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

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Journal of Experimental Marine Biology and Ecology

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

Michael S. Foster a,⁎, David R. Schiel b

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

⁎ Corresponding author. Tel./fax: +1 831 786 8853.E-mail addresses: mfoster@mlml.calstate.edu (M.S. F

david.schiel@canterbury.ac.nz (D.R. Schiel).

0022-0981/$ – see front matter © 2010 Elsevier B.V. Aldoi:10.1016/j.jembe.2010.07.002

a b s t r a c t

a r t i c l e i n f o

Article history:Received 28 April 2010Received in revised form 30 June 2010Accepted 5 July 2010

Keywords:Ecosystem collapseKelp forestsMacrocystisSedimentationSouthern CaliforniaWater quality

It is increasingly argued that human-induced alterations to food webs have resulted in the degradation ofcoastal ecosystems and even their “collapse.” We examined the evidence for this argument for Macrocystispyrifera (giant kelp) forests in southern California. Others have concluded that forests in this region collapsedbetween 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 butare currently sustained as a result of a commercial sea urchin fishery that began in the early 1970s. Ourexamination of the historical record, primary publications, and previously unpublished data showed thatthere was no widespread decline in the region between 1950 and 1970, but there were localised declines inmainland kelp forests near the rapidly growing cities of Los Angeles and San Diego. The preponderance ofevidence indicates that kelp losses were caused primarily by large increases in contaminated sewagedischarged 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 foodresources. The forests recovered when sewage treatment improved and sewage outfalls were relocated. Theeffects of fisheries were explored by correlation analysis between kelp canopy cover and commercial seaurchin landings, and among fisheries landings for sea urchins, abalone, sheephead and lobster. Thesecorrelations were generally insignificant, but were often confounded by differences in the spatial scale overwhich 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 kelpabundance, most likely because roe (gonad) development is directly related to food availability. A literaturereview showed that although sheephead and lobsters may control sea urchin abundance at small spatialscales within some sites, there is little evidence they do so over large areas. That abalone and sea urchinscompete, such that sea urchins increased as a result of abalone harvesting, is largely conjecture based ontheir similar habitat and food utilization. This study shows that kelp forests in southern California did notcollapse, 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 thatmanagement by species' protection or reserves will not be effective if poor habitat quality impacts the abilityof giant kelp to survive and thrive.

oster),

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© 2010 Elsevier B.V. All rights reserved.

1. Introduction

“Coincident with the increased discharge of sewage effluent, har-vestable kelp has virtually disappeared in the vicinity of WhitesPoint for a distance of two or three miles along the coast.”—Revelleand Wheelock (1954)

An increasing number of published papers argue that human-induced alterations to food webs have resulted in the degradationof coastal ecosystems worldwide (e.g., Steneck et al., 2002; Springeret al., 2003; Bellwood et al., 2004; Berkes et al., 2006; Worm et al.,2006) and even their “collapse” (Jackson et al., 2001). A centralargument for this is that overfishing over long time periods has causedthe loss of structural and functional components of these ecosystemscompared to their state hundreds or even thousands of years ago. Inthe case of coastal kelp ecosystems, the premises for this argumentare that that top predators such as sea otters, fish and lobsters havebeen fished to ecological extinction and that this has led to popula-tion explosions of herbivores, which results in deforestation (Esteset al., 1989; Jackson et al., 2001). Much of the evidence comes from

60 M.S. Foster, D.R. Schiel / Journal of Experimental Marine Biology and Ecology 393 (2010) 59–70

northern hemisphere kelp beds and forests, particularly in the north-western Atlantic (e.g., Mann and Breen, 1972; Steneck et al., 2004;but see Elner and Vadas, 1990) and the eastern Pacific along thewestern coast of North America. In the latter region there has beenconsiderable ecological and political interest in the predator-urchin-kelp interaction since North and Pearse (1970) first suggested it as apossible 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 interestbecause sea urchin predators, sea urchins and kelp are commonfeatures of temperate nearshore communities worldwide (Lawrence,1975), and the interaction could inform larger debates over thegenerality of predator or keystone control of prey populations (reviewin Power et al., 1996) and the occurrence and frequency of alternatestable states (Connell and Sousa, 1983). If kelp systems are now trulycollapsed due to the removal of predators, then there are considerableimplications for urgent management decisions to resurrect coastalkelp communities, such as through no-take reserves. Alternatively, if apredator–urchin–kelp interaction is not the primary factor affectingkelp forests, or if these communities have not actually collapsed, thenother forms of management may be more appropriate.

We review evidence from southern California about the status ofkelp forest communities through time and the effects of environ-mental influences (frequently termed “drivers”) that are known toaccount for variation in kelp forest structure, and relate these to whatis known about sea urchins, their competitors and predators. Weargue that it is of debatable value to surmise that pre-historical statesof kelp, lacking spatial and temporal resolution, represent a usefulnatural state against which to gauge present status, but rather that a“strong-inference” approach (Platt, 1964) allows not only fundamen-tal understanding of underlying processes, including human impacts,but also points to testable hypotheses at relevant scales. Elner andVadas (1990) argued for a similar approach in their analysis of seaurchin population dynamics in the northwestern Atlantic. Our pur-pose is not to dispute that degradation of kelp forests has occurred inmany places or that there have not been declines in the numbers andsizes of many large predators worldwide. Rather, we review the lit-erature and offer a contrasting perspective on the status of kelp forestsalong southern California coasts and the underlying processes thatdrive their structure.

1.1. Background

Kelp communities comprise the dominant biogenic habitat alongtemperate and boreal rocky shores worldwide (Mann, 1973). Giantkelp forests are particularly common along eastern Pacific shores,support communities of thousands of species, and are known to beaffected by awide range of influences that alter their local distribution,abundance and biogeographic distribution (North, 1971a; Foster andSchiel, 1985; Grahamet al., 2007). Because of their great ecological andcultural importance there are legitimate concerns about their statusand the factors that affect them, which has been the case since at leastthe early 1950s (Revelle and Wheelock, 1954). Giant kelp produces afloating surface canopy that can contain N60% of the entire plantbiomass (North and Hubbs, 1968) making assessment of abundancebased on canopy area from surveys at sea or from the air relativelysimple. Along the coast of California, some assessments of kelpcanopies are available beginning in the late 1800s. More comprehen-sive surveys were done occasionally from 1911 (Crandall, 1915), anddetailed knowledge of kelp forest structure began with the advent ofdiving studies in the 1940s (Andrews, 1945; Aleem, 1956; Dawsonet al., 1960; North and Hubbs, 1968). Therefore, there is considerableknowledge about the structure of these kelp forests for well over60 years. These studies pre-datedmuch of the intense fishing pressurethat has been exerted on coastal fishes (Leet et al., 2001) and urbanexpansion along the coastline and, therefore, offer at least decadal-

scale baselines, althoughwith varying spatial and temporal resolution,against which to gauge the current status of kelp forests.

A central question about gauging kelp-based ecosystems is howfar backwemust look for comparisons. The notion of sliding or shiftingbaselines iswell-recognised in both the ecological (Dayton et al., 1998)and fisheries (Pauly, 1995) literature, providing a necessary caveatto assessments about change because of the potentially degradedcondition of an ecosystem used as an initial comparison. However,reconstructions of an idyllic past state before human impacts pre-sumably occurred are also fraught with difficulties because they ob-scure temporal and spatial variability, which are known to confoundeven short-term assessments of change (e.g., Underwood, 1992), andthey are usually based on presumption and scant evidence.

Crucial to a deep historical perspective is the assumption that theonce intact trophic relationships of numerous species that have nowbeen overfished to ecological extinction had exerted influences thatproduced major structural and functional differences in kelp ecosys-tems before their presumed collapse (Jackson et al., 2001). Trophiccontrol of structure and function are, therefore, central to argumentsabout kelp forest status. There is an old, rich and varied literature fromkelp forests worldwide (Lawrence, 1975) that demonstrates thenumerous ways that sea urchins, the dominant herbivores of kelpcommunities, can remove large tracts of kelp, reducing them to a stateoften called “barrens” even though they can be highly speciose (Beginet 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 concernsand management intervention as far back as 1959 (North, 1959a,b).The nature and extent of these effects have been examined bothexperimentally and through surveys at hundreds of sites (Foster andSchiel, 1985, 1988).

The other link in the chain of trophic control, crucial to argumentsabout the effects of overfishing, is that of top predators on urchins.One argument is that some historical and published evidence (whichwe review in this paper) shows that southern California kelp forestscollapsed in the 1950s to 1970s due to sea urchin overgrazing, and thiswas an indirect result of overfishing of lobsters, sheepheadwrasse andabalones (Jackson et al., 2001). Kelp forests in this region now exist,therefore, only because a sea urchin fishery, which began in the early1970s, has functionally replaced these former predators and compe-titors. This has also been called a phase shift, referring to the alteredstate of kelp forests (Steneck et al., 2002). Although Steneck et al.(2002) state that pollution and other factors affected southernCalifornia kelp forest dynamics, they highlighted the roles of diversesea urchin predators and competitors acting as buffers from sys-tematic deforestation. These claims underpin important marine con-servation and management imperatives because if ecosystems havecollapsed and if the direct cause is overfishing, then amelioratingfishing effects with management strategies such as no-take reservesshould not only increase fish stocks but indirectly resurrect coastalkelp communities from their putative collapsed states (e.g., Jacksonet al., 2001; Lubchenco et al., 2007; Palumbi, 2008).

Many studies show that there is a great deal of natural variabilityin the areal extent of kelp forests associated with storms and longerterm climatic events such as El Niño–Southern Oscillation (Daytonand Tegner, 1984; Tegner and Dayton, 1987; Foster and Schiel, 1993;Edwards, 2004). Entire giant kelp canopies and much of the under-story assemblage can disappear or be severely reduced in abundancefollowing intense El Niño periods, and recovery can take several years.This background of variability from natural disturbances is not justthe noise against which human-induced changes must be gauged butis fundamental to the population dynamics and natural history of kelpforest communities (e.g., Dayton et al., 1999; Schiel and Foster, 2006).Algal-dominated communities worldwide are known to be affected bya wide range of multiple stressors from altered coastlines and land-use practices over many years, often decades (Airoldi, 2003). We do

61M.S. Foster, D.R. Schiel / Journal of Experimental Marine Biology and Ecology 393 (2010) 59–70

not dispute that predators can have large effects on kelp forestsdynamics 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 rela-tive influences and extent in different circumstances if understandingunderlying mechanisms leading to better management and restora-tion are goals. To achieve this, it is necessary to understand 1) thecurrent status of kelp communities, 2) the influences of trophic inter-actions on kelp forest structure relative to other sources of naturalvariability, and 3) human-induced factors that may influence struc-ture and function and the degree to which these have degraded kelpforests. We examine these by using data sets from fishery landings,surveys, monitoring and experiments in the kelp forests of southernCalifornia, an area that has been the focus of arguments about col-lapsed ecosystems.

2. Methods

Our sources of information were as comprehensive as possible.1) We reviewed the numerous publications and reports relating tothe status of kelp forests in southern California from the early 1900sonwards. These were by primary researchers working in kelp forestsat the time and contained many first-hand observations as well asdata. 2) Primary data setswere obtained frompublications and reportsthat are generally available, and are cited accordingly. 3) Archived(unpublished) data were obtained by making requests to publicagencies and private companies involved in fisheries assessment andkelp forest management. Collectively, these were used to constructan historical record of the status and factors affecting southernCalifornian kelp forests since the early 1900s. Although we have alsoconsidered thewider status of kelp forests in California,we focusedourefforts on studies of the Palos Verdes kelp forest off Los Angeles and thePoint Loma kelp forest off San Diego (Fig. 1). These two mainlandsouthern California kelp forests were those used primarily by others to

Fig. 1. Southern California: mainla

formulate their conclusions about trophic control of California kelpforest ecosystems (e.g., Jackson et al., 2001; Steneck et al., 2002).

Jackson et al. (2001) do not define “collapse.” We assume it wasused to characterize abundance when it fell below 10% of maximumsince Jackson was one of the authors of Worm et al. (2006) and thelatter used “catches dropping below 10% of the recorded maximum”

as the definition of fisheries collapse. We use giant kelp surfacecanopy area as the metric and the 10% figure as an indication ofcollapse, 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 ofCalifornia from Monterey Bay southwards, but are patchy and usuallydepend on the presence of rocky reef (Foster and Schiel, 1985). Thegenerally robust status of kelp forests along coastal California for thepast century can be surmised from the kelp harvesting record (Fig. 2)which shows the harvest generally stayed above 80,000 tonnes from1948 onwards. Kelp beds are numbered and licensed, and wereharvested when canopies were extensive. The primary processingplant was in San Diego, and most of the harvest came from southernCalifornia. As harvesting increased from the early 1940s, low periodsof harvest invariably coincided with El Niño events when the kelpsurface canopy cover was low, especially the repeated events after1980. However, except for large coastal runoff associated with highrainfall in 1979 (Fisk, 2010) followed by the prolonged and intenseEl Niño of 1982–83, themost intense after 1950when detailed recordsbecame available (NOAA, 2010), the kelp harvestwas high. The declinein harvest after 1990 was related to decreased kelp canopy after the1991–1992 El Niño and a deteriorating US market for kelp products asAsian alginate production dramatically increased (Bedford, 2001), andnot due to the collapse of kelp forests.

nd coast and offshore islands.

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

62 M.S. Foster, D.R. Schiel / Journal of Experimental Marine Biology and Ecology 393 (2010) 59–70

3.2. Kelp forest dynamics in southern California: evidence from canopysurveys and field observations

Southern California is the region from Point Conception to theMexican border, including the offshore islands (Fig. 1). Kelp canopiesalong the mainland in the region from Los Angeles southwardsgenerally declined in areal extent in the 1950s–1970s and those atPalos Verdes and Point Loma (Fig. 3) declined to near zero cover(North, 1970; Wilson et al., 1980; Meistrell and Montagne, 1983). Thecanopy area of forests in southern California from Point Conceptionsouth to the Palos Verdes peninsula, however, changed very little(Harger, 1983) and large declines were not noted around the offshoreChannel Islands (Neushul, 1965; Clarke and Neushul, 1967; Foster,1975; Foster pers. obs.). The forests in these latter sub-regions cancomprise 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 late1960s along the mainland from just south of Los Angeles (HuntingtonBeach) to the Mexican border was between 40 and 60% of the

Fig. 3. Giant kelp canopy area of the Palos Verdes and Point Loma kelp forests. Events indidischarge from Mission Bay and sewage discharge from San Diego Bay at Point Loma; 3, 19urchin fishery in southern California and advanced sewage treatment at Palos Verdes; 6, El NVerdes kelp canopy area from Wilson et al. (1977; 1911–1973) and J. Gully (pers. com.), LNorth (1975; 1911–1967), North and MBC Applied Environmental Sciences (2001; 1968–2

historical maximum, depending on which historical sample is used(North, 1970; North andMBC Applied Environmental Sciences, 2001).Since the 1980s there have been large, state-wide declines in canopycover followed by recovery as a result of El Niños, as reflected inharvest records (Fig. 2) and canopy dynamics where high resolutiondata are available (Fig. 3). Furthermore, this has ramifications for thewider kelp community. Dayton et al. (1999) found that inter-annualand inter-decadal variation in ocean climate affected giant kelp withconsequent large and lasting long term effects on understory kelpsvia canopy competition for light. The exceptions to recovery are afew large canopies just north of Santa Barbara which were producedby plants growing on sand; these forests have not recovered sincethey 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 ofsewage discharge, coastal development, sedimentation and El Niños

There is strong evidence that sewage discharges, sedimentationfrom coastal development, and the El Niño of 1957–1959, were the

cated by arrows: 1, installation of shallow sewage outfall at Palos Verdes; 2, sediment57–59 El Niño; 4, sewage discharge moved offshore at Point Loma; 5, beginning of seaiño events after 1959; 7, record southern California rainfall and runoff (Fisk, 2010). Palosos Angeles County Sanitation District (1974–2007). Point Loma kelp canopy area from000), and MBC Applied Environmental Sciences (2007; 2000–2006).

63M.S. Foster, D.R. Schiel / Journal of Experimental Marine Biology and Ecology 393 (2010) 59–70

main causes of severe declines inmainland kelp forests near themajormetropolitan areas of southern California, particularly those at PalosVerdes and Point Loma. Revelle andWheelock (1954) were the first toreport the effects of sewage discharge on the Palos Verdes kelp forest,followed by the extensive studies of W.J. North and colleaguesbeginning in 1957. These studies were stimulated by observationsthat “considerable kelp disappeared in the last 15 years in the vicinityof Los Angeles and San Diego,” and this “raised the question ofwhether increasing waste discharges have adversely influenced kelpand associated marine resources” (North, 1958).

Los Angeles and San Diego were the two most rapidly growingmetropolitan areas of southern California during this period. Primary-treated domestic and industrial wastes from the Los Angeles areawere discharged through ocean outfalls at 18 m–90 m depths locatedjust north (Hyperion outfall) and within the southern edge (offWhites Point; outfall completed in 1937; Fig. 1) of the kelp forest atPalos Verdes. Between 1940 and 1960, waste discharge through theseoutfalls increased from 152 to 950×106 l/day, and suspended solidsin the discharge increased from 15,000 to 125,000 tonnes/year(Meistrell and Montagne, 1983). The discharge included a variety ofchlorinated hydrocarbons and metals, including large amounts ofcopper and zinc, which are known to be toxic to the early life stages ofgiant kelp (Anderson and Hunt, 1988; Anderson et al., 1990). Waterand benthos quality were further reduced by sedimentation fromcoastal 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 Verdeskelp 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). Therewas also astrong, negative correlation between canopy cover at Palos Verdesbetween 1947 and 1979 and the amount of suspended solids dis-charged 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 industrialwastes from the San Diego area were discharged into San Diego Bayin 1952, turning Bay waters “from a dark blue to a murky greenishbrown hue” (Jamieson, 2002). The bay entrance is within the southernend of the historical distribution of the Point Loma kelp forest (North,1991). The former northern endwas near the entrance toMission Bay,which was extensively developed in the late 1940s. Considerabledredging was done, resulting in sediment plumes extending offshore.These, plus diversion of sand by the entrance jetties, may have con-verted former kelp habitat into soft bottom (North, 1964, 1991; Northand 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, midwaybetween the entrances of Mission and San Diego Bays.

Fig. 4. Palos Verdes giant kelp canopy area and mass emission rates of suspended solids fromand J. Gully (pers. com.), Los Angeles Country Sanitation District.

That the alteration in water and benthos quality was the primarydirect cause of the initial, very large declines in these kelp forestsis strongly indicated by the population biology of giant kelp relativeto conditions observed within the beds and patterns of bed decline atthe time. The microscopic gametophyte and young sporophyte stagesof giant kelp occur on the benthos where they depend on lighttransmitted through the water column and sediment-free substratesfor attachment, reproduction, growth and survival. Kelp forests nearthese large sewage outfalls experienced a decrease in benthic lightfrom discharged particulates in the water column (North, 1964;Wilson et al., 1980), increased sedimentation on the benthos, and anincrease in discharged nutrients that stimulate phytoplankton growthin the absence of a kelp canopy (Eppley et al., 1972). Sedimentationalso 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 cmthick at historic kelp forest depths between 14 and 20 m. Thesediment extended nearly 10 km along an area of the Palos Verdespeninsula formerly occupied by giant kelp, and sediment depthdecreasedwith distance from theWhites Point sewer outfall. Divers atthe time commonly noted thin coatings of organic material (leptopelor floc) and very fine sediment on benthic surfaces (North, 1967,1971b). The microscopic stages of giant kelp do not survive burialand abrasion by even a very thin sediment layer (Devinny and Volse,1978), and such sediments can also harbor high concentrations oftoxic chemicals (Schiff et al., 2000) that impede gametogenesis.

Kelp canopy cover throughout southern California was furtherreduced by the warm nutrient-depleted conditions associated withthe 1957–1959 El Niño. Giant kelp sporophyte recruitment is in-hibited and sporophytes deteriorate as a result of low nutrientsassociatedwith elevatedwater 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 ofsouthern California (Foster and Schiel, 1985; Edwards, 2004). North(1959a) reported temperatures of 20 °C or more persisting well intoNovember, 1958 with “harsh” effects on all kelp forests in southernCalifornia. Local kelp harvesters reported less surface canopy than atany time since harvesting began in 1929 (North, 1959a), and this isreflected in the kelp harvest (Fig. 2). By 1959, the canopy cover wasreduced 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 effectsof sewage discharge during this period are indicated by North's(1964) observation that, by late 1959, many kelp forests along thesouthern mainland of southern California had begun to recover fromthe 1957–1959 El Niño (Fig. 2) but those at Palos Verdes and PointLoma did not (Fig. 3), an apparent example of the effects of multiplestressors.

ocean sewage discharges off Whites Point. Data from Meistrell and Montagne (1983)

64 M.S. Foster, D.R. Schiel / Journal of Experimental Marine Biology and Ecology 393 (2010) 59–70

The recovery of kelp canopies after 1957–1959ElNiño in both PalosVerdes and San Diego was strongly coincident with improvements inwater quality (Figs. 3, 4). Between 1971 and 1981, suspended solids inthe sewage discharge at Palos Verdes were gradually reduced from375,000 to c.75,000 tonnes/year, during which time the kelp canopyincreased to c.2.5 km2 (Fig. 4; Wilson et al., 1980; Meistrell andMontagne, 1983). In 1963,most of thewastes formerly discharged intoSan Diego Bay were treated and then discharged through a newoffshore ocean outfall at a depth of 67 m (Jamieson, 2002). The kelpcanopy around Point Loma began to recover within a year (Fig. 3),covering nearly 6 km2 by 1970 (North, 1991). This is ~50% of thehistoric maximum, not 6% as estimated by Steneck et al. (2002). Since1970 most major periods of abrupt decline have been associated withlarge El Niño events and large storms (Fig. 3).

3.4. The importance of grazing by sea urchins: observations and fisherydata

There is strong evidence that grazing by sea urchins, especially thepurple sea urchin Strongylocentrotus pupuratus (Stimpson) and redsea urchin S. franciscanus (Agassiz), contributed to kelp canopydecline and slowed recovery during the 1950s–1970s at Palos Verdesand Point Loma (North, 1964; Leighton et al., 1966). North (1959b)hypothesized that active grazing on attached plants by sea urchinswas initially stimulated by a reduction in the supply of adult kelpplants, and therefore of algal drift, as a consequence of the lack ofreplacement due to declining water quality associated with seweroutfalls and coastal development. Divers in the 1950s noted highconcentrations of sea urchins (especially red urchins) and gastropods,but these were found primarily along the seaward edge of the fewshallow kelp stands that remained, apparently having migratedshoreward as their food resources declined in deeper water (North,1959b, 1964). North (1959b) further observed that areas obviouslydeforested by sea urchins also occurred at sites away from outfalls, butthese areas were small such that “one did not have to swim far to findrich algal growth.” Early investigators reported that sea urchinsseemed to subsist primarily on drift algae when algal abundance washigh but switched to active foraging on attached plants when algalabundance was low, creating barrens (e.g., North, 1967).

This interaction between algal abundance and sea urchin behav-iour has since been repeatedly observed in southern California (Deanet al., 1984; Ebeling et al., 1985; Harrold and Reed, 1985). Thesestudies showed that active grazing is induced by storms or hightemperatures and low nutrients that reduce drift algal abundance, andceases when suitable oceanographic conditions occur that facilitatealgal settlement and growth or when there is a decline in urchinabundance due to storms or disease. Declining food resources can alsolead to kelp mortality from grazing fishes that concentrate aroundremaining plants (North, 1968;Wilson et al., 1977). In the case of PalosVerdes 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 andinhibited new recruitment and growth, which led to continued activegrazing until sewage treatment and disposal were improved.

Based on the success of local-scale removals of sea urchins at PointLoma, Pearse et al. (1970) concluded that sea urchin grazing was thedirect cause of kelp disappearance. Large-scale removal of sea urchinsby hand or by spreading quicklime (calcium oxide) pellets began atPoint Loma in 1962. Removal efforts were concentrated in areas ofhigh 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 SanDiego sewer discharge well offshore (discussed above). That changesin water quality were most important is suggested by North's (1966)observations that vegetation also appeared in areas not treated withquicklime, and that the “Point Loma bed is still expanding largely by

natural processes.” These observations indicate that improvements inwater quality enabled giant kelp and other macroalgae to re-colonizeformer habitat and the removal of sea urchins increased the rate ofkelp recovery. Improvement in sewage treatment is further implicat-ed as the primary factor affecting forest dynamics by the timing of therecovery of giant kelp at Palos Verdes (Fig. 3). Sea urchin removals andother restoration efforts were started there in the early 1960s but theforest did not begin to recover until 1971 (North, 1964; Wilson et al.,1977) when sewage treatment improved. As at Point Loma, recoveryoccurred around removal and restoration sites, but also elsewhere(Wilson et al., 1977).

A commercial fishery for sea urchins began in southern Californiain the early 1970s (Fig. 5A). Comparing landings from the sea urchinfishery with kelp abundance could ostensibly provide insight into theinteraction of the fishery with the status of kelp beds. In practice suchcomparisons are difficult to interpret, in part because the spatialresolution of the data is not the same for each metric. Also, changes inlandings do not necessarily reflect changes in kelp abundances; forexample, urchin landings decline during El Niño years and whenprices decline (Kalvass and Rogers-Bennett, 2001) and, as notedabove, when drift algae are sparse sea urchins may switch to activeforaging and have a greater effect on attached plants. Nevertheless,any conclusions regarding fishing as a replacement for lost predatorycontrol 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 of1986–1989,most of thefisherywas in southern California (CFG, 2004).Both red and purple sea urchins can cause deforestation (review inFoster and Schiel, 1985). The sea urchin fishery in southern Californiawas comprised only of red urchins until 1989, after which relativelysmall catches of purple urchins were also taken (Kalvass and Rogers-Bennett, 2001; Parker and Ebert, 2001). As for kelp harvesting, thecatch of sea urchins generally decreased during years with strongEl Niños (Fig. 5A). The fishery increased steadily from 1976 to 1980,but this was well after the time when the kelp forest at Point Lomabegan recovery (Fig. 3). The recovery of the Point Loma kelp forest inthe 1960s, therefore, was not coincident with the removal of largenumbers of sea urchins for commercial purposes. The recovery at PalosVerdes began later than at Point Loma, just before the urchin fisherybegan. The kelp canopy at Palos Verdes did continue to expand as theurchin fishery developed (Figs. 3, 5A) but the expansion was highlycorrelated with improvements in water quality (Fig. 4).

The relationship between the urchin fishery and the cover of kelpforests after the 1982–83 El Niño, however, is equivocal. Although notin the same spatial scales, there is a significant correlation betweenthe combined Point Loma and Palos Verdes kelp canopy cover and thesize of sea urchin landings from the southern California mainland(r29=0.5316, p=0.003). The peak urchin fishery along the mainlandwas 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 between1989 and 2000, which generally coincided with reductions in kelpcanopies but also with repeated intense El Niño events. At Point Loma,where urchin landings and kelp data were both recorded in thesame coastal sections after 1978, there was a variable but revealingrelationship between urchins and kelp cover (Fig. 6). Both kelp coverand urchin landings were low during El Niño periods, and landingswere generally high when giant kelp was abundant. During periods ofintermediate canopy cover, however, urchin landings could also behigh. This may reflect minimum thresholds of kelp availability neededfor gonad development. The state of gonads, weather conditions andthe condition of the urchin roe market are all variable and tend toweaken the relationship between kelp cover and urchin harvest. Fur-thermore, the urchin fishery declined as permits were restricted in thelate 1980s (CFG, 2004).

Taken together, these data are useful mainly in showing that theyprovide little to no support for concluding that the sea urchin fishery

Fig. 5. A. Commercial landings of all abalone species in California and red sea urchin (Strongylocentrotus franciscanus) landings for southern California and southern Californiamainland. 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 to1981. Shaded bars are El Niño periods.

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is a driver of kelp abundance. The evidence is that the opposite is morelikely, that urchins are fishedwhen kelp forests are robust and providemore algal drift, supporting good gonad development and thereforegreater recovery of urchin roe.

3.5. The importance of competitors and predators of sea urchins: fisherydata

North and Pearse (1970) first suggested that abalone and seaurchins may compete for space and food such that a decrease in

Fig. 6. Giant kelp canopy area and commercial sea urchin catch at Point Loma, 1978–2004Department of Fish and Game. Data incomplete prior to 1978.

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

. Kelp canopy area from Fig. 3. Sea urchin data from P. Kalvass (pers.com.), California

66 M.S. Foster, D.R. Schiel / Journal of Experimental Marine Biology and Ecology 393 (2010) 59–70

During the 1980s, many abalone fishers began harvesting sea urchins.This is reflected in the negative correlation between urchin andabalone landings between 1971 and 2001 (r31=−0.4141, p=0.021),which were both mainly in southern California. Between 1975 and1981 they peaked together, but after 1984 the urchin fishery largelyreplaced the abalone fishery.

Tegner (1980) was the first to suggest that in the absence of seaotters, sheephead wrasse (Semicossyphus pulcher Ayres) and Califor-nia spiny lobsters (Panulirus interruptus Randall), might control seaurchin populations via predation in southern California. Statisticsfrom the California lobster fishery (Barsky, 2001) are unclear on thispoint. The lobster fishery persisted throughout the 20th Century, witha maximum of c. 420 tonnes and a minimum of c. 80 tonnes duringWWII (Fig. 5B). Lobsters were consistently fished at between 100 and400 tonnes annually after 1976, presumably reflecting their relativeabundance (cf., the abalone fishery data), but there was no obviousresponse in urchin landings that would shed light on competitiverelease. During this period, there was only a very weak negativecorrelation between lobster and southern California urchin landings(r24=−0.296, p=0.20), but this was mostly a reflection of fishersswitching from overfished abalone to sea urchins. There is no cor-relation between lobsters and any measure of kelp abundance orharvest.

Sheephead wrasse are common in nearshore waters of southernCalifornia and feed on a variety of prey, including sea urchins andlobsters (review in Cowen, 1983). Fishing them has been implicatedas a cause of increases in sea urchins (e.g., Jackson et al., 2001; Stenecket al., 2002). Sheephead were fished throughout the 20th Century,with annual landings generally below 50 tonnes, and a maximumcatch of 185 tonnes in 1998 (Fig. 5B; Stephens, 2001). Sheepheadlandings are not significantly correlated with that of urchins (r19=−0.3617, p=0.128).

Although none of these landing data are necessarily goodindicators of abundance of any species, which are fished for a varietyof reasons and intensities over many years, they offer little support forarguments of trophic control of kelp through fishing pathways and,consequently, allow the conjecture that oceanographic and physicaldrivers have had the major influences on kelp abundance over thepast century.

4. Discussion

Piecing together disparate data on long term processes is fraughtwith difficulties. The data themselves may not always have beencollected reliably, there were usually different purposes and reasonsfor their collection, and the spatial scales of data from different taxaare usually different. This leaves considerable room for variedanalyses and interpretation, based on how data were selected andoften on broad correlations. Extending these to conclusions aboutcausality is a considerable step further. Here we have tried to go backto all known sources of primary data, both published in the primaryliterature and in reports, many of which are well-cited in the liter-ature. Our premise is that the work and views of the researchers whocollected the data should be extensively reviewed and considered inany syntheses that offer new interpretations of the causes of pastevents. Two scenarios, one based on overfishing of predators andcompetitors and the other based on changes in water quality andoceanographic events, could produce roughly the same changes inkelp forest composition and extent over time, and roughly the sameoutcome. Considering all evidence is therefore crucial in determiningwhere the balance of probability lies.

4.1. Has collapse occurred?

The data considered here show that there have been periods oflarge declines in giant kelp canopy area in southern California, and

that these were invariably associated with El Niño events, and thewarm water, low nutrients and storms they bring to coastal waters(Figs. 2, 3). The data also show that there were severe declines in kelpforests near large metropolitan areas along the southern portion ofthe mainland coast of southern California, associated with coastaldevelopment, sedimentation and sewage outfalls. However, recoveryof kelp occurred when stressors were reduced, and there was noevidence for a general collapse of kelp forest structure or a permanentshift to a largely kelp-free state.

The data show that the canopy dynamics of kelp forests at PointLoma and Palos Verdes in the 1950s to 1970s, for example, were notindicative of canopy dynamics throughout southern California (dis-cussed above). By 1980, the kelp forests at Palos Verdes and Point Lomahad both recovered from the severe declines in the 1950s–1970s;however, neither kelp forest returned to their historicmaxima recordedprior to 1950 (Fig. 3). Kvitek et al. (2008) recently mapped the seaflooralong the Palos Verdes peninsula usingmulti-beam sidescan sonar, sub-bottom profiling, sediment sampling and underwater video. They thencompared the historic kelp distribution (from surveys in 1893 and1912) to that from surveys in 1989 and 1999, noted where kelp hadoccurred historically but not recently, and examined these areas forsubstratum changes. They concluded that most kelp losses occurredover areas now buried or “dusted” by sediment. Kvitek et al. (2008) alsofound considerable kelp loss at depths below 15 m, suggesting thatwater clarity and quality may still be limiting. Their work providesfurther evidence of alteredphysical drivers underpinning changes to theextent of kelp forests, that the area did not fully recover from thehabitatdegradation associated with increasing urbanization in the 1940s, and/or that present-day sedimentation and discharges from landmay still behaving adverse effects.

4.2. The relative importance of sea urchin grazing

The data and literature review clearly indicate that anthropogenicdegradation of water quality and the 1957–1959 El Niño were primarycauses of kelp declines in the 1950s–1970s, and very likely the indirectcause of associated grazing effects. Others have attributed the kelpdecline solely to increased sea urchin grazing resulting from intenseexploitationof seaurchin competitors (abalone) andpredators (lobstersand sheephead), and recovery to the subsequent fishing of the largestsea urchin species in the 1970s and 1980s (Jackson et al., 2001). Thisconclusion was based on citations of Tegner and Dayton (1991, 2000),who provided data on 1973–1987 fishery landings of red sea urchinsfrom Point Loma and other kelp forests near San Diego, and suggestedtrends in the data indicated that the fishery enhanced the recovery ofthe Point Lomakelp forest after the 1982–83 El Niño. Tegner andDayton(2000), however, concluded that potential ecosystem impacts of theurchin fishery had been obscured by grazing by non-exploited urchinspecies, and climatic effects on kelp resources. Moreover, neither papersuggested that the sea urchin fishery resulted in the recovery of the kelpforest at Point Loma or kelp forests elsewhere in southern California.

In addition to active sea urchin grazing from a reduction in drift algalsupply, poor water quality may have led to local increases in sea urchinabundance. North (1967) noted abundant, newly recruited sea urchinlarvae in organic-rich fine sediment layers near a sewer outfall, andPearse et al. (1970) showed that sea urchin recruits could survive andgrow by eating these fine sediment layers and by absorbing organiccompounds from thewater. The high sea urchin densities around seweroutfalls in the near absence of macroalgae may have resulted in partfromenhanced settlement and growth on sewage-derived surfacefilms.

4.3. The importance of competitors and predators of sea urchins

4.3.1. Sea ottersSea otters arewell 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) arguedthat this top-down or keystone effect of otters was most importantin structuring kelp communities in the Aleutian Islands and that theimportance of this interaction to kelp forest structure might applythroughout the present and former range of the sea otter including, aspreviously suggested by North and Pearse (1970), southern California.Both papers suggested phenomena other than sea otters might havebeen causing the then recent increase in sea urchins at some southernCalifornia sites because sea otters had been extinct in the region sincethe 1850s (review in VanBlaricom et al., 2001). Using a correlativeapproach similar to Estes et al. (1978) based on natural site differences,Foster and Schiel (1988) assessed the importance of the absence of seaotter predation to kelp forest structure in California by reviewing datafrom 224 sites where sea otters were absent. The sites were partitionedinto categories from completely forested with no large sea urchins todeforested (“barrens”)with abundant large sea urchins. They found thatless than 10% of sites were deforested state-wide, and less than 10%in southern California, showing that even in the absence of sea otterpredation on urchin populations, there were relatively few areas whereurchins had deforested kelp stands over large areas. They proposed analternative model of kelp forest dynamics in California based on mul-tiple effects that included predators as well as the demography anddynamics of giant kelp, which is highly prone to a wide range of dis-turbances and whose weedy life history characteristics facilitate rapidrecovery if water quality remains suitable (Schiel and Foster, 2006).

4.3.2. AbaloneNorth and Pearse (1970) suggested that the abalone fishery in

southern California may have indirectly led to an increase in seaurchins via competitive release. Lowry and Pearse (1973) found thatabalone were abundant in large crevices and sea urchins in smallcrevices at a site in central California. Since drift algal food wasabundant at the study site, they suggested abalone may out-competesea urchins for large crevice space. However, they also suggested thispattern could reflect the ability of sea otters to more easily extract seaurchins from large crevices, and that field experiments were neededto test these alternative hypotheses. They did not speculate about whysea 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 differentsize and species combinations in laboratory tanks over 2.5 years. Theyfound no significant differences among treatments, only trends thatlarge urchins grew better by themselves at all food levels, and all sizesof abalone grew better with urchins when food supply was moderateor in excess. Despite these results, they argued these trendsrepresented weak competition: “abalones may be able to decreaseurchins' relative fitness when food supplies are adequate throughinterference competition.” It could also be argued, however, that thelack of abalone would lead to an increase in drift algae and, therefore,decrease the probability of sea urchins switching from passive toactive foraging. Tegner and Levin (1982) also pointed out that urchinsmay have positive effects on abalone by providing habitat forjuveniles and benefiting growth, and Rogers-Bennett and Pearse(2001) found more juvenile abalone at sites with more red seaurchins. Moreover, observations of the effects of the sea urchin fisheryin northern California rocky habitats suggest the direction of thecompetition hypothesis may be reversed with sea urchins out-competing abalone (Tegner and Dayton, 2000).

To our knowledge these are the only data available from Californiato answer the sea urchin–abalone competition question. Jackson et al.(2001, Table 1) used data on declines in white abalone (Haliotissorenseni Bartsch) populations to support the argument that fishinghas greatly reduced abalone populations, thereby indirectly increas-ing sea urchins through a reduction in competition. Although whiteabalone populations in California have severely declined, this speciesis found at depths of 20–60 m (Haaker et al., 2001). If competitive

effects involving white abalone did exist, they could not affect mostCalifornia kelp forests which are most prolific above 20 m depth(Foster and Schiel, 1985). Commercial landings of abalone species thatare most common in southern California kelp forests (H. rufescensSwainson and H. corrugata Gray) remained high until 1969–1972(Haaker et al., 2001), suggesting populations were not severelyreduced in the 1950s–1970s. In summation, it seems unlikely thatabalone–sea urchin competition was significant and therefore shouldhave, at most, a very minor role in models of southern California kelpforest dynamics.

4.3.3. Sheephead and lobsterCalifornia sheephead occur primarily from southern California

south into México, and feed on a variety of prey, including sea urchinsand lobsters, in nearshore areas during the day and shelter at night(review in Cowen, 1983). California spiny lobsters have a similargeographic distribution, a broad diet including sea urchins, and shelterduring the day. They migrate into deeper water in winter (review inBarsky, 2001).

Tegner (1980) proposed that sheephead and spiny lobsters mightcontrol sea urchin populations because they eat sea urchins andbecause of patterns of urchin abundance and distribution at offshorelocations where sheephead were abundant. She noted, however, thatdeforestation by sea urchins had also been observed at these offshorelocations. Tegner and Dayton (1981) examined variation in sizefrequency distributions of live and dead sea urchins and made otherobservations at three sites in the Point Loma kelp forest, two withsheephead and spiny lobster and one without, to determine thepotential combined effects of these predators. Differences betweensea urchin size frequency distributions among sites with sheepheadand lobsters and the site without these predators were interpreted asindicating predator control, and this was further argued by Barry andTegner (1990). Their results were equivocal, however; the no-predator area was not replicated, the two treatments differed indepth and topography, and others have shown that urchin sizefrequency distributions similar to those found in these earlier studiescan also result from sporadic recruitment (Ebert et al., 1993).

Cowen (1983) provided the only experimental evidence forsheephead effects on sea urchins. He removed sheephead from a13,000 m2 site at the Channel Islands, and compared changes inlocations, numbers and size frequency distributions of red sea urchinsassociated with crevices to a control site. He found that the totalnumber of sea urchins and the number occurring outside of crevicesincreased in the experimental site over two years, and that urchinsplaced in the open in the control area quickly moved to shelter orwere eaten by sheephead. Cowen's (1983) data show that sheepheadeat red sea urchins and can affect their distribution relative tocrevices, but do not show that sheephead control red sea urchinabundance. He only sampled in and around crevices in a small portionof the experimental area and, as he pointed out, the increasedabundances could have resulted from sea urchin redistribution withinthe experimental area, with a net movement into the crevice areassampled. These experiments and surveys were done in kelp forestswhere sea urchins occurred mostly in and around crevices, presum-ably eating drift algae. It is not knownwhether the effect of sheepheadon sea urchin distribution would be amplified under conditions of lowdrift abundance when sea urchins begin to forage actively. It is clear,however, that drift supply alone can “regulate” deforestation byaltering sea urchin behaviour in areas where sheephead and lobsterare rare (Ebeling et al., 1985; Harrold and Reed, 1985; Reed, pers.com.). Moreover, Dean et al. (1984) found that the spatial distributionand movement of red sea urchin aggregations at a site with highsheephead densities (258/ha) “appeared to be unrelated to predationpressure from fishes or lobsters.” Dean et al. (1984) further concludedthat urchins exacerbate kelp declines initiated by other causes such asstorms and high temperatures/low nutrients.

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The primary data concerning lobster predation on sea urchins arefrom Tegner and Levin (1983) who found that lobsters maintained ona diet of sea urchins preferred to eat purple sea urchins and ate allsizes in the laboratory. They suggested this accounted for theunimodal size frequency distribution of live animals observed in thefield where lobsters and sheephead were common. Lobsters also atered sea urchins in these lab experiments, but mostly intermediatesizes, and Tegner and Levin (1983) suggested this was reflected in thebimodal size distribution of live red sea urchins in the field where thetwo predators were common. From these results and from landingdata for the lobster fishery, Tegner and Levin (1983) concludedlobsters and sheephead controlled sea urchin populations until fishingreduced their abundance in the 1950s, leading to increases in seaurchin populations and deforestation.

Lafferty (2004) and Behrens and Lafferty (2004) argued for lobstercontrol of sea urchin populations based on field survey data from theNorthern Channel Islands (Fig. 1). Comparisons were made of live seaurchin abundances and live and dead sea urchin size frequencydistributions between two sites within a no-take reserve on oneisland and areas where lobsters were fished on other islands. Lafferty(2004) found lobsters to be significantly more abundant and seaurchins significantly less abundant at reserve sites. Behrens andLafferty (2004) used size frequency distributions of live Strongylocen-trotus franciscanus and S. purpuratus and concluded that thesematched those suggested by Tegner and Levin (1983) to result fromlobster predation. However, this conclusion is only partially correct;comparing the data from these two studies shows they did match forS. purpuratus but not for S. franciscanus. Moreover, Behrens andLafferty (2004) did not mention that differences in size frequencydistributions may result from sporadic recruitment rather thanpredation (Ebert et al, 1993). Despite a lack of treatment replication,biogeographic differences between the reserve sub-sites and fishedsites (Murray et al., 1980; Hamilton et al., 2010) and potentialproblems with the interpretation of size frequency distributions,Behrens and Lafferty (2004) concluded their findings verified thehypothesis that spiny lobsters control sea urchin populations in thefield.

Taken together, these studies show that sheephead can affect seaurchin behaviour and that sheephead and lobster feeding increasessea urchin mortality. However, they constitute relatively weak,contradictory and inconsistence evidence for the hypothesis thatthese predators exert control over sea urchin populations in southernCalifornia at spatial scales larger than patches. It remains possible, ofcourse, that the removal of large lobsters and sheephead in the earlystages of the fisheries (Tegner and Levin, 1983) had a disproportion-ate effect on subsequent urchin population dynamics, although this isspeculative and is not reflected in any of the urchin data we havefound. Altogether, therefore, there is little evidential justification forthe statement by Jackson et al. (2001) that “ecological extinction” ofthese predators from fishing led to the disappearance of kelp forestduring the 1950s–1970s.

5. Conclusions

Giant kelp forests in southern California have not collapsed; thisclaim was based on extrapolating the dynamics of a few highlyimpacted mainland kelp forests to the entire region. Undoubtedly,there are multiple stressors affecting kelp forests and so the com-bination 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 andsedimentation on giant kelp, and recovery occurred when waterquality improved. A large El Niño and sea urchin grazing contributedto the declines, and the abundance of the latter may have beenincreased by discharged wastes. This scenario highlights improving

water quality as the most effective and direct management strategyfor the conservation of giant kelp forests in southern California.

Giant kelp is subjected to frequentdisturbance fromawidevariety ofnatural andhuman-induced influences. Given thehighdegree of canopyand biomass fluctuations seasonally and inter-annually it is remarkablethat this species, a key habitat engineer with which thousands of otherspecies are associated, is so robust and that kelp forests flourish invirtually all areas now that they occupied historically. The weedylife history of M. pyrifera enables relatively quick replenishment ofpopulations even after such major oceanographic events such as the1982–83 El Niño that deforested large areas of kelp forest along theentire California coast. The severe reduction of iconic species such asabalone, and large fishes and lobsters, is lamentable and requiresregulatory intervention to reverse. These reductions, however, do notappear to have led to the severe degradation of kelp forests generally,but effects on foodweb complexity, trophic interactions, carbon cyclingand other ecosystem processes are largely unknown. As pointed out intwo areas of the world, however, such reductions often leave fewvestiges or detectable effects on the wider ecosystem (Dayton et al.,1998; Schiel, 2006). Most people who have worked in or enjoyed kelpforestswould like to see the restoration of a full ensemble of species andinteractions, so decisions directly affecting giant kelp itself, rather thandiffuse and poorly understood interactions by some species, seems to usto be the most promising avenue of management. No level of species'protection or reserve statuswill be effective ifwater quality, coastal run-off, increased sedimentation, and contamination impact the ability ofgiant kelp to survive and thrive.

Acknowledgements

We thank P. Kalvass (California Department of Fish and Game) forproviding data on the sea urchin fishery, and J. Gully (Los Angeles CountySanitation District) and C. Mitchell (MBC Applied EnvironmentalSciences) for data on kelp canopy abundance. P. Dayton, J. Pearse,D. Reed, S. Schroeter and J. Steinbeck provided critical comments (eventhough they had reservations about some of our arguments). DRS thanksthe New Zealand Foundation for Research, Science and Technology(Coasts and Oceans OBI) and the Royal Society of New Zealand MarsdenFund, which support much of his coastal research. [ST]

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