Quinn and Dunham (1983) that ecology seeks to evaluate the relative importance of the many causes underlying an observed pattern. Overestimating and overvaluing our capacity to promote resilience through local action, however, diverts attention and resources from the issue of climate change.

11.3 Indicators of Degradation

McClanahan, Graham et al. (2011) showed that, as reefs in the Indian Ocean degraded, corals were the most resistant components and the last to decline . Planes et al. (2005) reported that when shock waves from nuclear testing extirpated the reef-fish assemblages of Mururoa Atoll in French Polynesia , the living coral assemblages remained intact and the habitat they provided facilitated recolonization of the fish. Likewise, the early losses of reef components in the Caribbean other than corals, such as fish stocks (Jackson and Johnson 2000; Jackson 2008), are not incompatible with the recent loss of acroporids and other coral species to diseases and other causes (cf. Woodley 1992). On reefs where living coral cover has recently declined, however, the loss of that coral and concomitant loss of physical structure have resulted in significant declines in reef fish, independent of any impacts of fishing pressure (Jones et al. 2004; Alvarez-Filip et al. 2009; Paddack et al. 2009).

Our perception of the extent of reef degradation clearly depends on how degradation is defined and which components are considered important from ecological, societal, or other viewpoints. Significant functional degradation is perceived to have occurred earlier if fish stocks are considered most important than if corals are considered to be the signal components of the reefs named for them (Jackson and Johnson 2000; Pandolfi et al. 2003; McClanahan, Graham et al. 2011). The first view, in which the ‘health’ of a reef hinges on the state of its fish assemblage, implies long periods of latent degradation at the ecosystem level. Undetected losses of resilience are expressed latterly by threshold phenomena in the coral assemblages, including mass mortalities and poor recovery from bleaching events and disease outbreaks. The second view, which focuses on the corals themselves as the bellwethers of reef condition, implies the alternative hypothesis that fish—especially herbivorous fish—are less important to maintaining coral dominance than was previously thought. Herbivory could be critical to recovery in situations in which macroalgae have the potential to monopolize the substratum and suppress populations of juvenile corals (Sammarco 1982; Hughes and Tanner 2000; Mumby 2006; Box and Mumby 2007; Mumby et al. 2007a; Idjadi et al. 2010; Adam et al. 2015). Even in those cases, however, fish are not necessarily the most important herbivores . Echinoids, especially Diadema antillarum, are often far more potent herbivores on Caribbean reefs (Sammarco 1982; Edmunds and Carpenter 2001; Idjadi et al. 2006; Idjadi et al. 2010). The latter observation is independent of whether or not overfishing artificially enhanced the abundance of Diadema prior to their regional mass mortality (Hay 1984; Carpenter 1986; Precht and Aronson 2006; Sandin and McNamara 2012).

If macroalgae pose a threat to the recovery of coral populations on most reefs, there is little difference in the two views beyond the semantic issue of the point at which a reef is said to be degrading or degraded. The two views differ markedly, however, if macroalgae generally do not threaten coral recovery. Although, as stated above, high abundances of macroalgae can suppress coral recruitment , Bruno et al. (2009, 2014) questioned the proposition that macroalgae dominate most Caribbean reefs in their current state of low coral cover (see also Côté et al. 2005, 2013). The implication is that the decline and recovery of coral populations are largely decoupled from fishing pressure, as has been demonstrated explicitly for several reef systems in the Caribbean (Aronson et al. 2012; Edmunds 2013).

11.4 The Role of Marine Protected Areas

The overfishing hypothesis in its extreme form leads directly to the idea that ecological problems in the sea would largely be solved if only we would control fishing pressure (e.g., Jackson et al. 2014). This presumption provides a clear rationale for continuing to set aside marine protected areas (MPAs) and continuing the protections afforded by existing MPAs, which are designed to control the exploitation of fish stocks. By virtue of their current design, however, MPAs are less effective or wholly ineffective at controlling terrigenous inputs of nutrients and sediments. And by virtue of their scale they do not address the root-causes of climate change: the size of the human population and greenhouse-gas emissions. Climate change is expressed on coral reefs through the direct impacts of increasing sea temperatures, decreasing carbonate saturation states, and rising sea levels, as well as ancillary effects that may include outbreaks of coral disease and increasing intensities of hurricanes (Kleypas et al. 2001; Gardner et al. 2005; Hoegh-Guldberg et al. 2007; Harvell et al. 2009; Anthony et al. 2011; and many others).

Although Mora et al. (2006) concluded that MPAs in their current form do not preserve the trophic cascades of predation and herbivory that putatively maintain coral populations, there are benefits to coral cover of protection from fishing and terrigenous input, as well as benefits of protection from fishing pressure alone (Houk et al. 2010; McCook et al. 2010; Selig and Bruno 2010). Evidence is rapidly mounting, however, that overfishing is not the primary threat to benthic assemblages on coral reefs. Protecting fish stocks does not necessarily reduce the cover of macroalgae , increase coral populations, or preserve or increase the topographic heterogeneity that is critical to maintaining and increasing those fish stocks (McClanahan et al. 2001, 2005; Aronson and Precht 2006; Bood 2006; Idjadi et al. 2006; Vroom et al. 2006; Coelho and Manfrino 2007; Kramer and Heck 2007; Bruno et al. 2009; Myers and Ambrose 2009; Stockwell et al. 2009; Dudgeon et al. 2010; Alvarez-Filip et al. 2011; Lowe et al. 2011; McClanahan, Huntington et al. 2011; Żychaluk et al. 2012; Bégin et al. 2016). The threats of continuing climate change and its collateral impacts loom large, raising questions about the potential of local management alone, or the phenotypic or evolutionary responses of corals and their zooxanthellae (Baker et al. 2008; Sammarco and Strychar 2009; Pandolfi et al. 2011; van Woesik and Jordán-Garza 2011), to reverse or even delay significantly the hemmorhagic damage that is already well underway (Donner et al. 2005; Hoegh-Guldberg et al. 2007; Donner 2009; Hoegh-Guldberg et al. 2011; Toth et al. 2014).

The claim that restoring herbivores will save coral populations by reducing the cover of macroalgae (Aronson 1990; Jackson et al. 2001; Pandolfi et al. 2003; Mumby 2006; Jackson et al. 2014) has been ‘augmented’ with the idea that MPAs will maintain the resilience of reefs, ‘buying time’ while we address climate change (Hughes et al. 2003; Bellwood et al. 2004; Hughes et al. 2007; Mumby et al. 2007a, 2007b; Hughes et al. 2010; Edwards et al. 2011). Even the latter concept is of questionable validity (McClanahan et al. 2005; Bood 2006; Graham et al. 2008; McClanahan 2008; Graham et al. 2011; Huntington et al. 2011). Existing MPAs do not enhance the resistance or resilience of reef assemblages to thermal stress (Hoegh-Guldberg and Bruno 2010; Selig et al. 2012). Previous exposure of corals to high-temperature conditions is a far better predictor of the persistence of coral populations during positive thermal anomalies than their status of protection (Thompson and van Woesik 2009; Selig et al. 2012; Grottoli et al. 2014).

Côté and Darling (2010) pointed out that disturbed reef assemblages, which replace more pristine (or less degraded) assemblages following perturbations , are by default more resilient because ‘recovery ’ to those early successional states requires little time and meets with little or no systemic resistance . It is cold comfort to be reminded of the inescapable, thermodynamic reality that the end of marine life will be the most stable state of all. As we labor to prevent that dreadful eventuality from accelerating into the present century, it is well to remember that the most resistant or resilient configurations are not necessarily the most desirable (Rogers 2013).

11.5 Parsimonious Explanations

Millennial-scale physical drivers, including natural trends in climate , often explain the historical limits to the growth and composition of Holocene coral assemblages more simply and more completely than hypotheses of human exploitation and other forms of interference. Fishing and terrigenous input are regional issues, but they are perpetrated and controlled locally. Climate change occurs at the largest spatial scales, but the resultant changes in parameters such as sea temperature and pH act at very small scales. They influence the coral holobiont, its physiological rates, and the microenvironment in which it lays down aragonite crystals, as well as rates of carbonate precipitation and submarine cementation from other biotic and abiotic processes (Kleypas et al. 1999; Macintyre and Aronson 2006; Manzello et al. 2008). Those microscale processes scale up to the level of the reef system and beyond, interacting with such second-order rates as the flux of nutrients into the system and their influence on carbonate deposition and bioerosion . Herbivory, predation , and other rates that ecologists view as critical to the healthy functioning, persistence, and resilience of reef systems (Jackson et al. 2001; Mumby et al. 2007b; Sandin et al. 2008) overprint the impacts of physical processes, driving the trajectories of benthic reef assemblages over a range of relevant time scales (Urban et al. 2012). The challenge is to determine how important those ecological interactions really are on ecological scales of decades to centuries, whether larger, millennial time scales have been more important than ecological scales, and which processes have been important on those millennial time scales.

Just because people were around when acroporid corals ceased building reef framework off present-day Fort Lauderdale 6000 years ago (Lighty et al. 1978; Toscano and Macintyre 2003; Banks et al. 2007) does not mean humans were responsible for their decline . Climatic cooling in the late Holocene was likely the primary cause of reef shutdown off the eastern coast of the Florida Peninsula, and a warming climate is now permitting the northward re-expansion of cold-sensitive coral taxa in the western Atlantic (Precht and Aronson 2004; Precht and Miller 2007; see Greenstein and Pandolfi 2008 for an example from western Australia ).

In a similar vein, human activities did not drive branching Porites corals to replace A. cervicornis ~500 years ago in the shallow zones of the uncemented, lagoonal , rhomboid shoals in Belize . As those reefs grew to sea level , the living coral assemblages relocated themselves to a new physical environment—shallower water—and their species composition changed to produce the observed shallowing-upward sequence (Aronson et al. 1998). The persistence of the coral assemblages and the geomorphology of the reefs are controlled by tectonic events, which occur every few millennia, cause catastrophic slope-failure, and wipe out a substantial proportion of the benthic communities (Aronson et al. 2012).

Some investigators have insisted on anthropogenic causes for the degradation of coral reefs, in spite of evidence that is at best equivocal. Much has been written about a connection between the advent of European agriculture in Barbados and the demise of populations of Acropora palmata there, but the supposition that the story is one of nutrient loading from agricultural runoff rests almost entirely on a mild suggestion in a paper by John Lewis (1984). Lewis attributed the late Holocene disappearance of Acropora palmata from inshore, fringing reefs along the west coast of Barbados to storm damage, successional processes, and possibly terrigenous runoff that resulted from land-clearing and sugar-cane cultivation beginning in the 1600s. Twenty-seven years later, Sala and Jackson (2011, p. 197) had this to say about Lewis’s (1984) results from Barbados:

Circumstantial evidence suggests that the problems in Barbados were due to deforestation of the island for sugarcane and the consequent runoff of sediments and human waste, as well as extreme overfishing to feed the burgeoning population.

Four years before that, however, Macintyre et al. (2007) had commented on Lewis’s (1984) interpretation of the late Holocene record of A. palmata in Barbados :

Formerly attributed to human activity, the demise of a bank–barrier reef off southeastern Barbados …is now thought to be largely the result of late Holocene , millennial-scale storm damage.

Macintyre et al. (2007) suggested that the vibrant growth of stands of A. palmata in a reef-crest habitat off the south coast had ceased long before the European colonization of Barbados . The primary cause of mortality was physical damage 3000–4500 cal BP, with agricultural runoff possibly accounting for the mortality of remnant colonies 300–400 cal BP (see also Toscano 2016). Roff et al. (2013) echoed this latter model in suggesting that terrigenous runoff combined with climatic perturbation drove a phase shift in a nearshore coral assemblage on the Great Barrier Reef. The loss of A. palmata from the south coast of Barbados , millennia before any European influence, could have been part of a regional drawdown of that species (Hubbard et al. 2005; Macintyre et al. 2007).

Lewis’s (1984) speculation about agriculture applies to inshore, fringing reefs . It is, therefore, not necessarily incompatible with the conclusions of Macintyre et al. (2007) about a bank–barrier reef further offshore , but populations of A. palmata in Barbados were not all killed by runoff . Furthermore, the idea that terrigenous input was to blame cannot fully account for the replacement of A. palmata by a vibrant assemblage of massive corals on the inshore reefs (Lewis 1960). A coral assemblage dominated by Orbicella spp. and other massive species is precisely what one would expect in a hurricane -dominated environment (Stoddart 1963; Porter et al. 1981; Woodley 1989), such as that envisioned by Macintyre et al. (2007). Overfishing , incidentally, had never been part of Lewis’s (1984) original scenario, nor did it figure in the interpretations of Macintyre et al. (2007) or Toscano (2016).

An argument for the runoff hypothesis would also have to explain why Barbados , which is a low, carbonate island, is the only known location in the Caribbean for which such a scenario has been suggested. Jamaica , for example, is a high island with a similar history of colonization and sugar-cane cultivation. The impacts of terrigenous runoff should have been accentuated compared to Barbados, yet there is no evidence for a mass mortality of A. palmata in Jamaica (or any other Caribbean island) 300–400 years ago. The best way to test the hypothesis would be through biogeochemical analysis (cf. Aronson et al. 2014).

11.6 Temporal Priority

Which regional or global driver is or was most important in the degradation of modern reefs is largely a consequence of temporal priority. Thermally induced bleaching has been secondarily important in the Caribbean only because bleaching episodes came after much of the Acropora had already been killed by white-band disease , in a regional outbreak that itself might have been thermally driven. The situation is reversed in the Indo-Pacific : bleaching has had a greater influence than coral disease because bleaching-induced mortality occurred on a large geographic scale prior to disease outbreaks (Buddemeier et al. 2004). It remains to be seen whether ocean acidification will have the opportunity to damage coral populations and coral reefs (Hoegh-Guldberg et al. 2007; Veron 2008; Hönisch et al. 2012) after the impacts of rising sea temperatures have taken their toll. Some state or rate is the primary limit to the growth of coral populations, the integrity of the assemblages those populations comprise, and the accretion of the reef frameworks on which they perch as a living veneer.

11.7 Conclusion

It should come as no surprise that coral reefs are highly sensitive to climate change. Narrow annual temperature ranges characterize shallow-benthic habitats at tropical and polar latitudes, compared to analogous habitats in the highly seasonal temperate zone (Fig. 11.2). Benthic ectotherms near their latitudinal extremes are adapted to the narrow seasonal temperature ranges to which they are normally exposed. Stenothermy limits their scope for phenotypic or evolutionary adjustment to warming temperatures. The tropical and polar benthos, therefore, are responding earlier and more strongly to warming sea temperatures than are temperate-marine biotas.

All biological interactions play out on the template of the physical environment. For every ecosystem , there is a range of larger scales at which physical drivers trump biotic interactions in determining its trajectory; it is just a matter of scaling up sufficiently to that range. Reefs through Phanerozoic time are no exception (Wood 1999, 2007; Veron 2008; Kiessling and Simpson 2011; Hönisch et al. 2012; Norris et al. 2013). The projected dynamical responses of reefs to large-scale physical drivers, both natural and anthropogenic , are the appropriate foundations of hypotheses against which to test the ecological effects of localized human activities.

Our point about modern coral reefs under human influence is that one does not have to scale up very much at all to discover the scales at which physical controls predominate, because the lower end of the range of scales at which physical drivers strongly influence the biotic milieu, or indeed overpower biological processes, is not very large. For coral populations and coral reefs, physical controls are primary at scales equal to, or only slightly larger than, the scales at which biotic interactions are measured and observed. A good example is the influence of anthropogenically warming temperatures on outbreaks of coral disease . Geologic al and ecological processes may be disjunct in some marine ecosystems, but for coral reefs they operate on scales that are very similar.

That is not to say that geology equals ecology . Some paleoecologists view the geologically rapid changes in sea level and sea temperature during the Pleistocene glaciations as disturbances in ecological time, precipitating the collapse of reef communities and requiring their subsequent reconstitution in other habitats or locations (Jackson 1992; Pandolfi 1996; Webster et al. 2004; Jackson and Erwin 2006; Pandolfi and Jackson 2006). But even when glacial/interglacial fluctuations were dramatic in geologic time, they were still slow compared to the turnover rates of the corals (see Kleypas 1997). In most cases, coral populations should have been able to alter their spatial and habitat distributions incrementally over long periods, but the endpoints of these incremental shifts displayed in the fossil record have been presumed to represent radical disassembly and reassembly (see discussion in Tager et al. 2010; see also Chap. 7).

Overfishing is a terrible problem with far-reaching consequences. There is more than enough sorrow to go around over the strangulation of marine ecosystems, and we wholeheartedly endorse efforts to protect life in the sea from the grotesque and irrational level of exploitation to which it is being subjected. Even worse for coral reefs, unfortunately, are rising temperatures, ocean acidification , and other potential impacts of climate change, such as predicted future increases in the intensity of hurricanes (see also Glynn 2011). Because these physical drivers operate at a global scale, they exert a powerful influence that is harder to control by a long shot than fishing pressure and terrigenous input, which are the feasible targets of existing MPAs and other local management strategies. Local actions to control fishing and runoff , along with a more strategic, integrative approach to the design and location of marine reserves that accounts for spatial variation in susceptibility to climate change (Riegl and Piller 2003; McClanahan et al. 2008, 2009; Mumby et al. 2011), could prolong the death-throes of coral populations; however, they will keep dying until government and society recognize climate change for the grave threat it is and address it on a geopolitical level. If and when the impacts of climate change can be mitigated or reversed, their impacts will no longer overwhelm local threats. Fishing , terrigenous input, and other localized problems will then more commonly become the limiting factors and will respond more strongly to the appropriate protective measures. Liebig’s Law of the Minimum, a nineteenth-century model from agronomy (Hooker 1917), has great value for understanding the challenges facing coral reefs and their human stewards.

The Law of the Minimum describes the serial limitation of different nutrients on crop yields. Liebig used the metaphor of a barrel with staves of different lengths, each stave representing a nutrient . The water-level in the barrel represented crop yield, and the shortest stave represented the limiting nutrient. Adapting the metaphor to coral reefs, the staves of the barrel in Fig. 11.3 represent the many factors that potentially limit the growth of coral populations and accretion of reef frameworks , both of which are represented by the water level. The shortest stave identifies the rate-limiting process. The water can be no higher than that stave, meaning that coral growth or framework accretion can be no greater than allowed by the rate-limiting process.

Fig. 11.3
figure 3

Liebig’s Law of the Minimum applied to coral reefs. The lengths of the barrel-staves were set somewhat arbitrarily for illustrative purposes

Full size image

We have chosen rising global temperatures as the primary limitation in Fig. 11.3, based on the foregoing discussion. If the temperature rise can be slowed or reversed, then warming will cease to be the rate-limiting process. The global-warming stave of the barrel will then be lengthened and will no longer be the shortest one. The shortest stave in the figure will then be disease , which as we have said could be linked to warming seas. Synergistic or antagonistic interactions among the drivers of coral-reef degradation (Mora et al. 2007) mean that lengthening some staves will lengthen or shorten certain other staves as well. If and when regional- and global-scale limitations are adequately addressed, then local-scale factors, such as nutrient loading and overfishing , will serially become the limiting staves. The upshot is that simultaneous actions at local, regional, and global levels are our only hope for saving coral reefs (see also Hoegh-Guldberg and Bruno 2010; Kennedy et al. 2013; Rogers 2013). Clearly, global problems will be more difficult to solve and take longer than local ones, but that is very different from saying we should focus on local management now because it will buy time for us to address the impacts of climate change in the future.

Planning to have reefs around for our children and our children’s children to enjoy—meaning that we want to keep them reasonably intact for a little over a century—is just not good enough. We need to conserve reefs on a millennial time scale by fighting climate change on a global spatial scale. It may sound perverse but, considering the jeopardy in which we humans have placed coral reefs, the biosphere, and our very existence, addressing climate change and being left with an exceedingly difficult set of local-scale, ecological problems to attack would be a blessing.