Loss of predators and the collapse of southern California kelp forests (?): Alternatives, explanations and generalizations

Loss of predators and the collapse of southern California kelp forests (?): Alternatives, explanations and generalizations

Journal of Experimental Marine Biology and Ecology 393 (2010) 59–70 Contents lists available at ScienceDirect Journal of Experimental Marine Biology...

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Journal of Experimental Marine Biology and Ecology 393 (2010) 59–70

Contents lists available at ScienceDirect

Journal of Experimental Marine Biology and Ecology j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / j e m b e

Loss of predators and the collapse of southern California kelp forests (?): Alternatives, explanations and generalizations Michael S. Foster a,⁎, David R. Schiel b a b

Moss Landing Marine Laboratories, 8272 Moss Landing Road, Moss Landing, California 95039, USA Marine Ecology Research Group, School of Biological Sciences, University of Canterbury, Private Bag 4800, Christchurch, New Zealand

a r t i c l e

i n f o

Article history: Received 28 April 2010 Received in revised form 30 June 2010 Accepted 5 July 2010 Keywords: Ecosystem collapse Kelp forests Macrocystis Sedimentation Southern California Water quality

a b s t r a c t It is increasingly argued that human-induced alterations to food webs have resulted in the degradation of coastal ecosystems and even their “collapse.” We examined the evidence for this argument for Macrocystis pyrifera (giant kelp) forests in southern California. Others have concluded that forests in this region collapsed between 1950 and 1970 as a result of sea urchin grazing driven by overfishing of sea urchin predators (sheephead wrasse and spiny lobsters) and competitors (abalone), and that the kelp forests recovered but are currently sustained as a result of a commercial sea urchin fishery that began in the early 1970s. Our examination of the historical record, primary publications, and previously unpublished data showed that there was no widespread decline in the region between 1950 and 1970, but there were localised declines in mainland kelp forests near the rapidly growing cities of Los Angeles and San Diego. The preponderance of evidence indicates that kelp losses were caused primarily by large increases in contaminated sewage discharged into coastal waters, sedimentation from coastal development, and the 1957–1959 El Niño. Increases in active sea urchin foraging were most likely a secondary effect following dwindling food resources. The forests recovered when sewage treatment improved and sewage outfalls were relocated. The effects of fisheries were explored by correlation analysis between kelp canopy cover and commercial sea urchin landings, and among fisheries landings for sea urchins, abalone, sheephead and lobster. These correlations were generally insignificant, but were often confounded by differences in the spatial scale over which the data were collected as well as factors other than simple abundance that affect the fisheries. However, where area-specific data were available, the landings of sea urchins generally tracked kelp abundance, most likely because roe (gonad) development is directly related to food availability. A literature review showed that although sheephead and lobsters may control sea urchin abundance at small spatial scales within some sites, there is little evidence they do so over large areas. That abalone and sea urchins compete, such that sea urchins increased as a result of abalone harvesting, is largely conjecture based on their similar habitat and food utilization. This study shows that kelp forests in southern California did not collapse, and that declines in some coastal sites were caused primarily by degradation of water quality, increased sedimentation and contamination, and unfavorable oceanographic conditions. We conclude that management by species' protection or reserves will not be effective if poor habitat quality impacts the ability of giant kelp to survive and thrive. © 2010 Elsevier B.V. All rights reserved.

1. Introduction

“Coincident with the increased discharge of sewage effluent, harvestable kelp has virtually disappeared in the vicinity of Whites Point for a distance of two or three miles along the coast.”—Revelle and Wheelock (1954)

⁎ Corresponding author. Tel./fax: + 1 831 786 8853. E-mail addresses: [email protected] (M.S. Foster), [email protected] (D.R. Schiel). 0022-0981/$ – see front matter © 2010 Elsevier B.V. All rights reserved. doi:10.1016/j.jembe.2010.07.002

An increasing number of published papers argue that humaninduced alterations to food webs have resulted in the degradation of coastal ecosystems worldwide (e.g., Steneck et al., 2002; Springer et al., 2003; Bellwood et al., 2004; Berkes et al., 2006; Worm et al., 2006) and even their “collapse” (Jackson et al., 2001). A central argument for this is that overfishing over long time periods has caused the loss of structural and functional components of these ecosystems compared to their state hundreds or even thousands of years ago. In the case of coastal kelp ecosystems, the premises for this argument are that that top predators such as sea otters, fish and lobsters have been fished to ecological extinction and that this has led to population explosions of herbivores, which results in deforestation (Estes et al., 1989; Jackson et al., 2001). Much of the evidence comes from

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northern hemisphere kelp beds and forests, particularly in the northwestern Atlantic (e.g., Mann and Breen, 1972; Steneck et al., 2004; but see Elner and Vadas, 1990) and the eastern Pacific along the western coast of North America. In the latter region there has been considerable ecological and political interest in the predator-urchinkelp interaction since North and Pearse (1970) first suggested it as a possible explanation for the dynamics of some mainland giant kelp (Macrocystis pyrifera (L.) C. Agardh) forests in southern California. Hypotheses concerning this interaction became of general interest because sea urchin predators, sea urchins and kelp are common features of temperate nearshore communities worldwide (Lawrence, 1975), and the interaction could inform larger debates over the generality of predator or keystone control of prey populations (review in Power et al., 1996) and the occurrence and frequency of alternate stable states (Connell and Sousa, 1983). If kelp systems are now truly collapsed due to the removal of predators, then there are considerable implications for urgent management decisions to resurrect coastal kelp communities, such as through no-take reserves. Alternatively, if a predator–urchin–kelp interaction is not the primary factor affecting kelp forests, or if these communities have not actually collapsed, then other forms of management may be more appropriate. We review evidence from southern California about the status of kelp forest communities through time and the effects of environmental influences (frequently termed “drivers”) that are known to account for variation in kelp forest structure, and relate these to what is known about sea urchins, their competitors and predators. We argue that it is of debatable value to surmise that pre-historical states of kelp, lacking spatial and temporal resolution, represent a useful natural state against which to gauge present status, but rather that a “strong-inference” approach (Platt, 1964) allows not only fundamental understanding of underlying processes, including human impacts, but also points to testable hypotheses at relevant scales. Elner and Vadas (1990) argued for a similar approach in their analysis of sea urchin population dynamics in the northwestern Atlantic. Our purpose is not to dispute that degradation of kelp forests has occurred in many places or that there have not been declines in the numbers and sizes of many large predators worldwide. Rather, we review the literature and offer a contrasting perspective on the status of kelp forests along southern California coasts and the underlying processes that drive their structure. 1.1. Background Kelp communities comprise the dominant biogenic habitat along temperate and boreal rocky shores worldwide (Mann, 1973). Giant kelp forests are particularly common along eastern Pacific shores, support communities of thousands of species, and are known to be affected by a wide range of influences that alter their local distribution, abundance and biogeographic distribution (North, 1971a; Foster and Schiel, 1985; Graham et al., 2007). Because of their great ecological and cultural importance there are legitimate concerns about their status and the factors that affect them, which has been the case since at least the early 1950s (Revelle and Wheelock, 1954). Giant kelp produces a floating surface canopy that can contain N60% of the entire plant biomass (North and Hubbs, 1968) making assessment of abundance based on canopy area from surveys at sea or from the air relatively simple. Along the coast of California, some assessments of kelp canopies are available beginning in the late 1800s. More comprehensive surveys were done occasionally from 1911 (Crandall, 1915), and detailed knowledge of kelp forest structure began with the advent of diving studies in the 1940s (Andrews, 1945; Aleem, 1956; Dawson et al., 1960; North and Hubbs, 1968). Therefore, there is considerable knowledge about the structure of these kelp forests for well over 60 years. These studies pre-dated much of the intense fishing pressure that has been exerted on coastal fishes (Leet et al., 2001) and urban expansion along the coastline and, therefore, offer at least decadal-

scale baselines, although with varying spatial and temporal resolution, against which to gauge the current status of kelp forests. A central question about gauging kelp-based ecosystems is how far back we must look for comparisons. The notion of sliding or shifting baselines is well-recognised in both the ecological (Dayton et al., 1998) and fisheries (Pauly, 1995) literature, providing a necessary caveat to assessments about change because of the potentially degraded condition of an ecosystem used as an initial comparison. However, reconstructions of an idyllic past state before human impacts presumably occurred are also fraught with difficulties because they obscure temporal and spatial variability, which are known to confound even short-term assessments of change (e.g., Underwood, 1992), and they are usually based on presumption and scant evidence. Crucial to a deep historical perspective is the assumption that the once intact trophic relationships of numerous species that have now been overfished to ecological extinction had exerted influences that produced major structural and functional differences in kelp ecosystems before their presumed collapse (Jackson et al., 2001). Trophic control of structure and function are, therefore, central to arguments about kelp forest status. There is an old, rich and varied literature from kelp forests worldwide (Lawrence, 1975) that demonstrates the numerous ways that sea urchins, the dominant herbivores of kelp communities, can remove large tracts of kelp, reducing them to a state often called “barrens” even though they can be highly speciose (Begin et al., 2004; Graham, 2004). Such areas can persist for many years (e.g., Chapman, 1981). This is also the case for southern California, where “overgrazing” by sea urchins caused considerable concerns and management intervention as far back as 1959 (North, 1959a,b). The nature and extent of these effects have been examined both experimentally and through surveys at hundreds of sites (Foster and Schiel, 1985, 1988). The other link in the chain of trophic control, crucial to arguments about the effects of overfishing, is that of top predators on urchins. One argument is that some historical and published evidence (which we review in this paper) shows that southern California kelp forests collapsed in the 1950s to 1970s due to sea urchin overgrazing, and this was an indirect result of overfishing of lobsters, sheephead wrasse and abalones (Jackson et al., 2001). Kelp forests in this region now exist, therefore, only because a sea urchin fishery, which began in the early 1970s, has functionally replaced these former predators and competitors. This has also been called a phase shift, referring to the altered state of kelp forests (Steneck et al., 2002). Although Steneck et al. (2002) state that pollution and other factors affected southern California kelp forest dynamics, they highlighted the roles of diverse sea urchin predators and competitors acting as buffers from systematic deforestation. These claims underpin important marine conservation and management imperatives because if ecosystems have collapsed and if the direct cause is overfishing, then ameliorating fishing effects with management strategies such as no-take reserves should not only increase fish stocks but indirectly resurrect coastal kelp communities from their putative collapsed states (e.g., Jackson et al., 2001; Lubchenco et al., 2007; Palumbi, 2008). Many studies show that there is a great deal of natural variability in the areal extent of kelp forests associated with storms and longer term climatic events such as El Niño–Southern Oscillation (Dayton and Tegner, 1984; Tegner and Dayton, 1987; Foster and Schiel, 1993; Edwards, 2004). Entire giant kelp canopies and much of the understory assemblage can disappear or be severely reduced in abundance following intense El Niño periods, and recovery can take several years. This background of variability from natural disturbances is not just the noise against which human-induced changes must be gauged but is fundamental to the population dynamics and natural history of kelp forest communities (e.g., Dayton et al., 1999; Schiel and Foster, 2006). Algal-dominated communities worldwide are known to be affected by a wide range of multiple stressors from altered coastlines and landuse practices over many years, often decades (Airoldi, 2003). We do

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not dispute that predators can have large effects on kelp forests dynamics in some areas (e.g., Aleutian Islands; Estes and Palmisano, 1974). Given the numerous factors that can impact kelp forests, however, it is both reasonable and necessary to determine their relative influences and extent in different circumstances if understanding underlying mechanisms leading to better management and restoration are goals. To achieve this, it is necessary to understand 1) the current status of kelp communities, 2) the influences of trophic interactions on kelp forest structure relative to other sources of natural variability, and 3) human-induced factors that may influence structure and function and the degree to which these have degraded kelp forests. We examine these by using data sets from fishery landings, surveys, monitoring and experiments in the kelp forests of southern California, an area that has been the focus of arguments about collapsed ecosystems. 2. Methods Our sources of information were as comprehensive as possible. 1) We reviewed the numerous publications and reports relating to the status of kelp forests in southern California from the early 1900s onwards. These were by primary researchers working in kelp forests at the time and contained many first-hand observations as well as data. 2) Primary data sets were obtained from publications and reports that are generally available, and are cited accordingly. 3) Archived (unpublished) data were obtained by making requests to public agencies and private companies involved in fisheries assessment and kelp forest management. Collectively, these were used to construct an historical record of the status and factors affecting southern Californian kelp forests since the early 1900s. Although we have also considered the wider status of kelp forests in California, we focused our efforts on studies of the Palos Verdes kelp forest off Los Angeles and the Point Loma kelp forest off San Diego (Fig. 1). These two mainland southern California kelp forests were those used primarily by others to

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formulate their conclusions about trophic control of California kelp forest ecosystems (e.g., Jackson et al., 2001; Steneck et al., 2002). Jackson et al. (2001) do not define “collapse.” We assume it was used to characterize abundance when it fell below 10% of maximum since Jackson was one of the authors of Worm et al. (2006) and the latter used “catches dropping below 10% of the recorded maximum” as the definition of fisheries collapse. We use giant kelp surface canopy area as the metric and the 10% figure as an indication of collapse, with the largest verified abundance as the reference point. 3. Results and data review 3.1. Kelp forest dynamics in California: evidence from harvesting records Large, offshore stands of giant kelp occur over most of the coast of California from Monterey Bay southwards, but are patchy and usually depend on the presence of rocky reef (Foster and Schiel, 1985). The generally robust status of kelp forests along coastal California for the past century can be surmised from the kelp harvesting record (Fig. 2) which shows the harvest generally stayed above 80,000 tonnes from 1948 onwards. Kelp beds are numbered and licensed, and were harvested when canopies were extensive. The primary processing plant was in San Diego, and most of the harvest came from southern California. As harvesting increased from the early 1940s, low periods of harvest invariably coincided with El Niño events when the kelp surface canopy cover was low, especially the repeated events after 1980. However, except for large coastal runoff associated with high rainfall in 1979 (Fisk, 2010) followed by the prolonged and intense El Niño of 1982–83, the most intense after 1950 when detailed records became available (NOAA, 2010), the kelp harvest was high. The decline in harvest after 1990 was related to decreased kelp canopy after the 1991–1992 El Niño and a deteriorating US market for kelp products as Asian alginate production dramatically increased (Bedford, 2001), and not due to the collapse of kelp forests.

Fig. 1. Southern California: mainland coast and offshore islands.

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Fig. 2. Annual harvest of giant kelp, Macrocystis pyrifera, along the California coast. Shaded bars are El Niño periods. Kelp data from Bedford (2001).

3.2. Kelp forest dynamics in southern California: evidence from canopy surveys and field observations Southern California is the region from Point Conception to the Mexican border, including the offshore islands (Fig. 1). Kelp canopies along the mainland in the region from Los Angeles southwards generally declined in areal extent in the 1950s–1970s and those at Palos Verdes and Point Loma (Fig. 3) declined to near zero cover (North, 1970; Wilson et al., 1980; Meistrell and Montagne, 1983). The canopy area of forests in southern California from Point Conception south to the Palos Verdes peninsula, however, changed very little (Harger, 1983) and large declines were not noted around the offshore Channel Islands (Neushul, 1965; Clarke and Neushul, 1967; Foster, 1975; Foster pers. obs.). The forests in these latter sub-regions can comprise 85% of the total kelp canopy area in southern California (based on mainland canopy cover data from the 1930s–1979 (Harger, 1983) and mainland and Channel Islands data from 1967 (Bedford, 2001)). Data from canopy maps show that canopy cover after the late 1960s along the mainland from just south of Los Angeles (Huntington Beach) to the Mexican border was between 40 and 60% of the

historical maximum, depending on which historical sample is used (North, 1970; North and MBC Applied Environmental Sciences, 2001). Since the 1980s there have been large, state-wide declines in canopy cover followed by recovery as a result of El Niños, as reflected in harvest records (Fig. 2) and canopy dynamics where high resolution data are available (Fig. 3). Furthermore, this has ramifications for the wider kelp community. Dayton et al. (1999) found that inter-annual and inter-decadal variation in ocean climate affected giant kelp with consequent large and lasting long term effects on understory kelps via canopy competition for light. The exceptions to recovery are a few large canopies just north of Santa Barbara which were produced by plants growing on sand; these forests have not recovered since they were removed by the intense storms of the 1982–1983 El Niño (Bedford, 2001). 3.3. Kelp forest dynamics at Palos Verdes and Point Loma: the effects of sewage discharge, coastal development, sedimentation and El Niños There is strong evidence that sewage discharges, sedimentation from coastal development, and the El Niño of 1957–1959, were the

Fig. 3. Giant kelp canopy area of the Palos Verdes and Point Loma kelp forests. Events indicated by arrows: 1, installation of shallow sewage outfall at Palos Verdes; 2, sediment discharge from Mission Bay and sewage discharge from San Diego Bay at Point Loma; 3, 1957–59 El Niño; 4, sewage discharge moved offshore at Point Loma; 5, beginning of sea urchin fishery in southern California and advanced sewage treatment at Palos Verdes; 6, El Niño events after 1959; 7, record southern California rainfall and runoff (Fisk, 2010). Palos Verdes kelp canopy area from Wilson et al. (1977; 1911–1973) and J. Gully (pers. com.), Los Angeles County Sanitation District (1974–2007). Point Loma kelp canopy area from North (1975; 1911–1967), North and MBC Applied Environmental Sciences (2001; 1968–2000), and MBC Applied Environmental Sciences (2007; 2000–2006).

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main causes of severe declines in mainland kelp forests near the major metropolitan areas of southern California, particularly those at Palos Verdes and Point Loma. Revelle and Wheelock (1954) were the first to report the effects of sewage discharge on the Palos Verdes kelp forest, followed by the extensive studies of W.J. North and colleagues beginning in 1957. These studies were stimulated by observations that “considerable kelp disappeared in the last 15 years in the vicinity of Los Angeles and San Diego,” and this “raised the question of whether increasing waste discharges have adversely influenced kelp and associated marine resources” (North, 1958). Los Angeles and San Diego were the two most rapidly growing metropolitan areas of southern California during this period. Primarytreated domestic and industrial wastes from the Los Angeles area were discharged through ocean outfalls at 18 m–90 m depths located just north (Hyperion outfall) and within the southern edge (off Whites Point; outfall completed in 1937; Fig. 1) of the kelp forest at Palos Verdes. Between 1940 and 1960, waste discharge through these outfalls increased from 152 to 950 × 106 l/day, and suspended solids in the discharge increased from 15,000 to 125,000 tonnes/year (Meistrell and Montagne, 1983). The discharge included a variety of chlorinated hydrocarbons and metals, including large amounts of copper and zinc, which are known to be toxic to the early life stages of giant kelp (Anderson and Hunt, 1988; Anderson et al., 1990). Water and benthos quality were further reduced by sedimentation from coastal development along the Palos Verdes peninsula (Wilson et al., 1980) and expansion of Los Angeles Harbor (Revelle and Wheelock, 1954) immediately south of Palos Verdes. By 1958, the Palos Verdes kelp forest consisted of a few small stands (c. 0.13 km2) in shallow (b8 m) water, 5–11 km north of the Whites Point sewer outfall (North, 1958, 1959b, 1960, 1975; Wilson et al., 1977). There was also a strong, negative correlation between canopy cover at Palos Verdes between 1947 and 1979 and the amount of suspended solids discharged from sewer outfalls (Fig. 4; r = −0.925; Wilson et al., 1980; Meistrell and Montagne, 1983). Over 150 × 106 l/day of minimally treated domestic and industrial wastes from the San Diego area were discharged into San Diego Bay in 1952, turning Bay waters “from a dark blue to a murky greenish brown hue” (Jamieson, 2002). The bay entrance is within the southern end of the historical distribution of the Point Loma kelp forest (North, 1991). The former northern end was near the entrance to Mission Bay, which was extensively developed in the late 1940s. Considerable dredging was done, resulting in sediment plumes extending offshore. These, plus diversion of sand by the entrance jetties, may have converted former kelp habitat into soft bottom (North, 1964, 1991; North and MBC Applied Environmental Sciences, 2001). By the late 1950s, the Point Loma kelp forest was reduced to a few stands (c. 1.6 km2), concentrated near the historical upper margin of the forest, midway between the entrances of Mission and San Diego Bays.

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That the alteration in water and benthos quality was the primary direct cause of the initial, very large declines in these kelp forests is strongly indicated by the population biology of giant kelp relative to conditions observed within the beds and patterns of bed decline at the time. The microscopic gametophyte and young sporophyte stages of giant kelp occur on the benthos where they depend on light transmitted through the water column and sediment-free substrates for attachment, reproduction, growth and survival. Kelp forests near these large sewage outfalls experienced a decrease in benthic light from discharged particulates in the water column (North, 1964; Wilson et al., 1980), increased sedimentation on the benthos, and an increase in discharged nutrients that stimulate phytoplankton growth in the absence of a kelp canopy (Eppley et al., 1972). Sedimentation also increased within affected kelp forests. Revelle and Wheelock (1954) noted “deposits of silt, slime and flocculate, suspended solids” in the vicinity of the Whites Point sewer outfall at Palos Verdes. Grigg and Kiwala (1970) reported fine grain sediment 0.1–1.2 cm thick at historic kelp forest depths between 14 and 20 m. The sediment extended nearly 10 km along an area of the Palos Verdes peninsula formerly occupied by giant kelp, and sediment depth decreased with distance from the Whites Point sewer outfall. Divers at the time commonly noted thin coatings of organic material (leptopel or floc) and very fine sediment on benthic surfaces (North, 1967, 1971b). The microscopic stages of giant kelp do not survive burial and abrasion by even a very thin sediment layer (Devinny and Volse, 1978), and such sediments can also harbor high concentrations of toxic chemicals (Schiff et al., 2000) that impede gametogenesis. Kelp canopy cover throughout southern California was further reduced by the warm nutrient-depleted conditions associated with the 1957–1959 El Niño. Giant kelp sporophyte recruitment is inhibited and sporophytes deteriorate as a result of low nutrients associated with elevated water temperatures (N17 °C) during El Niños (Jackson, 1977; Dean and Jacobsen, 1986; Deysher and Dean, 1986), and these effects are particularly strong in the warmer waters of southern California (Foster and Schiel, 1985; Edwards, 2004). North (1959a) reported temperatures of 20 °C or more persisting well into November, 1958 with “harsh” effects on all kelp forests in southern California. Local kelp harvesters reported less surface canopy than at any time since harvesting began in 1929 (North, 1959a), and this is reflected in the kelp harvest (Fig. 2). By 1959, the canopy cover was reduced to b0.03 km2 at Palos Verdes and b0.6 km2 at Point Loma (Fig. 3; North, 1960, 1975; Wilson et al., 1977). The continued effects of sewage discharge during this period are indicated by North's (1964) observation that, by late 1959, many kelp forests along the southern mainland of southern California had begun to recover from the 1957–1959 El Niño (Fig. 2) but those at Palos Verdes and Point Loma did not (Fig. 3), an apparent example of the effects of multiple stressors.

Fig. 4. Palos Verdes giant kelp canopy area and mass emission rates of suspended solids from ocean sewage discharges off Whites Point. Data from Meistrell and Montagne (1983) and J. Gully (pers. com.), Los Angeles Country Sanitation District.

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The recovery of kelp canopies after 1957–1959 El Niño in both Palos Verdes and San Diego was strongly coincident with improvements in water quality (Figs. 3, 4). Between 1971 and 1981, suspended solids in the sewage discharge at Palos Verdes were gradually reduced from 375,000 to c.75,000 tonnes/year, during which time the kelp canopy increased to c.2.5 km2 (Fig. 4; Wilson et al., 1980; Meistrell and Montagne, 1983). In 1963, most of the wastes formerly discharged into San Diego Bay were treated and then discharged through a new offshore ocean outfall at a depth of 67 m (Jamieson, 2002). The kelp canopy around Point Loma began to recover within a year (Fig. 3), covering nearly 6 km2 by 1970 (North, 1991). This is ~ 50% of the historic maximum, not 6% as estimated by Steneck et al. (2002). Since 1970 most major periods of abrupt decline have been associated with large El Niño events and large storms (Fig. 3). 3.4. The importance of grazing by sea urchins: observations and fishery data There is strong evidence that grazing by sea urchins, especially the purple sea urchin Strongylocentrotus pupuratus (Stimpson) and red sea urchin S. franciscanus (Agassiz), contributed to kelp canopy decline and slowed recovery during the 1950s–1970s at Palos Verdes and Point Loma (North, 1964; Leighton et al., 1966). North (1959b) hypothesized that active grazing on attached plants by sea urchins was initially stimulated by a reduction in the supply of adult kelp plants, and therefore of algal drift, as a consequence of the lack of replacement due to declining water quality associated with sewer outfalls and coastal development. Divers in the 1950s noted high concentrations of sea urchins (especially red urchins) and gastropods, but these were found primarily along the seaward edge of the few shallow kelp stands that remained, apparently having migrated shoreward as their food resources declined in deeper water (North, 1959b, 1964). North (1959b) further observed that areas obviously deforested by sea urchins also occurred at sites away from outfalls, but these areas were small such that “one did not have to swim far to find rich algal growth.” Early investigators reported that sea urchins seemed to subsist primarily on drift algae when algal abundance was high but switched to active foraging on attached plants when algal abundance was low, creating barrens (e.g., North, 1967). This interaction between algal abundance and sea urchin behaviour has since been repeatedly observed in southern California (Dean et al., 1984; Ebeling et al., 1985; Harrold and Reed, 1985). These studies showed that active grazing is induced by storms or high temperatures and low nutrients that reduce drift algal abundance, and ceases when suitable oceanographic conditions occur that facilitate algal settlement and growth or when there is a decline in urchin abundance due to storms or disease. Declining food resources can also lead to kelp mortality from grazing fishes that concentrate around remaining plants (North, 1968; Wilson et al., 1977). In the case of Palos Verdes and Point Loma, however, poor water and benthic quality (discussed above), and not storms or other oceanographic events, most likely caused the initial, severe reduction in algal abundance and inhibited new recruitment and growth, which led to continued active grazing until sewage treatment and disposal were improved. Based on the success of local-scale removals of sea urchins at Point Loma, Pearse et al. (1970) concluded that sea urchin grazing was the direct cause of kelp disappearance. Large-scale removal of sea urchins by hand or by spreading quicklime (calcium oxide) pellets began at Point Loma in 1962. Removal efforts were concentrated in areas of high sea urchin densities around the margins of existing kelp stands, and often resulted in an increase in stand area (North, 1966, 1967). However, these increases occurred coincident with moving the San Diego sewer discharge well offshore (discussed above). That changes in water quality were most important is suggested by North's (1966) observations that vegetation also appeared in areas not treated with quicklime, and that the “Point Loma bed is still expanding largely by

natural processes.” These observations indicate that improvements in water quality enabled giant kelp and other macroalgae to re-colonize former habitat and the removal of sea urchins increased the rate of kelp recovery. Improvement in sewage treatment is further implicated as the primary factor affecting forest dynamics by the timing of the recovery of giant kelp at Palos Verdes (Fig. 3). Sea urchin removals and other restoration efforts were started there in the early 1960s but the forest did not begin to recover until 1971 (North, 1964; Wilson et al., 1977) when sewage treatment improved. As at Point Loma, recovery occurred around removal and restoration sites, but also elsewhere (Wilson et al., 1977). A commercial fishery for sea urchins began in southern California in the early 1970s (Fig. 5A). Comparing landings from the sea urchin fishery with kelp abundance could ostensibly provide insight into the interaction of the fishery with the status of kelp beds. In practice such comparisons are difficult to interpret, in part because the spatial resolution of the data is not the same for each metric. Also, changes in landings do not necessarily reflect changes in kelp abundances; for example, urchin landings decline during El Niño years and when prices decline (Kalvass and Rogers-Bennett, 2001) and, as noted above, when drift algae are sparse sea urchins may switch to active foraging and have a greater effect on attached plants. Nevertheless, any conclusions regarding fishing as a replacement for lost predatory control must consider the status of the fishery. Sea urchin fishery data are commonly reported state-wide (e.g., Kalvass and Rogers-Bennett, 2001) although with the exception of 1986–1989, most of the fishery was in southern California (CFG, 2004). Both red and purple sea urchins can cause deforestation (review in Foster and Schiel, 1985). The sea urchin fishery in southern California was comprised only of red urchins until 1989, after which relatively small catches of purple urchins were also taken (Kalvass and RogersBennett, 2001; Parker and Ebert, 2001). As for kelp harvesting, the catch of sea urchins generally decreased during years with strong El Niños (Fig. 5A). The fishery increased steadily from 1976 to 1980, but this was well after the time when the kelp forest at Point Loma began recovery (Fig. 3). The recovery of the Point Loma kelp forest in the 1960s, therefore, was not coincident with the removal of large numbers of sea urchins for commercial purposes. The recovery at Palos Verdes began later than at Point Loma, just before the urchin fishery began. The kelp canopy at Palos Verdes did continue to expand as the urchin fishery developed (Figs. 3, 5A) but the expansion was highly correlated with improvements in water quality (Fig. 4). The relationship between the urchin fishery and the cover of kelp forests after the 1982–83 El Niño, however, is equivocal. Although not in the same spatial scales, there is a significant correlation between the combined Point Loma and Palos Verdes kelp canopy cover and the size of sea urchin landings from the southern California mainland (r29 = 0.5316, p = 0.003). The peak urchin fishery along the mainland was 4500 tonnes in 1989, within a few years of peak canopy cover (Figs. 3, 5A). There was a c. 75% decline in the urchin fishery between 1989 and 2000, which generally coincided with reductions in kelp canopies but also with repeated intense El Niño events. At Point Loma, where urchin landings and kelp data were both recorded in the same coastal sections after 1978, there was a variable but revealing relationship between urchins and kelp cover (Fig. 6). Both kelp cover and urchin landings were low during El Niño periods, and landings were generally high when giant kelp was abundant. During periods of intermediate canopy cover, however, urchin landings could also be high. This may reflect minimum thresholds of kelp availability needed for gonad development. The state of gonads, weather conditions and the condition of the urchin roe market are all variable and tend to weaken the relationship between kelp cover and urchin harvest. Furthermore, the urchin fishery declined as permits were restricted in the late 1980s (CFG, 2004). Taken together, these data are useful mainly in showing that they provide little to no support for concluding that the sea urchin fishery

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Fig. 5. A. Commercial landings of all abalone species in California and red sea urchin (Strongylocentrotus franciscanus) landings for southern California and southern California mainland. B. Commercial landings of California spiny lobster (Panulirus interruptus) and sheephead wrasse (Semicossyphus pulcher) in southern California. Data from Haaker et al. (2001; abalone), CFG (2004; sea urchins), Barsky (2001; lobster) and Stephens (2001; sheephead). Red sea urchin data not available for sub-areas of southern California prior to 1981. Shaded bars are El Niño periods.

is a driver of kelp abundance. The evidence is that the opposite is more likely, that urchins are fished when kelp forests are robust and provide more algal drift, supporting good gonad development and therefore greater recovery of urchin roe. 3.5. The importance of competitors and predators of sea urchins: fishery data North and Pearse (1970) first suggested that abalone and sea urchins may compete for space and food such that a decrease in

abalone might lead to an increase in sea urchins. The abalone fishery was strong throughout most of the 20th Century except during World War II, but there was a dramatic decline after 1979 (Fig. 5A) due to overfishing in southern California and to the combined effects of overfishing and sea otter predation in central California (Haaker et al., 2001). A cursory look at landing data might suggest that as abalone were becoming overfished and less abundant, sea urchins became more abundant (as reflected in landing data; Fig. 5A). However, urchins represented an essentially new fishery which was exploited as abalone became less available (Kalvass and Rogers-Bennett, 2001).

Fig. 6. Giant kelp canopy area and commercial sea urchin catch at Point Loma, 1978–2004. Kelp canopy area from Fig. 3. Sea urchin data from P. Kalvass (pers.com.), California Department of Fish and Game. Data incomplete prior to 1978.

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During the 1980s, many abalone fishers began harvesting sea urchins. This is reflected in the negative correlation between urchin and abalone landings between 1971 and 2001 (r31 = − 0.4141, p = 0.021), which were both mainly in southern California. Between 1975 and 1981 they peaked together, but after 1984 the urchin fishery largely replaced the abalone fishery. Tegner (1980) was the first to suggest that in the absence of sea otters, sheephead wrasse (Semicossyphus pulcher Ayres) and California spiny lobsters (Panulirus interruptus Randall), might control sea urchin populations via predation in southern California. Statistics from the California lobster fishery (Barsky, 2001) are unclear on this point. The lobster fishery persisted throughout the 20th Century, with a maximum of c. 420 tonnes and a minimum of c. 80 tonnes during WWII (Fig. 5B). Lobsters were consistently fished at between 100 and 400 tonnes annually after 1976, presumably reflecting their relative abundance (cf., the abalone fishery data), but there was no obvious response in urchin landings that would shed light on competitive release. During this period, there was only a very weak negative correlation between lobster and southern California urchin landings (r24 = − 0.296, p = 0.20), but this was mostly a reflection of fishers switching from overfished abalone to sea urchins. There is no correlation between lobsters and any measure of kelp abundance or harvest. Sheephead wrasse are common in nearshore waters of southern California and feed on a variety of prey, including sea urchins and lobsters (review in Cowen, 1983). Fishing them has been implicated as a cause of increases in sea urchins (e.g., Jackson et al., 2001; Steneck et al., 2002). Sheephead were fished throughout the 20th Century, with annual landings generally below 50 tonnes, and a maximum catch of 185 tonnes in 1998 (Fig. 5B; Stephens, 2001). Sheephead landings are not significantly correlated with that of urchins (r19 = −0.3617, p = 0.128). Although none of these landing data are necessarily good indicators of abundance of any species, which are fished for a variety of reasons and intensities over many years, they offer little support for arguments of trophic control of kelp through fishing pathways and, consequently, allow the conjecture that oceanographic and physical drivers have had the major influences on kelp abundance over the past century. 4. Discussion Piecing together disparate data on long term processes is fraught with difficulties. The data themselves may not always have been collected reliably, there were usually different purposes and reasons for their collection, and the spatial scales of data from different taxa are usually different. This leaves considerable room for varied analyses and interpretation, based on how data were selected and often on broad correlations. Extending these to conclusions about causality is a considerable step further. Here we have tried to go back to all known sources of primary data, both published in the primary literature and in reports, many of which are well-cited in the literature. Our premise is that the work and views of the researchers who collected the data should be extensively reviewed and considered in any syntheses that offer new interpretations of the causes of past events. Two scenarios, one based on overfishing of predators and competitors and the other based on changes in water quality and oceanographic events, could produce roughly the same changes in kelp forest composition and extent over time, and roughly the same outcome. Considering all evidence is therefore crucial in determining where the balance of probability lies.

that these were invariably associated with El Niño events, and the warm water, low nutrients and storms they bring to coastal waters (Figs. 2, 3). The data also show that there were severe declines in kelp forests near large metropolitan areas along the southern portion of the mainland coast of southern California, associated with coastal development, sedimentation and sewage outfalls. However, recovery of kelp occurred when stressors were reduced, and there was no evidence for a general collapse of kelp forest structure or a permanent shift to a largely kelp-free state. The data show that the canopy dynamics of kelp forests at Point Loma and Palos Verdes in the 1950s to 1970s, for example, were not indicative of canopy dynamics throughout southern California (discussed above). By 1980, the kelp forests at Palos Verdes and Point Loma had both recovered from the severe declines in the 1950s–1970s; however, neither kelp forest returned to their historic maxima recorded prior to 1950 (Fig. 3). Kvitek et al. (2008) recently mapped the seafloor along the Palos Verdes peninsula using multi-beam sidescan sonar, subbottom profiling, sediment sampling and underwater video. They then compared the historic kelp distribution (from surveys in 1893 and 1912) to that from surveys in 1989 and 1999, noted where kelp had occurred historically but not recently, and examined these areas for substratum changes. They concluded that most kelp losses occurred over areas now buried or “dusted” by sediment. Kvitek et al. (2008) also found considerable kelp loss at depths below 15 m, suggesting that water clarity and quality may still be limiting. Their work provides further evidence of altered physical drivers underpinning changes to the extent of kelp forests, that the area did not fully recover from the habitat degradation associated with increasing urbanization in the 1940s, and/ or that present-day sedimentation and discharges from land may still be having adverse effects. 4.2. The relative importance of sea urchin grazing The data and literature review clearly indicate that anthropogenic degradation of water quality and the 1957–1959 El Niño were primary causes of kelp declines in the 1950s–1970s, and very likely the indirect cause of associated grazing effects. Others have attributed the kelp decline solely to increased sea urchin grazing resulting from intense exploitation of sea urchin competitors (abalone) and predators (lobsters and sheephead), and recovery to the subsequent fishing of the largest sea urchin species in the 1970s and 1980s (Jackson et al., 2001). This conclusion was based on citations of Tegner and Dayton (1991, 2000), who provided data on 1973–1987 fishery landings of red sea urchins from Point Loma and other kelp forests near San Diego, and suggested trends in the data indicated that the fishery enhanced the recovery of the Point Loma kelp forest after the 1982–83 El Niño. Tegner and Dayton (2000), however, concluded that potential ecosystem impacts of the urchin fishery had been obscured by grazing by non-exploited urchin species, and climatic effects on kelp resources. Moreover, neither paper suggested that the sea urchin fishery resulted in the recovery of the kelp forest at Point Loma or kelp forests elsewhere in southern California. In addition to active sea urchin grazing from a reduction in drift algal supply, poor water quality may have led to local increases in sea urchin abundance. North (1967) noted abundant, newly recruited sea urchin larvae in organic-rich fine sediment layers near a sewer outfall, and Pearse et al. (1970) showed that sea urchin recruits could survive and grow by eating these fine sediment layers and by absorbing organic compounds from the water. The high sea urchin densities around sewer outfalls in the near absence of macroalgae may have resulted in part from enhanced settlement and growth on sewage-derived surface films. 4.3. The importance of competitors and predators of sea urchins

4.1. Has collapse occurred? The data considered here show that there have been periods of large declines in giant kelp canopy area in southern California, and

4.3.1. Sea otters Sea otters are well known to consume large quantities of sea urchins when urchins are available and, by reducing sea urchin grazing, may

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indirectly lead to an increase in kelp. Estes and Palmisano (1974) argued that this top-down or keystone effect of otters was most important in structuring kelp communities in the Aleutian Islands and that the importance of this interaction to kelp forest structure might apply throughout the present and former range of the sea otter including, as previously suggested by North and Pearse (1970), southern California. Both papers suggested phenomena other than sea otters might have been causing the then recent increase in sea urchins at some southern California sites because sea otters had been extinct in the region since the 1850s (review in VanBlaricom et al., 2001). Using a correlative approach similar to Estes et al. (1978) based on natural site differences, Foster and Schiel (1988) assessed the importance of the absence of sea otter predation to kelp forest structure in California by reviewing data from 224 sites where sea otters were absent. The sites were partitioned into categories from completely forested with no large sea urchins to deforested (“barrens”) with abundant large sea urchins. They found that less than 10% of sites were deforested state-wide, and less than 10% in southern California, showing that even in the absence of sea otter predation on urchin populations, there were relatively few areas where urchins had deforested kelp stands over large areas. They proposed an alternative model of kelp forest dynamics in California based on multiple effects that included predators as well as the demography and dynamics of giant kelp, which is highly prone to a wide range of disturbances and whose weedy life history characteristics facilitate rapid recovery if water quality remains suitable (Schiel and Foster, 2006). 4.3.2. Abalone North and Pearse (1970) suggested that the abalone fishery in southern California may have indirectly led to an increase in sea urchins via competitive release. Lowry and Pearse (1973) found that abalone were abundant in large crevices and sea urchins in small crevices at a site in central California. Since drift algal food was abundant at the study site, they suggested abalone may out-compete sea urchins for large crevice space. However, they also suggested this pattern could reflect the ability of sea otters to more easily extract sea urchins from large crevices, and that field experiments were needed to test these alternative hypotheses. They did not speculate about why sea urchins were more abundant in small crevices. Tegner and Levin (1982) examined the effects of food (giant kelp) supply on the growth of red sea urchins and red abalone in different size and species combinations in laboratory tanks over 2.5 years. They found no significant differences among treatments, only trends that large urchins grew better by themselves at all food levels, and all sizes of abalone grew better with urchins when food supply was moderate or in excess. Despite these results, they argued these trends represented weak competition: “abalones may be able to decrease urchins' relative fitness when food supplies are adequate through interference competition.” It could also be argued, however, that the lack of abalone would lead to an increase in drift algae and, therefore, decrease the probability of sea urchins switching from passive to active foraging. Tegner and Levin (1982) also pointed out that urchins may have positive effects on abalone by providing habitat for juveniles and benefiting growth, and Rogers-Bennett and Pearse (2001) found more juvenile abalone at sites with more red sea urchins. Moreover, observations of the effects of the sea urchin fishery in northern California rocky habitats suggest the direction of the competition hypothesis may be reversed with sea urchins outcompeting abalone (Tegner and Dayton, 2000). To our knowledge these are the only data available from California to answer the sea urchin–abalone competition question. Jackson et al. (2001, Table 1) used data on declines in white abalone (Haliotis sorenseni Bartsch) populations to support the argument that fishing has greatly reduced abalone populations, thereby indirectly increasing sea urchins through a reduction in competition. Although white abalone populations in California have severely declined, this species is found at depths of 20–60 m (Haaker et al., 2001). If competitive

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effects involving white abalone did exist, they could not affect most California kelp forests which are most prolific above 20 m depth (Foster and Schiel, 1985). Commercial landings of abalone species that are most common in southern California kelp forests (H. rufescens Swainson and H. corrugata Gray) remained high until 1969–1972 (Haaker et al., 2001), suggesting populations were not severely reduced in the 1950s–1970s. In summation, it seems unlikely that abalone–sea urchin competition was significant and therefore should have, at most, a very minor role in models of southern California kelp forest dynamics. 4.3.3. Sheephead and lobster California sheephead occur primarily from southern California south into México, and feed on a variety of prey, including sea urchins and lobsters, in nearshore areas during the day and shelter at night (review in Cowen, 1983). California spiny lobsters have a similar geographic distribution, a broad diet including sea urchins, and shelter during the day. They migrate into deeper water in winter (review in Barsky, 2001). Tegner (1980) proposed that sheephead and spiny lobsters might control sea urchin populations because they eat sea urchins and because of patterns of urchin abundance and distribution at offshore locations where sheephead were abundant. She noted, however, that deforestation by sea urchins had also been observed at these offshore locations. Tegner and Dayton (1981) examined variation in size frequency distributions of live and dead sea urchins and made other observations at three sites in the Point Loma kelp forest, two with sheephead and spiny lobster and one without, to determine the potential combined effects of these predators. Differences between sea urchin size frequency distributions among sites with sheephead and lobsters and the site without these predators were interpreted as indicating predator control, and this was further argued by Barry and Tegner (1990). Their results were equivocal, however; the nopredator area was not replicated, the two treatments differed in depth and topography, and others have shown that urchin size frequency distributions similar to those found in these earlier studies can also result from sporadic recruitment (Ebert et al., 1993). Cowen (1983) provided the only experimental evidence for sheephead effects on sea urchins. He removed sheephead from a 13,000 m2 site at the Channel Islands, and compared changes in locations, numbers and size frequency distributions of red sea urchins associated with crevices to a control site. He found that the total number of sea urchins and the number occurring outside of crevices increased in the experimental site over two years, and that urchins placed in the open in the control area quickly moved to shelter or were eaten by sheephead. Cowen's (1983) data show that sheephead eat red sea urchins and can affect their distribution relative to crevices, but do not show that sheephead control red sea urchin abundance. He only sampled in and around crevices in a small portion of the experimental area and, as he pointed out, the increased abundances could have resulted from sea urchin redistribution within the experimental area, with a net movement into the crevice areas sampled. These experiments and surveys were done in kelp forests where sea urchins occurred mostly in and around crevices, presumably eating drift algae. It is not known whether the effect of sheephead on sea urchin distribution would be amplified under conditions of low drift abundance when sea urchins begin to forage actively. It is clear, however, that drift supply alone can “regulate” deforestation by altering sea urchin behaviour in areas where sheephead and lobster are rare (Ebeling et al., 1985; Harrold and Reed, 1985; Reed, pers. com.). Moreover, Dean et al. (1984) found that the spatial distribution and movement of red sea urchin aggregations at a site with high sheephead densities (258/ha) “appeared to be unrelated to predation pressure from fishes or lobsters.” Dean et al. (1984) further concluded that urchins exacerbate kelp declines initiated by other causes such as storms and high temperatures/low nutrients.

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The primary data concerning lobster predation on sea urchins are from Tegner and Levin (1983) who found that lobsters maintained on a diet of sea urchins preferred to eat purple sea urchins and ate all sizes in the laboratory. They suggested this accounted for the unimodal size frequency distribution of live animals observed in the field where lobsters and sheephead were common. Lobsters also ate red sea urchins in these lab experiments, but mostly intermediate sizes, and Tegner and Levin (1983) suggested this was reflected in the bimodal size distribution of live red sea urchins in the field where the two predators were common. From these results and from landing data for the lobster fishery, Tegner and Levin (1983) concluded lobsters and sheephead controlled sea urchin populations until fishing reduced their abundance in the 1950s, leading to increases in sea urchin populations and deforestation. Lafferty (2004) and Behrens and Lafferty (2004) argued for lobster control of sea urchin populations based on field survey data from the Northern Channel Islands (Fig. 1). Comparisons were made of live sea urchin abundances and live and dead sea urchin size frequency distributions between two sites within a no-take reserve on one island and areas where lobsters were fished on other islands. Lafferty (2004) found lobsters to be significantly more abundant and sea urchins significantly less abundant at reserve sites. Behrens and Lafferty (2004) used size frequency distributions of live Strongylocentrotus franciscanus and S. purpuratus and concluded that these matched those suggested by Tegner and Levin (1983) to result from lobster predation. However, this conclusion is only partially correct; comparing the data from these two studies shows they did match for S. purpuratus but not for S. franciscanus. Moreover, Behrens and Lafferty (2004) did not mention that differences in size frequency distributions may result from sporadic recruitment rather than predation (Ebert et al, 1993). Despite a lack of treatment replication, biogeographic differences between the reserve sub-sites and fished sites (Murray et al., 1980; Hamilton et al., 2010) and potential problems with the interpretation of size frequency distributions, Behrens and Lafferty (2004) concluded their findings verified the hypothesis that spiny lobsters control sea urchin populations in the field. Taken together, these studies show that sheephead can affect sea urchin behaviour and that sheephead and lobster feeding increases sea urchin mortality. However, they constitute relatively weak, contradictory and inconsistence evidence for the hypothesis that these predators exert control over sea urchin populations in southern California at spatial scales larger than patches. It remains possible, of course, that the removal of large lobsters and sheephead in the early stages of the fisheries (Tegner and Levin, 1983) had a disproportionate effect on subsequent urchin population dynamics, although this is speculative and is not reflected in any of the urchin data we have found. Altogether, therefore, there is little evidential justification for the statement by Jackson et al. (2001) that “ecological extinction” of these predators from fishing led to the disappearance of kelp forest during the 1950s–1970s.

5. Conclusions Giant kelp forests in southern California have not collapsed; this claim was based on extrapolating the dynamics of a few highly impacted mainland kelp forests to the entire region. Undoubtedly, there are multiple stressors affecting kelp forests and so the combination of their effects is greater than any one factor acting alone. Two large kelp forests did precipitously decline in the 1950s–1970s, but this was primarily due to the direct effects of sewer discharges and sedimentation on giant kelp, and recovery occurred when water quality improved. A large El Niño and sea urchin grazing contributed to the declines, and the abundance of the latter may have been increased by discharged wastes. This scenario highlights improving

water quality as the most effective and direct management strategy for the conservation of giant kelp forests in southern California. Giant kelp is subjected to frequent disturbance from a wide variety of natural and human-induced influences. Given the high degree of canopy and biomass fluctuations seasonally and inter-annually it is remarkable that this species, a key habitat engineer with which thousands of other species are associated, is so robust and that kelp forests flourish in virtually all areas now that they occupied historically. The weedy life history of M. pyrifera enables relatively quick replenishment of populations even after such major oceanographic events such as the 1982–83 El Niño that deforested large areas of kelp forest along the entire California coast. The severe reduction of iconic species such as abalone, and large fishes and lobsters, is lamentable and requires regulatory intervention to reverse. These reductions, however, do not appear to have led to the severe degradation of kelp forests generally, but effects on food web complexity, trophic interactions, carbon cycling and other ecosystem processes are largely unknown. As pointed out in two areas of the world, however, such reductions often leave few vestiges or detectable effects on the wider ecosystem (Dayton et al., 1998; Schiel, 2006). Most people who have worked in or enjoyed kelp forests would like to see the restoration of a full ensemble of species and interactions, so decisions directly affecting giant kelp itself, rather than diffuse and poorly understood interactions by some species, seems to us to be the most promising avenue of management. No level of species' protection or reserve status will be effective if water quality, coastal runoff, increased sedimentation, and contamination impact the ability of giant kelp to survive and thrive. Acknowledgements We thank P. Kalvass (California Department of Fish and Game) for providing data on the sea urchin fishery, and J. Gully (Los Angeles County Sanitation District) and C. Mitchell (MBC Applied Environmental Sciences) for data on kelp canopy abundance. P. Dayton, J. Pearse, D. Reed, S. Schroeter and J. Steinbeck provided critical comments (even though they had reservations about some of our arguments). DRS thanks the New Zealand Foundation for Research, Science and Technology (Coasts and Oceans OBI) and the Royal Society of New Zealand Marsden Fund, which support much of his coastal research. [ST] References Airoldi, L., 2003. The effects of sedimentation on rocky coast assemblages. Oceanogr. Mar. Biol. 41, 161–236. Aleem, A.A., 1956. Quantitative underwater study of benthic communities inhabiting kelp beds off California. Science 123, 183. Anderson, B.S., Hunt, J.W., 1988. Bioassay methods for evaluating the toxicity of heavy metals, biocides and sewage effluent using microscopic states of giant kelp Macrocystis pyrifera (Agardh): a preliminary report. Mar. Environ. Res. 26, 113–134. Anderson, B.S., Hunt, J.W., Turpen, S.L., Coulon, A.R., Martin, M., 1990. Copper toxicity to microscopic stages of giant kelp Macrocystis pyrifera: interpopulation comparisons and temporal variability. Mar. Ecol. Prog. Ser. 68, 147–156. Andrews, H.L., 1945. The kelp beds of the Monterey region. Ecology 26, 24–37. Barry, J.P., Tegner, M.J., 1990. Inferring demographic processes from size-frequency distributions: simple models indicate specific patterns of growth and mortality. Fish. Bull. 88, 13–19. Barsky, K.C., 2001. California spiny lobster. In: Leet, W.S., Dewees, C.M., Klingbeil, R., Larson, E.J. (Eds.), California's Living Marine Resources: A Status Report. California Department of Fish and Game, Sacramento, pp. 98–100. Bedford, D., 2001. Giant kelp. In: Leet, W.S., Dewees, C.M., Klingbeil, R., Larson, E.J. (Eds.), California's Living Marine Resources: A Status Report. California Department of Fish and Game, Sacramento, pp. 277–281. Begin, C., Johnson, L.E., Himmelman, J.H., 2004. Macroalgal canopies: distribution and diversity of associated invertebrates and effects on the recruitment and growth of mussels. Mar. Ecol. Prog. Ser. 271, 121–132. Behrens, M.D., Lafferty, K.D., 2004. Effects of marine reserves and urchin disease on southern California rocky reef communities. Mar. Ecol. Prog. Ser. 279, 129–139. Bellwood, D.R., Hughes, T.P., Folke, C., Nyström, M., 2004. Confronting the coral reef crisis. Nature 429, 827–833. Berkes, F., Hughes, T.P., Steneck, R.S., Wilson, J.A., Bellwood, D.R., Crona, B., Folke, C., Gunderson, L.H., Leslie, H.M., Norberg, J., Nyström, M., Olsson, P., Österblom, H., Scheffer, M., Worm, B., 2006. Globalization, roving bandits, and marine resources. Science 311, 1557–1558.

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