The Lasting Effects of Fisheries on Maine's Kelp Forest Ecosystem

BULLETIN OF MARINE SCIENCE. 89(1):31–55. 2013 MOTE SYMPOSIUM INVITED PAPER http://dx.doi.org/10.5343/bms.2011.1148

ECOSYSTEM FLIPS, LOCKS, AND FEEDBACKS: THE LASTING EFFECTS OF FISHERIES ON MAINE’S KELP FOREST ECOSYSTEM

Robert S Steneck, Amanda Leland, Douglas C McNaught, and John Vavrinec

Abstract Ecosystems can “flip” and, as a result of reinforcing feedback mechanisms, “lock” into alternative stable states. We studied this process in a kelp-forest ecosystem in Maine, USA, for nearly four decades and found two stable states: one dominated by green sea urchins and crustose coralline algae and the other by erect fleshy macroalgae. Herbivory by urchins drives algal deforestation, but declined after fishing for sea urchins began in 1987. As the fishery expanded northeastward, so did phase shifts to macroalgae. By 2000, macroalgae dominated nearly all of coastal Maine. Monitoring newly settled sea urchins between 1996 and 2002 revealed sites with thousands of settlers per square meter per year, but virtually none survived to become adults. Algal succession to densely branched morphologies may create nursery habitat for settling crabs that prey on settling sea urchins. Experiments intended to restore herbivory to prefishing levels, through translocation of 51,000 adult sea urchins over two consecutive years at multiple release sites (with controls), resulted in complete urchin mortality both years as a result of predation by large migratory Cancer borealis Stimpson, 1859 crabs. Population densities of this crab increased fivefold coastwide soon after the macroalgal phase shift. Persistent absence of urchins (even in no-take reserves) probably resulted from predation on newly settled and/or adult urchins. Fisheries-induced declines in herbivory may therefore have improved recruitment potential for predatory crabs that then became the region’s new apex predator. Cascading sequential processes of herbivory, recruitment, and predation create reinforcing feedback, effectively locking this ecosystem into alternative stable states.

Fisheries managers are often surprised when species dynamics fail to meet their equilibrium-based expectations. For example, the collapse of Atlantic cod, Gadus morhua Linnaeus, 1758, on Canada’s Grand Banks caught fisheries managers off guard, so they abruptly closed the entire area to fishing in 1990 (Hutchins 2000). They expected full recovery to take no more than a decade, but more than 20 yrs later, stocks remain largely depleted (Shelton et al. 2006). When alternative population, community, or ecosystem states of fished stocks become stable or lag in their recovery (i.e., undergo “hysteresis” sensu Scheffer et al. 2001), the problem transcends academic curiosity and becomes a socioeconomic crisis because many people depend on fisheries for their livelihood. Surprisingly long recoveries can also create crises of confidence in the science we apply to marine resources. Such surprises are anathema to traditional single-species management, which requires relatively predictable population (e.g., maximum sustainable yield) and ecosystem behaviors. As we move toward ecosystem-based management for the oceans (see McLeod and Leslie 2009), understanding processes that can create

Bulletin of Marine Science 31 © 2013 Rosenstiel School of Marine & Atmospheric Science of the University of Miami 32 BULLETIN OF MARINE SCIENCE. VOL 89, NO 1. 2013 sudden changes in, and effective “locking” of, marine ecosystem structure and function is more important than ever. Fisheries effects on marine ecosystems are pervasive (Jackson et al. 2001), but they can also provide unique learning opportunities. Often, the timing, areal extent, and magnitude of fishing effects are well documented, but when fishing or any extraction of natural resources depletes species that are strong ecosystem drivers (i.e., highly interactive species), its impacts extend well beyond those species and affect the entire ecosystem (Soulé et al. 2003). Here, we focus on the sea-urchin fishery and its impact on kelp forest ecosystems in the Gulf of Maine. Globally, plant-herbivore interactions between sea urchins and kelp forests (including all other associated erect foliose macroalgae, hereafter simply called “macroalgae”) in cold-water marine coastal zones have been extensively studied and reviewed (Estes and Palmisano 1974, Lawrence 1975, Dayton 1985, Steneck et al. 2002, 2008). Well known are both the strong inverse relationship between sea urchins and macroalgae and the depletion of these herbivores by the sea-urchin fishery (Andrew et al. 2002, Berkes et al. 2006). Changes in community dominance between kelp forested and deforested states are called “phase shifts” (sensu Steneck et al. 2002), a term adopted from coral-reef studies (Done 1992, Hughes 1994). The term describes a sudden shift in dominance but not the cause or the duration of the shift. Less well under- stood is whether this fishery can trigger a profound and durable change in kelp-forest ecosystem structure and function, thus maintaining an alternative stable state. Alternative stable states have attracted considerable scientific attention for nearly half a century. On this topic, theory (e.g., Lewontin 1969, May 1977) preceded em- pirical or experimental evidence. Subsequent research illustrated the potential for alternative stable states, but some research also heightened the controversy over ex- actly how they should be defined and what would constitute proof of their existence (Petraitis and Latham 1999, Petraitis and Dudgeon 2004). From recent introspection, three primary criteria emerged as necessary, but rarely met, conditions for alternative stable ecosystem states. First is that transitions between states occur suddenly and persist in the same physical habitat. Second, they must result from a pulse perturbation (or experiment) and not a press perturbation. Finally, the ecosystem study must be long enough to demonstrate that the alternative state is self-sustaining and that its effects occur over a large enough area to qualify as an ecosystem-scale phe- nomenon (Petraitis and Latham 1999, Petraitis and Dudgeon 2004). Although those criteria are necessary to document alternative stable states, they do not necessarily reveal why these ecosystems “flipped” and “locked” into alternative states. Answering those questions requires understanding reinforcing feedback mechanisms (Petraitis and Latham 1999) that prevent or slow the reversal of these complex ecosystems (i.e., create hysteresis). This understanding requires knowledge of the ecosystem drivers (usually ecological processes) and how and why these drivers have changed fundamentally. It is that challenge that we address in the present study. We suggest that Maine’s kelp forest–sea urchin ecosystem meets the three criteria and illustrates alternative stable states well. We further suggest that the sea-urchin fishery was effectively a pulse perturbation (i.e., a large-scale herbivore removal) ap- plied throughout the region over a relatively short period of time, after which ecosystem function changed fundamentally, and this changed ecosystem has persisted long enough to meet strict definitions of stability (Connell and Sousa 1983). We draw from several lines of evidence to suggest that fisheries-induced changes to the ecological STENECK ET AL.: ECOSYSTEM FLIPS, LOCKS, AND FEEDBACKS 33 process of herbivory have directly or indirectly flipped and locked this ecosystem into an alternative state. Specifically, we draw heavily from a spectrum of quantita- tive and experimental studies conducted at specific sites on the Maine coast over the past 36 yrs by us and by others. Our central point is derived from the integration of these studies. We present rudimentary methods for each key category of information we present, but because detailed methods cannot be fully described here, we suggest interested readers refer to the sources (McNaught 1999, Vavrinec 2003, Leland 2002). With this integration, we developed a novel understanding of how differently this ecosystem behaves as a result of fishing pressures on it.

Loss of Large Predatory Finfish: Slow Drivers Set the Stage in the Gulf of Maine

Rocky shores of the western North Atlantic are among the most species-depau- perate shallow marine ecosystems in the world (Witman et al. 2004). This condition is particularly conspicuous within the region’s kelp-forest ecosystems, where very few predatory fishes, a single herbivorous species of sea urchin, Strongylocentrotus droebachiensis (O. F. Müller, 1776), and relatively few kelp taxa define this ecosystem (Steneck et al. 2002). Such simplified food webs often develop strong interaction strength among species, possibly because of the limited number of alternative trophic pathways (Polis and Strong 1996). These strong interactions may help explain why keystone species and trophic cascades are most pronounced and most common in low-diversity benthic marine ecosystems (reviewed by Shurin et al. 2002). The well-documented long history of fisheries-induced reductions of large predatory finfish (Jackson et al. 2001, Myers and Worm 2003) effectively set the stage for the changes in recent decades that we describe. Large predatory finfish such as Atlantic cod, haddock, Melanogrammus aeglefinus (Linnaeus, 1758), and Atlantic wolfish, Anarhichas lupus Linnaeus, 1758, probably dominated the North Atlantic’s coastal food webs before the arrival of humans (Steneck 1997, Jackson et al. 2001, Steneck et al. 2004, Estes et al. 2013). Archaeological evidence suggests that these and other finfish were important for thousands of years in coastal Maine despite possible localized and temporary depletions occurring more than 500 yrs ago, before European contact (Bourque et al. 2008). Historical studies conclude that predator biomass, especially of Atlantic cod, remained high through the 1800s (Rosenberg et al. 2005). Cod were described as “flourishing” in 1918 along coastal Maine (Steneck and Carlton 2001), but suffered a significant decline when spawning aggregations were heavily fished in the 1930s (Steneck 1997, Ames 2004). Despite depletion, cod continued to show considerable top-down cascading effects along Canada’s Scotian Shelf (Frank et al. 2005) and elsewhere (Worm and Myers 2003) in recent decades, but these coastal-zone extirpations have continued over the last 20 yrs, leaving large stretches of coastal Maine virtually free of large predatory finfish (e.g., Atlantic cod is illustrated in Fig. 1). Taken together, all 14 species of demersal predatory fishes caught today on the coast of Maine comprise only 2% of Maine’s fisheries landings by mass (MDMR 2010a). Globally, few if any mobile benthic invertebrates can persist in marine habitats dominated by large predatory finfishes without shelter protection, which therefore limits their population size. The protracted serial decline of Atlantic cod and other large predatory groundfishes may therefore have relaxed the ceiling that limits prey 34 BULLETIN OF MARINE SCIENCE. VOL 89, NO 1. 2013

Figure 1. Nineteen-year trends in the abundance of Atlantic cod. Modified from Sosebee and Cadrin (2006). Circles represent numbers of fish caught in National Marine Fisheries Service autumn trawl surveys. Smaller black squares along the coast mark our four study regions. population size. Although that decline may have been a prerequisite for subsequent population increases in prey species, it will not necessarily trigger them (especially if the prey species is regulated by demographically important feedback among other species in the ecosystem, as we will describe below). Three ecologically or economi- cally important benthic invertebrate species were shown in experiments to suffer high rates of attack from predators and were therefore likely candidates to respond positively to the decline in coastal predators: the green sea urchin, S. droebachiensis (Vadas and Steneck 1995); the American lobster, Homarus americanus H. Milne- Edwards, 1837 (Wahle and Steneck 1992, Steneck and Carlton 2001, Butler et al. 2006); and cancer crabs, especially Cancer borealis Stimpson, 1859, the Jonah crab (Witman and Sebens 1992, Leland 2002, all reviewed by Estes et al. 2012). Against this well-documented background, we considered some lasting consequences to ecological interactions and feedback resulting from functionally altered trophic cascades under kelp-forested and deforested states.

Maine’s Sea Urchin Fishery: A Pulse Perturbation Triggers a Kelp-Forest Phase Shift

In nonupwelling regions such as the North Atlantic, kelp-forest ecosystems are deforested primarily by sea-urchin grazing (Lawrence 1975, Steneck et al. 2002). When sea urchins are abundant, kelp and other erect fleshy algae are rare, and sub- strates are dominated by herbivore-resistant crustose coralline algae (CCA). The green sea urchin, S. droebachiensis, is the only herbivore capable of controlling algal abundance in the North Atlantic (Steneck and Dethier 1994, Steneck et al. 2002, Norderhaug and Christie 2009). STENECK ET AL.: ECOSYSTEM FLIPS, LOCKS, AND FEEDBACKS 35

Figure 2. Fifty-year trends in sea-urchin fishery landings, sea-urchin abundance, rates of herbivory, and percent cover of erect foliose algae. The gray vertical bar indicates the phase- shift period coinciding with the maximum harvest of sea urchins and greatest change in each measure. Data are from McNaught (1999), Steneck and Carlton (2001), Vavrinec (2003), and Maine Department of Marine Resources (MDMR 2010a). Note that sea-urchin fishery landings have been recorded since 1964 but were so low that they appear to be zero in A. Variance (SE). 1 from Adley (1964, 1965).

The fishery for the green sea urchin in Maine began in 1987 as a result of converg- ing socioeconomic and ecological factors. The Japanese appetite for “uni” (urchin roe) drove an ever-expanding global fishery for sea-urchin species (Andrew et al. 2002, Berkes et al. 2006). The global reach of the fishery was catalyzed by the good exchange rate between the Japanese yen and the US dollar (Botsford et al. 2004). Maine’s green sea urchin was both highly abundant and a highly valued food for Japanese palettes (Chenoweth 1994). Maine’s sea-urchin fishery began along its southwestern coast near human population centers and spread from there northeastward toward the Canadian border (Andrew et al. 2002, MDMR 2010b). The effects of fisheries-induced sea-urchin depletion and the resulting changes in rates of herbivory and algal community structure in Maine were profound. They are well illustrated at Pemaquid Point, a sentinel site in midcoastal Maine that had been studied for at least three decades before the initiation of Maine’s significant sea-urchin fishery (Fig. 2). Sea-urchin landings began to be recorded in 1964 but were so low as to be indistinguishable from zero (Fig. 2A). The significant statewide sea- urchin landings exported to Japanese markets, beginning in 1987, sharply increased to a peak in 1993 before declining almost as rapidly (Fig. 2A). Demographic impacts 36 BULLETIN OF MARINE SCIENCE. VOL 89, NO 1. 2013 of this fishery were profound. Stock assessments indicated that sea-urchin biomass declined 90% statewide between 1986 and 2001 (Chen and Hunter 2003). In 1983, well before the start of the export fishery, we began quantifying sea-urchin herbivory by measuring the rates at which tethered, standardized strips of kelp were consumed (Steneck 1997, Steneck and Carlton 2001). We began in 1990 at the Pemaquid sentinel site, where we recorded a rapid decline in herbivory (Fig. 2C) in concordance with the decline in sea-urchin population densities there (Fig. 2B; McNaught 1999). The result was a marked increase in the percent cover of foliose macroalgae that has persisted though 2010 (Fig. 2D). The precipitous decline in sea-urchin abundance and rise of macroalgae was evident over hundreds of kilometers along most of coastal Maine, but it was neither syn- chronous nor homogenous. Maine’s phase shift occurred earliest at the southwest sites, where the sea-urchin fishery began, and progressed patchily toward northeastern Maine (Fig. 3), roughly matching the coastwide progression of the sea-urchin fishery (Andrew et al. 2002). Within sites, phase-shifted patches grew and coalesced until a site became predominantly dominated by macroalgae (RSS pers obs). Between 1994 and 2000 some patches and study sites in the Pemaquid region had shifted phase while others had not (Fig. 3, McNaught 1999). In the years before 2000, two study sites in the Pemaquid region that were only 250 m apart (called “Kresge” and “Pemaquid Point”) were largely urchin free and dominated by foliose algae (shown in Fig. 2) while the other site was dominated by sea urchins and CCA (McNaught 1999, Steneck and Carlton 2001, Vavrinec 2003). This patchiness is reflected in the high among-site variance in this region (Fig. 3, McNaught 1999), but by 2002 all of the sites in this region had shifted phase to macroalgae (Fig. 3). Similarly, the Mt. Desert region showed steadily declining urchin biomass in mostly CCA-dominated sites until about 2003, when the phase shift to macroalgae occurred (Vavrinec 2003). No obvious trends were revealed during the relatively short period when the Jonesport region was studied, but formerly CCA-dominated “barrens” have since shifted phase to macroalgae because of reduced sea-urchin densities there (McNaught pers obs 2009–2011). In sum, the asynchrony in phase shifts from west to east (Fig. 3) matches the spatiotemporal pattern of sea-urchin harvesting (MDMR 2010b), further sug- gesting this fishery caused the decline in sea-urchin abundance regionally and at site level (Fig. 3). Although the chronology and northeastern progression are evident (Fig. 3), we cannot assume that the timing or stability of alternative states we describe for coastal Maine applies throughout the western North Atlantic. For example, kelp forests along the eastern shore of Nova Scotia have oscillated between kelp-forested and deforested states three times since the mid-1960s (Johnson and Mann 1993, reviewed by Steneck et al. 2002) without any sign that those phase shifts reflect alternative stable states. Also, the duration of CCA-dominated deforested states varies regionally. For example, the midcoast of Maine in the 1960s was described as dominated by coralline algae that were heavily “scraped” by sea urchins (Adey 1964, 1965), whereas areas adjacent to the New Hampshire border were described as having high sea-urchin abundance and a decline in macroalgae decades later, during the 1980s (Harris and Tyrrell 2001). These differences in the timing and stability of phase shifts may be related to the ecological (process-related) feedback mechanisms that work to stabi- lize them (described below). STENECK ET AL.: ECOSYSTEM FLIPS, LOCKS, AND FEEDBACKS 37 ) in the four study regions.year The in which eachregion underwentphase a shift from sea-urchin barrens to macroalgal cover is in parentheses. −2 (Data from McNaught 1999 with additional data from RS Steneck unpublished). Figuretrends 3. Averaged from multiple sites in each fourof study regions (identified by squares along the coast) biomass in of sea urchins (bold line,wet g m mass 38 BULLETIN OF MARINE SCIENCE. VOL 89, NO 1. 2013

Stability of Alternative States After Phase Shifts

Although regionally patchy, site-specific phase shifts in dominance were relatively sudden and protracted after the shift in algal abundance (Fig. 2). This pattern was best documented at Pemaquid Point, where CCA dominated rocky habitats for 20–35 yrs before the phase shift in 1993–1995. The site has remained macroalga- dominated for the last 16 yrs (Fig. 2). Similar rapid change followed by stability was observed elsewhere. We documented stability of the macroalgal phase shift in the York region after the early increase of fishing pressure there (Fig. 3; McNaught 1999, MDMR 2010b). One decade later (2002–2003), landings had declined by three orders of magnitude, where they remain today. The consistently low catch per unit effort indicates the persistently low sea-urchin population densities that have resulted in abundant macroalgae since the phase shift (Fig. 3). The most definitive test of the stability of the urchin-free state not confounded by fishing pressure was from no-take reserves that were established in 1999 in each of our study regions. Regular monitoring of sea-urchin abundance (both newly settled population densities and overall per-area biomass) and algal percent cover was conducted annually at control and reserve sites from 1996 to 2002 in each of our four regions. Areas closed and open to fishing showed similar trends in general but mostly showed declines in urchin abundance (Vavrinec 2003). The pattern was most obvious at two reserves in the York region, where settlement of sea urchins increased after the closure to fishing while overall sea-urchin biomass declined (Fig. 4). Demonstrating stability of alternative states after a phase shift is important, but it does not help us understand why such fundamental changes to the structure and function of this ecosystem occurred and why this ecosystem has locked into its alternative state.

Ecological Processes That Lock Alternative States

Relatively few ecological processes may be important in delaying the reversion of ecosystems to their original state in benthic marine communities (i.e., in impos- ing hysteresis) or in preventing it. These include low settlement of larvae resulting from availability of too few reproductive individuals or low fertilization success and therefore too few larvae (i.e., Allee effects; Knowlton 1992) or too few larvae arriving to nursery habitats from reproductive populations (i.e., larval connectivity effects; Cowen et al. 2007). If no shortage of competent settlers is evident, however, then the most logical explanation for low population densities is high rates of postsettlement mortality. High Post-settlement Mortality for Sea Urchins.—To determine population densities of recently settled sea urchins at our study regions, we deployed standardized sea-urchin settlement collectors (Harris et al. 1994) in the form of artificial plastic grass (Astroturf, Monsanto Corp.) attached to 20- × 9-cm polyvinyl chloride plates bolted to the benthos at four study sites in each of the four study regions shown in Figure 3. Twenty-eight collectors were deployed at each site from 1996 to 2002 from June through July, when the majority of sea-urchin settlement occurs in coastal Maine (for detailed methods, see McNaught 1999, Vavrinec 2003). STENECK ET AL.: ECOSYSTEM FLIPS, LOCKS, AND FEEDBACKS 39

Figure 4. Sea urchin abundance in no-take reserves. (A) Density of newly settled sea urchins (2 mo old) and (B) total sea-urchin biomass at two no-take reserve sites (East Point, open circles; Seal Head, filled circles) in the York region. The dotted line indicates the point in time at which the reserves were closed to urchin fishing. Error bars are SE.

Densities of newly settled sea urchins were greatest in the York region (McNaught 1999, Vavrinec 2003), where we recorded lowest population density and biomass of adult sea urchins after the phase shift in 1994 (Fig. 3). Regional patterns of settlement declined monotonically from the York region toward the northeast at our monitored sites (McNaught 1999, Vavrinec 2003), but in the two southwestern regions, where sea-urchin fishing in Maine began and where we recorded the earliest decline in urchin abundance (i.e., the York and Pemaquid regions; Fig. 3), we also recorded the highest population density of newly settled sea urchins [averaging 7914 (SE 1885) m−2 at York and 2457 (SE 604) at Pemaquid]. Despite the protracted period (1996–2002) of high urchin settlement, even slightly larger and older urchins were virtually absent (Fig. 3). This pattern held at two sites in the York region closed to sea-urchin fishing (Fig. 4). In sum, mortality immediately after settlement is nearly complete and results in undetectably low population densities of larger sea urchins. Habitat Architecture and the Recruitment Potential of the Benthos for Crab Micropredators and Settling Sea Urchins.—Early postsettlement mortality drives demography for a variety of marine benthic invertebrates (Hunt and Scheibling 1997). Studies of newly settled decapods in Maine determined that the architecture of small shelter-providing habitats creates a demographically important 40 BULLETIN OF MARINE SCIENCE. VOL 89, NO 1. 2013

Figure 5. Algal communities before and after the phase shift at Pemaquid Point, Maine. (A) Sea-urchin and crustose-coralline-alga-dominated ledge (photo 1985). (B) Close-up of grazing chiton (ca. 1 cm length) and newly settled sea urchin in circle (photo 1975). (C) Kelp Saccharina sp. (photo 1995). (D) Filamentous and corticated macroalgae, including Chondrus crispus in upper right corner (Photo 2011). refuge from predatory fishes for both lobsters (Wahle and Steneck 1991, 1992) and crabs (Palma et al. 1998, 1999). Scheibling and Hamm (1991) studied predation on juvenile green sea urchins (10–15 mm diameter or at least 1 yr old) and found in caging experiments that cobblestone habitats create spatial refugia for urchins that decrease predation by Atlantic rock crabs, Cancer irroratus Say, 1817, C. borealis, and lobsters. After the phase shift to macroalgae in the mid-1990s, epibenthic habitat architecture changed profoundly (Figs. 2, 3, 5–7). As a result, survivorship of newly settled sea urchins (i.e., 2 mo or less postsettlement or 0.3 to 1 mm diameter) was signifi- cantly lower in macroalgae than in adjacent CCA-dominated “barrens” habitats or undetectable (McNaught 1999; e.g., see circle in Fig. 5B). Field caging and laboratory studies determined that predation by small predatory crustaceans, primarily newly settled C. irroratus, were most responsible for this early postsettlement mortality. Laboratory studies in Nova Scotia also found high rates of later postsettlement mortality among urchins exposed to predation from a wider variety of decapod micropredators including cancrid crabs, Crangon shrimp, but largely Pagurus hermit crabs (Scheibling and Robinson 2008). They also reported mortality from littorinid gastro- pods, but this study did not examine these mortality effects in situ or in algal habitats (Scheibling and Robinson 2008). Of these potential predators, only cancrid crabs were commonly found in algal habitats at the 10-m study sites (McNaught 1999). At the time of settlement both crab megalope and settling sea urchins are too small for directed swimming or much postsettlement movement (Palma et al. 1998, McNaught 1999, Scheibling and Robinson 2008). The receptivity of the habitat into which they settle, called “recruitment potential of the environment” (sensu Steneck and Dethier 1994), therefore regulates their postsettlement survival. It can be demographically important given sufficient settlement of competent larvae. In the macroalgal communities we studied, high population densities of micropredator crabs (hundreds per square meter; McNaught 1999) created a very low recruitment STENECK ET AL.: ECOSYSTEM FLIPS, LOCKS, AND FEEDBACKS 41 potential for settling sea urchins, but such high density occurs only in the dense algal habitats where they themselves are safe from fish predators. Increasing macroalgal abundance therefore increases the recruitment potential for crabs, but newly settled crabs decrease the recruitment potential for settling sea urchins. Algal Succession Improves Habitat Architecture for Micropredators.— More than a decade after the initial phase shift (gray bar in Fig. 2), continuing succession resulted in algae of functionally different morphology, habitat architecture, and carrying capacity for micropredators. Specifically, dominance shifted rapidly from CCA (Fig. 5A,B) to kelp and other macroalgae, but over the next decade algal succession continued, changing both species composition and the dominant algal morphology (Figs. 5, 6). Algal functional groups are polyphyletic assemblages that have evolved convergent anatomical, morphological, and ecological characteristics (Steneck and Dethier 1994). Two key morphological characteristics of algae as habitat for microcrustaceans are the volume occupied by the algal thallus and interstitial spaces created by its branches (Fig. 6; Hacker and Steneck 1990). Both crustose algae and kelp (leathery macrophytes) have no, or only modestly dissected, interstitial space and thus little habitat for microcrustaceans, but filamentous and corticated macrophytes (Figs. 5D) create the preferred habitat architecture for small and larger microcrustaceans, re- spectively (Fig. 6; Hacker and Steneck 1990). This distinction was conspicuous during suction sampling of foliose macroalgal habitats and adjacent habitats dominated by encrusting coralline algae. McNaught (1999) determined that high densities of newly settled predatory C. irroratus were primarily found among the corticated macroalgal and filamentous algal habitats. Although C. irroratus were the dominant micropredator in the 1990s (McNaught 1999), C. borealis were also collected from shallow beds of macroalgae in New Hampshire, near the southwesternmost coast of Maine (Williams and Wahle 1992). The filamentous and corticated-macroalgal functional groups have been increasing steadily over the past decade (Fig. 6). Prominent among the corticated macrophytes is the alga Chondrus crispus (Linnaeus) J. Stackhouse, which has steadily increased in abundance from the southwest toward the northeast. By 2004 it was abundant over roughly half of coastal Maine (Fig. 7). The increase in dominance of C. crispus is probably related to its unique growth characteristics. Regardless of their parental source, sporelings of this alga germinate, coalesce, and form persistent crustose holdfasts that can monopolize primary substrate and gradually dominate other algae competitively (Maggs and Cheney 1990, Santelices 2004). Surveys conducted from 1996 to 2004 recorded C. crispus abundance ranging between 30% and 50% cover. At the Pemaquid site, Chondrus abundance remained low until 2003, but by 2011 it had increased to 40% cover (Fig. 6; RSS unpubl data). The relatively recent phase shift in the Mt. Desert region produced only a modest 1.5% cover of C. crispus by 2004 (Fig. 7). A possible consequence of algal succession is an increase in the carrying capacity for crab micropredators. Before the phase shift to macroalgae (Figs. 3, 5), newly settled Cancer crabs were primarily found in cobblestone habitats, which comprise only a small fraction of shallow coastal zones (ca. 2%–10% of coastal Maine; Wahle and Steneck 1991). This situation obviously changed as phase shifts to finely branched filaments and corticated macrophytes such as Chondrus progressed up the coast of Maine during the mid-1990s to mid-2000s (Figs. 3, 6, 7) providing settling crabs 42 BULLETIN OF MARINE SCIENCE. VOL 89, NO 1. 2013

Figure 6. Habitat architecture of benthic marine algae, its suitability for microcrustaceans, and trends in the abundance of heavily branched morphologies over the past 35 yrs. Upper panels: habitat architecture of algal morphological functional groups defined by the volume of the algal thallus (abscissa) and interstitial volume (ordinate). Lower panel: temporal trend in abundance of corticated macroalgae and filamentous algae (i.e., most occupied habitats for large and small microcrustaceans; after Hacker and Steneck 1990) at Pemaquid Point, Maine (10 m depth). Error bars are SE. Gray bar identifies the timing of the algal phase shift (i.e., Fig. 2). ample opportunities to hide (Fig. 5C,D) especially on macroalga-covered ledge habitats. Macroalga-dwelling crab micropredators that limit recruitment and postsettlement survival of sea urchins remain the best explanation for the failure of high densities of newly settled sea urchins to result in higher urchin population densities. This conclusion seems corroborated by studies in Norway by Christie et al. (2009), who determined that myriad crustaceans including crabs use macroalgae as habitat STENECK ET AL.: ECOSYSTEM FLIPS, LOCKS, AND FEEDBACKS 43

Figure 7. Large-scale abundance of Chondrus crispus along coastal Maine at 5- and 10-m depths in 2004. Each dive site selected by means of a random stratified sampling design is marked with a dot. because of their “value as a refuge from predators (Martin-Smith 1993).” Perhaps not coincidentally, the York region, where C. crispus dominated the benthos in 1997, was the only region where young-of-the-year C. irroratus and C. borealis were found (Palma et al. 1999). Others concluded that “recent settlers of S. droebachiensis are vulnerable to predation by small crustaceans, particularly crabs” (Scheibling and Robinson 2008). This predation could help explain the widespread observation that, even without a sea urchin fishery, adult sea urchins were hardly ever observed in foliose macroalga–dominated communities (Norderhaug and Christie 2009). Maine’s sea-urchin fishers quickly noticed that, when abundance of red algae such as Chondrus increased, the sea-urchin fishery would probably not recover. Accordingly, sea-urchin fishers in Maine refer to the phase shift to filamentous and corticated macroalgae as “red death.”

Attempts to Break the Lock: Relocating Adult Sea Urchins

To determine whether the alternative sea urchin–free stable state could be broken, we devised an experiment in which we relocated adult sea urchins from distant relict populations to recreate preharvested functional population densities (20–50 m−2; Fig. 2B; Leland 2002). We hypothesized that locally elevating rates of herbivory would reduce abundance of benthic macroalgae, thereby reducing density of micropredators by exposing them to predation and therefore increasing the recruitment potential for settling sea urchins. We therefore expected urchin recruitment to recover in these patches after a decline in macroalgal abundance. In each of 2 yrs, we placed collected sea urchins on isolated islands of subtidal rocky habitat near Cape Elizabeth, Maine. Release plots were surrounded by sediment (a barrier to emigration). In 2000, we in- troduced 3000 adult urchins per plot to nine plots, for a total of 27,000 transplanted individuals, and in 2001, we did the same for eight plots, for a total of 24,000. In both years, we monitored changes in algal, urchin, and urchin-predator populations about 44 BULLETIN OF MARINE SCIENCE. VOL 89, NO 1. 2013

Figure 8. The average abundance of sea urchins and the crab Cancer borealis on nine plots (A, 2000) and eight plots (B, 2001) to each of which 3000 adult urchins were experimentally relocated at the point indicated by the gray triangle. Error bars represent SE. every 2 wks at the release plots and an equal number of control plots. Subsamples of 186 of the relocated urchins were brought to a holding facility in the laboratory as controls for handling and other possible sources of mortality. The results in both years were striking because of the rate at which sea urchins declined after the arrival of large adult C. borealis crabs. In 2000, urchin abundance declined rapidly after relocation (Fig. 8A). After 1 wk, censuses revealed 50% urchin mortality, and we observed large numbers of C. borealis feeding on the relocated sea urchins. Weekly surveys of crab abundance beginning at that point revealed that the highest crab densities (0.8 m–2) were in plots with relocated urchins in late August, but they declined steadily as sea-urchin abundances declined (Fig. 8A). We observed no urchins on any plot 4 wks after reintroduction. In contrast, crab densities on plots without relocated urchins showed no trend and remained low, at 0.25 crabs m–2, throughout sampling in 2000. In 2001, handling protocol was identical to that of 2000, but larger urchins were relocated (all >50 mm diameter), and the release occurred 3 mo earlier (Fig. 8B). The decline in sea-urchin abundance in 2001 also began in August, which in 2001, was several months after relocation (Fig. 8B), and was coincident with the influx of C. borealis. Again, divers observed crabs feeding on urchins at the time urchin density began to decline (Fig. 2B; AVL and JV pers obs). After the 2001 release, urchin densities STENECK ET AL.: ECOSYSTEM FLIPS, LOCKS, AND FEEDBACKS 45

Figure 9. The abundance trends of Cancer borealis by size classes (CW, carapace width) on (A) urchin-relocation plots and (B) control plots without urchins. Arrow indicates date of urchin relocation. Note the rapid and sudden influx of large crabs on the urchin-relocation plots only. remained high (2100 to 2900 urchins per plot; 32–45 urchins m−2) from May through mid-July, and crab density was low (0.1 crabs m−2) on all plots. In August, crab density sharply rose to 0.78 crabs m–2 on the relocation plots, but no such increase was recorded on urchin-free control plots, where crab density remained consistently low (0.2 crabs m–2; Fig. 9). Pooled over all sampling dates, crab abundances were signifi- cantly higher only on the urchin-addition plots (F1,9 = 17.1; P = 0.003). Large C. borealis (75–100 mm carapace width) migrated specifically to the urchin- release plots but not to control plots (Fig. 9), and their densities peaked at 0.4 crabs m–2 in August but were less than 0.1 m–2 in all other months. Population density of the relatively rare C. borealis with carapace width >100 also peaked in August only at the urchin release plots. We conclude that large C. borealis can detect and move specifically to discrete patches of sea-urchin prey. They then leave the area after that resource is depleted (Fig. 8). Although the increase in crab density corresponded to seasonal warm tem- peratures, the crabs’ departure did not. In 2001, water temperature and crab density both increased in late August (to 14.5 °C and 0.78 crabs m−2; Leland 2002), but crab 46 BULLETIN OF MARINE SCIENCE. VOL 89, NO 1. 2013 densities declined by early October even though water temperature remained high (13–14 °C) that year until nearly mid-October (Fig. 8; Leland 2002). The survival of the urchins relocated to the laboratory and maintained in predator- free aquaria as controls for handling effects remained high in 2001. Between 4 May and 25 September water temperature in the six aquaria ranged from 12.6 to 19.6 °C, and urchin survival averaged 97.2% (SE 1.6). Average per capita predation rates between 14 July and 16 August in nine crab- inclusion cages at the laboratory ranged from 1.5 to 2.37 (SE 0.66) urchins per crab per day. At those rates of consumption, the observed crab population densities at the relocation plots were sufficient to account for the mortality of all 51,000 of the relocated sea urchins (Fig. 8).

The Rise of Crabs: A New Apex Predator?

Apex predators are the highest-order carnivores capable of exerting demographic effects on their prey. Historically, populations of crabs and lobsters may have been limited by large predatory finfish (Witman and Sebens 1992, Steneck 1997, Steneck and Carlton 2001). In fact, crab remains are conspicuous in their absence from Maine archaeological sites dating between 4500 and 500 yrs BP (Steneck 1997, Steneck et al. 2004, Bourque et al. 2008). Those sites were dominated by remains of large predatory finfish such as cod, which averaged over 1 m in length for thousands of years before European contact and were therefore formidable predators on crabs (Jackson et al. 2001). Later, 19th-century biologists obtained C. borealis study specimens from the stomachs of cod caught in coastal zones near our relocation experiments (Smith 1879), but today those predators are virtually absent (Jackson et al. 2001, Fig. 1) or no longer function as apex predators (sensu Steneck et al. 2004). Many predatory invertebrates have therefore been released from predator controls (Worm and Myers 2003, Butler et al. 2006). The largest decapod in the North Atlantic, the American lobster, is a relatively benign predator in coastal Maine today because it receives most of its nutrition from lobster trap bait (Grabowski et al. 2010), but crabs are different. Cancer borealis, the largest, is a formidable predator of sea urchins as was evident in our laboratory and field studies (e.g., Fig. 8; Leland 2002) and corroborated by others (e.g., Siddon and Witman 2004). It may be the new apex predator in coastal Maine. Although adult crabs may be apex predators, during settlement they require protection from small visual predators like fishes (Palma et al. 1998, Palma and Steneck 2001). Because C. borealis recruits to shallow beds of macroalgae (Williams and Wahle 1992), the phase shift to macroalgae could increase the recruitment potential of the benthos for them. Accordingly, the west-to-east phase shift that began in the mid-1990s (Fig. 3) could have had local demographic effects on C. borealis. Evidence for local increase in this species at 10 m depth was revealed by a massive coastwide scuba survey of large decapods conducted during the summer of 1997 (see Steneck and Wilson 2001 for methodological details), which revealed that C. borealis was most abundant in boulder habitats in western regions such as Pemaquid and declined toward the northeast (Fig. 10). This pattern corresponded to that of phase shifts to algae at the time (Fig. 2). Although the southwesternmost region, York, harbored lower population densities (possibly because of the lower sample size, i.e., fewer quadrats), that region also had the highest population density of newly settled (young of the STENECK ET AL.: ECOSYSTEM FLIPS, LOCKS, AND FEEDBACKS 47

Figure 10. Substrate-specific population densities of Cancer borealis in 1997 at five study regions in coastal Maine (all at 10 m depth). Region abbreviations: YORK (York region), PEM (Pemaquid Point), PEN (Penobscot Bay), MDI (Mt. Desert Island), JON (Jonesport). Number of study sites in the region is given below each region abbreviation. nd, no data. Error bars are SE. year) C. borealis (Palma et al. 1999). Importantly, that region was also dominated by macroalgae at the time (RSS pers obs; Fig. 3). The sharp increase in C. borealis population densities during the phase shift to macroalgae in the mid-1990s (Fig. 11) was evident in standardized groundfish fall bottom-trawl surveys (US National Marine Fisheries Service semiannual groundfish trawl surveys for the Gulf of Maine) conducted over the last 35 yrs in the three statistical reporting zones (numbered 511–513) covering the coastal shelf span from northeastern Maine to southern Maine. Between 1973 and 1995 density averaged 0.25 (SE 0.05) crabs per tow, but it increased markedly after the algal phase shift beginning in 1996. The higher but variable crab population density average recorded during 1996–2011 was 1.47 (SE 0.2) crabs per tow, more than a fivefold increase over the average for the previous two decades (Fig. 11).

Alternative Stable States: Cascading Processes Reinforce Ecosystem “Flip” and “Lock”

Changes in the kelp-forest ecosystems along the coast of Maine over the past half century (Fig. 2) fulfill the three criteria for alternative stable states set forth by Petraitis and Dudgeon (2004). That is, changes occurred in the same environment, they resulted from a pulse perturbation, and they occurred at a large scale and for sufficient duration to be self-sustaining. We suggest with our studies on the coast of 48 BULLETIN OF MARINE SCIENCE. VOL 89, NO 1. 2013

Figure 11. Temporal trend in Cancer borealis abundance in standardized groundfish trawl surveys. The gray bar indicates the period when most coastal regions in Maine underwent phase shifts to algae (see Fig. 2). National Marine Fisheries Service trawl data for coastal statistical areas along the coast of Maine (areas 511–513). Error bars are SE. The three horizontal lines reflect averages over the period (SE).

Maine that these alternative states meet even the most stringent definition of stability (Connell and Sousa 1983). The rapidity of the “flip” may have resulted from reinforcing feedback between cascading sequential processes of herbivory and recruitment. The reinforcing feedback locks the system into one of two alternative states illustrated in a conceptual model (Fig. 12). When sea urchins dominate, high rates of herbivory (Fig. 2C) result in low algal biomass and reduced habitat architecture (Figs. 2D, 3, 6). Herbivore-resistant CCA dominate under these heavily grazed conditions (Fig. 5A,B; Steneck and Dethier 1994) but offer no protection for recruiting crabs from abundant small predatory fishes (see e.g., Wahle and Steneck 1992), but CCA are good nursery habitats for settling sea urchins, and because they are not preferred prey for resident fishes (Ojeda and Dearborn 1989, 1991, Rowley 1989, Leinaas and Christie 1996, McNaught 1999, Meyer and Byers 2011), the recruitment potential of CCA habitats for settling sea urchins is high. Despite the abundance of young green sea urchins in CCA-dominated communities (e.g., see circle in Fig. 5B), few if any newly settled or older sea urchins live adjacent to macroalgal habitats (Keats 1991, McNaught 1999, Hereu et al. 2005, Christie et al. 2009). The reinforcing feedback may therefore operate at very small scale. Herbivory locally reducing macroalgal habitats may effectively increase urchin recruitment (left side of Fig. 12) thereby perpetuating high rates of herbivory. The sea-urchin fishery (indicated by the square on right side of Fig. 12) rapidly created a large-scale sea urchin–removal experiment. The boom and bust of the fishery created the pulse perturbation to the system (Fig. 2A). In less than a decade, sea- urchin harvesters had vacated the area, but the urchin-free phase persisted. Increases in complexity of habitat architecture increase the recruitment potential of the benthos for crabs, but the newly recruited crabs reduce the recruitment potential of the benthos for sea urchins (Fig. 12). Reduced sea-urchin recruitment results in fewer adult sea urchins and lower herbivory, further increasing the macroalgal habitat architecture (Figs. 2–7, McNaught 1999). This hypothesized improvement of crab nursery habitat increased rapidly during the 1990s and slightly preceded STENECK ET AL.: ECOSYSTEM FLIPS, LOCKS, AND FEEDBACKS 49

Figure 12. A conceptual model illustrating cascading processes hypothesized to flip, lock, and reinforce alternative-stable-state feedback. Coralline = crustose coralline algae. The square represents sea-urchin fishery impact. Asterisks represent key driving processes. Note that crab abundance is a proxy for crab predation. the step increase in adult C. borealis abundance (Fig. 11). Although the role of macroalgae as nursery habitat for this species of crab remains to be demonstrated, the increase in the crab’s population is clear (Fig. 11), and the resulting demonstrably high rates of predation on adult sea urchins could easily account for their continued absence in this ecosystem (see e.g., Figs. 8, 9). Although reports of localized mass mortality in Maine’s sea urchins are common, reports of observed crab predation are not, perhaps because C. borealis predation is fleeting. Our results show that they are highly mobile predators, capable of detecting and consuming high population densities of sea urchins and then rapidly depart- ing. This fleeting but important predator-prey interaction illustrates how easily this association could go undetected. The patchy mortality of urchins could also be the result of disease, but none has been found that is capable of such rapid and patchy epidemiology. Survival of relocated control urchins indicates they were disease free. A surprising feature of North Atlantic sea urchin–algal interactions is how patchy and absolute the shifts between states are in both space and in time. The earliest in situ observation of coastal kelp forest communities in the Gulf of Maine described them as a mosaic having widely spaced but conspicuous CCA–sea urchin patches (Lamb and Zimmerman 1964). During the next two decades, sea urchins and their resultant herbivory increased—widening and coalescing CCA patches reached peak abundance in the 1980s (Steneck et al. 2002). At that time CCA dominated shallow 50 BULLETIN OF MARINE SCIENCE. VOL 89, NO 1. 2013 rock habitats (Steneck 1986, 1990), but a rare scattering of small, isolated urchin-free algal patches (10–20 m in diameter) remained (RSS pers obs). Again, the all-or-none nature of sea-urchin abundance in these patches is striking. Several Norwegian researchers have also observed that, in the eastern North Atlantic, “Adult green sea urchins are seldom found in dense kelp forest (Hjörleifsson et al. 1995, Skadsheim et al. 1995) but may accumulate along the edge of kelp forests or in clearings [patches] within the kelp forest” (Norderhaug and Christie 2009). Similar observations of urchin-deforested patches have been described from Ireland, the UK, Japan, Chile, and New Zealand (reviewed by Steneck et al. 2002). Such widespread but distinct CCA-dominated patches with urchins stand in stark contrast to adjacent areas where urchins are virtually absent. We suggest that the sequential cascading processes related to herbivory and recruitment that we describe in Maine create reinforcing feedback capable of operating at very small spatial scales (i.e., a few meters). Of course the “locks” we describe are not permanent. They probably represent “hysteresis,” from the Greek word for “lag,” but for fisheries managers, protracted lags meeting the definition of alternative stable states represent a significant high- stakes assessment and management challenge. In 1995, Maine’s sea urchin fishery was second only to that for the American lobster in value. At that time the fishery supported over 1500 full-time urchin fishers. Today, the sea-urchin fishery has declined 84% in value, and no full-time fishers remain (MDMR 2010a,b). Nevertheless, we should not lose sight of the big picture. The Gulf of Maine’s coastal ecosystem evolved under strong top-down control by predatory finfish (Estes et al. 2013). The recent rise of crabs as apex predators is therefore only possible because of earlier fisheries-induced declines in their predators (Steneck 1997, Steneck et al. 2004; Fig. 1). Arguably, the ultimate means of breaking the lock of crab dominance may be to manage for increased abundance, diversity, and body size of predatory finfish. Presumably, without crab predation, storms or other physical disturbances would eventually create alga-free patches in which the recruitment potential for sea urchins would be high. With time, herbivory would expand CCA patches, and as they coalesced the regions would revert to a more diverse landscape including patches of sea urchins, CCA, kelp, macroalgae, and associated organisms. The larger point is that marine ecosystems within which commercially valuable species live are complex and can have nonlinear behavior. We cannot use classic fisheries-management approaches such as managing for maximum sustainable yield for highly interactive species (Soulé et al. 2003), especially when managed species have strong (e.g., predator-prey) interactions. We need to understand the drivers of the system and manage them with greater precaution because the consequences of a flip and a lock can be costly, and recovery (i.e., unlocking the system) may be difficult or impossible to achieve over the period of decades.

Acknowledgments

We are indebted to many people for their encouragement, guidance, and help in devel- oping this paper. R Vadas was largely responsible for getting R Steneck actively involved in sea urchin research in the 1970s. C Pfister helped gather data on algae and urchins in the 1980s. Numerous people helped us in the field in the 1990s and 2000s, including E Annis, C Cleaver, R Russell, G Welch, J Higgins, and R Downs. Field assistants also included B Smith, STENECK ET AL.: ECOSYSTEM FLIPS, LOCKS, AND FEEDBACKS 51

A Haupt, R Beas, M Baird, S Parks, L Harmon, S Nayyar, E Horan, R Howell, C Harbison, J Kassakian, B Laberge, D Drisco, and S Zimsen. Team leader for the massive Chondrus surveys was T Suskewitz. Urchin harvesters included J Berke, L Gardner, B Preney, R Odlin, J Reidy, and P Fischer. This research benefited from conversations with L Watling, R Wahle, P Yund, and R Vadas. L Jacobson from NOAA’s Northeast Fisheries Science Center in Woods Hole, MA, provided the federal groundfish survey data. R Vadas, J Steneck, and C Cleaver commented on earlier drafts of this paper. The National Science Foundation (OCE83151136, OCE8600262), NOAA’s National Undersea Research Center at the University of Connecticut, Maine’s Department of Marine Resources, Maine’s Sea Urchins Zone Council, Maine Sea Grant, a Pew Fellowship in Marine Conservation, and the University of Maine’s Association of Graduate Students funded this research. We are grateful to all.

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Date Submitted: 4 January, 2012. Date Accepted: 4 April, 2012. Available Online: 13 November, 2012.

Addresses: (RSS) School of Marine Sciences, University of Maine, Darling Marine Center, Walpole, Maine 04573. (AL) Environmental Defense Fund, 1875 Connecticut Ave., NW, Washington, DC 20009. (DCM) Department of Environmental and Biological Sciences, University of Maine, Machias, Machias, Maine 04654. (JV) Marine Sciences Laboratory, Pacifi c Northwest National Laboratory, 1529 West Sequim Bay Road, Sequim, Washington 98382. Corresponding Author: (RSS): Email: .