Marine Sanctuaries Conservation Series MSD-04-2

U.S. Department of Commerce April 2004

National Oceanic and Atmospheric Administration

National Ocean Service

Office of Ocean and Coastal Resource Management

Marine Sanctuaries Division

 Sebastes chrysomelas, black-and-yellow rockfish

A Review of the Ecological Effectiveness of Subtidal Marine

Reserves in Central California

Part I:

Synopsis of Scientific Investigations
About the Marine Sanctuaries Conservation Series

The National Oceanic and Atmospheric Administration’s Marine Sanctuary Division (MSD)

administers the National Marine Sanctuary Program. Its mission is to identify, designate,

protect and manage the ecological, recreational, research, educational, historical, and aesthetic

resources and qualities of nationally significant coastal and marine areas. The existing marine

sanctuaries differ widely in their natural and historical resources and include nearshore and

open ocean areas ranging in size from less than one to over 5,000 square miles. Protected

habitats include rocky coasts, kelp forests, coral reefs, sea grass beds, estuarine habitats, hard

and soft bottom habitats, segments of whale migration routes, and shipwrecks.

Because of considerable differences in settings, resources, and threats, each marine sanctuary

has a tailored management plan. Conservation, education, research, monitoring and

enforcement programs vary accordingly. The integration of these programs is fundamental to

marine protected area management. The Marine Sanctuaries Conservation Series reflects and

supports this integration by providing a forum for publication and discussion of the complex

issues currently facing the National Marine Sanctuary Program. Topics of published reports

vary substantially and may include descriptions of educational programs, discussions on

resource management issues, and results of scientific research and monitoring projects. The

series will facilitate integration of natural sciences, socioeconomic and cultural sciences,

education, and policy development to accomplish the diverse needs of NOAA’s resource

protection mandate.
A Review of the Ecological Effectiveness of Subtidal Marine

Reserves in Central California

Part I:

Synopsis of Scientific Investigations

Richard M. Starr1, Mark H. Carr2, Jennifer Caselle3, James A. Estes4,

Caroline Pomeroy2, Craig Syms2, David A. VenTresca5, Mary M. Yoklavich6

1

 University of California Sea Grant Extension Program

2

 University of California Santa Cruz

3

 University of California Santa Barbara

4

 U.S. Geological Survey

5

 California Department of Fish and Game

6

 National Marine Fisheries Service

April 2004

 

Silver Spring, Maryland

April 2004

U.S. Department of Commerce

Donald L. Evans, Secretary

National Oceanic and Atmospheric Administration

VADM Conrad C. Lautenbacher, Jr. (USN-ret.)

Under Secretary of Commerce for Oceans and Atmosphere

National Ocean Service

Richard W. Spinrad, Ph.D., Assistant Administrator

Marine Sanctuaries Division

Daniel J. Basta, Director
DISCLAIMER

Report content does not necessarily reflect the views and policies of the National Marine

Sanctuary Program or the National Oceanic and Atmospheric Administration, nor does the

mention of trade names or commercial products constitute endorsement or recommendation for

use.

REPORT AVAILABILITY

Electronic copies of this report may be downloaded from the National Marine Sanctuaries

Program web site at www.sanctuaries.nos.noaa.gov. Hard copies may be available from the

following address:

National Oceanic and Atmospheric Administration

Marine Sanctuaries Division

SSMC 4, N/ORM62

1305 East-West Highway

Silver Spring, MD 20910

COVER PHOTO CREDITS

Sebastes chrysomelas, black-and-yellow rockfish: Daniel W. Gotshall

SUGGESTED CITATION

Starr, R. M., M. H. Carr, J. Caselle, J. A. Estes, C. Pomeroy, C. Syms, D. A. VenTresca, and M.

M. Yoklavich. 2004 A Review of the Ecological Effectiveness of Subtidal Marine Reserves in

Central California, Part I: Synopsis of Scientific Investigations. Marine Sanctuaries Conservation

Series MSD-04-02. U.S. Department of Commerce, National Oceanic and Atmospheric

Administration, Marine Sanctuaries Division, Silver Spring
CONTENTS

Introduction................................................................................................. 1

Chapter 1. No-Take Reserves In Central California Kelp Forests: Metrics of Human

Impact or The Tip Of The Iceberg?...................................................... 2

James A. Estes and Michelle J. Paddack

Chapter 2. Baseline In Situ Surveys and Landing Creels of Nearshore Fishes Within

the Vicinity of Big Creek Ecological Reserve...........................................24

David A. VenTresca, Marty L. Gingras, Mike Donnellan,

Jennifer Fisher, Briana Brady, Howatt King, and Chad King

Chapter 3. Deepwater Habitat And Fish Resources Associated With The Big Creek

Ecological Reserve..........................................................................43

Mary Yoklavich, Gregor Cailliet, Robert N. Lea, H. Gary Greene,

Richard M. Starr, Jean De Marignac, and Jeff Field

Chapter 4. Anacapa Island Ecological Reserve: A Research Review. ........................... 44

Jennifer Caselle

Chapter 5. Marine Reserves In The Monterey Bay National Marine Sanctuary:

 What We Do And Don’t Know......................................................... 51

Mark H. Carr and Craig Syms

Chapter 6. Effectiveness Of Marine Reserves: Socio-Economic Considerations.............. 73

Caroline Pomeroy

Summary................................................................................................... 88

Acknowledgements....................................................................................... 89

Literature Cited............................................................................................ 89

Appendix 1. Additional Literature Associated With Some Central California

 Marine Protected Areas................................................................ 109
1

INTRODUCTION

Marine reserves, often referred to as no-take MPAs, are defined as areas within which

human activities that can result in the removal or alteration of biotic and abiotic components of

an ecosystem are prohibited or greatly restricted (NRC 2001). Activities typically curtailed

within a marine reserve are extraction of organisms (e.g., commercial and recreational fishing,

kelp harvesting, commercial collecting), mariculture, and those activities that can alter

oceanographic or geologic attributes of the habitat (e.g., mining, shore-based industrial-related

intake and discharges of seawater and effluent). Usually, marine reserves are established to

conserve biodiversity or enhance nearby fishery resources. Thus, goals and objectives of marine

reserves can be inferred, even if they are not specifically articulated at the time of reserve

formation.

In this report, we review information about the effectiveness of the three marine reserves in

the Monterey Bay National Marine Sanctuary (Hopkins Marine Life Refuge, Point Lobos

Ecological Reserve, Big Creek Ecological Reserve), and the one in the Channel Islands National

Marine Sanctuary (the natural area on the north side of East Anacapa Island). Our efforts to

objectively evaluate reserves in Central California relative to reserve theory were greatly

hampered for four primary reasons; (1) few of the existing marine reserves were created with

clearly articulated goals or objectives, (2) relatively few studies of the ecological consequences

of existing reserves have been conducted, (3) no studies to date encompass the spatial and

temporal scope needed to identify ecosystem-wide effects of reserve protection, and (4) there are

almost no studies that describe the social and economic consequences of existing reserves.

To overcome these obstacles, we used several methods to evaluate the effectiveness of

subtidal marine reserves in Central California. We first conducted a literature review to find out

what research has been conducted in all marine reserves in Central California (Appendix 1). We

then reviewed the scientific literature that relates to marine reserve theory to help define criteria

to use as benchmarks for evaluation. A recent National Research Council (2001) report

summarized expected reserve benefits and provided the criteria we used for evaluation of

effectiveness. The next step was to identify the research projects in this region that collected

information in a way that enabled us to evaluate reserve theory relative to marine reserves in

Central California. Chapters 1-4 in this report provide summaries of those research projects.

Contained within these chapters are evaluations of reserve effectiveness for meeting specific

objectives. As few studies exist that pertain to reserve theory in Central California, we reviewed

studies of marine reserves in other temperate and tropical ecosystems to determine if there were

lessons to be learned from other parts of the world (Chapter 5). We also included a discussion of

social and economic considerations germane to the public policy decision-making processes

associated with marine reserves (Chapter 6). After reviewing all of these resources, we provided

a summary of the ecological benefits that could be expected from existing reserves in Central

California. The summary is presented in Part II of this report.
2

CHAPTER 1

No-Take Reserves in Central California Kelp Forests:

Metrics of Human Impact or the Tip of the Iceberg?

James A. Estes1 and Michelle J. Paddack2

1

U.S. Geological Survey

University of California Santa Cruz, Center for Ocean Health

100 Shaffer Road, Santa Cruz, California 95060

2

Rosenstiel School of Marine & Atmospheric Science

Division of Marine Biology & Fisheries

4600 Rickenbacker Causeway, Miami, Florida 33149

Abstract

Most evidence for the effects of marine reserves comes from tropical nearshore ecosystems.

We attempted to evaluate the effects of marine reserves on temperate kelp forest systems by

contrasting the population structure (density and size distribution) of 10 species of epibenthic

fishes and several aspects of the associated ecosystems between three marine reserves and

adjacent exploited areas in Central California. Densities of fishes were 12-35% greater within

the reserves but this difference was not statistically significant. Habitat features explained only

4% of the variation in fish density and did not vary consistently between reserves and nonreserves. The average length of rockfish (genus Sebastes) was significantly greater in 2 of the 3

reserve sites, as was the proportion of larger fish. Population density and size differences

combined to produce substantially greater biomass and therefore reproductive potential per unit

of area within the reserves. The magnitude of these effects seems to be influenced by the

reserve’s age. While our results demonstrate that current levels of fishing pressure influence

kelp forest rockfish populations, differences between the reserves and adjacent non-reserves are

surprisingly small. We discuss a number of reasons why the influences of fishing on kelp forest

ecosystems may be greater, or at least different, than our findings indicate. Potentially

confounding influences include the very small size of the reserves, effects of historical fishing,

poaching, spillover effects on adult and larval populations from reserve to non-reserve habitats,

and the possibility that catastrophic phase shifts induced by human disturbances have altered

both reserve and non-reserve areas.
3

Introduction

Marine resources worldwide are showing signs of degradation (Dayton et al. 1995, Botsford

et al. 1997, Lauck et al. 1998, Pauly et al. 1998). The problem may be even worse than we

imagine, owing to the largely unappreciated importance of top-down forcing processes (Pace et

al. 1999, Estes et al. 2001), past losses of large vertebrates and other consumers (Jackson et al.

2001), and the tendency of many ecological systems to behave in chaotic and non-linear ways

(Scheffer et al. 2001). Given these potential difficulties and their associated uncertainties, how

do we assess and redress the impacts of fisheries and other anthropogenic effects on marine

ecosystems? In response to the increasing evidence that management has failed to achieve

fishery sustainability (Dayton et al. 1995, Botsford et al. 1997, Ralston 1998), there is growing

interest in the use of no-take areas (marine reserves) as a fisheries management tool (Bohnsack

1993, Castilla and Fernandez 1998, Dayton 1998). Marine reserves may serve as buffers against

catastrophic declines caused by the synergistic interactions between exploitation and

environmental extremes (Bohnsack 1993), as well as protect against the inherent risk of

uncertainty in fisheries management (Lauck et al. 1998). Additionally, they may aid in

sustaining and possibly enhancing stocks (Murray et al. 1999).

Beneficial effects of marine reserves on fish and invertebrate populations have been

demonstrated in numerous studies (National Research Council 2001, Halpern in press). These

effects include increased abundance and increased individual size and age in targeted

populations. No take reserves may also enhance habitat quality, such as recovery of corals

(Roberts and Polunin 1993, McClanahan 1997a), species diversity (Cole et al. 1990, Russ and

Alcala 1996), and community stability (Castilla and Durán 1985, Roberts and Polunin 1993,

Dayton et al. 1995). Reserve effects may extend beyond reserve boundaries through spillover of

adults and/or larvae to fishing grounds (Attwood and Bennett 1994, Holland and Brazee 1996,

Castilla and Fernandez 1998).

The evidence that marine reserves enhance population abundance and individual size of

exploited species is unequivocal. However, it really shouldn’t surprise us that protection results

in measurable effects, given the large number of fisheries that have declined or collapsed

worldwide (Botsford et al. 1997, Pauly et al. 1998). The absence of any such effects would be

far more surprising and remarkable. The more important question is whether or not marine

reserves, in their present forms, can be used to achieve the goals of marine conservation and

fisheries management? The answer depends on the extent to which measured effects of marine

protected areas capture the full impacts of human exploitation. If marine protected areas in their

current forms return these systems to near pristine conditions, then the course of effective

conservation and management is relatively simple. If, on the other hand, pristine conditions were

far different from those achieved or achievable by marine reserves, then the proper course of

conservation and management is far less certain.

There are 103 marine protected areas along the California coast, only 11 of which receive

protection from all take (McArdle 1997). Unfortunately, most of California’s marine protected

areas lack baseline biological information, without which their effects are difficult to evaluate.

Understanding these effects is crucial if marine reserves are truly going to be incorporated into

fisheries and ecosystem management. Thus, we set out in the mid 1990s to assess the effect of
4

marine reserves on populations of exploited fish species and their habitat in Central California

kelp forests. We did this by contrasting three existing marine reserves with nearby areas in which

fishing was permitted. This approach has been used elsewhere to evaluate the influence of

fishing on reef fish populations (see for example, Buxton and Smale 1989, Cole et al. 1990,

Grigg 1994, McClanahan 1994), although relatively few of these have been conducted in coldtemperate regions (Bohnsack 1998) (for cold-temperate examples, see -- South Africa: Buxton

and Smale 1989, Bennett and Attwood 1991, Buxton 1993, Attwood and Bennett 1994; Chile:

Castilla and Durán 1985, Castilla 1996, Castilla and Fernandez 1998; USA: Palsson and

Pacunksi 1995, Rogers-Bennett et al. 1995, Palsson 1998), and only a small number of other

studies have considered kelp forest fishes (Cole et al. 1990, Palsson and Pacunski 1995, Palsson

1998, Babcock et al. 1999). Cold water/kelp forest systems differ from tropical reefs in

numerous ways, several of which may influence their response to fishing. Compared with coral

reefs, kelp forests are more productive (Duggins et al. 1989), have a lower fish species diversity

(Ebeling and Hixon 1991), and support fish faunas with higher overall trophic status because of

the paucity of herbivorous species (Gaines and Lubchenco 1982), at least in the northern

hemisphere (Choat 1982).

The purpose of this paper is two-fold. First, we contrast three marine protected areas with

nearby non-reserve areas in the Monterey Bay National Marine Sanctuary. These findings are a

synopsis of the results reported by Paddack and Estes (2000). Second, we consider the

implications of these findings to fisheries management and the conservation of kelp forest

ecosystems in Central California and elsewhere. We do this in two ways: by evaluating the

potential influences of adult fishes and their reproductive products both within the reserves and

in nearby non-reserve areas; and by identifying a variety of historical and modern processes that

could have led to substantial differences between the measured effects of the existing marine

reserves and the structure of pristine ecosystems.

Materials and Methods

Field research was done from 1994 through 1996 in or near three marine reserves (referred

to hereafter as areas--Fig. 1) within the Monterey Bay National Marine Sanctuary: Hopkins

Marine Life Refuge (all fishing prohibited since 1984); Point Lobos State and Ecological

Reserve (designated as a State Reserve in 1963 which prohibited invertebrate collection but

allowed commercial and recreational fishing, all fishing prohibited since 1973), and Big Creek

Marine Ecological Reserve (all fishing prohibited since January 1994). Non-reserve areas

consisted of at least two sites nearby each reserve (Table 1). All sampling was done in habitats

with a giant kelp (Macrocystis pyrifera) surface canopy at about 14 m depth over rocky

substrates with moderate rock relief.
5

 Figure 1. Location of study areas. Non-reserve sites (listed in Table 1) marked as dots.

Reserves shaded.

N

Green Gables

Mono Lobo S.M.

Vicente

Slate

Lopez

MBA

Macabe

Santa Cruz

California

Monterey

Pt. Lopez

Pt. Sur

10 km

Big Creek Marine 

Ecological Reserve

Pt. Lobos 

State Reserve

Hopkins Marine 

Life Refuge

3637’ °

121°57’

1 km

5 km

1 km
6

Table 1. List of Sites for each Reserve/Adjacent Non-Reserve (From Paddack and Estes 2000).

Area Site Location/ Depth Average and

Range

Hopkins Marine Life

Refuge

“Lead line” (HMLR-LD) Mid-reserve (off Bird Rocks)

9.1 m / 5.5-12.1 m

Hopkins East (HMLR-E) At eastern edge of reserve

10.3 m/9.7-11.8 m

Hopkins West (HMLR-W) Western-facing side of reserve

10 m/8.8-12.1 m

Hopkins: Adjacent NonReserveMonterey Bay Aquarium

(MBA)

Reef in front of Monterey Bay

Aquarium (0.2 km from eastern

reserve boundary) 10.6 / 7.6-12.1 m

Macabe Beach SW off Macabe Beach (0.9 km from

eastern reserve boundary)

11.2 m / 7.3-13.3 m

Green Gables In front of Green Gables B&B

(0.6 km from western reserve

boundary) 12.1 m / 9.1-13.0 m

Pt Lobos Marine Reserve Cypress Cove toward western point of Reserve

12.1 m / 8.5-18.5 m

Whalers Cove near the mouth of Whalers Cove

10.6 m / 10.6-17.0 m

Pt Lobos: Adjacent NonReserveMono Lobo NE of Whalers Cove (0.2 km from

reserve boundary)

12.7 m / 12.7-18.2 m

South Monastery off southern end of Monastery

Beach (0.5 km from reserve

boundary) 13.6 m / 7.0-17.0 m

Big Creek Marine

Ecological Reserve

Big Creek Cove Off main beach

13.9 m / 11.8-15.2 m

Square Black Rock 1 mile north of cove

15.2 m / 8.2-20.9 m

Big Creek : Adjacent NonReserveSlate Rock Off Esalen (3.2 km from northern

reserve boundary)

13.6 m / 12.1-23.3 m

Vicente Creek 1st creek south of Big Creek (1.2 km

from southern reserve boundary)

13.6 m / 10.3-21.2 m

Lopez Point south of Vicente Creek (3.2 km from

southern reserve boundary)

19.1 m / 17.0-21.2 m
7

Species

We focused on fishes that are both common in Central California kelp forests and exploited

by commercial and recreational fisheries. This included six rockfish species --Sebastes

atrovirens (kelp rockfish), S. carnatus (gopher rockfish), S. caurinus (copper rockfish), S.

chrysomelas (black and yellow rockfish), S. nebulosus (china rockfish), and S. miniatus

(vermilion rockfish); and four others --Ophiodon elongatus (lingcod); Hexagrammos

decagrammus, (kelp greenling); Semicossyphus pulcher (sheephead), and Scorpaenichthys

marmoratus (cabezon). Population analyses focused on the rockfishes. Young-of-the-year

(YOY) were not surveyed as their numbers are seasonal, they utilize different habitats than

adults (often recruiting into the giant kelp canopy; Carr 1989), and are difficult to identify to

species in the field.

Fish counts

The density of targeted fish species was estimated from fish counts by scuba divers within

50 m x 4 m transects. All transects were located from randomly selected origins and compass

bearings. Two divers simultaneously counted the number of fish encountered in a swath 2 m

wide and 1 m above the bottom on either side of the transect line.

Habitat surveys

Habitat surveys were conducted to determine the degree of similarity among sites in

topography, turf algal species composition and abundance, and kelp density (Fig. 2). This

information was in turn used to 1) discern whether habitat variation influenced fish populations,

and 2) help assess whether the differing assemblages of exploited species had any influence on

their associated ecosystems. Algal abundance and cover were measured because algae provide

recruitment habitat for rockfish, shelter fishes from predators, and provide a substrate for prey

(Love et al. 1991).

Divers counting fish were followed along the transect by two habitat surveyors. One

surveyor measured the percent cover of turf algae by placing a 1 m2 quadrat at a randomly

chosen distance along each 5 m segment of the transect line. Turf algae were categorized as

fleshy reds (e.g., Gigartina spp., Gelidium robustum, Rhodymenia californica), articulated

corallines (Corallina spp., Bosiella spp., Calliarthron spp.), encrusting corallines

(Lithothamnium spp. and Lithophyllum spp.), and epibenthic browns (Dictyopteris spp.,

Dictyoneurum californicum, Desmarestia ligulata, Cystoseira osmundacea), and the percent

cover for each category was visually estimated (Dethier et al. 1983). The second diver counted

sporophytes >1 m high of giant kelp (Macrocystis pyrifera) and other stipitate brown algae

(Laminaria spp., Pterygophora californica, Eisenia arborea) in a swath 1 m to either side of the

transect. The second habitat surveyor also classified the substrate in each 5 m segment as being

predominately sand, cobble, flat rock, low boulders (<1 m high), medium boulders (1-3 m high),

or pinnacles (>3 m high). Bottom depth was recorded at 5 m intervals along the transect line.

The relationship between fish density and each of the habitat factors was evaluated using a

stepwise linear regression.
8

 Figure 2. Substrate comparisons between reserve and non-reserve areas.

a. Hopkins

0

10

20

30

40

50

60

70

80 Reserve

Non-Reserve

b. Pt. Lobos

0

10

20

30

40

50

60

70

80

c. Big Creek

0

10

20

30

40

50

60

70

80

Sand Flat Rock Cobble Low Medium Pinnacle

 Boulder Boulder
9

Fish sizes

Total length of individual fishes was estimated in situ to the nearest centimeter with a diverheld Plexiglas ruler mounted perpendicularly at the end of a 1 m long rod (the rod and flexible

ruler allowed divers to measure fish in crevices). Transects to measure fish sizes were conducted

in the same areas as fish counts, but on different days in order to not bias either the count or size

data. Two divers swam a square course starting at the boat’s anchor, measuring every targeted

fish species that was encountered within 1 m of the sea floor. The accuracy of this method was

evaluated by collecting some of the fish with a pole spear in non-reserve areas after they had

been measured in situ. This was done for 32 individual fish of 3 species (S. carnatus, S.

chrysomelas, and S. atrovirens), ranging in total length from 14 to 32 cm. In situ measurements

of fish length ranged from < 1 cm to > 2 cm actual length (average difference was +0.13 cm ±

0.98 SD).

Total length (TL) measurements were converted to standard lengths (SL) using SL/TL ratios

provided for each species by Lea et al. (1999). A mass-standard length regression, based on

>500 individuals of S. atrovirens and S. chrysomelas collected near Monterey, California

between 1984 and 1986 (Larson et al., unpublished data), was then used to convert standard

length to biomass. We did not extrapolate these biomass estimates to other species as small

differences in length/weight relationships can translate into large differences in biomass for equal

numbers of fish. Fish biomass per unit area for each site was calculated by summing the product

of mass for each size class of fish, the proportion of fish in the corresponding size class, and fish

density.

Length-specific fecundity (LSF) was calculated for S. atrovirens and S. chrysomelas using

relationships with standard length (SL) provided by Romero (1988) and Zaitlin (1986)

respectively (S. atrovirens--LSF=2.1x10-5SL4.134, r

2

=0.788; S. chrysomelas--LSF=1.36x10-

8

SL5.59, r

2

=0.92). We use the term "reproductive potential" to describe the number of eggs

produced per area of habitat by a population. Reproductive potential was estimated for each

species and site by summing the product of the fecundity for each size class, the proportion of

fish measured in the corresponding size class, and fish density (Table 2). For this computation,

fish density was determined by dividing the fish counts by 2, assuming that rockfish have a

50:50 sex ratio (as there is no evidence to the contrary for inshore rockfishes).
10

Table 2. Total length at which rockfish populations reach 50% sexual maturity (from Wyllie

Echeverria 1987) and proportion of fish estimated to be sexually mature from this study

(From Paddack and Estes 2000).

___________________________________________________________________________

Size (& age) Reserve Non-Reserve

Fish Species

at 50%

maturity Area

n, size range

(TL, cm)

 %

mature

n, size

range

 %

mature

S. carnatus 17 (4 yr) Hopkins 28, 15-35 96 15, 10-30 87

Pt. Lobos 18, 16-37 94 55, 10-35 82

Big Creek 123, 9-35 93 85, 9-32 94

S. chrysomelas 15-16 (3 yr) Hopkins 44, 18-36 100 29, 11-31 88

Pt. Lobos 11, 26-35 100 25, 10-30 92

Big Creek 11, 15-34 100 6, 13-26 83

S. atrovirens no data

(assume 16 yr)

Hopkins 165, 10-42 96 100, 10-33 89

Pt. Lobos 28, 16-39 100 42, 17-34 100

Big Creek 41, 20-45 100 39, 16-37 100

S. caurinus 32-34 (4-6 yr.) Hopkins 6, 23-30 0 5, 25-30 0

Pt. Lobos 5, 37-44 100 3, 19-26 0

Big Creek 6, 15-43 50 8, 28-43 38

S. nebulosus 27 (4 yr.) Hopkins none none

Pt. Lobos none 1, 24 0

Big Creek 1, 26 0 none

S. miniatus 37-38 (5 yr.) Hopkins none none

Pt. Lobos 1, 21 0 2, 30-35 0

Big Creek 5, 33-40 60 4, 40-48 100

___________________________________________________________________________
11

Results

Habitat analyses

Frequency distributions of substrate types did not differ within areas between reserve and

non-reserve sites (Kolmogorov Smirnov 2 sample tests: Hopkins P=0.15, Pt. Lobos P=0.26, Big

Creek P=0.36). Frequency distributions of percent cover for articulated coralline, encrusting

coralline, brown, and foliose red algae varied considerably between each reserve/non-reserve

pair (Fig. 3). G-tests showed that 8 of 12 possible pairwise comparisons (3 areas x 4 algal

categories) between reserve and non-reserve sites differed significantly. However, there were no

consistent trends for any algal class between reserve versus non-reserve areas (Fig. 3). Neither

giant kelp (ANOVA, F1,2=1.24, P=0.38) nor epibenthic stipitate kelp (ANOVA, F1,2=3.99,

P=0.18) densities differed significantly between reserve and non-reserve sites (Table 3). A

multiple regression showed significant correlations between articulated coralline algae (p=0.01),

brown algal ground cover (p=0.03), and stipitate kelps (p=0.009) and rockfish density.

Collectively however, these factors described only 4% of the variation in fish density (r2=0.041).

_______________________________________________________________________________

Table 3. Mean counts (No. 10 m-2 ±SD) of canopy (Macrocystis pyrifera) and understory

(Laminaria spp., Pterygophora californica, Eisenia arborea) kelps. (From Paddack and

Estes 2000).

Area Canopy

F1,2=1.24, P=0.38

Understory

F1,2=3.99, P=0.18

Reserve Non-Reserve Reserve Non-Reserve

Hopkins 1.6 ± 1.27 1.6 ± 0.71 1.2 ± 1.91 0.4 ± 0.62

Pt. Lobos 1.7 ± 1.77 1.3 ± 0.78 5.3 ± 1.39 2.4 ± 2.11

Big Creek 3.6 ± 1.82 0.7 ± 0.60 4.9 ± 3.87 4.4 ± 3.45

______________________________________________________________________________
12

Figure 3. Average percent cover of turf algae in reserve and non-reserve areas

for each of four categories (*** - P<0.001, ** - P<0.01, *- P<0.05).

a. Hopkins

0

10

20

30

40

50

60

70

Reserve

Non-Reserve

b. Pt. Lobos

% Frequency

0

10

20

30

40

50

60

70

c. Big Creek

0

10

20

30

40

50

60

70

**

*

*

*

*

*

*

*

*

* **

*

Encrusting

corallines

Articulated

corallines

Foliose reds Turf browns
13

Fish Species Composition

Fish species composition differed somewhat among the three study areas, due to the absence

or very low density of certain species at some sites, but this did not appear to be due to reserve

status (Fig. 5). Cluster analysis indicated that species composition was most similar between

each reserve/non-reserve pair (Fig. 4b). At Pt. Lobos and Big Creek, S. carnatus was the most

common species whereas S. atrovirens was the most common species at Hopkins. S. nebulosus

was rare, found only in the Big Creek sites and the non-reserve site at Pt. Lobos. Semicossyphus

pulcher (a labrid fish near the northern end of its geographic range in Central California--Miller

and Lea 1972) occurred at Big Creek, was rare at Pt. Lobos, and was absent from our samples at

Hopkins.

Figure 4. Cluster diagrams based on a) habitat variables and b) species

composition. Dissimilarity measure is 1 minus the Pearson

product-moment correlation coefficient. (H= Hopkins, L=

Pt. Lobos, B= Big Creek, I= Inside reserve, O= Outside reserve).

0.0 0.1 0.2 0.3 0.4

HI

HO

LI

LO

BI

BO

A. Habitat

Dissimilarity Index

B. Fish Species Composition

0.0 0.05 0.10 0.15 0.20

HI

HO

LI

LO

BI

BO
14

Figure 5. Species composition of surveyed fish species at each reserve/nonreserve site.

a. Hopkins

S. atrovirens

S. chrysomelas

S. carnatus

H. decagrammus

S. marmoratus

O. elongatus

S. caurinus

S. miniatus

S. nebulosus

S. pulcher Reserve

Non-Reserve

b. Pt. Lobos

S. atrovirens

S. chrysomelas

S. carnatus

H. decagrammus

S. marmoratus

O. elongatus

S. caurinus

S. miniatus

S. nebulosus

S. pulcher

c. Big Creek

Percent Composition

60 40 20 0 20 40 60

S. atrovirens

S. chrysomelas

S. carnatus

H. decagrammus

S. marmoratus

O. elongatus

S. caurinus

S. miniatus

S. nebulosus

S. pulcher
15

Fish Density

Fish density was 12-35% greater in the reserves than the adjacent non reserves (Fig. 6,

Table 4), but these results were not statistically significant (two-way blocked ANOVA,

F1,2=12.751, P=0.07). It should be noted, however, that the power of this test is very low (1-β

<0.20 for effect size of 0.12 and 1-β = 0.41 for an effect size of 0.25).

____________________________________________________________________________

 Figure 6. Fish densities (mean ± SEM) in reserve and non-reserve areas for all 

species combined (n=number of transects per area).

Hopkins Pt. Lobos Big Creek

Average Fish Density per 10 m-2

0.0

0.2

0.4

0.6

0.8

1.0

1.2 Reserve

Non-Reserve

n=31

n=12

n=6

n=15

n=12

n=11
16

Table 4. Density of each fish species (No. 10 m-2 ± SD). The F statistics below each fish species

test for overall reserve effects. (From Paddack and Estes 2000)

Fish species

Hopkins

Reserve

NonReserve

Pt. Lobos

Reserve

NonReserveBigCreek

Reserve

NonReserve

Sebastes

carnatus 0.12±0.17 0.04±0.06 0.28±0.26 0.28±0.22 0.33±0.27 0.22±0.23

F1,2=3.40, P=0.21

S. chrysomelas 0.13±0.13 0.08±0.12 0.17±0.18 0.17±0.13 0.05±0.10 0.10±0.23

F1,2=0.01, P=0.93

S. caurinus 0.02±0.04 0 0.04±0.07 0.04±0.09 0.03±0.05 0.01±0.03

F1,2=2.87, P=0.23

S. atrovirens 0.33±0.41 0.23±0.30 0.24±0.15 0.18±0.30 0.20±0.30 0.17±0.15

F1,2=8.99, P=0.10

S. miniatus 0 0 0.02±0.03 0 0 0.04±0.04

F1,2=0.34, P=0.62

S. nebulosus 0 0 0 0.003±0.01 0.004±0.01 0.002± 0.01

F1,2=0.13, P=0.75

Scorpaenichtys

marmoratus 0.03±0.04 0.03±0.04 0.02±0.03 0.03±0.04 0.02±0.04 0.01±0.03

F1,2=0.20, P=0.70

Semicossyphus

pulcher 0 0 0.01±0.02 0 0.09±0.19 0.04±0.07

F1,2=1.25, P=0.38

Ophiodon

elongatus 0.01±0.02 0.01±0.02 0.04±0.06 0.01±0.02 0.02±0.05 0.02±0.02

F1,2=2.60, P=0.25

Hexagrammos

decagrammus 0.04±0.06 0.05±0.06 0.04±0.06 0.02±0.03 0.07±0.07 0.04±0.06

F1,2=2.18, P=0.28
17

Population Structure

Size-frequencies were combined within sites for those rockfish species having similar

maximum sizes (i.e., S. carnatus, S. chrysomelas, S. atrovirens, S. caurinus; Fig. 7). For the two

areas protected the longest (Hopkins, 12 yr; Pt. Lobos, 23 yr), average lengths were significantly

greater in reserves than non-reserves (Hopkins--t=9.29, df=390, P<0.001; Pt. Lobos--t=7.10,

df=191, P<0.001). Mean lengths at Big Creek (protected for 1 yr.at the time of our study) did

not differ significantly (t=0.51, df=328, P=0.304). Kolmogorov-Smirnov two sample tests for

each reserve/non-reserve comparison also showed significantly different length-frequency

distributions (i.e., population structures) at Hopkins (P<0.001) and Pt. Lobos (P<0.001), but not

Big Creek (P=0.99). Pooled Kolmogorov-Smirnov tests resulted in a significant difference

between reserves and non-reserves overall (P=0.0002). The populations of non-reserve fish at

Hopkins and Pt. Lobos were dominated by small size classes, whereas at Big Creek the size class

distributions were similar between reserve and non-reserve sites (Fig. 7). This pattern held for

each of the common rockfish species.

Biomass

Biomass density (g fish 10 m-2) estimates for S. atrovirens and S. chrysomelas at both

Hopkins and Pt. Lobos were >2 times higher in the reserve than non-reserve sites while at Big

Creek there was no discernible difference between the reserve and non-reserve sites (Fig. 8).

Reproductive Potential

While the length-fecundity equations are power functions typical of mass-length allometries

(Zaitlin 1986, Romero 1988), they diverge considerably between S. atrovirens and S.

chrysomelas for fish >20 cm standard length. Based on these length-fecundity relationships and

size frequency distributions, estimated reproductive output for both species was greater in the

reserve than the non-reserve sites at Hopkins and Pt. Lobos, but similar at Big Creek (Fig. 9).

The comparatively low estimates for Big Creek result from the general lack of larger individuals

in both reserve and non-reserve sites.
18

 Figure 7. Length frequency distributions of rockfish species combined (S. atrovirens, S.

chrysomelas, S. carnatus, S. caurinus) in the reserve and non-reserve areas at Hopkins,

Pt. Lobos, and Big Creek.

a. Hopkins

n=243

mean=28.0 +4.9

0

4

8

12

n=149

mean=23.2 +5.2

0

4

8

12

Non-Reserve

b. Pt. Lobos

n=63

mean=29.7 +5.7

% Frequency

0

4

8

12

Reserve

n=131

mean=23.9 +5.1

0

4

8

12

c. Big Creek

n=187

mean=26.1 +5.75

0

4

8

12

n=143

mean=26.4 +5.75

Total Length (cm)

0 4 8 12 16 20 24 28 32 36 40 44 48

0

4

8

12
19

Figure 8. Average biomass (±SEM) per unit area (10 m2) for S. atrovirens and

S. chrysomelas in reserve/non-reserve areas.

Fish Biomass (g 10 m-2)

0

20

40

60

80

100

120

140

160

Reserve

Non-Reserve

 a. Sebastes atrovirens

Hopkins Pt. Lobos Big Creek

0

20

40

60

80

100

120

140

160 b. Sebastes chrysomelas
20

Figure 9. Estimated reproductive potential (mean ± SEM) for S. atrovirens and S.

chrysomelas in reserve and non-reserve areas.

 a. Sebastes atrovirens

Reproductive potential (no. eggs produced 10 m-2)

0

5000

10000

15000

20000

25000

30000

35000

40000 Reserve

Non-Reserve

 b. Sebastes chrysomelas

Hopkins Pt. Lobos Big Creek

0

5000

10000

15000

20000

25000

30000

35000

40000
21

Discussion

Our findings, like those of so many others who have studied the influences of marine

reserves (National Research Council 2001, Halpern in press), provide a clear but surprisingly

subtle sign of human impact. Are these data an accurate metric of human impact or are they only

showing us the tip of the proverbial iceberg? They might represent the “tip of an iceberg” if 1)

reserves are too small to allow buildup of biomass or retain viable populations of rare species, 2)

high spillover occurs from reserves to non-reserves, 3) recovery in the reserves from

overexploitation is still ongoing, 4) poaching is a significant factor, or 5) earlier impacts of

human exploitation resulted in large-scale changes in ecosystem function that influence reserve

and non-reserve areas similarly.

The possibility that the reserves we studied are too small to capture the full influence of

human take is especially worthy of consideration. Conservation biologists have been quibbling

over optimum reserve size for decades (e.g., the SLOSS debate; Diamond 1975, Simberloff and

Abele 1976). A renewed interest in this topic has led some terrestrial ecologists and conservation

biologists to conclude that small reserves are inadequate to preserve biodiversity, owing largely

to a growing realization that top-down forcing processes are important in many natural systems

(Pace et al. 1999, Estes et al. 2001) and large areas of habitat are necessary to maintain viable

populations of large, apex predators (Soulé and Terborgh 1999). For instance, when home range

size is large relative to reserve size, a species’ probability of extinction within a reserve or

protected area may be high (Woodroffe and Ginsberg 1998). Furthermore, if their normal

movements frequently take individuals beyond a reserve boundary and the risk of being caught

in that area is high, stock increases within the reserve are unlikely to realize their full potential.

Small or isolated reserves also are more likely to depend upon external sources of larvae, making

them vulnerable to recruitment over-fishing in heavily fished areas (DeMartini 1993, Jennings et

al. 1996, Roberts 1997, Carr and Raimondi 1998).

The sizes of most coastal marine reserves worldwide are miniscule compared with their

terrestrial analogues, and the reserves we studied are no exception (Carr et al., in press). Might

larger reserves show more dramatic or even qualitatively different effects of human exploitation

than their small counterparts? Halpern (in press) surveyed the literature on marine reserves and

found no evidence for an effect of reserve size on species diversity, density of exploited

populations, or trophic disfunction. While this study was based on an impressively long list of

case studies, an effect of reserve size could be masked by profound and large-scale effects of

historical overfishing (Jackson et al 2001) and the extremely small size of most existing marine

reserves. Thus, even the largest marine reserves may be too small to capture a reserve size effect.

This possibility seems especially likely to us, given that historical overfishing targeted the large

apex predators (Pauley et al. 1998), the spatial scale of key forcing processes is greater in the sea

than it is on land, and nearly all marine reserves are very small compared with their terrestrial

counterparts.

Several studies have shown increased fish abundance in exploited areas nearby marine

reserves, a phenomenon attributed to adult spillover from the reserves (Polunin and Roberts

1993, Attwood and Bennett 1994, Russ and Alcala 1996, Wantiez et al. 1997). This effect also

may decrease rockfish density differences between exploited and protected areas by enhancing
22

populations in exploited areas and reducing populations in protected areas. While there is some

evidence that rockfish move from areas of high to low population density (Matthews 1985), they

also have been shown to have small home ranges (Miller and Geibel 1973, Larson 1980).

Rockfish density and biomass also may be slow to recover from fishing pressure (Holland and

Brazee 1996, Russ and Alcala 1996, Gunderson 1997, McClanahan 1997a) as these species are

typically long-lived, slow growing, and late to mature (Love et al. 1990, Leaman 1991).

Poaching, which undoubtedly occurs in Central California, may also help to explain the modest

differences in fish populations between reserves and adjacent protected areas.

The length-frequency data from our study provide additional and statistically stronger

support for rockfish population differences between protected and unprotected kelp forest sites.

Increased fish sizes within reserves also have been demonstrated in many other studies (Roberts

and Polunin 1993, Dufor et al. 1995). But here again, do these patterns properly reflect the full

effect of human impact on fish size or might truly unexploited populations contain much larger

fish? Examples of fish from aboriginal midden sites that are far larger than those known from

modern or more recently historic populations seems to support this latter possibility. Because

reproductive output often scales to fish size and biomass, this same caveat applies to our

calculations of reserve effects on the supply of young fish.

The perception of understanding human impacts on the kelp forest ecosystems of Central

California by contrasting reserve/non-reserve areas must be further tempered by other sources of

uncertainty. One of these is the lack of any real historical information on the structure and

function of these systems before the time of human contact. This is especially important for the

New World, which was not peopled until about 14,000 ybp. The influence of human exploitation

on coastal marine systems is more extreme and has a deeper history than many people realize or

even care to admit (Jackson et al. 2001). Worldwide, spectacularly abundant populations of large

animals inhabited coastal waters and in many cases exerted important functional influences on

these systems. Many of these large animals have been so depleted that their ecological roles are

extinct, even if the species themselves still survive. Central California kelp forests lack Steller’s

sea cows (Hydrodamalis gigas), a large herbivorous mammal that was common in the area until

late in the Pleistocene and may have exerted an importance grazing influence on kelp forests

(Estes et al. 1989). Sea otters (Enhydra lutris) are another functionally important species in

Central California kelp forests although they have arguably recovered to historic levels in

Monterey Bay. The pervasive influence of sea otters as predators on benthic invertebrates, and

the cascading effects of this interaction on kelp populations (Estes and Palmisano 1974) may

explain the absence of any evident top-down influence of the marine reserves in our study as the

otters forage across the reserves and their adjacent non-protected sites. Nonetheless, other

important consumers, especially the larger coastal fishes, are probably reduced or absent

compared with earlier times. The functional ghosts left by these species may have preconditioned

their associated ecosystems for further change, which we now consider natural (Dayton et al.

1998, Jackson 2001). An even greater potential for change through time relates to the fact that

many natural systems undergo catastrophic phase-shifts due to non-linearities in the forcing

functions (Scheffer et al. 2001). Phase shifts of this nature are known to occur in kelp forest

ecosystems (Konar and Estes in press) and probably take place in other coastal marine

ecosystems as well (Petraitis and Dudgeon 1999). A highly significant consequence of this
23

dynamic is that small disturbances can cause large change while large counter-disturbances to

the altered system may result in no effect at all (May 1977, Scheffer et al. 2001).

In sum, while our research provides unequivocal evidence for both a detrimental effect of

human take on marine resources of the Monterey Bay National Marine Sanctuary and thus a

beneficial influence of marine reserves on these same resources, these findings may mean more

than is apparent. The extremely small size of these reserves, their relatively young age, and

uncertainties relating to both an appropriate historical baseline and the chaotic features of

ecosystem behavior might very well render as grossly incorrect the naïve and simplistic view that

pristine populations and ecosystems have somehow been re-created by removing the influence of

human exploitation from our coastal marine reserves.

Acknowledgements

We wish to acknowledge assistance from the following individuals in the conduct of our

field work and preparation of the manuscript: Stian Allessandrini, Giacomo Bernardi, Mark Carr,

Nicole Crane, Eric Danner, Dan Doak, Matt Edwards, Ross Hein, Ralph Larson, Bob Lea,

Christian MacDonald, Tim McClanahan, Gary Russ, Enric Sala, Joshua Sladek Nowlis, John

Pearse, and Tim Tinker. Special thanks go to Paul Dayton, who the editor for our paper in

Ecological Applications, held us to high standards. We are grateful to the Oceanic Society, the

Packard Foundation, Monterey Bay Regional Studies Program (MBRS), Mildred E. Mathias

Natural Reserve Systems Grant, Meyers Oceanographic and Marine Biology Trust, and the

UCSC Institute of Marine Sciences for funding and support.
24

CHAPTER 2

Baseline In Situ Surveys And Landing Creels Of Nearshore Fishes

Within The Vicinity Of Big Creek Ecological Reserve

David A. VenTresca, Marty L. Gingras, Mike Donnellan,

 Jennifer Fisher, Briana Brady, Howatt King, and Chad King

Marine Region

California Department of Fish and Game

20 Lower Ragsdale Drive, Monterey, California 93940

Abstract

The decline in the success and quality of Central California’s nearshore rockfish fisheries

over the past decades and the rapid expansion of the live/premium fish fishery, which targets

shallow water kelp forest fishes, have underscored the need for a revised management strategy

for this resource. Marine protected areas have been suggested as an alternative or additional

management tool to create sustainable fisheries. The overarching goal of the study is to provide a

benchmark or “line in the sand” of data collected in a newly established marine protected area.

Because both fishery independent and dependent data are crucial in the evaluation of the reserve

as a fishery management tool, we collected data from in situ surveys of fish populations within

and adjacent to the Big Creek Ecological Reserve (BCER) and monitored the landings of

commercial and recreational fisheries targeting fishes in the vicinity of BCER.

We examined numbers of all fish observed during random transects conducted within and

adjacent to BCER. Counts north of BCER in 1995 and 1998 were significantly different (41%

declined); however, counts within and south of BCER were not statistically different between

1995 and 1998. A high percentage of 1998-99 Big Sur commercial skiff fishery landings were

composed of cabezon, and black-and-yellow and gopher rockfishes. This study documented that

4.5 times more Commercial Passenger Fishing Vessels (CPFV) trips were made to the Cape San

Martin region as were made to the Point Sur region. Most species taken in the Point Sur region

were larger than those taken in the Cape San Martin region; however, there was not a significant

difference in the CPUE between the two areas.
25

Introduction

Marine reserves, established as a fisheries management tool, are intended to manage

exploited stocks by protecting populations of sexually mature species vulnerable to overfishing

and thus insuring a continual source of new recruits. Reserves have been reported to enhance

fisheries in other parts of the nation and world; however, limited information is available to

evaluate their effectiveness on California's sport and commercial rockfish fisheries. Rockfish are

an important and heavily exploited component of sport and commercial fisheries in California.

The estimated value of the sport fishery to California’s economy is about one billion dollars

annually (Lenarz 1987), while the annual ex-vessel value of commercial fisheries for rockfish

exceeds ten million dollars (CDFG unpubl. data). In 1995, California Commercial Passenger

Fishing Vessels (CPFV) reported landing 1,174,991 rockfish; 62% were landed in Central

California (CDFG unpubl. data).

The success and quality of Central California’s nearshore rockfish sport fisheries have

declined over the past decades, particularly in areas close to ports (Miller and Gotshall 1965,

VenTresca and Lea 1984, Reilly et al. 1993, Karpov et al. 1995, Mason 1995). In the northern

and Central California sport fishery, between the late 1950s and early 1980s, the average weight

per rockfish declined by 13 percent, or 0.43 pounds per fish; average weight decreased for 12 of

16 major rockfish species (Karpov et al. 1995). In Monterey Bay, from 1987 to 1991, the average

length of several rockfish species sampled in the CPFV fishery was below the average size at

50% sexual maturity (Reilly et al. 1993 and Reilly et al. 1998). Presently, fishery managers are

concerned with rapid increase in fishing effort and landings of the emerging commercial

live/premium fish fishery that has developed coast-wide for rockfish, cabezon, lingcod, and kelp

greenling during recent years. The rapid expansion of this fishery is of greatest concern in the

nearshore, shallow, rocky habitats where many of these species are concentrated. As stocks

proximal to port become depleted, fishers have expanded their range further from port and into

deeper waters. Many of these nearshore areas, especially along the Big Sur coast, had received

little fishing pressure and until recently had functioned as de facto reserves.

Limited movement, relatively long life span, and late maturity of rockfishes, coupled with

heavy sport and commercial fishing pressure, has resulted in removal of many mature fish from

reefs proximal to fishing ports (Lea et al. 1999). Consequently, commercial and sport fishers are

traveling farther from port to maintain a quality catch. Once large reproductive adults are

removed, continued fishing pressure prevents remaining fish from reaching the size at which

sexual maturity occurs. Reproductive potential is also affected by natural phenomena, such as an

El Niño event. During these events growth rates and survival are often reduced (Bailey and Incze

1985, VenTresca et al. 1995).

Recognizing that a change in the current resource management approach is warranted,

alternative regulatory and enhancement strategies are of paramount importance. Rockfish appear

to be an excellent candidate for enhancement using a reserve management approach because in

addition to being residential and long-lived they are extremely fecund and have a lengthy larval

stage. Reserves, utilized as a rockfish management tool, would protect critical spawning stock

biomass and ensure a continual recruitment supply to fished areas via larval dispersal and

emigration. However, to determine the effectiveness of a reserve to enhance rockfish fisheries,
26

baseline information is first needed to demonstrate that densities and sizes of fish populations

within the reserve are greater than adjacent fished areas and that absolute numbers of species of

concern within the reserve are sufficient to provide a reasonable/significant source of larvae.

Marine fisheries reserves offer potential benefits to coastal fisheries and marine resource

management, including: the enhancement and restoration of fishery yields through larval/egg

transport; "spillover" from reserves to surrounding areas; protection of reproductive potential of

targeted species; maintenance of biological diversity; and increased social and economic benefits

to local communities (Russ 1985, Davis 1989, Bohnsack 1990, Roberts and Polunin 1991,

Rowley 1994, Roberts et al. 1995). Additionally, reserves may increase the resiliency of heavily

exploited fisheries during episodic environmental anomalies by serving as a buffer against

drastic declines and hastening the recovery of these fisheries by supplying harvested stocks with

new recruits (Carr and Reed 1993). Although reserves provide an excellent opportunity for in

situ testing of concepts related to the protection and enhancement of fish populations, studies

demonstrating a significant increase in fish population densities and average length within a

reserve are few (Rowley 1992, Dugan and Davis 1993, Palsson and Pacunski 1995, Roberts et al.

1995, Russ and Alcala 1989, and Paddack and Estes 2000). Studies showing enhancement or

"spillover" are also limited but are probably most critical in convincing public user groups of the

value of reserves (Rougharden and Iwasa 1986, Battershill 1993, Rowley 1992).

Although marine reserves have been reported to enhance fisheries in other parts of the

nation and world, limited information is available to evaluate their effectiveness on California’s

sport and commercial rockfish fisheries. Because improperly designed reserves may endanger a

fishery by providing a false sense of protection, determining the effectiveness of a reserve is of

utmost importance (Carr and Reed 1993, Yoklavich 1998). The establishment of the Big Creek

Ecological Reserve (BCER) in Monterey County in January 1994 provided an excellent

opportunity to evaluate the effects of a reserve on the nearshore rockfish resource; however,

without first obtaining baseline information on species composition, densities, and length

frequencies of rockfish populations within and adjacent to BCER, determining change in

population parameters or future benefits to adjacent and distant fisheries will be difficult at best.

Rockfishes are an excellent candidate species group for this study. They are long-lived,

residential, fecund, and have an open ended reproductive system that distributes offspring

spatially. Resident populations of large sexually mature rockfishes within the reserve would be a

continual source of larvae that in turn would be carried via nearshore currents to other sections of

the coastline (Chelton et al. 1982, Wyllie-Echeverria 1987, VenTresca et al. 1996). A future

management strategy might include a system of coastal reserves that would reinforce and

enhance the reproductive strategies of rockfish.

Baseline information on species composition, densities, and size frequencies of rockfish

populations within and adjacent to BCER are crucial to evaluate the effectiveness of this reserve.

In addition, resource managers will be unable to evaluate marine reserves as an alternative

management tool for rockfish. Furthermore, assessing the status of nearshore fish

populations/stocks in the vicinity of BCER is timely and relevant due to the recent dramatic

increase in commercial hook-and-line landings of nearshore species throughout California. The

extensive kelp forests along the Big Sur coast, until recently were one of the last remaining
27

unexploited marine habitats in California. Now this area is being intensely fished to provide for

the increasing demand in the lucrative live fish market. As fishing effort increases on these

formerly pristine stocks, it becomes imperative to evaluate the effectiveness of a Central

California coast reserve to protect and enhance rockfish stocks.

Most studies of marine protected areas have focused on assessing changes in adjacent

fisheries or changes in the population structure of fishes within the reserves and/or adjacent

areas. Both of these sources of information are crucial in the evaluation of the reserve as a

fishery management tool. We monitored both the landings of a commercial skiff fishery and the

CPFV fisheries that target fishes near BCER. We also used random transects and permanent

stations sampling of in situ fish population of fishes within and adjacent to BCER. The

overarching goal of this study is to provide a benchmark or “line in the sand” of data collected on

a newly established marine protected area. Our hope is that we have done an adequate job so that

future studies can statistically document temporal changes in nearshore fish populations within

and adjacent to BCER.

Methods

Population parameters of selected fish species in nearshore habitats within and adjacent to

BCER were assessed and analyzed utilizing two approaches: 1) in situ subtidal-reef fish

population data were collected using standardized scuba techniques in summer/fall 1995-98 and

2) species composition, catch-per-unit-effort (CPUE) and length frequencies of fishes taken in

commercial skiff and CPFV fisheries in the areas adjacent to BCER were collected from fall

1997 through winter 1999.

In Situ Fish Density Estimates

The BCER subtidal study site, located 50 miles south of Monterey, Monterey County, was

divided into three areas: 1) BCER, which encompasses 2.5 na. mi. of coastline from Oyster

Catcher Point in the south, northward to Rat Creek; 2) north of BCER from the northern BCER

boundary, northward to Slate Rock, a distance of 4 na. mi.; and 3) south of BCER from the

southern BCER boundary, southward to Lopez Point, a distance of 4 na. mi. (Figure 1). Total

number of fish counted, i.e., the sum of the counts of both scuba divers surveying the transect

were use for the analysis. Total area surveyed by each transect was 320 m2.

Optimal conditions for obtaining in situ visual density estimates of subadult and adult fish

populations using scuba along the Central California coast occur after the spring plankton

blooms during the summer, and before winter storms in the fall (personal observation). Due to

the remote location of the Big Creek study area and in an attempt to collect a high number of

observations during similar conditions of underwater visibility and surge, the Department’s R/V

MAKO was utilized for 2-3 week survey cruises. The R/V MAKO provided a platform to

accommodate 7-8 divers and an air compressor to fill scuba tanks. Divers made 3-4 dives per

day, commencing after 0900 hours to optimize the underwater visibility.
28

Figure 1. Commercial Passenger Fishing Vessel and Commercial Skiff Fishery fishing sites and

scuba survey area (in vicinity of Big Creek Ecological Reserve), Monterey County.
29

A survey team consisted of two divers, each equipped with standard personal scuba gear, a

plastic slate that securely held data recording sheets and a 10-m retractable transect (Ugortez et

al. 1997). Divers were trained in underwater species identification and assessment of size of

sexual maturation. Scuba transects were deployed in 10-20 m depths in kelp forest habitats.

Target fishes for this study were nearshore subadult and adult fish species that are harvested

commercially and/or recreationally in Central California nearshore rocky bottom areas. These

species include: Sebastes atrovirens (kelp rockfish), Sebastes carnatus (gopher rockfish),

Sebastes chrysomelas (black-and-yellow rockfish), Sebastes caurinus (copper rockfish), Sebastes

nebulosus (China rockfish), Sebastes melanops (black rockfish), Sebastes miniatus (vermilion

rockfish), Sebastes serranoides (olive rockfish), Hexagrammos decagrammus (kelp greenling),

Ophiodon elongatus (lingcod), and Scorpaenichthys marmoratus (cabezon). All of these species

are readily quantifiable using benthic scuba transects. Several infrequently observed species such

as: California sheephead (Semicossyphus pulcher), kelp bass (Paralabrax clathratus), treefish

(Sebastes serriceps), and wolf-eel (Anarrhichthys ocellatus) were also recorded when observed.

Landing of Commercial Skiff Fishery

Commercial skiff fishery landings along the Big Sur coast were monitored and sampled

during 1997-1999. Sampling procedures were, in part, based on criteria developed by the

California Cooperative Commercial Groundfish Survey (Thomas et al. 1995). We interviewed

commercial skiff fishermen as they returned from fishing, on the beach at Mill Creek State Park

or BCER, Monterey County (Figure 1). For each returning fishermen, the sampler recorded date,

port of landing, boat number, gear, estimated sampled weight and total weight of catch, CDFG

Block number, site-specific fishing location, time, and depth fished. Because fishermen wanted

to transport the live fish to market as soon as possible they did not permit the sampler to weigh

the catch with a scale. Fishers used plastic laundry baskets and 5-gallon buckets to transport live

and dead fish, respectively, from their skiffs to cars. The average weight of fish in a full basket

was 65-75 pounds and 30-35 pounds in a full 5-gallon bucket. Fishers and the sampler estimated

total weight and sampled weight of catch independently. Agreement between estimated weights

was good. Fish in the sample subset were individually identified to species, their total length was

measured on a measuring board, and it was noted whether they were dead or alive.

Catch of CPFV Fishery

Sport angler catch aboard CPFVs departing from San Simeon Cove, San Luis Obispo

County, was sampled to obtain species composition and size data from remote areas along the

Big Sur coast. Commercial passenger fishing vessels from San Simeon occasionally conducted

1- and 2-day long-range sport fishing trips that fished from Piedras Blancas north to Point Sur

and encompassed the vicinity of BCER (Figure 1). In 1997-98, a sampler was placed on these

trips, space and weather permitting, to observe and record fish caught, using sampling

methodology established in 1987 by the Department of Fish and Game’s (CDFG) Central

California Marine Sport Fish Survey (Reilly et al. 1998). This effort provided a cost-efficient

supplementary database of fish sizes and fishing effort of nearshore sport fishes in the vicinity of

BCER, an area that historically has received less sport and commercial fishing pressure than

areas proximal to Monterey and Morro Bay.
30

Samplers were trained in marine fish species identification and were equipped with foul

weather gear, gloves, clipboard, waterproof data sheets, fish-length measuring board, lead

pencils, and field guides to California marine fishes. The sampler contacted the landing to secure

passage on the vessel prior to the trip. CDFG vessel number, port code, departure time, type of

fishing trip (offshore, nearshore, surface, bottom, mix), number of paid and free anglers, and type

of fishing tackle used were recorded on a standard sampling form.

When the vessel arrived at the first fishing location, the sampler chose a subset of anglers to

observe throughout the trip and recorded this number (usually less than 15). The sampler

recorded bottom depth, the time when fishing lines were lowered, the number of observed and

total anglers, and either latitude and longitude, LORAN coordinates, or compass bearings and

coastline features. When the last observed fishing line was raised, signifying the end of a "drift",

time and depth were recorded and the process was repeated throughout the day.

Samplers observed anglers in the stern half of the vessel, where a larger sample size could

be obtained. An assumption in this sampling methodology, proven statistically in 1993

(Wilson-Vandenberg et al. 1996), was that catch, effort, and CPUE data from observed anglers in

the stern of the vessel were representative of all anglers on the vessel. Samplers identified to

species and counted each fish caught by all observed anglers. The ultimate fate of each observed

fish was recorded as either kept, released, used as bait, or unknown. If a fish was released, the

sampler attempted to determine if it survived or died. The combined catch by species for all

observed anglers was recorded on one data sheet; individual catch per angler was not recorded.

When fishing had ceased for the day, the sampler measured total length in mm of as many

observed kept fishes as possible by marking the length of each fish on a plastic measuring board,

keeping all species separated. Not all observed kept fishes were measured due to refusal of an

angler to have his/her catch examined, early filleting by the deck hand, or hazardous working

conditions caused by inclement weather. If time permitted, fishes kept by unobserved anglers

also were measured and their lengths were recorded separately from observed fishes' lengths.

Results

In Situ Density Estimates Derived From Random Transects

From 1995 to 1998, 27, 22, 15, and 76 random transects, respectively, were conducted in the

Big Creek study area for a total of 140 randomly selected transects. Areas were sampled

unequally within and among years. Data for counts of fish per transect north of BCER passed

tests of normality and equal variances; however, counts per transect data within and south of

BCER did not pass these tests; therefore, analysis was confined to examination of annual mean

counts per transect within each area separately for the four sampling years (1995-98). The nonnormal distribution of count data within and south of BCER is most likely due to anomalous high

counts in 1997 (Figure 2).
31

Figure 2. Counts of all species observed during random transects within, north, and south

of the Big Creek Ecological Reserve (BCER) by area, Monterey County, 1995-98.

Annual means of numbers of fish observed per transect north of BCER decreased 41%

(from 27.0 to 15.9 fish per transect) from 1995 to 1998 (Figure 2). Within BCER, except for the

high 1997 value of 26.7 fish per transect, annual mean counts were within 3.0 fish per transect

for the four-year period. From 1995 to 1998, mean numbers of fish per transect increased 19

percent from 16.0 to 19.1, respectively. South of BCER, annual mean counts of fish per transect

were within 2.1 fish, except for the high 1997 value. The percent difference over the 4-year

period was similar to BCER, with a 20 percent increase observed.

A two-sample t test was conducted separately for each area to determine significance of

differences between annual mean counts of fish per transect in 1995 compared to 1998. Counts

north of BCER in 1995 and 1998 were significantly different (p=0.002); however, counts within

and south of BCER were not statistically different (p=0.453 and p=0.669, respectively) between

1995 and 1998.

0

10

20

30

40

1995

1996

1997

1998

North BCER BCER South BCER
32

Commercial Skiff Landings

From August 1997 to December 1999, landings of 149 commercial skiff fishing trips were

sampled on the beach at Mill Creek State Park and BCER, Monterey County (Table 1). When

commercial fishers launched from and returned to Mill Creek, their fishing sites ranged from

Fuller’s (36° 12.70’N), which is 9.2 na. mi. north of the northern boundary of BCER to Gorda

Point (35°53.75’N), which is 13.6 na. mi. south of the southern boundary of BCER (Figure 1).

When they launched from and returned to BCER their fishing sites ranged from Fuller’s to

Limekiln (36°00.75’N), which is 2.5 na. mi. south of the southern boundary of BCER. Reported

depth of fishing ranged from 5-600 feet (1.5-181.8 m); however, the majority of fishing occurred

between 20-40 feet (6.0-12.1 m). Although rod-and-reel was occasionally used, most fishermen

used “stick gear” and carried 10-15 units in their skiffs. Each stick gear unit was comprised of a

five-foot long, weighted one inch PVC pipe to which five to eight circle hooks were attached.

Each unit had a line and buoy attached. Fishermen deployed each unit separately and retrieved

them within 1-3 hours.

The total catch of the 149 landings weighed 15,015 pounds. Fifty-three percent (8,017

pounds) of the total catch, representing 5,151 fish of 24 species, were sampled for species

composition and length. Common and scientific names of fishes are listed in Table 2. From

1997 through 1998, we sampled 14%, 54%, and 65% of the landings, respectively. For the threeyear sampling period the majority of the landings occurred from July through October. During

these months, sea conditions are more conducive to launching and landing a small skiff along the

Big Sur Coast.

From 1997 to 1998, blue rockfish declined from comprising 32% of the sampled landings to

2% and in 1999 they comprised less than 2% (Figure 3). Cabezon comprised 7% of the sampled

landings in 1997, dramatically increased to 42% in 1998 (6-fold increase), and then declined

slightly to 38% in 1999. Black-and-yellow rockfish showed a steady increase in percent

composition from 1997 (15%), to 1998 (20%), to 1999 (32%). Gopher rockfish comprised 16%

of the sampled landings in both 1997 and 1998 and declined to 9% in 1999. Grass rockfish

increased in percent composition of sampled landing from 1997 (5%) to 1998 (13%) and then

declined in 1999 (11%).
33

Table 1. Monthly Big Sur commercial nearshore skiff sampling effort and landings, 1997-99.

_____________________________________________________________________________

# Boats # Fish Total Total

Year Month Sampled Measured Landed (Lbs) Sampled (Lbs)

1997 8 6 150 0 .

9 8 194 325 .

10 15 304 1635 285

Totals 29 648 1960 285

1998 6 14 474 1380 165

7 13 639 1850 900

8 15 748 1975 1520

9 10 470 1565 1045

10 3 188 360 250

Totals 55 2519 7130 3880

1999 4 11 395 905 810

5 6 168 455 420

7 18 565 1565 1170

8 8 309 815 575

9 10 172 1235 395

10 5 69 365 175

11 6 132 375 217

12 1 35 210 90

Totals 65 1845 5925 3852

Grand Totals 149 5,012 15,015 8,017

_____________________________________________________________________________
34

Table 2. Common and scientific names of species observed in landings of Big Sur commercial

nearshore skiff or commercial passenger fishing vessel fisheries, 1997-99.

Common Name Scientific Name

Bank rockfish Sebastes rufus

Black rockfish Sebastes melanops

Black-and-yellow rockfish Sebastes chrysomelas

Blue rockfish Sebastes mystinus

Bocaccio Sebastes paucispinis

Brown rockfish Sebastes auriculatus

Canary rockfish Sebastes pinniger

Chilipepper Sebastes goodei

China rockfish Sebastes nebulosus

Copper rockfish Sebastes caurinus

Cowcod Sebastes levis

Flag rockfish Sebastes rubrivinctus

Gopher rockfish Sebastes carnatus

Grass rockfish Sebastes rastrelliger

Greenspotted rockfish Sebastes chlorostictus

Greenstriped rockfish Sebastes elongatus

Kelp rockfish Sebastes atrovirens

Olive rockfish Sebastes serranoides

Quillback rockfish Sebastes maliger

Rosy rockfish Sebastes rosaceus

Speckled rockfish Sebastes ovalis

Squarespot rockfish Sebastes hopkinsi

Starry rockfish Sebastes constellatus

Treefish Sebastes serriceps

Vermilion rockfish Sebastes miniatus

Widow rockfish Sebastes entomelas

Yelloweye rockfish Sebastes ruberrimus

Yellowtail rockfish Sebastes flavidus

Cabezon Scorpaenichthys marmoratus

California halibut Paralichthys californicus

California sheephead Semicossyphus pulcher

Kelp greenling Hexagrammos decagrammus

King salmon Oncorhynchus tshawytscha

Lingcod Ophiodon elongatus

Ocean whitefish Caulolatilus princeps

Pacific bonito Sarda chiliensis

Pacific mackeral Scomber japonicus

Pacific sanddab Citharichthys sordidus

Rock greenling Hexagrammos superciliosus

Rock sole Lepidopsetta bilineata

Sand sole Psettichthys melanostictus
35

Figure 3. Species composition of landings of Big Sur commercial skiff fishery, 1997-

99.

1997

Species # Observed Percent

 Black 10 2

 Black-and-yellow 95 15

 Blue 209 31

Bocaccio 13 2

Cabezon 45 7

Chilipepper 10 2

 Gopher 109 16

 Grass 33 5

 Kelp 12 2

 Olive 14 2

* Other 23 4

 Vermilion 52 8

 Widow 23 4

Total Observed = 648

* Other = species less than 2% of total catch.

 Blue

31%

 Gopher

16%

Black-and-

 yellow

15% Vermilion

8%

Cabezon

7%

 Grass

5%

 Widow

4%

 Olive

2%

 Black

2% Other

4%

Chilipepper

2%

 Kelp

2%

Bocaccio

2%

Cabezon

42%

Black-and-

 yellow

20%

 Gopher

16%

Kelp greenling

2%

 Olive

2%

 Grass

13%

Other

3%

 Blue

2%

1998

Species # Observed Percent

 Black-and-yellow 541 20

 Blue 45 2

Cabezon 1099 42

 Gopher 430 16

 Grass 343 13

Kelp greenling 44 2

 Olive 65 2

* Other 91 3

Total Observed = 2658

* Other = species less than 2% of total catch.

1999

Species # Observed Percent

 Black-and-yellow 589 32

Cabezon 688 38

 Gopher 173 9

 Grass 196 11

Lingcod 46 2

* Other 95 5

 Vermilion 58 3

Total Observed = 1845

* Other = species less than 2% of total catch

Cabezon

38%

Black-and-

 yellow

32%

 Grass

11%

 Gopher

9%

Other

5%

Lingcod

2%

 Vermilion

3%
36

Commercial Passenger Fishing Vessel Catch

From July 1997 to July 1998, twelve 1-day and three 2-day CPFV trips (total of 18 days)

were observed. All trips departed from San Simeon, San Luis Obispo County and headed north

to as far as Hurricane Point (Figure 1). During four days, fishing occurred in the Point Sur area,

which extended from Ventura Rocks (2 na. mi. north of Point Sur, Latitude-36° 20.24’N) to

Pfeiffer Point (5 na. mi. south of Point Sur, latitude-36° 12.31’N). During the remaining 12 days,

fishing took place in the Cape San Martin area, which extended from Gamboa Point (3 na. mi.

north of Lopez Point, latitude-36° 02.60’N) to Piedras Blancas (latitude-35° 41.14’N). Fishing

occurred in water depths ranging from 6 fathoms (10.9 m) to 90 fathoms (163.6 m). From

collected data we could not discern a significant difference between depths fished in the two

areas. The average number of drifts that occurred for one day of fishing was 12 and the range

was from 7 to 26 (Table 3). Total fishing time per day averaged approximately 4 hours and the

average number of fish taken per day was 223. The sampler observed on average 10 anglers per

day and each angler caught an average of 5.7 fish per angler-hour. We calculated daily CPUE for

the 18 days of observed fishing. Fourteen days were spent in the Cape San Martin area and four

days in the Point Sur area. Due to the insufficient sample size we were unable to analyze this

data with a student t-test; therefore, we use a non-parametric equivalent Kruskal-Wallis, using

alpha-0.05 and found no significant difference (p=0.089) in CPUE between the two areas.

A total of 4020 fish were observed taken during the 18 days of sampling. Three thousand

and forty-eight fish, representing 34 species, were observed taken during 12 days of fishing in

the Cape San Martin region and 972 fish, representing 21 species, were observed taken during 4

days of fishing in the Point Sur region. Blue rockfish ranked number one in species composition

and comprised 26% of the total catch of fishes taken during fishing in the Cape San Martin

region; whereas, blue rockfish ranked third and comprised 11% of the catch in the Point Sur

region (Figure 4). Yellowtail rockfish ranked second and comprised 12% of the catch at Cape

San Martin, but were in the first rank and comprised 20% of the catch from Point Sur. Olive

rockfish ranked third and comprised 11% of the Cape San Martin catch; however, at Point Sur

they shared the sixth rank with widow rockfish and only comprised 6% of the catch. Bocaccio

ranked sixth and comprised 7% of the catch in the Cape San Martin region; whereas, they were

the second most abundant species taken at Point Sur and comprised 16% of the catch.

Fishery length data are usually assumed not to be normally distributed. We tested the total

length of fish for each species taken in the Point Sur and Cape San Martin areas for assumptions

of equal variances and normal distribution and, as expected, these test failed for the majority of

species. We then performed a non-parametric equivalent (Kruskal-Wallis) for each species using

alpha=0.05. There was a highly significant difference between Point Sur and Cape San Martin in

the total lengths for 8 of the 11 species tested (Figure 5 and Table 4). Mean total lengths of blue,

canary, copper, olive, vermilion, widow, and yellowtail rockfishes and lingcod were larger in the

Point Sur region compared to the Cape San Martin region.
37

Table 3. Monthly fishing summary of catches from commercial passenger fishing vessels

fishing in the Cape San Martin and Point Sur regions of the Central California coast,

1997-98.

Fishing Total # Fish Total Avg Number

Year Date Region Trip Drifts Observed Minutes Anglers Mean Variance

1997 July 25 Cape San Martin 1 7 102 171 5 7.54 3.56

August 1 Cape San Martin 2 11 153 186 6 8.52 16.30

August 8 Cape San Martin 3 18 160 201 8 6.57 12.55

August 15 Cape San Martin 4 10 229 204 12 5.96 23.76

August 22 Cape San Martin 5 9 207 175 12 5.53 6.66

August 29 Cape San Martin 6 9 129 266 6 4.64 3.50

September 12 Cape San Martin 7 12 618 281 12 10.84 60.01

September 19 Cape San Martin 8 8 175 203 7 7.92 7.16

October 24 Cape San Martin 9 11 267 250 12 6.19 8.92

December 27 Point Sur 10 A 19 266 315 12 4.09 11.60

December 28 Cape San Martin 10 B 14 250 199 12 6.13 33.57

1998 January 24 Point Sur 11 A 26 323 381 11 4.48 19.26

January 25 Cape San Martin 11 B 8 148 153 12 3.42 7.73

May 9 Point Sur 12 A 19 241 324 8 5.05 7.78

May 10 Point Sur 12 B 10 142 217 8 4.95 27.04

May31 Cape San Martin 13 14 172 289 12 2.71 3.87

June 28 Cape San Martin 14 8 147 235 12 3.07 2.25

July 26 Cape San Martin 15 11 291 305 10 4.86 12.10

18 224 4020 72 hr 35 min 176 - -

- 12.4 223.3 4 hr 2min 10 5.69 -

Sum Total

Total Average per Day

Catch per Angler Hour
38

Cape San Martin

Species # Observed Percent

Blue 759 26

Bocaccio 201 7

Canary 58 2

Gopher 127 4

Lingcod 181 6

Olive 332 11

* Other 288 10

Rosy 57 2

Starry 136 5

Vermilion 141 5

Widow 296 10

Yellowtail 338 12

Grand Total = 2914 

 

* Other = Species less than 2% of the total catch.

 

Bocaccio

16%

Yellowtail

20%

Yelloweye

3%

Vermilion

6%

Olive

6%

Widow

6%

Copper

8%

Lingcod

11%

Greenspotted

2%

Starry

2%

Other

3% Canary

6%

Blue

11% 

 

Point Sur 

Species # Observed Percent

Blue 113 11

Bocaccio 164 16

Canary 58 6

Copper 79 8

Greenspotted 19 2

Lingcod 108 11

Olive 66 6

* Other 37 3

Starry 21 2

Vermilion 63 6

Widow 66 6

Yelloweye 26 3

Yellowtail 200 20

Grand Total = 1020 

* Other = Species less than 2% of the total catch.

Widow

10%

Other

10%

Bocaccio

7%

Lingcod

6%

Starry

5%

Vermilion

5%

Rosy

2%

Olive

11%

Yellowtail

12%

Blue

26% Gopher

4%

Canary

2%

 

Figure 4. Species composition of fishes sampled aboard commercial passenger fishing

vessels fishing in the Cape San Martin and Point Sur regions of the central

California coast, 1997-98.
39

Total Length (mm)

200 400 600 800 1000

Cape San Martin

Point Sur

n = 759

n = 113

n = 201

n = 164

n = 58

n = 58

n = 48

n = 79

n = 127

n = 2

n = 43

n = 19

Blue

rf.

Gopher

rf.

Bocaccio

Canary

rf.

Greenspotted

rf.

Copper

rf.

Total Length (mm) 200 4 6 800 10 Cape San Martin Point Sur Olive rf. Yellowtail Vermilion Widow Lingcod Yelloweye n = 332 n = 66141 n = 63296 n = 226 n = 338 n = 200 n = 181 n = 108

Figure 5. Total length of 12 species sampled aboard commercial passenger fishing

vessels fishing in the Cape San Martin and Point Sur regions of the Central California

coast, 1997-98. Dotted line represents the mean, solid line represents the median,

boxed area is the interquartile range (50% of values), and the upper and lower

“whiskers” represent 90th and 10th percentile points, respectively.
40

Table 4. Results of Kruskal-Wallis non-parametric test of species’ total lengths between

sampled catch aboard commercial passenger fishing vessels fishing in the Cape San Martin

and Point Sur regions of the Central California coast, 1997-98.

 Kruskal-Wallis Significant Difference

Species p Value Between Areas

Blue rockfish <0.001 Yes

Bocaccio 0.561 No

Canary rockfish <0.001 Yes

Copper rockfish 0.001 Yes

Greenspotted rockfish 0.082 No

Olive rockfish 0.033 Yes

Vermilion rockfish <0.001 Yes

Widow rockfish <0.001 Yes

Yelloweye rockfish 0.975 No

Yellowtail rockfish <0.001 Yes

Lingcod <0.001 Yes
41

Discussion

The decline in the success and quality of Central California’s nearshore rockfish fisheries

over the past decades and the rapid expansion of the live/premium fish fishery, which targets

shallow water kelp forest fishes, have underscored the need for revised management strategy for

this resource. Marine protected areas have been suggested as an alternative or additional

management tool to create sustainable fisheries. There is little doubt that eliminating fishing in a

designated area will help maintain the area’s natural biological diversity and provide a “heritage

area” for aesthetic and scientific purposes. However, before managers can advocate marine

reserves as an effective management tool, they must first be able to demonstrate that the biomass

and reproductive potential of target species within reserves is significantly greater than in fished

areas.

The overarching goal of the study is to provide a benchmark or “line in the sand” of data

collected in a newly established marine protected area. Our hope is that we have done an

adequate job so that future studies can statistically document temporal changes in nearshore fish

populations within and adjacent to BCER. Because both fishery independent and dependent data

are crucial in the evaluation of the reserve as a fishery management tool, we collected data from

in situ surveys of fish populations within BCER and monitored the landings of commercial and

recreational fisheries targeting fishes in the vicinity of BCER.

We also examined numbers of all fish observed during random transects conducted within

and adjacent to BCER. Randomly collected data is more representative of the surveyed areas

than data collected at permanent stations; however, it is more variable due to differences in

habitat. Significant differences among years and areas were noted for randomly collected data.

We suspect high fish density estimates in 1993 and 1997 were related to the El Niño conditions

during these years. The last quarter of 1993 was the later portion of the 1992-93 major El Niño

events along the Central California coast (Hayward et al. 1994 and Lynn et al. 1995) and 1997

was a minor warm water event (Norton et al. 1999). El Niño conditions have been documented to

reduce primary productivity, affect species distribution and abundance, promote recruitment

failure for some species of rockfish, and cause poor growth and condition of adult rockfish

(Lenarz et al. 1995 and VenTresca et al. 1995). During these warm water periods we observed

“more fish out in the open looking for food” compared to non El Niño years. This variability

among years underscores the necessity incorporating environmental parameters with in situ

estimates of fish densities for spatial and temporal comparisons. Excluding 1997 from the 1995-

98 within area annual comparisons shows little change within and south of BCER and a

significant decline north of BCER from 1995-1998. We suspect that the decline of numbers of

fish per transect observed north of BCER from 1995 to 1998 is due to increased fishing pressure

from the commercial live/premium fish fishery in this area. The area south of BCER has been

fished for many years by the commercial small skiff fishery that used launch sites located south

of BCER.

 Landings of the Big Sur commercial skiff fishery reveal a concentration on species

commonly observed in kelp forest. A high percentage of 1998-99 landings were composed of

cabezon and black-and-yellow and gopher rockfishes. It will be extremely interesting to see if

these species maintain a high percentage in future landings and if the densities of these species
42

changes in future in situ surveys. The majority of the CPFV fishery occurred in areas not

adjacent to BCER. This study documented that 4.5 times more trips, which departed from San

Simeon, San Luis Obispo County, were made to the Cape San Martin region as were made to the

Point Sur region. Personnel communications with the CPFV industry confirmed that the Cape

San Martin region receives more fishing pressure than does the Point Sur region. Most species

taken in the Point Sur region were larger than those taken in the Cape San Martin region;

however, there was not a significant difference in the CPUE between the two areas.

The experimental design of this survey, like many previous studies of California's central

coast nearshore reef fishes, stratified random sampling of fish densities using the canopy of the

giant kelp, Macrocystis pyrifera, as a proxy for reefs/rocky bottom strata. The increased

availability of precise high-resolution maps of the seafloor and habitat, generated from acoustic

remote-sensing methods and enhanced navigation tools (dGPS) makes it possible to accurately

locate and randomly sample specific habitats. In the future, in situ surveys samples to estimate

fish densities can be allocated in specific strata to minimize variability. We recognize that an a

priori stratification of sampling effort may be a more efficient and statistically powerful

approach (Cuff and Coleman 1979, McCormick and Choat 1987, Jolly and Hampton 1990,

Smith and Gavaris 1993, and VenTresca et al. 2002).
43

CHAPTER 3

Deepwater Habitat And Fish Resources Associated With

The Big Creek Ecological Reserve

Mary Yoklavich1, Gregor Cailliet2, Robert N. Lea3, H. Gary Greene2,

Richard M. Starr4, Jean De Marignac2, And Jeff Field2

1 NOAA, NMFS, SWFSC, Santa Cruz Laboratory, 110 Shaffer Road, Santa Cruz, CA 95060

2 Moss Landing Marine Laboratories, 8272 Moss Landing Road, Moss Landing, CA 95039

3

 California Dept. Fish and Game, 20 Lower Ragsdale Drive, Monterey, CA 93940

4 UC Sea Grant Extension Program, 8272 Moss Landing Road, Moss Landing, CA 95039

Abstract

Big Creek Ecological Research Reserve (BCER), located off Central California, has been

closed to fishing since January 1994. We used side scan sonar and an occupied submersible to

collect baseline information on species-habitat relationships, density, and species and size

composition of fishes inside and outside BCER. Forty-three dives were made in Fall of 1997 and

1998, in depths from 20-250 m. From 142 video transects, we identified over 70,000 fishes of 82

taxa, including 36 species of rockfishes. About 93% of the 25,159 fishes inside BCER were

rockfishes comprising at least 20 species. Young-of-the-year rockfishes dominated rock outcrops

in 20-90 m depth inside and outside BCER. Four distinct fish assemblages were associated with:

1) fine smooth sediment in deep water; 2) bedrock with uneven surface in deep water; 3) sand

waves and shell hash in shallow water; and 4) boulders and organic habitats on rock in shallow

water. There were no significant differences in fish density among locations (inside and outside

BCER), depths, or between years. Density was significantly higher in high relief rock habitat

than in low relief soft and mixed sediments, regardless of location. There were no consistent

patterns of larger fishes inside compared to outside the protected area. We recommend

development of a monitoring program to continue these surveys after increased time of

protection and with increased assessment effort in the appropriate habitats of economically

valuable species. In addition, extending the boundaries of BCER seaward would protect habitats

and fishes in water depths >100 m.

Detailed information about this project can be found in the following publications:

Yoklavich, M., G. Cailliet, R.N. Lea, H.G. Greene, R. Starr, J. De Marignac, and J. Field. 2002.

Deepwater habitat and fish resources associated with the Big Creek Ecological Reserve.

California Cooperative Fisheries Investigation Report. CalCOFI MS2002-03.

Yoklavich, M., G. Cailliet, R.N. Lea, H.G. Greene, R. Starr, J. De Marignac, and J. Field. 2001.

Deepwater habitat and fish resources associated with a Marine Reserve: Implications for

Fisheries Management. Part 1. California Sea Grant College Program. Marine Ecological

Reserves Research Program Research Results R/BC 1.
44

CHAPTER 4

Anacapa Island Ecological Reserve: A Research Reveiw

Jennifer Caselle

University of California Santa Barbara, Marine Science Institute

Santa Barbara, CA 93106

Abstract

The Natural Area of Anacapa Island Ecological Reserve was established by the Fish and

Game Commission in 1978 and consists of 12 hectares along the north shore of East Anacapa

Island. This area is closed to all fishing. Several biological monitoring programs operating in

the Northern Channel islands (Channel Islands National Park, Partnership for Interdisciplinary

Studies of Coastal Oceans (PISCO) and Channel Islands Research Program (CIRP)) include the

Anacapa reserve. In this chapter I review existing studies for the Anacapa reserve which may

shed light on the following questions: 1) Are target species larger or more abundant in the

Anacapa reserve?, 2) Does the reserve serve as a control to evaluate changes in nearby fished

areas?, 3) Does the Anacapa reserve promote ecosystem function?, and 4) Is the reserve large

enough to protect target species?
45

Anacapa Island Ecological Reserve

Background to the Natural Area of the Anacapa Island Ecological Reserve.

Anacapa Island is in the Northern Channel islands and consists of three islands separated by

small passages. The Natural Area of Anacapa Island Ecological Reserve was established by the

Fish and Game Commission and consists of 12 hectares along the north shore of East Anacapa

Island. This area was closed to all fishing in 1978 and is one of the smallest no-take areas in

California (Mc Ardle 1997). In addition, a small area on the north side of West Anacapa island

is designated as a brown pelican fledging area. No entry is allowed in this area from January 1

through October 31. The brown pelican fledging area covers approximately 4000 ft of shoreline

and extends out to a water depth of approximately 120 ft.

Existing monitoring programs that include Natural Area of Anacapa Island Ecological Reserve

The Channel Islands National Park has maintained a monitoring program for kelp forests

throughout the Northern Channel islands (Santa Barbara, Anacapa, Santa Cruz, Santa Rosa and

San Miguel Islands) since 1982 (Davis et al. 1997). The kelp forest monitoring program (KFM)

samples a variety of organisms annually at 16 sites spread throughout these islands. Two of

these KFM sites are located in the reserve at Anacapa. The protocol includes several techniques

for monitoring a variety of organisms including macro algae (kelps, understory algae), macro

invertebrates (urchins, lobsters, starfish, etc.), recruitment of selected invertebrates (urchins,

abalone, etc.) and fishes (Davis et al. 1997).

PISCO (Partnership for Interdisciplinary Studies of Coastal Oceans) established rocky reef

monitoring sites at Anacapa and Santa Cruz islands in 1999. PISCO also focuses on kelp forest

habitats and uses complementary but not identical protocols to the KFM. Several sites are

monitored by both KFM and PISCO including 2 sites in the Anacapa reserve. Important

differences in the protocols include the following: PISCO monitors on randomly placed transects

within each site, KFM uses a fixed transect, PISCO species lists are more detailed than KFM and

PISCO monitors physical and chemical oceanography with fixed moorings at the sites as well as

monthly settlement dynamics for invertebrates and fishes.

The Channel Islands Research program (CIRP) has been conducting marine biological

studies around the eight islands in the Southern California Bight each year since 1980, with

preliminary cruises dating back to 1978. Cruises include: water temperature studies, especially

during El Nino - La Nina, mantis shrimp studies, echinoderm and fish surveys at Santa Catalina

and San Clemente Islands; kelp/urchin community studies at Anacapa Island; eelgrass ecosystem

surveys, echinoderm aggregations and disease around the northern islands; intertidal monitoring

projects in cooperation with various agencies seeking information on marine issues of concern in

California. CIRP does not specifically target the marine reserve at Anacapa for long-term

monitoring.
46

Review of research at the Anacapa Ecological reserve

In general, very few studies have been conducted or published that documented research on

Anacapa Island. There are even fewer studies that focused on the Anacapa reserve to test

specific goals or functions of marine reserves. This is likely due to its extremely small size

and/or the lack of other reserve areas nearby for comparison. However, there are a few studies

(both published and unpublished) that have given some insight into several aspects of the

functioning of the Anacapa reserve. For the rest of this document, the term Anacapa reserve

refers to the Natural Area of the Anacapa Island Ecological Reserve unless other specified.

Are target species larger or more abundant in the Anacapa reserve?

Increases in the size and the abundance of organisms inside reserves relative to outside of

reserves have been well documented in the literature (Halpern in press). More and larger

individuals will have an important effect on potential production in a reserve. The effects of

high density are relatively straightforward. Barring any density-dependent effects, more fish

produce more eggs. However, this effect is greatly magnified when the fish are larger. For fish,

the relationship between body length and fecundity (number of eggs) is not linear. A small

increase in body size results in a disproportionately large increase in egg production. If a reserve

is large enough to contain the movements of adults within the boundaries yet allow larval

dispersal out, then the reserve is likely to export young to outside areas and increase fisheries

yields outside the reserve. The presence of large individuals also has simple conservation

benefits. People enjoy observing large fish and other marine organisms. The presence of large

individuals, especially in areas where they no longer exist (e.g., Southern California) could also

benefit tourism and provide additional recreational opportunities.

Tests of this effect are less common in Southern California reserves than other parts of the

world. In one study, the abundance and size structure of fish populations were compared inside

and outside of 5 no-take marine reserves in Southern California (Tretault, in prep). Two of these

five sites were the Anacapa reserve and the Pelican closure at Anacapa where fishing is

prohibited during the pelican breeding season (January through October). For two harvested fish

species, kelp bass (Paralabrax clathratus) and California Sheephead (Semicossyphus pulcher),

both the density of all sizes combined and of only harvestable sized fish were higher inside of

both protected areas relative to outside. The average density of harvestable sized kelp bass was

17 times higher inside the protected areas than outside. Harvestable California sheephead were

on average 11 times more abundant inside than outside. She also measured the density of several

common but non-harvested species (Garibaldi (Hypsypops rubicundus), Seniorita (Oxyjulis

californica) and Rock wrasse (Halichoeres semicintus)). In contrast to the results for the

harvested species, non-harvested species did not show significant increases in density or size

according to protection status. In fact, seniorita were significantly more abundant outside of the

pelican closure than inside. Importantly, habitat characteristics of the reserve and control areas

were not significantly different.

Data on red sea urchin abundance collected by the National Park Service showed that there

has been a decline in the abundance of large individuals, and suggests that fishing may be the

cause. Since 1985, the abundance of harvestable sized red urchins (>82.5mm test diameter) has
47

declined more than 5% per year at fished sites on Santa Rosa and San Miguel islands (the sites

contributing most to the overall catch) relative to non-fished reserve sites on Anacapa (S.

Schroeter and D. Reed, unpublished analysis of KFM data). Similar declines were not observed

in the abundance of red sea urchin recruits (<25mm test size).

Does the reserve serve as a control to evaluate changes in nearby fished areas?

Although the use of marine reserves for fisheries enhancement is widely discussed (and

debated), their use for testing stock assessment methods or traditional management strategies has

received less attention (NRC 2001). No-take areas may be critical for providing baseline data

with which to test and improve fisheries management. By strictly controlling the effects of

human impacts, marine protected areas can be used to focus more intense local management

designed to increase yield and allow research to help define sustainability and protect against

uncertainty (Dayton et al. 2000). One study has used the reserve at Anacapa to directly compare

between two methods of stock assessment for an emerging fishery.

Schroeter et al. (2001) used fishery independent data (long-term abundance from the NPS

KFM program) to evaluate the status of the newly emerging dive fishery on warty sea cucumber

(Parastichopus parvimensis) in Southern California. They employed a BACI design, using

abundance data from inside a marine reserve (Anacapa) and outside (the other KFM sites) both

before and after the onset of fishing. They then compared the results of this stock assessment

using fishery-independent data with a stock assessment based on CPUE (catch per unit effortfishery dependent data).

Monitoring data showed that the abundance of warty sea cucumber decreased throughout

the islands at the fished sites following the onset of the fishery. At two unfished sites in the

reserve at Anacapa, the abundance of sea cucumbers showed no significant change but tended to

increase. All seven fished sites used in the BACI analysis (those sites that met the assumptions

of BACI) showed significant declines relative to the unfished sites following the onset of fishing.

These declines ranged from 33% to 83%. Stock assessment based on CPUE data differed

dramatically from that based on the monitoring data. CPUE did not decline at any of the 6

islands during the 3-6 year period after the onset of fishing despite the general decline in

abundance at the sites.

The combination of a no-take marine reserve with substantial and long-term monitoring data

inside and out made it possible for these authors to a) test the coherence of stock assessments

based on fishery dependent data versus fishery independent data and b) disentangle the

potentially confounding effects of natural changes in populations from those due to fishing

mortality.

The authors point out that no-take areas can provide beneficial information even for wellestablished fisheries. By comparing the trends in populations before and after establishment of a

marine reserve we can begin to understand the resiliency of stocks to recover in the absence of

fishing mortality and the extent to which a fishery has caused impacts to stock. This knowledge

is impossible to get by reducing fishing pressure throughout the spatial range of a fishery, but

will require spatial set asides large enough for populations to recover.
48

Does the Anacapa reserve promote ecosystem function?

One goal of a marine reserve is to promote healthy ecosystems by protecting biodiversity

and allowing all members of the community to be present and function ecologically. Marine

systems have been changed drastically by fishing and other anthropogenic influences. In many

systems the higher-order predators and other important species have been virtually eliminated.

In many others, the benthic habitat itself has been disrupted. Understanding the changes in

marine ecosystems is especially difficult due to logistic constraints, the lack of long-term data

and the potentially lasting consequences of rare or episodic events (Dayton et al. 2000). Yet,

without natural systems or benchmarks to compare the observed changes, the difficulties are

magnified. Without pristine systems important questions cannot be studied--for example, how

the ecosystem roles of various species can be assessed, how they can be managed in a

sustainable manner, and how we can evaluate resilience or relative rates of recovery. (Dayton et

al. 2000).

Here I review two studies that have investigated the community or ecosystem dynamics in

and around the Anacapa reserve. The Anacapa reserve contains, among other habitats, temperate

kelp forests. Although kelp bed dynamics are complex, the interactions between macroalgae,

grazers such as urchins, and predators on those grazers are widely recognized as controllers in

these systems (Tegner and Dayton 2000). Both studies reviewed here investigate those

dynamics.

A community analysis using NPS KFM data from the marine reserve at Anacapa concluded

that the ecosystem within the reserve has a more “natural balance” than ecosystems exposed to

fishing pressure (Sladek Nowlis in press). Two of the major urchin predators in Southern

California kelp beds are the spiny lobster and the California sheephead (Tegner and Levin 1983).

Both of these predators were more abundant in the marine reserve at Anacapa than other KFM

sites outside the reserve. Spiny lobster densities inside were 10 times what they were outside,

while California sheephead were 1 1/2 times more abundant inside than outside. There are three

main urchin species in the Channel islands; red sea urchins that are harvested by humans and

white and purple urchins, which are not harvested commercially. All three are algal grazers.

This analysis showed that the non-targeted purple and white urchins have grown dramatically in

abundance in fished areas, probably as a result of lack of competitors or predators. The density

of the non-harvested, white sea urchins has increased approximately 4 times in the reserve

compared to 15 times at sites outside the reserve since 1983. As a consequence, the large canopy

forming giant kelp, Macrocystis, which provides the majority of the structure in a healthy kelp

forest, has all but disappeared in sites outside of the reserve, while increasing in the reserve sites.

Sladek Nowlis (in press) concludes that by protecting even small areas from the effects of

fishing, ecosystems are able to achieve a level of health and natural balance not possible in areas

that continue to be fished.

In another analysis of the KFM data, Lafferty and Kushner (2000) investigated the roles of

predation and disease in the population regulation of purple sea urchins throughout the Channel

Islands, including the Anacapa reserve. They found that urchin abundance was not explained by

recruitment patterns. There was an inverse relationship between urchin density and invertebrate
49

urchin predator abundance (Spiny lobsters and sunflower stars) but not for fish predators. They

also investigated the prevalence of disease in the urchin populations and found that outbreaks of

bacterial disease tended to occur in high-density populations but disease was not related to water

temperature. In the western sites where lobsters are rare and the major invertebrate predator is

the sunflower star, urchin densities tend to be low and disease was rare. In the eastern sites

where the major predator is the spiny lobster, which are fished heavily, urchin abundance is high

and urchin disease is prevalent. The exception to the pattern in the eastern sites was the marine

reserve at Anacapa. Presumably due to protection, spiny lobsters are more abundant in the

reserve, urchins were less abundant, and disease was rare.

Is the reserve large enough to protect target species?

Among many potential fisheries benefits of protected areas, two relate to flux of individuals

between fished and unfished areas. One possible but largely unproved benefit is that reserves

protect a critical minimum spawning stock and maintain or enhance fishery yields at a large scale

by larval export out of reserves (“recruitment” effect). Although there is mounting evidence that

both population sizes and individual sizes can be larger inside of reserves (Halpern in press),

evidence for successful recruitment of larvae produced by these protected fish is lacking. A

second potential benefit is the maintenance or enhancement of fishery yields in areas adjacent to

a reserve by movement of adults or juveniles across the boundaries after a period of initial

growth in the reserve (the “spillover” effect). Transfer rates across reserve boundaries depend

not only on the permeability of those boundaries but also on the size of the reserve relative to the

movements and home range sizes of the target species. 

Surprisingly little is known about normal fish movements and home range sizes, and

evidence for a spillover effect is minimal. Since acceptance of reserves (and hence effectiveness

in terms of enforcement) often depends on strong community support, locally increased catches

(spillover) may be more convincing to anglers than regional recruitment benefits. In either case,

the ability of a reserve to provide protection depends on consistent use of the protected area by

individual fish. Thus, knowledge of the movement and activity patterns of target species is

critical.

The author (Caselle) is currently conducting a study of fish movements at the Anacapa

reserve. Using commercial live-fish trap gear, we capture and tag (with external colored tags)

fish throughout the reserve. We divided the reserve into 3 movement blocks (blocks 1-3). Each

has a linear distance along the shoreline of roughly 750 m. Within each of the 3 blocks in the

reserve, we fish as similarly as possible to how commercial fishermen would fish normally. That

is, each trapset location is not predetermined, nor for the most part is the effort we put into each

block. We use a different color fish tag for each block. We assess movement by recapturing the

fish (which gives exact distance between captures) and by visually resighting using SCUBA

surveys (which gives estimates of movement from block to block). The majority of the trapped,

retrapped, and resighted individuals are California sheephead.

The average distance between trap recaptures for California sheephead was 115 m. The

minimum distance between recaptures was 18m and the maximum was 860 m. There was no

relationship between distance traveled and either body size or time at large. We resighted 417
50

California sheephead to date. Most fish (80-100%) were resighted in the block in which they

were tagged. The small percentage that moved from one block to another, usually moved to an

adjacent block. This shows that movements on the scales of 500-750 m alongshore are not

common.

These movement estimates are in line with the home range size of California sheephead

estimated from ultrasonic telemetry on Catalina island (Christopher Lowe, unpublished data). In

that study, home range areas (not linear distances) ranged from 1000’s of square meters to

approximately 70,000 m2.
51

CHAPTER 5

Marine Reserves in the Monterey Bay National Marine

Sanctuary: What We Do and Don’t Know

Mark H. Carr and Craig Syms

Department of Ecology and Evolutionary Biology

University of California Santa Cruz, Institute of Marine Sciences

100 Shaffer Road, Santa Cruz, California 95060

Abstract

Rapidly growing global consideration for the development of coastal networks of

marine reserves has focused attention on their potential application in the coastal

waters of California. This focus stems in part from recent State and federal legislation

necessitating development and implementation of some form of marine reserve

network throughout California’s coastal waters, and from the potential for past and

present research in three existing marine reserves (Big Creek Ecological Reserve,

Point Lobos Ecological Reserve, and Hopkins Marine Life Refuge) to inform

managers on the potential value of marine reserves as tools for conservation and

fisheries management. Simultaneous with this increased interest has been a rapid

increase in theoretical and analytical studies on the design and evaluation of marine

reserves, as well as empirical accounts of reserve effects throughout the world. Few

studies of California’s marine reserves exist. All of these studies have focused solely

on the response of selected fish species to protection within reserves. From the few

studies conducted in the three reserves encompassed within the Monterey Bay

National Marine Sanctuary (MBNMS), few general conclusions can be drawn. To

date, there is some evidence that fishes inside two of three reserves within the

MBNMS appear to have greater potential larval production than nearby populations

outside of reserves inhabiting comparable areas of reef habitat. Studies in the third

reserve, Big Creek Ecological Reserve, were carried out only within the first four

years subsequent to establishment. Given the episodic nature of fish recruitment, it is

not surprising that marked differences were not observed. The magnitude of the

reserve effect on increased potential of larval production appears to be positively

related to the age of the reserve. To date, little information exists on community- or

ecosystem-wide responses to protection by reserves and only one study has examined

the state of fish populations for more than two years. Moreover, the lack of sampling

over several years necessary to test for trajectories in populations inside and outside

reserves makes attributing causal inferences to reserve effects equivocal. To develop a

better understanding for the potential of reserves as tools for conservation and fisheries

management, it is clear that ecosystem-wide responses have not been examined and

remain unknown, the temporal persistence of differences between reserve and nonreserve populations is unknown, and unequivocal ascription of differences between

reserve and non-reserve populations to causal reserve effects is tenuous. Better

understanding of the consequences of reserve establishment requires both creation of

new reserves in concert with well-designed monitoring studies over many years.
52

I. Why the current interest in marine reserves within the MBNMS?

Marine reserves are designated areas within which human activities that can result in the

removal or alteration of biotic and abiotic components of an ecosystem are prohibited or greatly

restricted (NRC 2001). This definition corresponds with what is typically referred to in the

literature as “no-take” marine protected areas (MPAs) or marine reserves (MRs). Activities

specifically curtailed within a marine reserve are extraction of organisms (e.g., commercial and

recreational fishing, kelp harvesting, commercial collecting), mariculture, and those activities

that can alter oceanographic or geologic attributes of the habitat (e.g., mining, shore-based

industrial-related intake and discharges of seawater and effluent, respectively). Local examples

of marine reserves and such restrictions include Hopkins Marine Life Refuge, Point Lobos State

Ecological Reserve and Big Creek Ecological Reserve (McArdle 1997, Brown 2000).

Current and urgent interest in the proposed development of a network of marine reserves

within the MBNMS stems from three contemporaneous events. At the global scale, national

governments and conservation oriented non-governmental organizations (NGO’s) are promoting

networks of marine reserves as an approach for conservation akin to more familiar terrestrial

reserves. This impetus is motivated both by recognition of the global paucity (< 1%) of reserves

in the coastal marine environment (Kelleher 1999) relative to terrestrial environments, and

growing evidence for overfishing and related detrimental impacts to marine ecosystems (NRC

1995, Vitousek et al. 1997, Botsford et al. 1997, Jackson et al. 2001). Focus on marine reserves

as a tool for ecosystem and fisheries conservation has been stimulated by a growing literature on

their potential positive consequences in the scientific literature (Plan Development Team 1990,

Agardy 1997, Allison et al. 1998, Bohnsack 1998, Lauck et al. 1998, Murray et al. 1999, NRC

2001, Roberts et al. 2001). These arguments have been made on conceptual inferences, theory

based on fishery models, and patterns or results derived from observational studies of marine

reserves (i.e., empirically based inferences).

At the national scale, Executive Order 13158 signed by retiring U.S. President Clinton

directed the federal government to work with public and private partners to significantly

strengthen and expand the national system of MPAs. This goal includes analysis of the existing

MPAs in U.S. waters and consideration for the development of a nation-wide network of MPAs.

This Order also established the national Marine Protected Area Center

(http://mpa.gov/welcome.html) to implement these objectives. Because of parallel activities in

the state of California and growing interest in other west coast states, the Center is following

closely processes currently underway here. Simultaneously, National Marine Sanctuaries along

the west coast are examining the potential role of marine reserves and other MPAs as they relate

to their mandated objective of protecting habitat within Sanctuaries.

At the regional level, the Pacific Fisheries Management Council has also been examining

the potential applicability of marine reserves (see “marine reserves” link at

http://www.pcouncil.org/). Following the recommendations of an ad-hoc committee (see Parrish

et al. 2000), the Council adopted six fishery management objectives that might be addressed by

marine reserves as a supplemental tool for management of groundfish fisheries. The objectives

that marine reserves might help address, ranked in order of descending priority, include; (1)

Stock Rebuilding: assist in rebuilding overfished stocks and maintaining them at productive
53

levels (2) Biological Productivity: enhance long-term biological productivity, (3) Economic

Productivity: assist in achieving long-term economic production, while minimizing short-term

negative economic impact on all users, (4) Insurance: provide protection for the resource, as a

hedge against the realities of management uncertainty and the effects of natural environmental

variability, (5) Habitat Protection: conserve and protect EFH, and (6) Research and Education:

provide unfished areas for research that will serve as controls for assessment of the effects of

long-term environmental variations and the potential habitat alterations due to fishing, and also

increase our understanding of the role marine reserves may play in fishery management.

At the state level, two recent legislative acts prompted consideration of marine reserves for

both conservation and fisheries management. The Marine Life Management Act was enacted in

1999. The Act (MLMA) stipulates several new fisheries management and conservation

objectives for California's marine living resources (Weber and Heneman 2000). Rather than

focusing on single fisheries management, the MLMA calls for an ecosystem-wide approach to

management. This includes better understanding of how fishery stocks respond to natural

changes in ecosystems (e.g., decadal oscillations, El Nino-La Nina events) as well as the impact

of reducing fished stocks and populations on ecosystem structure, function and services. As such,

the MLMA requires management approaches that consider not only species taken commercially

or recreationally, but to all marine wildlife and their habitats. The Act also requires development

of Nearshore Fishery Management Plans (NFMPs) initially targeting 13 coastal finfishes. These

NFMPs are to also consider more precautionary approaches to fisheries management than has

been demonstrated by past management approaches. For these NFMP objectives and the broader

ecosystem-based approach to management, the California Department of Fish and Game is to

consider any existing or new management approaches, including marine reserves. More

information on the objectives, process and state of implementing the Act is available at

http://www.dfg.ca.gov/mrd/mlma/index.html.

Concurrently, the Marine Life Protection Act (MLPA) requires consideration of marine

protected areas (MPAs) as one of several complimentary management approaches for conserving

nearshore marine ecosystems. More information on the objectives, process and state of

implementing the Act is available at http://www.dfg.ca.gov/mrd/mlpa/. Six goals of MPAs

established in Fish and Game Code Section 2853(b) include the use of MPA’s to:

(1) protect the natural diversity and abundance of marine life, and the structure, function, and

integrity of marine ecosystems,

(2) help sustain, conserve, and protect marine life populations, including those of economic

value, and rebuild those that are depleted,

(3) improve recreational, educational, and study opportunities provided by marine ecosystems

that are subject to minimal human disturbance, and to manage these uses in a manner

consistent with protecting biodiversity,

(4) protect marine natural heritage, including protection of representative and unique marine life

habitats in California waters for their intrinsic value,

(5) ensure that California's MPAs have clearly defined objectives, effective management

measures, and adequate enforcement, and are based on sound scientific guidelines, and

(6) ensure that the state's MPAs are designed and managed, to the extent possible, as a network.
54

II. Why evaluate the effectiveness of marine reserves?

Management agencies are motivated to evaluate how well a marine reserve achieves the

objectives it was created to accomplish for several reasons. Now, more than ever in the history

of resource management, particularly in the state of California, management agencies are being

held accountable for policy decisions and are compelled to provide both clear statements of

reserve objectives and the means by which effectiveness will be measured. Indeed, like many

other recently created marine reserve programs, the MLPA identifies potential “sunset clauses”

that require assessment of reserve effectiveness every three years for the continued

implementation of a reserve.

The accountability mentioned above stems from two concerns. First is the limited financial

and human resources available to any resource agency to develop and implement management

strategies. When limited, resources allocated to one management plan preclude their allocation to

another. This becomes problematic when resources are allocated to the less effective of

alternative approaches. Thus any management approach comes with a potential cost if either it

does not succeed to achieve its objective or does so less efficiently than an alternative approach.

Second is the perceived, potential, or realized impacts that regulations have on stakeholders, who

for this reason demand accountability for regulatory policies.

Another concern for determining reserve effectiveness is to avoid a false sense of security or

achievement when a reserve is assumed to be achieving its objects, but in fact is not. This is

exacerbated if other regulations are relaxed because of the presumed precautionary role reserves

are intended to achieve. Failure of any management approach, due either to poor design or

evaluation, can potentially endanger the resources it was designed to protect. Moreover, poor

evaluation can jeopardize the future of a management approach if it fails to identify the real

value of a management approach or leads to an incorrect interpretation that the approach has

failed. The sooner the benefit (or cost) of a management approach is evaluated and recognized,

the more quickly that approach can be targeted for (or steered clear of) allocation of resources.

Reserve design is unlikely to evolve, through adaptive management, to become more effective if

the relative effectiveness of different designs are not determined and compared. As such, the

design of realistic and achievable conservation targets, and the measurement of their

effectiveness, will be crucial to the successful establishment of new MPA sites and to their longterm success through adaptive management (Carr and Raimondi 1999, Murray et al. 1999).

III. Approaches to evaluating reserve effectiveness

Understanding the various approaches to evaluating reserve effectiveness and their relative

costs and benefits is critical both to interpreting the value (and shortcomings) of past and existing

studies and the design of future reserve evaluation programs. Evaluating the effectiveness of a

marine reserve requires clearly identified goals and objectives, some knowledge of the many

sources of uncertainty—process, model, causal, and measurement—both in reserves achieving

objectives and in our ability to accurately evaluate them, and a well-designed evaluation program

(Syms and Carr 2001). The design and scope of an evaluation program requires objective-based

effectiveness parameters (i.e. response variables: population abundance or size distribution,

species composition or diversity, habitat condition), targets (e.g., specified levels or directions of
55

each of these parameters or response variables), limits (acceptable deviations from specific

targets), as well as a spatial (over what area a target is to be realized) and temporal (how soon

and for how long a target is to be met) context (Syms and Carr 2001).

Reserve goals and objectives come from conceptual and theoretical inferences as well as

empirically based inferences drawn from results detected in existing reserves (see Section II

below). The effectiveness targets defined by these objectives may be of three forms. They may

be absolute values to be attained over some defined spatial area and temporal period. For

example, there may be some reason to target a particular density or abundance of a species

within a reserve. Alternatively, a target may be a relative value, such as some percent increase in

abundance or density within a reserve relative to non-reserve populations. Or, a target may be a

rate of change in the difference between reserve and non-reserve effectiveness parameters over

time (i.e. a pre-defined trajectory of the difference between reserve and non-reserve populations).

There are strong arguments for any of these three forms of effectiveness parameters depending

on the reserve objective and the particular effectiveness parameter.

Two critical components of an effectiveness parameter are the spatial and temporal scale of

the parameter. For example, some parameters may be restricted to within the boundary of

reserve (e.g., increased larval production) and others may be manifested over a far greater spatial

expanse (e.g., larval dispersal to and replenishment of fished populations outside a reserve).

Similarly, some parameters may be expected to response rather rapidly after reserve

establishment (e.g., change in population size structure of a fast growing species within a

reserve) while others may take many years to fruition (e.g., the increased recruitment of a slow

growing species into a catchable stock outside a reserve). For realistic spatial and temporal

expectations and effectiveness targets, as well as appropriately designed sampling programs,

some estimate of the spatial and temporal scales of an effectiveness parameter must be made.

The design and scope of an evaluation program will also depend on the timing and duration

of the sampling or monitoring program relative to the establishment of a reserve. Three

approaches exist, of which the worst-case scenario represents the vast majority of evaluation

studies. This scenario occurs when only one reserve has been established long before an

evaluation program is initiated. In this case, differences between the reserve and non-reserve

treatment levels are confounded by all other site differences and cannot be unequivocally

attributed to protection afforded by the reserve. It will never be clear whether observed

differences (reserve vs. non-reserve) were caused by the reserve or if these differences already

existed before the reserve was established.

If instead, one or more reserves is to be evaluated and sampling can be initiated at the

proposed reserve site(s) and non-reserve “control” site(s) prior to reserve establishment, then

inferences about reserve effects become much stronger. Two approaches are commonly used

when ‘Before’ and ‘After’ data are available. The Impact vs Reference Site (IVRS) approach

treats reserves and controls as formal randomized experimental replicates, and hence makes

inferences about ‘reserve’ effects in general. IVRS requires that sites are truly independent and

sites are assigned randomly to either reserves or control treatments (Stewart-Oaten and Bence

2001). In practice, often these conditions do not hold and so the alternative Before-AfterControl-Impact (BACI) sampling design is used. BACI requires that reference sites be as similar
56

to reserves as possible, and is based on the model that temporal differences in sites are

attributable to reserve effects. Consequently, BACI approaches make site-specific statements of

reserve effectiveness.

BACI designs have been used more frequently in the literature, in particular to test for single

coastal environmental impacts, and a rich literature on this design and analysis exists (StewartOaten and Murdoch 1986, Stewart-Oaten and Bence 2001, and references therein). Non-reserve

control sites at varying distances from the reserve (spatial gradient approach) may be

incorporated into this design to examine the spatial extent of reserve effects. Alternatively, the

effectiveness of a reserve at protecting a species targeted for exploitation can be determined by

employing a BACI approach before and after exploitation begins as long as monitoring of that

species in and out of the reserve has been conducted over that period. An excellent example of

this approach is provided by Schroeter et. al. (2001) in which the status of an exploited

invertebrate was monitored in reserves and fishing grounds before and after initiation of the

fishery. This example demonstrates not only how a reserve can protect a population of a targeted

species, despite strong declines on fishing grounds, but also the role of reserves for assessment of

stocks and the impact of fishing. If instead evaluation sampling cannot be initiated prior to, but

near, the time of reserve establishment, trends in the difference between reserve and non-reserve

sites can be compared to determine if the sites are changing in predicted ways (i.e. increasing

differences over time in density and mean size of individuals within reserves relative to nonreserves).

The design, scope and inferences drawn from an evaluation program will also be strongly

influenced by the design of the MPAs to be evaluated (e.g., the number, size, distribution and

environmental conditions). If only one reserve is to be evaluated, any inferences regarding the

effectiveness of that reserve are largely constrained to only that reserve and cannot be

generalized to reserves (more importantly, potential reserves) in general given the great

environmental and biotic heterogeneity of the coastal marine environment. Any environmental

characteristics (species composition, geologic or oceanographic conditions) unique to that

reserve preclude generalizing how reserves in other areas would respond to protection. This is

particularly true with respect to effectiveness targets that are relative differences between reserve

and non-reserve sites because the relative differences (or trajectories) will depend on the

magnitude of human impacts (e.g., fishing catch) outside the reserve. If multiple reserves and

non-reserves can be sampled simultaneously, broader inferences regarding reserve effectiveness

can be made (general reserve effects rather than the effect of a specific reserve). Moreover,

environmental, design and management differences among reserves can be evaluated relative to

one another. Such an approach is critical to the adaptive management of reserves. Thus, the

design, implementation, analysis, inferences, and success of studies conducted to evaluate

effectiveness of a marine reserve will be influenced greatly by, and therefore must consider, all

of the criteria identified above.
57

II. Consequences of reserve establishment: theoretical and empirical inferences

 II.a. Conceptual inferences

Many postulated consequences (hence, objectives) of reserves are simply intuitive and based

on our limited understanding of the population, community and ecosystem ecology of marine

organisms. Summaries of the hypothesized benefits of marine reserves include the Plan

Development Team’s NOAA-NMFS Technical Memorandum (1990), Bohnsack (1998), Murray

et al. (1999), and the NRC (2001). Because these hypothesized or realized benefits translate into

reserve objectives, they in turn identify parameters by which to measure the effectiveness of a

reserve. These parameters are summarized in Table 1 and follow from the following

“logic/argument”.

Conservation -- For conservation purposes, protection of an intact ecosystem (i.e. biological

communities and their geologic and oceanographic environment) contributes to the persistence

of that ecosystem’s ecological integrity (e.g., species interactions and physical-biological

interactions), structure (e.g., species composition and relative abundance) and function (e.g.,

productivity, nutrient and mineral storage and cycling, habitat structure and integrity). Intact

ecosystems in turn contribute to the persistence of the communities and populations that

constitute it by providing them resources and maintaining species interactions that determine or

regulate populations either locally or regionally. By protecting entire ecosystems it is presumed

that, biodiversity, including rare and endangered species, is better protected. Species densities or

abundances achieved in reserves in the absence of fishing mortality or habitat-altering activities

are presumed to have stronger ecological interactions and effects more characteristic of their

ecological roles in more “natural” or “pristine” ecosystems (e.g., Jackson et al. 2001). The

greater integrity of such ecosystems is believed to lend them more stable (i.e. less variable over

time), resistant (i.e. requiring stronger natural or anthropogenic perturbations, including invasive

species, to cause them to change) and resilient (i.e. more likely to and quicker to return to a preperturbation state). Thus, several characteristics of ecosystems (e.g., productivity, nutrient

cycling, habitat quality) are targets for management and parameters of reserve effectiveness

(Table 1).

The same benefits are presumably conveyed to communities and populations that constitute

protected ecosystems. Thus, attributes of natural communities including species diversity,

community structure, as well as natural densities of species that are either ecologically important

in maintaining community structure or whose persistence reflects intact community/ecosystem

structure are targets for assessing whether a reserve is influencing the state or trajectory of a

community (Table 1). Species interactions of particular importance are those that have

community-wide consequences including the effects of keystone predators in contributing to the

maintenance of diversity, cascading trophic interactions, and the persistence and natural densities

of habitat forming species. In coastal temperate reef systems, examples of predators that have

strong cascading trophic effects are common (see recent reviews by Babcock et al. 1999,

Pinnegar et al. 2000, and Carr et al. in press).

At the population level, reduced fishing mortality and habitat protection within a reserve

should lead to increases in both density (and abundance) and the average size of fishes in a
58

population. Combined, these two attributes should lead to increased larval production of a

protected population relative to populations that are growth or recruitment overfished. These

higher population densities can be more resistant and resilient to the combined effects of natural

perturbations (e.g., storms, El Ninos) and human impacts. As for ecosystem and community

objectives, restrictions on activities that destroy or diminish habitat quality also enhance

persistence of populations within reserves. Because the genetic diversity of any population

(hence it’s potential ability to adapt to changes in environmental conditions) increases with the

number of individuals in that population, protecting populations within reserves increases greater

genetic diversity and the ability of that population to persist in the face of environmental

perturbations and change (e.g., diseases). Thus, both population and environmental attributes

constitute critical effectiveness parameters (Table 1). Importantly, because of the great dispersal

potential of offspring (spores, eggs, larvae) produced by many marine species, populations

protected within a reserve can contribute to the replenishment (hence, conservation) of

populations of these species well beyond the border of reserves (Carr et. al. in press, Shanks et

al. in press, Kinlan and Gaines in review).

Fisheries conservation/management -- One role of marine reserves posited for fisheries

conservation overlaps greatly with the goals of conservation reserves. That is the precautionary

role, or security of spawning stocks and ecosystems protected from fishing effects in light of the

difficulty of regulating fishing effort through most traditional management approaches (Larkin

1977, Ludwig, et al. 1993, Botsford, et al. 1997, Guenette, et al. 1998, Lauck et al. 1998). Thus,

the protection and sustainability of populations of fished species within reserves is a

fundamental objective of reserves created for the conservation of fisheries (Table 2).

In addition, because of the great dispersal potential of reproductive propagules (spores,

eggs, larvae) produced by many marine organisms (referred to as “larval export”), and the

potential movement of benthic juvenile and adult fishes from within to outside reserves (referred

to as “spillover”), some benefits accrued by populations protected within reserves can be

exported to populations outside reserves. This has been a leading consideration for the potential

application of marine reserves for the conservation or management of fisheries. Propagules

transported to populations outside reserves can replenish exploited populations to counter

“recruitment overfishing” (when the reproductive potential of a population has been reduced to

the extent that larval production and recruitment are diminished). Because most marine fishes

exhibit a positive non-linear increase in fecundity with body size (mass), fish allowed to survive

and grow to older ages and larger sizes have a disproportionately higher per-capita production

of larvae (Bagenal 1978, Thresher 1984). Indeed, size-fecundity relationships for rockfishes

(Genus Sebastes) exhibit this non-linear relationship (Love et al. 1990, Love and Johnson 1999).

(Note that the linear relationship described by Gunderson et al. (1980) reflects the narrow size

range of fishes sampled.
59

Table 1. Effectiveness parameters for individual and networked conservation reserves.

I. Species population parameters

Abundance

Density

Size structure

Age structure

Size specific fecundity

Larval production (product of density and size specific fecundity)

Spawning biomass

Population stability

Population resilience

Population resistance

Genetic diversity (within and between populations throughout network)

Demographic rates (reproduction, mortality, immigration and emigration)

Mean individual growth rates

Local population viability estimates

Larval dispersal (to assess extent to which reserve populations are self-replenishing)

Connectivity of larval dispersal with other reserves

Species-specific habitat quality and abundance

II. Community parameters

A. Focal species (e.g., rare, endangered, keystone, indicator, umbrella and flagship species)

- All or subset of species population parameters identified above

- Emphasis on interaction strengths and effects of keystone and exploited predator species

B. Community-wide

Species composition

Species richness

Relative densities of species

Species diversity

Trophic richness

Trophic diversity

Trophic structure

Guild structure and dynamics

Species redundancy

Species interactions and strengths (e.g., competition, predation, parasitism, mutualism)

Community stability and dynamics (e.g., resistance, resilience, constancy and

persistence)

Spatial relationships of populations

Community function (e.g., primary and secondary productivity)

Breadth of resource use (e.g., dietary breadth of predators)
60

Complementarity

Genetic diversity and structure

Threshold effects—potential alternative stable states

III. Ecosystem parameters

Habitat structure (size, shape, spatial arrangement of habitats)

Habitat richness

Habitat diversity

Habitat representativeness

Physical (structural) complexity (of abiotic and biotic substrata)

Interactions between biogenic physical structures and species that alter them.

Productivity (C gm fixed / area / time; total and by trophic level)

Nutrient and matter cycling and fluxes (e.g., rates of change, rates of cycling, fluxes,

nutrient ratios, nitrogen fixation)

Detrital production and export.
61

Table 2. Effectiveness parameters for individual and networked fishery reserves

I. Population parameters

A. Local (within reserve)

Abundance

Density

Size structure

Age structure

Size specific fecundity

Larval production (product of density and size specific fecundity)

Spawning biomass

Mean individual growth rates

Demographic rates (reproduction, mortality, immigration and emigration)

Population stability and dynamics (e.g., resistance, resilience, constancy and persistence)

Genetic diversity (within and between populations throughout network)

Local population viability estimates

Larval dispersal (to assess extent to which reserve populations are self-replenishing)

Density, dynamics, and stability of by-catch species

B. Regional (outside reserve)

Larval production and export rate (from inside to outside reserve)

Larval dispersal and recruitment patterns (outside reserves)

Emigration (i.e. “spillover”) and immigration of benthic stages inside and outside of

reserves

Stock stability and dynamics (e.g., resistance, resilience, constancy and persistence)

Fishery yield

II. Community parameters

A. Local (within reserve)

By-catch assemblage composition, structure, dynamics, and stability

Density, dynamics, and stability of resource requirements for exploited species

B. Regional (outside reserve)

Community stability, to extent that reserves contribute to regional stock abundance and

stability, and exploited species influence community structure

III. Ecosystem parameters

A. Local (within reserve)

Abundance and quality of spawning, recruitment and other habitat requirements

Abundance and quality of other ecosystem-based resource requirements

B. Regional (outside reserve)

Ecosystem stability, to extent that reserves contribute to regional stock abundance and

stability, and exploited species influence ecosystem structure

Ecosystem stability, to extent that reserves contribute to production and export of

ecosystem components (e.g., larval export and replenishment of biogenic habitat)
62

Thus, individuals protected in reserves can contribute disproportionately (relative to their

numbers) to larval production and recruitment to fished populations. Moreover, there is growing

evidence in the fishery literature that older females produce young of greater physiological

condition (Zastrow et al. 1987, Chambers and Legett 1989, Buckley et al. 1991) that translates

into higher growth and survival of those larvae (Blaxter and Hempel 1963, Houde 1987, Hislop

1988, Monteleone and Houde 1990). Indeed, this “maternal effect” of female age on larval

condition and survival has been demonstrated in laboratory studies of the black rockfish,

Sebastes melanops (Steve Berkeley, personal communication). This may increase further the

proportionate contribution of larval recruitment to fished populations by older females protected

in reserves. These contributions to larval recruitment become more important when considering

how quickly populations rebound from disturbances. If populations of larger, older females in

reserves are less vulnerable to disturbances (because the greater stability and resiliency of larger

populations and intact ecosystems), and fished populations outside reserves are recruitment

overfished, populations in reserves can produce and export large numbers of recruits to fished

populations after disturbances or during episodic environmental opportunities for recruitment.

Thus, reserves may act to hasten the rate that fished populations rebound from perturbations or

take advantage of episodic environmental conditions advantageous for recruitment. For all these

reasons, export or emigration to, and many predicted responses of, fished populations outside of

reserves are potential parameters of effectiveness of fisheries reserves (Table 2).

If individuals in a reserve grow to sizes they otherwise would typically not achieve in a

fished population, movement of these individuals from reserves to fished populations (i.e.

“spillover”) acts to increase the average size of fishes caught in the fishery adjacent to a reserve.

This replenishment of larger individuals in a fished population counters “growth overfishing”

(reductions in the average size of fish in a fishery caused by removal of larger individuals). The

spatial extent of this benefit depends on the range of individual movement of a species. The

movement of individuals (larvae, juveniles or adults) from reserves to fished populations

increases the genetic diversity of fished populations and provides the benefits of greater genetic

diversity to fished populations described in the Conservation section above.

Another important potential application of reserves for both fisheries conservation and

management is their role as “reference” or “control” sites to assess the influence of fishing

activities on marine populations and ecosystems. It is extremely difficult to tease apart the

effects of fishing and natural variation in the environment from one another without populations

that are not subjected to one or the other. Comparison of populations (and ecosystems) subjected

to and protected from fishing may allow the perceived or hypothesized effects of fishing to

actually be tested (and therefore supported or refuted). This is critical to understanding how

different levels of fishing mortality and related activities influence stocks and their ecosystems.

II. b. Theoretical (model-based) inferences

Conservation -- A large number of models developed for conservation reserves have focused on

issues of design theory rather than exploring reserve effects. There is a surprising dearth of

mathematical models developed to predict or understand how communities or ecosystems would

respond to protection by marine reserves. Perhaps to some extent this reflects the variety and

complexity of coastal ecosystems, which might require more specific mechanisms, knowledge of
63

the relative openness of the various populations, and therefore more specific (not necessarily

more detailed) models to be useful. For example, some models have identified the importance of

the strength and stage at which density dependence occurs (traits that vary widely among species

in a community) in determining how a population will respond to relaxed fishing mortality

(Hastings and Botsford 1999, Walters et al. 1999, St. Mary et al. 2000).

Fisheries conservation/management -- In contrast to the paucity of models developed to explore

the effects of conservation reserves, a large number of models have been constructed to examine

the potential of reserves for fisheries purposes. One recent summary of these models is provided

in the NRC (2001) review of marine protected areas (Table 6.3, Appendix G.). Reserve models

can be categorized by four primary objectives; (1) their role as a precautionary means to

ameliorate accidental overfishing, (2) their potential contribution to ameliorate growth

overfishing by exporting older life stages to a fishery (i.e. spillover), (3) their effect on fishery

yield and how their effect relates to and compares with more traditional approaches to

controlling fishing effort, and (4) predictions of socio-economic consequences to a fishery. This

last category (socio-economic consequences) is dealt elsewhere in this report.

Precautionary management -- Several recent essays have recognized the difficulties of traditional

management approaches to controlling fishing effort and catch. These difficulties stem in large

part on problems of stock assessment, the great natural variability of marine fish populations, and

the uncertain political structure of management decisions (Larkin 1977, Lauck et al. 1998,

Hilborn and Walters 1992, Mangel 2000a). Thus, many models have examined the potential role

of reserves in contributing to more precautionary approaches to fisheries management and the

long-term sustainability of fisheries (Goodyear 1993, Mace and Sissenwine 1993, Mace 1994,

Mann et. al. 1995, Lauck et al. 1998, Roughgarden 1998, Soh et al. 1998, Foran and Fujita 1999,

Guenette and Pitcher 1999, Guenette et al. 2000, Mangel 1998, 2000a, 2000b, 2000c, NRC 1999,

2001). These models identify that the proportion of fished stocks protected within reserves

depends on rates of fishing mortality and that the larger the portion protected within a reserve,

the more precaution afforded the fishery. These models predict reasonable to very high

proportions (20-70%) of a stock, depending on the species, are typically necessary to be set-aside

within reserves in order to contribute to sustainability when fishing mortality varies from

moderate to very high (i.e. the “scorched earth” scenario) rates, respectively. Because these

models are developed to examine a reserve’s precautionary role, these worst-case scenarios are

often applicable.

“Spillover” -- A small subset of fisheries reserve models have focused on potential effects of

fish emigration from a reserve (i.e. “spillover”) to a fishery (Polacheck 1990, Russ et al. 1992,

DeMartini 1993). These models suggest that emigration of adults into fished areas can lead to

increases in yield per recruit, but these increases occur only when emigration rates are

moderately high and fishing mortality outside the reserve is high, but regulated. Thus, these

models imply that knowledge of rates of bi-directional movement of exploited species into and

out of reserves, and of the various factors that contribute to movement rate, direction and

distance (e.g., fish size or age, density, habitat attributes or quality) can allow inferences of the

contribution of “spillover” to yields (and increased fish sizes) in adjacent fisheries. These

models also indicate that the contribution to increased yield is constrained to areas close to the

reserve. Therefore, sampling designs to detect the degree and spatial extent of the spillover
64

effect should incorporate an appropriate sampling gradient (e.g., for a linear coastline, sampling

concentrated toward the reserve boundary by distributing samples by the square root of distance

from the boundary). These studies have two additional important implications for assessing the

effectiveness of existing reserves in the MBNMS. First, is that the effect of a reserve in

exporting larger individuals (i.e. “spillover”) to a fishery depends on (1) the size of an individual

reserve relative to movement patterns and ranges of a targeted species, (2) the degree of growth

overfishing outside a reserve, and (3) the intensity of fishing within the range of spillover from a

reserve. To date, no studies have directly examined spillover from a reserve within MBNMS,

nor has information on the degree of growth overfishing or location and intensity of fishing

outside and adjacent to reserves been collected or examined. However, information on fishing

adjacent to Big Creek Ecological Reserve is being collected (John Smiley, pers. comm.). But

also critical to this evaluation is information on movement ranges of targeted species and how

that relates to the size of existing reserves. Without such information it is very difficult to

estimate rates of spillover and contributions of existing reserves to their potential in countering

growth overfishing and the size distribution of catches in a fishery.

Fishery yield -- Most fishery reserve models have examined the effects of reserves in

contributing to total fishery yield as well as its sustainability and reducing its temporal

variability. Most have compared the effects of varying total reserve size with levels of effort

control by traditional management approaches. Overall, these models indicate that the effect of

adding reserves on total yield is essentially the same as decreasing fishing mortality (reviewed by

Botsford et. al., in review). Mangel (1998, 2000b) demonstrated that yield depended on the

product of fishing mortality and area not in reserves, not on the specific values of each.

Predictions of how reserves contribute to increasing yield vary widely among these models.

Some models suggest that comparable yields can be achieved through traditional effort control

and reserves (Mangel 1998, 1999, 2000a, Hastings and Botsford 1999). Several studies indicate

that reserves can increase yields only when stocks have been heavily overfished (Quinn et al.

1993, Holland and Brazee 1996, Sladek Nowlis and Roberts 1997, 1999, Guenette and Pitcher

1999). These models suggest that the proportion of a stock necessary to be set aside within

reserves to enhance yields ranges from 20 to 50% (NCR 2001, Table 6-3). Thus, predictions of

the magnitude to which a reserve will influence yield from stocks outside reserves will depend

not only on the proportion of a stock protected in a reserve, and larval production and export, but

also on the state of the fishery and continued level of exploitation. Thus, any study measuring

the relative effectiveness of a reserve in protecting or enhancing populations either inside or

outside a reserve will benefit greatly from (if not absolutely require) good estimates of spatial

and temporal patterns of fishing mortality. Optimally, unequivocal conclusions regarding the

effect of a reserve on fishery yield are drawn from comparisons of yield between independent

fished stocks before and after, as well as with and without, presence of a marine reserve (i.e. a

BACI design with or without replicate stocks). To date, no such comparisons exist. The spatial

scale of the comparison (i.e., independent fish stocks) and the temporal scale that would provide

sufficient statistical power to distinguish these effects will require more information and much

longer continuous studies than currently exist, especially along the west coast of the United

States.

Fishery variability -- Some models have suggested that reserves can contribute to the resiliency

of a stock (Foran and Fujita 1999), or dampen the magnitude of temporal variability (Sladek
65

Nowlis and Yoklavich 1998, Sladek Nowlis and Roberts 1999). Determining to what extent a

reserve actually contributes to resiliency (how fast population size returns to a pre-disturbed

level following a disturbance) or dampen temporal variability of a fished population is very

difficult. As in the case of assessing reserve effects on yield, unequivocal conclusions are drawn

from comparison of these variables (resiliency and variability) between independent fished

stocks before and after, as well as with and without, presence of a marine reserve (i.e. a BACI

design). To date, no such comparisons exist. The spatial scale of the comparison (i.e.

independent fish stocks) and the temporal scale that would provide sufficient statistical power to

distinguish these effects will require more information than currently exists, especially along the

west coast of the U. S.

One very important conclusion and implication from all of the above models considering

reserves for fishery conservation or management is that measurable (detectable) effects of

reserves on fisheries yield and variability require total reserve areas (across the entire network)

that are far larger than the existing reserves throughout the MBNMS, California, and the West

Coast. This is very important when interpreting assessments of reserve effects from the existing

reserves within the MBNMS. Likewise, a large body of models has been developed to examine

and optimize the design of fisheries reserves. These models examine the effects of reserve

design (specifically, the overall size of a reserve network, individual reserve area, and the

number and spacing of reserves in a network).

II. c. Empirically-based inferences

Conservation -- There is widespread empirical evidence of the predicted increases in population

density, biomass and size frequency of exploited species to protection within reserves. Recent

syntheses of examples include Jones et al. (1992), Rowley (1994), Roberts and Hawkins (2000),

NRC (2001), and Halpern (in press). Halpern’s review is the most comprehensive, having

surveyed 89 separate studies, 41% of which were in temperate ecosystems. Halpern examined

three population level variables (density, biomass, and average size of organisms) inside and

outside of reserves or before and after reserve establishment. Relative to non-reserve references

(either before establishment or areas outside reserves), 63% of reserves had higher density, 90%

of reserves had higher biomass, 80% of reserves had larger organisms, and 59% of reserves had

higher diversity (all Chi-square analyses significant at p<<0.001). This pattern was consistent

across the four functional groups examined (carnivorous fishes, herbivorous fishes,

planktivorous fishes/invertebrate eaters, and invertebrates), with the exception of invertebrate

biomass and size. A small number of reserves had lower values for these three population

measures (7%, 0%, and 2% of reserves had lower density, biomass, and organism size,

respectively). Although there was great variance in the magnitude of these effects among

reserves, density doubled, biomass nearly tripled, and organism size increased by 20 to 30%

relative to the values for unprotected areas. Thus, these three desired and predicted responses of

species to protection within a reserve are reasonable parameters by which to examine and

evaluate reserve effectiveness.

In addition to the 3 studies conducted within marine reserves within the MBNMS--described

in detail in this report--specific examples from temperate reserves along the West Coast include

Palsson’s (1998) surveys of rockfishes at 7 sites in Puget Sound, two of which were no-take
66

marine reserves. Results varied among species, largely corresponding to fishing intensity.

Heavily and moderately fished species like copper and black rockfish, respectively, were at

significantly greater densities within reserves. Quillback rockfish exhibited a similar result for

larger individuals. In contrast, densities of lightly fished species (brown and Puget Sound

rockfish) did not differ significantly between reserve and fished sites. Length frequency

distributions of copper and quillback rockfish were also shifted to larger sizes within reserves. In

a separate study, lingcod density was 3 times greater in a Puget Sound reserve compared to

fished areas (Palsson and Pacunski 1995). Another example of increased density of an exploited

temperate reef fish within reserves is provided by Martell et al. (2000). The density of spawning

lingcod, Ophiodon elongatus, was significantly greater in two small reserves than in adjacent

fished areas within the Straits of Georgia.

Similar patterns of increased density and individual size has been demonstrated for

exploited species on temperate New Zealand rocky reefs (Jones et al. 1992). For example,

McCormick and Choat (1987) recorded that abundance of the red moki, Cheilodactylus

spectabilis, was six times higher in a New Zealand marine reserve than adjacent sites. Similarly,

increased density and sizes of snapper (Pagurus auratus) and spiny lobster (Jasus edwardsii)

occurred in New Zealand reserves (Babcock et. al. 1999). Snapper were 6 to 9 times more dense

and 50% longer within reserves. Lobster were 1.6 to 3.7 times more dense and had an average of

1.6 cm longer carapace length. In the first before-after comparison of a temperate reserve in

New Zealand, Cole et. al. (1990) also found similar increases in abundance of all three of these

previously mentioned species as well as the commercially important blue cod (Parapercis

colias), and decreases in 3 others. Commercially taken snails (Concholepas) in the temperate

intertidal of Chile increased in density 5 to 14 times and doubled in body size in reserves relative

to exploited areas (Castilla and Duran 1985). Rock lobster (Jasus) and a reef fish abundance

increased by 1 and 2 orders of magnitude on temperate rocky reefs within reserves in Tasmania

(Edgar and Barrett 1999). The more persistent (and increased) size of populations of the warty

sea cucumber within reserves in spite of simultaneous declines on fishing grounds in the Channel

Islands of California provide another example of protection afforded species within reserves on

temperate rocky reefs (Schroeter et al. 2001). These documented responses from other temperate

ecosystems suggest that increases in densities of lobster, finfish and other species seem

reasonable to anticipate and test for in California, particularly in heavily fished regions.

Community and ecosystem responses have focused on changes in species diversity and

cascading effects of ecologically influential species protected from fishing. In Halpern’s (in

press) comprehensive review mentioned in the preceding section, he also examined differences

in species diversity, inside and outside of reserves or before and after reserve establishment.

Fifty-nine percent of reserves had higher species diversity (Chi-square analysis: p<<0.001) than

their corresponding non-protected references. This pattern was also consistent across all four

functional groups examined (carnivorous fishes, herbivorous fishes, planktivorous

fishes/invertebrate eaters, and invertebrates). Diversity decreased within reserves in 10% of the

studies reviewed. Although variable among studies, on average, diversity increased by 20 to

30% relative to the values for unprotected areas.

Actual examples of the cascading effects created either by reduced human predation or by

increases in predator density in temperate reserves are few (reviewed by Steneck 1998, Castilla
67

1999, Pinnegar et. al. 2000) at least in part because very few studies have tried to assess them.

Those that are most conspicuous involve the role of predators that influence major grazers

(especially sea urchins) that in turn influence the cover and density of macroalgae (Pinnegar et

al. 2000). One particular example provided by Babcock et al. (1999), was the cascading

consequences of increased densities of snapper and lobster in reserves. With increased densities

of these two sea urchins predators within reserves, urchin densities declined and macroalgal

cover increased significantly, relative to fished areas outside reserves. Similar consequences of

the cascading effects of increased densities of urchin predators within reserves has been

described in coral reef systems (Sala et. al. 1998, McClanahan 2000). These general responses

have two important implications for measuring potential responses of communities and

ecosystems inside reserves established in areas where sea urchin predators or sea urchins

themselves are heavily fished. Community responses to protection within the MBNMS and

adjacent temperate rocky reef ecosystems might exhibit broader cascading responses to

protection, but the response is likely to vary markedly with regional variation in community

structure. In southern areas (e.g., the Channel Islands and Southern California Bight) where

urchin predators including sheephead and lobster are fished, effects of protection from fishing is

likely to cause cascading effects on urchins and macroalgal assemblages. Predicted increases in

macroalgal stands in turn is likely to influence the recruitment of many reef fishes (Carr 1989,

1991, 1994, Carr and Reed 1992). In contrast, throughout Central California and the MBNMS,

sea otters appear to limit sea urchin density and protection of finfishes or invertebrates within

reserves is less likely to influence urchin and kelp densities. Further north, in areas of Northern

California where human take of urchins is the primary determinant of local sea urchin density,

increased urchin densities within reserves is likely to reduce macroalgal stands, an effect

opposite of that predicted for Southern California. Here, another important indirect effect of

protection of sea urchins wthin reserves is their apparent influence on recruitment of abalone.

Rogers-Bennett and Pearse (2001) describe significant increases in abalone recruitment in the

presence of sea urchins within reserves. This influence of sea urchins on abalone recruitment

appears to occur over a broad geographic range, including Southern California (Tegner and

Dayton 1977). Hence, conceptual and empirically-based inferences suggest that such indirect

effects are critical measures of reserve effectiveness with respect to ecosystem and biodiversity

objectives in the MBNMS.

Fisheries conservation/management -- Global impacts of reserve implementation on adjacent

fisheries have been recently reviewed elsewhere (Rowley 1994, Roberts and Hawkins 2000,

NRC 2001). We focus this discussion on examples from coastal temperate examples from the

northern hemisphere that bear more directly on potential effects of reserves within the MBNMS.

Mentioned previously, an example of temperate reserves both protecting populations of a

stock declining on fishing grounds and providing a more accurate assessment of fishery trends

and impacts on the stock was provided by Schroeter et al. (2001). These investigators compared

long-term monitoring trends of the warty sea cucumber (Parastichopus parvimensis) within and

outside reserves prior to and during exploitation of that species at the Channel Islands off Santa

Barbara, California. The assessment demonstrated that populations within reserves remained

constant (or increased!) as populations on fishing grounds declined. Any differences in

cucumber densities between these sites prior to the fishery were significantly less than

differences in density over the 7-year period following initiation of the fishery.
68

Although fishery closures on Georges Bank were established to enhance recovery of

finfishes, populations of the scallop, Placopecten magellanicus, increased 14-fold over a 4-year

period within reserves relative to adjacent trawled areas (Murawski et al. 2000). In late 1994,

three large areas on Georges Bank and in Southern New England, totaling 17,000 km, were

closed year-round to any gears capable of retaining groundfish (trawls, scallop dredges, gill nets,

hook fishing). Four years later, by the end of the study, total and harvestable scallop biomasses

were 9 and 14 times denser, respectively, in closed than in adjacent open areas. Moreover,

spatial patterns of subsequent recruitment of young scallops outside the reserve strongly

suggested that spawning stocks protected within the reserves were largely responsible for larval

export and replenishment of adjacent fished populations.

“Spillover” -- More empirical studies on the movement of adult fishes relative to reserve

boundaries have been conducted on coral reef fishes in the tropics than on temperate rocky reef

fishes. Recent and notable examples include McClanahan and Mangi (2000) and Roberts et al.

(2001). One rather obvious but important pattern that emerges from these tropical studies are

that individuals whose home ranges straddle reserve boundaries move frequently in and out of

reserves, while individuals whose home ranges are either entirely within or outside the reserve,

may rarely cross reserve boundaries (Zeller and Russ 1998). Also critical to predicting and

interpreting rates of spillover is knowledge of how emigration responds to the local density of

conspecifics (i.e. density-dependent movement). Predicted increases in survival and density

within reserves, relative to continued depletion of conspecifics outside reserves creates a

potential density gradient that could induce individuals to emigrate to adjacent fishing grounds

(St. Mary et al. 2000, Sanchez Lizaso et. al. 2000). Of temperate zone studies, Attwood and

Bennett (1994) used a tag-recapture approach to estimate emigration of galjoen (Coracinus

capensis) from a surf-zone reserve in South Africa. Over 5.5 years, roughly 9% of the tagged

fish were recaught, of which roughly 18% had emigrated from the reserve to the adjacent fished

areas. However, movement in the opposite direction (immigration into the reserve) was not

estimated. Thus, knowledge of the distribution, relative movement rates, and factors that

contribute to movement (e.g., density and habitat quality) are critical information for predicting

and interpreting spillover from a reserve.

Fishery yield – We are unaware of any empirical studies in temperate oceans that have explicitly

examined effects of marine reserves on fisheries yield; none have done so along the west coast of

the United States. Theoretical studies indicate that total fisheries yield will only increase when

large reserves are present in a highly fished population (Hastings and Botsford 1999), and

modeling studies suggest that local increases in yield will be dependent on reserve size and

movement rates (Polacheck 1990, DeMartini 1993, Attwood and Bennett 1995). Empirical

studies have generally interpreted changes in abundance with distance from reserve boundaries

as potential increases in yield (eg. Cole 1994). However, yield increases have been recorded

adjacent to a tropical marine reserve in the Philippines Islands (Alcala and Russ 1990), and

reported from St. Lucia (tropical Caribbean) and sub-tropical Florida (Roberts et al. 2001, but

see an alternative interpretation by Tupper et al. 2002). These empirical studies demonstrate that

increases in yield are generally localized, and should be compared to overall fisheries yield. It is

unlikely that existing marine reserves in California contribute substantially to fisheries yield. The

small size of existing reserves would probably make any contribution to total yield insignificant.
69

The small boundary interface of most reserves with fished areas, in combination with the small

size of any abundance differences between reserve and adjacent non-reserve areas, would in all

likelihood preclude any local perception of yield increases.

III. What have past and present studies in existing California reserves addressed and

concluded?

Only three major studies have explicitly addressed the effectiveness of subtidal marine

reserves within the MBNMS. All three studies have focused solely on the response of selected

fish species to protection within reserves, and are more thoroughly summarized elsewhere in this

report. In general, these studies have by and large reinforced the most common and conspicuous

pattern of responses of exploited species to protection within a reserve. Estes and Paddack (see

also Paddack and Estes 2000) compared fish density and sizes inside and out of three reserves

within the MBNMS (Big Creek Ecological Reserve, Point Lobos Ecological Reserve, and

Hopkins Marine Life Refuge) and found a statistically non-significant trend of higher densities

of fishes in reserves compared to fished areas outside reserves. However, the average length of

rockfish (genus Sebastes) was significantly greater in 2 of the 3 reserve sites, as was the

proportion of larger fish. In combination, the greater population density and sizes combined to

produce substantially greater biomass and therefore potential larval production per unit of area

within the reserves. Importantly, the magnitude of these effects seemed to be correlated with the

age of the reserve.

The second and third studies examined fish assemblages inside and outside of Big Creek

Ecological Reserve. VenTresca et al. conducted diver surveys the first 4 years (1995-1998)

following reserve establishment. They detected no temporally consistent pattern of difference in

density or size of fishes inside and outside of the reserve. The third study (Yoklavich et al.)

examined fish assemblages inside and outside of the reserve 3-4 years (1997-1998) after reserve

establishment. Again, no significant difference in density of any species inside and outside of

reserves was detected. Moreover, they did not detect any significant difference in size of seven

economically valuable species. These comparisons were of combined densities across species.

Individual species comparisons were not made.

None of the three studies conducted in MBNMS reserves to date found significant

differences in fish densities between reserve and non-reserve areas. However, Paddack and Estes

found weak evidence that fish were larger in two of the three reserves. The other two studies

were conducted within 4 years of reserve establishment, and it is unlikely that any effects of

protection would have had time to be manifested. It is likely that even small increases in the

number of large fish will exert a disproportionate effect on the potential larval production of the

protected population (refer to section II.a. Conceptual inferences: Fisheries

conservation/management). Given the extended larval durations (1-4 mo) and dispersal potential

for most of the fishes that demonstrate increased potential larval production, export to and

replenishment of populations outside reserves is very likely. However, given the small size and

limited number (i.e. low probability of connectivity) of reserves in this region, it is unclear and

perhaps unlikely that these reserves function as a network that could buffer the effects of any

overfishing from surrounding areas (Carr and Reed 1993, Botsford et al. 2001).
70

IV. Filling in the gaps: critical directions for future research

To develop a better understanding for the potential of reserves as tools for conservation and

fisheries management, it is clear that far more monitoring effort will be required. In addition to

more information on the responses of targeted species, community-level and ecosystem-wide

responses need to be examined. Of the many effectiveness parameters identified from the

literature (Tables 1-3), few have been explicitly assessed in these studies. Without continuous

monitoring, temporal persistence of differences between reserve and non-reserve populations

will remain unknown. With the present system of reserves and monitoring effort, unequivocal

ascription of differences between reserve and non-reserve populations to causal reserve effects is

tenuous. Better understanding of the consequences of reserve establishment requires both

creation of new reserves in concert with well-designed monitoring studies over many years.

Evidence from a great variety of theoretical and empirical studies across a wide diversity of

coastal ecosystems indicates that reserves have the potential to achieve many conservation and

some fisheries management objectives. If reserves are to be developed further and successfully

used as a management tool, it is necessary to develop a structured and well supported monitoring

program, which clearly identifies a set of effectiveness parameters that measure how well a

reserve achieves a stated objective.

Evaluation design and analysis

All three of the published evaluation studies on MBNMS MPA’s have been “snapshots” of

population and community states after implementation of reserves. Consequently, inferences

about reserve ‘effects’ are weak. However, even if a suitable formal approach is employed

(BACI or impact vs reserve analysis) it is still not sufficient to just measure the response

variables when evaluating MPA effectiveness. A range of other features of the habitat, reserve

configuration, and exploitation pattern alter both empirical and theoretical evaluation (Table 3).
71

These effects will strongly influence the ability of monitoring designs to make statements of

effect. BACI analyses, for example, require that the underlying dynamics of reserve and nonreserve, with the exception of the reserve ‘effect’, be correlated. Conversely, an IVRS approach

requires that the dynamics of reserve and non-reserve be independent.

Marine populations and communities are subject to a range of forcing processes, such as

oceanographic and climate regimes, that can alter both population and community dynamics and

might alter the synchrony of reserve and non-reserve areas. Other variables might alter the

strength of interactions between populations in different reserves. For example, habitat quality

and structure might alter the replenishment and survivorship patterns of ‘response’ organisms.

This extra variability, in combination with events such as natural or anthropogenic disturbances,

might obscure or confound effectiveness measures.

The history and ability of the reserve to actually protect the organisms is also important to

incorporate into any analysis. Temporal lags in species population dynamics might define a

temporal limit to reasonably expect a protection effect. The size and shape of the reserve and the

permeability of its boundaries to emigration, in combination with the degree of enforcement are

important to establishing the size of the stock that is actually protected from fishing mortality.

Exploitation patterns are one of the least appreciated sets of variables that should be

included in MPA evaluation. Many models assume complete removal of stock from fished areas,

but this is rarely the case. It is important to measure the actual extraction rate, the location and

temporal structure of extraction, and other issues such as gear type, bycatch, and gear

Table 3. Covariates required to assess MPA effectiveness

Habitat variables Reserve characteristics Exploitation variables

1. Physiographic habitat

structure

2. Oceanographic environment

3. Biotic habitat structure

4. Connectivity with other

biotic and abiotic habitats

5. Climatic variation

6. Proximity of essential

habitats (e.g., nursery

grounds)

7. Potential for other threats

(e.g., pollution)

8. Natural disturbance regime

1. Size

2. Shape

3. Age

4. Enforcement effectiveness

5. Reserve network

configuration

6. Edge permeability

1. Total fishing effort

2. Spatial location of fishing

effort

3. Targeted species

4. Effect of other effort control

regulations in managing

fishery

5. Gear type and effects on

habitat

6. Bycatch composition and

abundance

7. Temporal distribution of

fishing effort

8. Amount of area fishable by

different gear types
72

disturbance issues. If fishing mortality is comparatively low, then it is unrealistic to expect that

an MPA will have significantly higher abundances of an exploited species. This of course does

not argue against MPA’s as a management tool, but suggests that their effectiveness lies not in

augmenting a fishery, but more as an insurance policy.

One of the most important classes of covariables are those that correspond to the linkages

between the MPA and surrounding habitats. Modeling approaches are sensitive to boundary

conditions, i.e., assumptions about limitations to values that go beyond the model domain. In

nature, boundaries are defined by oceanography and physical habitat structure. In addition, many

models make assumptions about connectivity (e.g., a common larval pool, or an assumed larval

dispersal distance) that are unlikely to be realistic or precisely estimated. Connectivity in marine

systems operates on at least two levels. Oceanographic connectivity is of central importance to

understanding linkages between stocks in different locations. Features such as eddies and

longshore currents will have very different effects on larval connectivity. Landscape connectivity

is important for identifying critical bottlenecks in ontogeny (e.g., proximity of reserve to nursery

or spawning habitats) and recognizing variability due to permeability of habitat types between

reserve and non-reserve areas.

Acknowledgements

We thank Meghan Sullivan and E. J. Dick for their assistance in the development of this

report and Steve Berkeley for information on maternal effects on larval condition. We were

supported jointly by the Monterey Bay Sanctuary Foundation and the Partnership for

Interdisciplinary Studies of Coastal Oceans: A Long-Term Ecological Consortium (PISCO)

funded by the David and Lucile Packard Foundation.
73

CHAPTER 6

Effectiveness of Marine Reserves:

Socio-Economic Considerations

Caroline Pomeroy

University of California Santa Cruz, Institute of Marine Sciences

Earth and Marine Science Building, A316, Santa Cruz, California 95060

Abstract

Marine reserves require careful consideration of a range of socio-economic, or more

accurately, human dimensions factors, including how people and communities perceive, value,

and use the marine environment, and the informal and formal rules and structures that influence

these values and uses. The rationale for attending to these is both legal and practical, and follows

from the understanding that 1) marine reserves generate social and economic impacts which both

individuals and society must bear; 2) human responses and adaptations to marine reserves have

ecological as well as socio-economic implications; and 3) attention to the human dimensions can

afford resources that are essential to marine reserves processes and outcomes. Public

participation and social science research are complementary means for addressing the human

dimensions of marine reserves. Public participation is useful for identifying and addressing

social, economic, and ethical concerns and issues associated with marine reserves, and

generating trust, support, and other social resources to assist in their design, management and

evaluation. It cannot, however, substitute for systematic collection and analysis of socioeconomic data to inform the process. To date, however, limited social science research has been

done, with the majority focused on identifying the necessary and sufficient conditions for the

establishment of marine reserves. Although the theory of marine reserves posits that social and

economic benefits will outweigh their costs, little research to test this hypothesis, or to inform

and evaluate the human dimensions of marine reserve design, implementation, management and

outcomes more generally, has been done. Among the four marine reserves discussed in this

report, only the Big Creek Ecological Reserve has been the focus of directed social science

research. Basic questions about changes in use patterns, perceptions, attitudes and beliefs,

economic values, costs and benefits, as well as social and economic relationships and

institutions, can be used to guide social science research on marine reserves effectiveness.

Reliable and valid answers to these questions, however, require baseline information on the

human dimensions of marine reserves, monitoring of these attributes throughout the marine

reserve process, and insuring their integration into management as it changes over time.
74

Introduction

As interest in marine reserves as a fishery and broader ocean management tool has grown,

so has the recognition of the need to consider their human, as well as their ecological,

dimensions (Fiske 1992, Alder et al. 1994, Cocklin et al. 1998). The rationale for attending to

these is practical as well as legal. The human dimensions include not only the potential social,

cultural and economic impacts of marine reserves, but also the social, cultural, economic and

political context in which they are designed, implemented and managed. Because marine

reserves are established to serve social and economic, as well as ecological, goals, attention to

the human dimensions throughout the process (i.e., design, implementation, management and

evaluation) is necessary in order to evaluate their effectiveness in achieving these goals.

Moreover, given that people interact with the marine environment within and outside marine

reserves, and that marine reserve boundaries are permeable, understanding the human

dimensions is also central to evaluating the ecological effectiveness of marine reserves.

To date, there has been limited social scientific study of the human dimensions - the socalled “socio-economic considerations” - of marine resource management in general and marine

reserves in particular. A growing literature, however, demonstrates the relevance of these

considerations, and how failure to explicitly address them has led to negative outcomes in

ecological and socio-economic terms (Fiske 1992; Alder 1994, 1996; White et al. 1994, 2002;

Cocklin et al. 1998, Pomeroy and Beck 1999, Russ and Alcala 1999). The key findings of this

work point to the critical participation and multifaceted role of the full range of stakeholders

throughout the marine reserve process. Most of this work has focused on evaluating marine

reserve processes to determine the social, cultural and political conditions for their successful

establishment. Little social science research has been done to inform the design, implementation,

management and evaluation of marine reserves. This research focuses on use patterns;

perceptions, attitudes and beliefs; economic values, costs and benefits; social and economic

relationships; and formal and informal governance institutions (e.g., regulations and norms,

respectively).

Yet, there is a critical disconnect between social science research and policy, linked in part

to the tendency for marine reserve processes to proceed more rapidly than relevant social science

research (Farrow 1996, Badalamenti 2000). This type of social science research is often carried

out as an afterthought, if at all, with inadequate time, personnel and financial resources and little

or no provision for its integration into the process. In most cases, insufficient social scientific

baseline information is generated, making it difficult to develop appropriate measures to evaluate

the human dimensions, and how these interact with the ecological dimensions, to influence the

effectiveness of marine reserves. This is the case for the four marine reserves addressed in this

report.
75

The first part of this chapter provides an overview of socio-economic considerations for

marine reserves. The next section presents the legal and practical rationales for addressing them

in the marine reserve process. The third section focuses on the diversity of socio-economic

considerations and their relevance throughout the process. The status of social science research

that has been conducted on the four marine reserves that are the subject of this report is noted for

each topic. The fourth section discusses the social science methods that can be used to generate

scientific information, and briefly discusses measures of effectiveness for marine reserves. The

last section presents a brief summary and conclusions regarding the relevance of the human

dimension, and the critical need for information on this integral aspect of the marine reserve

process.

Rationale for addressing the human dimension in the marine reserve process

There are both practical and legal rationales for investigating, understanding and integrating

the human dimensions in the marine reserve process. In practical terms, marine reserves are a

management tool directed toward governing - and often changing - human behavior. To govern

human behavior, it is essential to identify patterns of behavior and what drives them, determine

how and why they are problematic, and develop appropriate mechanisms that can facilitate

change. Also, because people respond in diverse ways to management interventions, attention to

the social, cultural and economic factors that influence behavior is critical, to help insure desired

outcomes and prevent unintended negative consequences (Davis and Tisdell 1995). And, because

marine reserve goals and objectives are shaped by human perceptions, values and beliefs,

understanding these attributes is critical to effective marine reserve design and evaluation.

A second practical argument is that the design, implementation, management and evaluation

of marine reserves constitute a process that requires extensive information and other resources

over time. Obtaining this information through scientific research and existing managementrelated data collection can be prohibitively costly and difficult. Incorporating the human

dimension in this case involves making use of the diverse resources that participants can bring to

the marine reserve process including their knowledge and experience, insights into potential

management hurdles, and other resources (Crosby et al. 2000). Active engagement of

stakeholders in the marine reserve process provides opportunities to elicit this information and

address information gaps and misconceptions held by all stakeholders (i.e., managers, scientists,

resource users, environmentalists, the broader public). It thereby can facilitate the development

of mutual trust, shared goals and understanding. These, in turn, can enhance support for the

marine reserve process and voluntary compliance with management decisions (Fiske 1992).

In legal terms, federal and state mandates require evaluation of the social and economic

impacts of marine reserves. The overarching federal statutory requirement is found in the

National Environmental Policy Act (NEPA), which mandates the integrated use of the social

sciences in assessing impacts on the human environment through attention to "the natural and
76

physical environment and the relationship of people with that environment" (40 CFR 1508.14;

ICGP 1994). Federal environmental impact statements are likened to "full-disclosure

procedure[s] for federal decision-makers, who are then expected to consider the negative as well

as the positive implications of potential courses of action, and the unintended as well as the

intended consequences, before they proceed" (ICGP 1994).

Although marine reserves are often construed as a broad ocean management tool, they are

explicitly or implicitly a fishery management tool, because they directly affect fishing activity.

The 1976 Magnuson Fishery Conservation and Management Act (MFCMA) established legal

requirements for considering the human dimensions of fishery management, especially in the

national standards for fishery management [MFCMA section 301(a)]. The MFCMA requires the

use of "a systematic, interdisciplinary approach which will ensure the integrated use of the

natural and social sciences ... in planning and decision-making" [NEPA section 102(2)(a)] in the

preparation and implementation of federal fishery management plans (FMPs). Because marine

reserves are a "system for limiting access to the fishery in order to achieve optimum yield"

[MFCMA section 303(b)(6)], the Secretary of Commerce and fishery management councils are

required to consider their economic and social impacts. This requirement was strengthened by

1990 amendments to the MFCMA, which specified that FMPs must "assess, specify, and

describe the likely effects of conservation and management measures on participants in the

affected fishery, and the effects on participants in other fisheries that may be affected directly or

indirectly" [MFCMA section 303(a)(9)]. The 1996 Sustainable Fisheries Act added national

standard 8, which requires conservation and management measures to minimize adverse

economic impacts on fishing communities, consistent with the Act's conservation requirements

(NOAA 1997). Fishing communities are defined as “substantially dependent on or substantially

engaged in the harvest or processing of fishery resources to meet social and economic needs”

(NOAA 1997). As a result, ecological , economic and social impact assessments are necessary to

meet MFCMA and NEPA requirements, based on the rationale that "the more comprehensive the

information base and analysis, the more objective and defensible will be the decision-making

process" (ICGP 2001).

In California, legal mandates to consider the human dimensions in the design and use of

management tools such as marine reserves are found in the California Environmental Quality

Act (CEQA) and the Marine Life Management Act (MLMA). In passing CEQA, the Legislature

noted that "it is the policy of the state to ... require governmental agencies at all levels to consider

qualitative factors as well as economic and technical factors and long-term benefits and costs, in

addition to short-term benefits and costs and to consider alternatives to proposed actions

affecting the environment" [CEQA § 21001 (g)]. Following the MLMA, California FMPs are

required "to summarize information on economic and social factors in the fishery" [7080(e)]. If

an FMP includes new management measures, it must analyze their anticipated effects on

fishermen as well as coastal communities and businesses that rely on the fishery [7083(b)]. The

MLMA directs FMP developers and managers to minimize adverse impacts on small-scale
77

fisheries, coastal communities, and local economies [7056(j)] (although these concerns are

secondary to the broader mission of fostering sustainable fisheries), and to allocate increases or

restrictions on catches fairly among recreational and commercial fishermen [7072(c)] (Weber

and Haneman 2001).

Socio-economic considerations for the marine reserve process

The features of the human environment that are relevant to marine reserves include use

patterns; perceptions, attitudes and beliefs; costs and benefits; social and economic relationships;

and formal and informal institutions related to marine resources, their use and management.

Understanding these as they interact with the biophysical environment is critical to all stages of

the marine reserve process, and to the achievement of marine reserve goals and objectives.

Use patterns

Use patterns comprise the spatial and temporal aspects of human activities in the marine

environment. These activities include both consumptive (e.g., fishing, kelp harvesting, oil and

gas drilling) and non-consumptive (e.g., diving, whale-watching, shipping) activities. Use

patterns vary considerably within and among groups, across locations and over time, and are

influenced by a range of environmental, economic, socio-cultural and regulatory factors.

Research on use patterns associated with the four Central California marine reserves addressed in

this report has been done only for Big Creek Ecological Reserve (Pomeroy in press, Wilcox and

Pomeroy in press).

Understanding use patterns and the forces that underlie them is critical to the effective

design, management and evaluation of marine reserves (Walters 2000). Information on use

patterns can be used to identify potential marine reserves sites. For example, if a marine reserve

is to be used to protect a particular habitat type, information on use patterns can inform the

selection of sites that seek to minimize and equitably distribute the displacement of users while

maximizing habitat protection. Consideration of use patterns, and the factors that influence them,

is also important for anticipating potential changes in use patterns such as displacement of

consumptive users following the establishment of marine reserves. Such shifts in activity can

result in negative and perhaps unintended ecological, social and economic consequences such as

localized habitat damage and overfishing, conflict over access to open sites, increased financial

and safety risks as fishermen travel to more distant or hazardous areas, and undesired changes in

the distribution of costs and benefits of marine resource management.

Information on use patterns before and after marine reserve implementation is critical for

evaluating the effectiveness of marine reserves, in both ecological and socio-economic terms.

The permeability of marine reserve boundaries means that activities, conditions and outcomes

within a reserve (and changes therein) will influence those outside its boundaries, and vice verse
78

(Pomeroy 1999, Walters et al. 2000). For example, fishermen may aggregate at the edge of a

marine reserve, enabling them to take advantage of potential spillover, but precluding the

broader distribution or recruitment of fish from the reserve to the areas beyond the reserve edge,

as McClanahan and Kaunda-Arara (1996) documented in Kenya. They argue that fishermen

experienced high catches at the boundary of the marine reserve, but that the concentration of

fishing effort likely prevented the broader distribution of the spillover from the reserve into the

larger surrounding area. These ecological impacts also have socio-economic implications

through the differential distribution of costs and benefits among resource users who fish the

reserve edge and those who do not. Shifts in use patterns can have broader implications for

marine reserve effectiveness if they result in excessive concentration of fishing in areas that

remain open. The interaction between this human dimension and the ecological dimensions of

marine reserves is evident in the concern that locating a reserve in a sink area may increase

fishing pressure on source populations and thereby prompt or exacerbate population decline

(Crowder et al. 2000 in NRC 2001).

Knowledge of use patterns and how they change is also critical for evaluating the

effectiveness of marine reserves in socio-economic terms. Most marine reserves will displace

fishing effort, which in turn will have social and economic impacts on resource users and related

communities. Fishermen may find adequate substitute sites to replace those lost to marine

reserves, although their operating, safety and social psychological costs may increase as a result

of having to transit further to and from these sites. Because the quality of fishing sites varies in

time and space, not all areas that remain open will afford commensurate fishing opportunities

(Kenchington 1995, Hilborn 2001). Moreover, quality fishing areas that remain open may

become crowded, and result in social conflict and other increased costs of fishing. In the Central

California context, legal mandates require the evaluation of these costs and benefits and their

distribution, and the design of reserves that seek to allocate these equitably.

Non-consumptive use patterns also warrant attention. Understanding non-consumptive use

patterns prior to marine reserve establishment can inform site selection and design, and provide a

baseline against which to evaluate marine reserve outcomes in terms of desired recreation,

tourism and educational benefits. Monitoring shifts in these use patterns can facilitate detection

of problems such as crowding, which can diminish the social and economic benefits to these

users and result in ecological damage (Davis and Tisdell 1995, Boersma and Parrish 1999,

Crosby et al. 2000, Dobrzynski and Nicholson 2001).

Changes in use patterns on the water can have shoreside repercussions as well. Marine

reserves, to the extent they limit or displace fishing, may result in losses of fishery-related

activity for ice and bait providers, harbors (which depend on commercial activities to qualify for

federal dredging funds), and other providers of local goods and services. Losses of local fresh

fish supply can also negatively affect local restaurants, markets and consumers who may in turn

become dependent on external supplies of seafood. These losses may be offset - or more than
79

compensated for - by increases in non-consumptive activities. Alternatively, they may be

exacerbated if these activities are already near or at their ecological or socio-economic carrying

capacity (Davis and Tisdell 1995). A second issue is the extent to which increases in nonconsumptive uses can be supported by existing or latent infrastructure and providers of goods

and services. Rapid increases in the demand for such infrastructure, goods and services may

provide a welcome stimulus to coastal communities, or they may overtax them. Reserve-induced

shifts in fishing effort to other regions may generate demand for and stimulate - or overburden -

coastal communities in those areas. Across sites, these shifts also can lead to the redefinition of

community identity, for example, from fishing towns to recreation centers, or from non-fishing

towns to fish producing centers.

Perceptions, attitudes and beliefs

Perceptions, attitudes and beliefs are components of individuals' views of the world that

shape their preferences, choices and actions. In the context of marine reserves, perceptions,

attitudes and beliefs of particular interest include those related to the marine environment, its

management and use, as well as marine reserves per se (Fiske 1992, Crosby et al. 2000). Among

the four Central California reserves, directed research on perceptions, attitudes and beliefs has

only been done with regard to Big Creek Ecological Reserve (Pomeroy and Beck 1998, Pomeroy

in press).

Shared perceptions of a problem (or opportunity) in the marine environment, positive

attitudes toward management, and beliefs that marine reserves are an appropriate management

tool can translate into support for them. A process that fully engages stakeholders can engender

these, and a sense of ownership in the process and its outcomes (Crosby et al. 2000). Such

support facilitates, and reduces the costs of, design and implementation (Alder 2002).

Perceptions, attitudes and beliefs that differ may result in opposition to marine reserves that can

be costly (Fiske 1992, White et al. 1994, Crosby et al. 2000). At the same time, divergent

opinion and action can catalyze more careful planning and action to address these differences,

resulting in a more robust outcome in ecological, social and economic terms.

It is also critical to consider the diversity of expectations for marine reserves. Unrealistically

high expectations of marine reserve benefits can lead to dissatisfaction with marine reserve

outcomes, and unfounded rejection of marine reserves as a management tool. With knowledge of

the potential mismatch between expectations and likely outcomes, efforts can be made to inform

people and help make these expectations more realistic (Wolfenden et al. 1994). Acute concerns

about negative social and economic impacts and their distribution also require consideration

(Suman et al. 1999). Failure to address these can result in unrealistic expectations, opposition

and reduced compliance to marine reserves once they are established (Fiske 1992, Alder 1996,

Suman 1997, Suman et al. 1999).
80

Economic values, costs and benefits

Economics of marine reserves pertains to market and non-market values, benefits and costs,

and positive and negative impacts of marine reserves, which allocate access to marine places and

resources. Although marine reserves are theorized to generate substantial social and economic

benefits (Hannesson 1998, Sanchirico and Wilen 1999), limited attention has been directed

toward their systematic, empirical evaluation to date (Badalamenti 2000, Alder 2002). Where

such evaluations have been done, they have focused primarily on net economic impacts (e.g.,

Dixon et al. 1993), have been limited by a lack of adequate baseline information (e.g.,

McClanahan and Kaunda-Arara 1996, Roberts et al. 2001), or have not adequately addressed

interactions with factors such as other (non-reserve) regulations (NOAA et al. 2002). A notable

exception is the research on Apo and Sumilon Island marine reserves in the Philippines (Alcala

and Russ 1990; White et al. 1994, 2002; Russ and Alcala 1999). Study of the economics of the

four Central California marine reserves has been limited to work done at Big Creek (Pomeroy in

press), and that directed more toward the larger fishery rather than the Big Creek reserve per se.

Part of the challenge in assessing costs and benefits follows from the difficulty of assigning

market values to the intangible aspects of the marine environment. Economic costs, benefits and

impacts such as landings of fish or increases in non-consumptive diving may be readily

measured and assigned a market value. Other qualities such as the aesthetic or social

psychological value of fishing for sport, subsistence or income, or existence value (knowing that

an area in the marine environment is protected from direct human disturbance) are not readily

assigned a market value. Moreover, it is difficult to tease out the economic effects of marine

reserves from those of other forces in the marine environment and the interactions among them.

A separate and seldom considered economic dimension of marine reserves is the cost of the

marine reserve process, from design through evaluation. These costs include time and funding to

support agency and other stakeholder participation, the collection, analysis and reporting of

quality biophysical and social scientific information, and the integration of scientific and local

knowledge throughout the process. These costs are not trivial, especially given agencies' limited

staff, considerable workload, and often scant financial resources, and the fact that other

stakeholders often participate without compensation for the time, travel costs, foregone income

and the considerable knowledge and social resources that they contribute.

Understanding the costs and benefits of marine reserves, and the distribution of these

impacts, is essential to reserve design. Together with data on use patterns, this information can

be used to design marine reserve options that seek to minimize negative socio-economic impacts

while maximizing ecological and other socio-economic benefits. This approach was used by a

group of commercial fishermen involved in the Channel Islands Marine Reserves Working

Group (MRWG) process. In March 2001, several fishermen, each representing a different fishery

or gear type, held an informal meeting where they drafted a marine reserve scenario to present to
81

the MRWG. They drew upon their own ecological and social knowledge and information

generated during the MRWG process, and were guided by ecological and socio-economic goals

and objectives developed by the MRWG. The result of their efforts was a proposed network of

marine reserves that addressed several (but not all) of the MRWG’s ecological goals while

seeking to minimize costs to and distribute them equitably among commercial fishery sectors.

Although their proposal was not formally considered by the MRWG, a modified version is now

before the Fish and Game Commission as one of six marine reserve alternatives.

Social and economic relationships

Social and economic relationships consist of the linkages among those who use and

otherwise value the marine environment and associated shoreside socio-economic systems. Such

linkages are evident, for example, in the social, cultural and economic ties between (commercial,

recreational and subsistence) fishermen on the one hand, and receivers, processors, harbors,

support businesses and their communities on the other. Other users of the marine environment

likewise have extensive shoreside linkages. These linkages have been documented for the fishery

associated with the Big Creek reserve (Pomeroy and Beck 1998, Pomeroy in press), but not the

other three Central California reserves.

These linkages convey information that can be used to inform, support or oppose the marine

reserve process. While not all community members may be willing or able to participate in the

marine reserve process1, well positioned representatives can convey information between

members of their social and economic networks and other participants in the process. In addition,

the social and economic forces manifest in these relationships influence attitudes, perceptions,

beliefs, use patterns and other behaviors related to marine reserves. Attention to these

relationships (along with other socio-economic considerations) contributed to the successful and

locally supported designation of the Fagatele Bay National Marine Sanctuary in Samoa (Fiske

1992) and Apo Island Reserve in the Philippines (White et al. 1994). The neglect of these

relationships contributed to the failure to establish a National Marine Sanctuary at La Parguera,

Puerto Rico (Fiske 1992), to the failure of a marine managed area in St. Lucia (Sandersen and

Koester 2000), and to the demise (at least temporarily) of the reserve at Sumilon Island,

Philippines (Alcala and Russ 1990; White et al. 1994, 2002; Russ and Alcala 1999).

These relationships also constitute the network through which the positive and negative

impacts of marine reserves circulate. Inattention to these linkages can result in lost opportunities

to realize broad benefits, or in greater than expected costs as negative impacts reverberate among

individuals and communities.

 1

In all except the most localized situations, it is impractical for all of those with an interest in the marine

reserve process to participate in all aspects of it.
82

Formal and informal governance institutions

Institutions are the shared norms, rules and strategies that manifest the social and economic

relationships discussed in the previous section and govern individual and collective behavior

(Ostrom 1990). Formal institutions include the structures of government (e.g., legislatures,

agencies) and associated rules and regulations (e.g., NEPA, MFCMA). The jurisdictions,

mandates and actions of these institutions vary, overlap and at times conflict with one another.

Informal governance institutions include locally devised rules and shared understandings that

govern behavior complementary to, in the absence of, or in spite of formal government. Formal

and informal institutions may be complementary, compatible or contradictory, but inevitably

interact with one another. In the context of marine reserves, these interactions may facilitate and

support, or hinder and undermine the effectiveness of marine reserves (Fiske 1992, Johannes

1998, Pomeroy and Beck 1999, Pomeroy 1999). For example, the cooperative arrangement at

Big Creek Reserve, California, where the manager of the (terrestrial) Landels-Hill Big Creek

(LHBC) Reserve and local fishermen established an informal marine reserve, is an example of an

informal, local institution (Pomeroy and Beck 1999, Pomeroy 1999). It was formalized as a state

Marine Ecological Reserve in 1994, and is co-managed by the LHBC reserve manager (through

UC Santa Cruz) and the state Department of Fish and Game (DFG). A memorandum of

understanding between UC Santa Cruz and DFG provides for the LHBC reserve manager’s

oversight of day-to-day operations, and thereby recognizes the cooperative arrangement between

the fishermen and the reserve manager (Pomeroy and Beck 1999). The institutional arrangements

at the other three Central California reserves have not been studied.

Informal institutions can serve as mechanisms for support and management of marine

reserves, although not all localities have the capacity to fulfill these functions (Johannes 1998,

King and Faasili 1999). On the other hand, efforts to establish marine reserves may be perceived

as a threat (or contradictory) to local institutions, and lead to resistance or efforts to undermine

them (Fiske 1992, White et al. 1994, Alcala and Russ 2000). Knowledge and understanding of

the formal and informal institutions that govern resource management and use and how they do

or might interact with one another are important and often neglected components of the marine

reserve process. In the U.S., for example, different federal and state agencies have authority over

different places and activities in the marine environment. Legally and practically, efforts to

establish marine reserves must work within these institutional arrangements. Failure to do so can

result in incompatibilities between marine reserve design and existing management that

undermine the effectiveness of both. Rules that govern the use of marine resources, whether in

the form of marine reserves or traditional fishery management, may interact with one another to

limit the extent to which those affected can adapt (Pomeroy and Hunter 2001, Pomeroy in press).

Coordination of marine reserves with other management measures is important in order to guard

against excessive effort and user conflicts in areas that remain open (NRC 2002), and to help

insure that activities within marine reserves are conducted in a way that is consistent with their

goals and objectives.
83

Approaches and methods for incorporating the human dimension

Two general approaches are useful for incorporating socio-economic considerations into the

marine reserve process: public participation and social scientific research. These two

mechanisms differ in their utility and their limitations, but can complement one another.

Forms of public participation vary in terms of the nature and extent of communication

between government and the larger public. Common mechanisms used in marine resource

management include public hearings and the solicitation of written comments to obtain input and

feedback on proposed management actions. These methods are consistent with centralized, topdown systems in which government retains full authority and responsibility for management,

informing the public about its decisions once they have been made. However, several factors

have led to the growing interest in more cooperative forms of governance, and the use, for

example, of small group meetings and workshops in which government consults with nongovernment stakeholders, giving the latter a greater say in management decisions, with more

positive ecological and socio-economic outcomes (Berkes 1989, Pinkerton 1989, White et al.

1994, Johannes 1998). These forms of public participation can be effective for eliciting

perceptions, values, beliefs, local ecological knowledge, opinions and other information from a

broad range of stakeholders throughout the marine reserve process. They are also valued because

they can enhance communication, generate mutual understanding and trust, and provide social

and other resources to facilitate and enhance marine reserve design, management and evaluation.

However, public participation alone is not sufficient for addressing socio-economic

considerations, nor is it a substitute for social scientific research for marine reserves (ICGP 1994,

Cocklin et al. 1998). Public participation does not afford full representation of all relevant

stakeholders, nor does it provide systematic, reliable and valid information on the full range of

human dimensions topics discussed above. Social science research can address these limitations,

and provide critical information on socio-economic considerations for marine reserves.

Several social science approaches and methods can be used to inform the marine reserve

process. Archival research entails the systematic review and analysis of previous studies,

landings and other quantitative data, and other historical materials. It is useful for assessing

trends (e.g., in use patterns) and developing an understanding of the environmental, socioeconomic and regulatory context (e.g., relationships and institutions). Archival research alone,

however, usually cannot provide a complete understanding of these dimensions, because the data

were not generated with the marine reserve process in mind, and quickly become outdated. For

example, in the Dry Tortugas, Florida and the California Channel Islands marine reserve

processes, existing landings data (which illustrate the distribution of fishing effort) had been

collected at too gross a scale to enable meaningful evaluation of marine reserve alternatives.

Moreover, the data were insufficient to inform assessment of potential interactions between local
84

fisheries and marine reserves. In the Channel Islands case, a recent study of the social and

economic organization of the California market squid fishery (Pomeroy and FitzSimmons 2001)

afforded considerable information on general use patterns, social and economic relationships,

attitudes, perceptions and beliefs, informal and formal institutions and the economics of the

fishery. This work had also generated social resources (e.g., fishing industry participants' and

others' good will, active support and constructive and timely input into the research effort)

essential to designing and conducting research to inform the MRWG process, even with

inadequate time and funding. Similarly, archival information was insufficient to efforts to

understand the Big Creek reserve process and the reserve's performance. Directed field research

was needed to address these critical information needs.

An important complement to archival research, field research entails the systematic

collection and analysis of primary data using methods and tools such as surveys, ethnography

(i.e., in-depth interviews and observation) and focus groups. Each of these has its strengths and

weaknesses. Mail, phone and in-person surveys, which usually employ statistical sampling and a

highly structured set of questions to enable analytical generalization, are useful for collecting

quantitative data from large numbers of people (Yin 1989, Babbie 1998). While surveys are

viewed as being more cost-effective than other field data collection methods, they tend to

produce less valid and reliable information (Babbie 1998). Ethnography entails the use of indepth interviews and observation (Spradley 1980) to gain in-depth understanding of, in this case

marine reserves, in local context. Ethnographic research tends to produce more reliable and valid

results than survey research, but requires considerable time and effort (Spradley 1980). Focus

groups and workshops, which bring individuals together to discuss a well defined set of topics,

can generate information on shared understandings about issues and ideas for resolving them.

They are vulnerable, however, to the effects of interpersonal and group dynamics (Babbie 1998),

as when more vocal or powerful members of the group dominate the discussion and thereby

suppress input from others.

The results of social science data analyses are reported in a variety of ways, and may be

descriptive, explanatory or predictive of individual and collective processes and potential

behaviors. In the marine reserve process, case studies can be done to describe and explain current

(social, economic, political) situations and trends. Social and economic impact assessments that

build upon such contextual understanding can be used to predict and compare marine reserve

scenarios' potential and actual outcomes. These outcomes are expressed in terms of 1) absolute

or relative benefits and costs, in social, economic and ecological terms, 2) how behavior and

associated impacts change with reserve implementation, and 3) how these are distributed within

and among human and ecological systems. Cost-benefit analyses can be used to weigh the costs

and benefits of marine reserves prior to and following implementation. Cost-effectiveness

analyses can be used to select the least costly option to achieve a given set of objectives (Crosby

et al. 2000).
85

Social science research to address marine reserve information needs requires adequate time,

funding and social resources. The researcher uses these to build working relationships with study

participants, develop research tools, carry out quality data collection and analysis, and prepare

results that are responsive to management and public needs and concerns. In addition, the

collection, analysis and reporting of human dimensions data to inform marine reserve processes

raise social and ethical issues that must be addressed. Information on use patterns, economics

and other aspects of the human environment is personal, potentially proprietary and sensitive.

Social scientists are obligated to operate by principles of ethical research, whereby they insure

that participation is voluntary and anonymous, and individuals’ information remains confidential.

However, access to and use of this aggregated information for the marine reserve process raise

social and ethical issues that need to be explicitly addressed in the public arena.

Measures of effectiveness for marine reserves: The human dimensions

Just as it is vital to evaluate marine reserves for their ecological effectiveness, it is also

critical that they be evaluated for their socio-economic effectiveness. Effectiveness may pertain

to the marine reserve process itself (i.e., the means), or its outcome (i.e., the ends). Goals and

objectives of marine reserves in most cases include some combination of maintaining ecosystem

functions and conditions, maintaining (or increasing) resource abundance and diversity,

promoting sustainability (in ecological, social and economic terms), and providing opportunities

for (non-consumptive) recreation, tourism, education and scientific research.

Criteria for evaluation - or measures of effectiveness - should be simple, measurable, costeffective, and reflective of these goals and objectives (Alder et al. 2002). They should also be

clearly defined and understood in common by all participants. Alder et al. (2002) adapted

RAPFISH (Pitcher et al. 1998), a fishery management evaluation model that uses

multidimensional scaling, to develop a marine protected area (MPA) evaluation model

(MPAEM), and pilot tested it at 20 MPA sites. To apply the model, they identified and defined

attributes associated with MPAs such as maintenance of habitat, biodiversity and resources,

sustainable exploitation, economic benefits, and social features (e.g., equity, stewardship,

management resilience, efficiency). They then had marine reserve managers and scientists

knowledgeable of the test cases evaluate the MPAs on each attribute using an ordinal score to

indicate whether that attribute had improved, declined or remained unchanged since reserve

establishment. The ordinal scores on each of several attributes were then arrayed for comparison,

and combined to provide a qualitative evaluation of marine reserve effectiveness. Although

Alder et al. (2002) report that this pilot test demonstrated the potential utility of the model, they

caution that it still needs to be evaluated by other stakeholders involved in those MPA processes.

The authors note, however, that its utility for stakeholders other than managers and scientists has

yet to be demonstrated.
86

The foregoing discussion of the socio-economic considerations for marine reserves suggests

several attributes that could be integrated with Alder et al.’s MPAEM to evaluate marine reserve

effectiveness from a human dimensions perspective. It is possible to measure marine reserve

effectiveness by asking stakeholders and expert observers how they would rate the reserve(s) on

each attribute after marine reserves have been in place for some time. Although perceptions of

effectiveness are critical, a more valid and robust evaluation would also entail the development

of baseline human dimensions data well before marine reserve implementation, and continued

monitoring and consideration throughout the marine reserve process.

MPAEM constitutes one approach to marine reserve evaluation, and can stimulate more

directed attention to the need for tools and measures to evaluate MPA processes and outcomes

from a human dimensions perspective. There is a critical need for measures of effectiveness that

explicitly address both qualitative and quantitative phenomena and the interactions between

them, as well as the interactions between marine reserves’ effectiveness in achieving ecological

and socio-economic goals. In addition, these measures need to address marine reserve processes

and outcomes as they interact with and compare to other marine resource management strategies.

Summary and Conclusion

Commonly viewed as a management tool, marine reserves are also a management process

that requires careful consideration of a range of socio-economic, or more accurately, human

dimensions factors. Among these are how people and communities perceive, value and use the

marine environment, and the informal and formal rules and structures that influence these values

and uses. The rationale for attending to these is practical as well as legal. Marine reserves

generate social and economic impacts which both individuals and society must bear. Moreover,

human responses and adaptations to marine reserves have ecological as well as socio-economic

implications. Attention to the human dimensions can facilitate adequate planning to take

advantage of the opportunities and mitigate the challenges that marine reserves pose. NEPA and

other federal and state statutes provide the legal foundations and some guidance for addressing

the human dimensions for marine reserves in Central California.

Public participation and social science research are complementary means for addressing

social and economic considerations for marine reserves. Public participation is especially useful

for addressing social and economic concerns associated with marine reserves, and generating

trust, support and other social resources to assist in their design, management and evaluation. It

is not, however, a substitute for systematic collection and analysis of socio-economic

information to inform the process. The social sciences offer diverse methods that can be adapted

to particular stages in the process and associated information needs. Yet to date, limited social

science research has been done, with the majority of this work focused on evaluating marine

reserve processes to identify the necessary and sufficient conditions for the establishment of
87

marine reserves. Little research to inform and evaluate marine reserve design, implementation,

management and outcomes from a human dimensions perspective has been done. This is

especially the case in Central California, where directed social science research has only been

done on the Big Creek Ecological Reserve, and that hindered by a lack of adequate baseline

information.

Measures of effectiveness relevant to the human dimensions of marine reserves are not well

developed, even as the theory of marine reserves posits that social and economic benefits will

outweigh their costs. Basic questions about changes in use patterns, perceptions, attitudes and

beliefs, economic values, costs and benefits; social and economic relationships and institutions

can be used as the basis for developing such measures of effectiveness. Reliable and valid

answers to these questions, however, require baseline information on the human dimensions of

marine reserves, monitoring of these attributes throughout the marine reserve process, and

insuring their integration into management as it changes over time.

Acknowledgements

This chapter is based upon my recent social science research on the California squid (M.

FitzSimmons, Co-PI), wetfish (S. Hackett, co-PI) and salmon fisheries, the cooperative

arrangement at Big Creek Marine Ecological Reserve, and the potential impacts of marine

reserves at the Channel Islands. California Sea Grant, the Marine Ecological Reserves Research

Program, the National Marine Fisheries Service, NOAA's Special Projects Office, the California

Seafood Council, and UC Santa Cruz' Institute of Marine Sciences have provided funding

support for these projects. I'm grateful to the many fishing industry participants, resource

managers, other stakeholders and colleagues for their insights into fisheries, fishery management

and marine reserves. The material presented here also was informed by involvement in a variety

of marine reserve workshops and efforts to develop a social science research strategy for marine

protected areas. I thank Chris Dewees, Mark Helvey and Monica Hunter for their review and

helpful suggestions on this manuscript.
88

SUMMARY

The exceedingly small size of existing marine reserves in Central California prevents them

from achieving many of the goals and benefits attributed to marine reserves in the scientific

literature. The number of fish and invertebrates inhabiting existing reserves is small, compared

to the total population sizes of species in Central California. Existing reserves in Central

California protect a variety of shallow water habitats and species, but do not provide reserve

benefits for animals living in deeper water, unless they reside in existing reserves during a

portion of their life. The older marine reserves in Central California show some of the primary

benefits associated with protection from exploitation, including modest increases in size and

abundance of fishes, but it is difficult to assess the degree to which these benefits represent

pristine conditions. This is to be expected, as the primary fish species inhabiting these reserves

(rockfishes) are slow growing and exhibit sporadic recruitment. Also, new scientific theories

suggest that substantially altered habitats may or may not return to pre-existing states after the

disturbance has been removed.

Marine reserves in other temperate and tropical oceans, and theoretical models of marine

reserves, show substantial conservation and some potential fishery benefits. For these reasons,

we expect marine reserves created in Central California for conservation purposes would accrue

many of the benefits predicted by reserve theory. The extent to which reserves in Central

California would successfully benefit fisheries, however, would depend on a large number of

social and biological factors, such as social acceptance of reserves, fishery effort shifts, catch

regulations, enforcement levels, the proportion of a stock protected in a reserve, rates of

movement and larval production of protected species, and reserve size and location. Currently,

only a small proportion of fished species are protected in reserves. To be an effective fishery

management tool, more area would need to be placed in reserve status, but not so much as to

preclude viable fisheries. If marine reserves are to be developed and successfully used in Central

California as a tool for fisheries management, however, they will need to be integrated into

existing fishery management processes. A structured and well-supported monitoring program,

which clearly identifies a set of effectiveness parameters, will also need to be established to

measure how well reserves achieve stated objectives.

Effective natural resource management requires public participation and buy-in to

management goals, objectives, and regulations. Thus, just as it is vital to evaluate marine

reserves for their ecological effectiveness, it is also critical that they be evaluated for their socioeconomic values. In this respect, the use of marine reserves is a public policy decision that must

be made with consideration of human activities. For marine reserves to be an effective public

policy tool in Central California, human use patterns, perceptions, attitudes, and beliefs will need

to be incorporated into the design process. Information about social and economic costs and

benefits should also be incorporated to maximize the effectiveness of a reserve system.

Ultimately, an understanding of how people interact with the biophysical environment is integral

to the design and development of marine reserve goals and objectives.
89

ACKNOWLEDGEMENTS

The Monterey Bay National Marine Sanctuary provided funds to produce this report. We

thank Andrew DeVoegelaere and Erica Burton of the MBNMS research program for their

guidance and support for producing this report. We also thank the home institutions of the

authors for providing the support for the authors to be involved in this project. This work was

supported in part by the University of California Sea Grant Extension Program, the University of

California Santa Cruz, the University of California Santa Barbara, the U.S. Geological Survey,

the California Department of Fish and Game, the National Marine Fisheries Service, and Moss

Landing Marine Laboratories, and the Friends of Moss Landing Marine Labs. Support was also

provided for M. Carr, C. Syms, and J. Caselle by the Partnership for Interdisciplinary Studies of

Coastal Oceans (PISCO), funded by the David and Lucile Packard Foundation.

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APPENDIX 1

Additional Literature Associated

With California Marine Protected Areas

Hopkins Marine Life Refuge

Brawley, S.H. 1989. Factors affecting recruitment of Fucus distichus: Timing of fertilization and

polyspermy. Journal of Phycology. 25(2)suppl:16.

Cooper,J., Wieland,M., and A.Hines. 1977. Subtidal abalone populations in an area inhabited by sea

otters. Veliger. 20(2):163-167.

DeBevoise, A.E. 1975. Predation on the chiton Cyanoplax hartwegii (Mollusca: Polyplacophora).

Veliger. 18(Suppl.):47-50.

Fadlallah, Y.H. 1982. Reproductive ecology of the coral Astrangia lajollaensis: Sexual and asexual

patterns in a kelp forest habitat. Oecologia. 55(3):378-388.

Holts, L.J. and K.A. Beauchamp. 1993. Sexual reproduction in the corallimorpharian sea anemone

Corynactis californica in a Central California kelp forest. Marine Biology. 116(1):129-136.

Pearse, J.S. and A.H. Hines. 1987. Long-term population dynamics of sea urchins in a Central

California kelp forest: Rare recruitment and rapid decline. Marine Ecology Progress Series.

39(3):275-283.

Lowry, L.F. and J.S. Pearse. 1973. Abalones and sea urchins in an area inhabited by sea otters. Marine

Biology. 23(3):213-219.

Lyman, B.W. 1975. Activity patterns of the chiton Cyanoplax hartwegii (Mollusca: Polyplacophora).

Veliger. 18(Suppl.):63-69.

Miller, D.J. and J.J. Geibel. 1973. Summary of blue rockfish and lingcod life histories; a reef

ecology study; and giant kelp, Macrocystis pyrifera, experiments in Monterey Bay,

California. Calif. Dept. Fish and Game Fish Bulletin 158. 137 pp.

Paddack, M.J. and J.A. Estes. 2000. Kelp forest fish populations in marine reserves and

adjacent exploited areas of Central California. Ecological Applications 10(3):855-870.

Russo, A.R. 1984. Space partitioning within populations of sea anemones (genus: Anthopleura) in the

California intertidal zone. Internationale Revue der gesamten Hydrobiologie. 69(4):521-528.

Sagarin, R.D., Barry, J.P., Gilman, S.E., and C.H. Baxter. 1999. Climate-related change in an intertidal

community over short and long time scales. Ecological Monographs. 69(4):465-490.
110

Schaeffer, T.N., Foster M. S., Landrau, M.E. Walder, R.K. 1999. Diver disturbance in kelp forests.

California Fish & Game. 85(4). 170-176.

Seiff, S.R. 1975. Predation upon subtidal Tonicella lineata of Mussel Point, California (Mollusca:

Polyplacophora). Veliger. 18(Suppl.):54-56.

Smith, A.M. 1992. Alternation between attachment mechanisms by limpets in the field. Journal of

Experimental Marine Biology and Ecology. 160(2):205-220.

Tomanek, L. and G.N. Somero. 1997. The effect of temperature on protein synthesis in snails of the

genus Tegula from the sub- and intertidal zone. American Zoologist. 37(5):188A.

Tomanek, L. and G.N. Somero. 1998. Features of a lethal heat shock: Impairment of synthesis of heat

shock proteins 70 and 90 during recovery in snails of the genus Tegula from the sub- and intertidal

zone. American Zoologist. 38(5):159A.

Watanabe, J.M., Cox, L.R. 1975. Spawning behavior and larval development in Mopalia lignosa and

Mopalia muscosa (Mollusca: Polyplacophora) in Central California. Veliger. 18(Suppl.):18-27.

Williams, R. 1975. Nitrogenous materials released from Mopalia muscosa (Gould, 1846), an intertidal

chiton. Veliger. 18(Suppl.):128.

Unpublished references related Hopkins Marine Life Refuge:

Ammann, A. J. 2001. Evaluation of a standard monitoring unit for the recruitment of fish in

Central California, M.A. Thesis, University of California, Santa Cruz. 92 pp.

Paddack, M.J. 1996. The influence of marine reserves upon rockfish populations in Central

California kelp forests. M.S. Thesis, University of California, Santa Cruz. 40 pp.

Pacific Grove Marine Gardens Fish Refuge

Nelson, P. A. 2001. Behavioral ecology of the young-of-the-year kelp rockfish, Sebastes

atrovirens Jordan and Gilbert (Pisces: Scorpaenidae). Journal of Experimental Marine

Biology and Ecology 256:33-50.

Unpublished references related to Pacific Grove Marine Gardens Fish Refuge:

Lea, R.N. 1978. California Dept. of Fish and Game Cruise Report 78-KN-12. Central California

Marine Sportfish Survey DJ F25R 6 pp. (hook-and-line surveys)

Lea, R.N. 1979. California Dept. of Fish and Game Cruise Report 79-X-3. Central California

Marine Sportfish Survey DJ F25R 7 pp. (hook-and-line surveys)
111

Miller, D. J., J. J. Geibel, and J.L. Houk. 1974. Results of the 1972 skindiver assessment

survey. Pismo Beach to Oregon. Calif. Dept. Fish and Game, Marine Resources Technical

Report No. 23: 61 pp.

Pearse, J.S. and L.F. Lowry. 1974. An annotated species list of the benthic algae and

invertebrates in the kelp forest community at Point Cabrillo, Pacific Grove, California.

Coastal Marine Laboratory, University of California, Santa Cruz. Tech Rep. 1: 73 p.

VenTresca, D. A. Summary report of 1960 Central California Council of Divers Free-diving

competition spearfish meet in Point Pinos area. California Dept. Fish and Game, Monterey.

VenTresca, D. A., J. L. Houk, et al. 1996. Early life history studies of nearshore rockfishes and

lingcod off Central California, 1987-92. Calif. Dept. fish and Game Mar. Res. Div. Admin.

Rept. 96-4:77.

Wilson-Vandenberg, D. 1987-1998. Unpublished, confidential site-specific location data of

catch and catch-per-unit effort from onboard observations of Commercial Passenger Fishing

Vessels (CPFV- recreational fishing) from Point Año Nuevo to Purisima Point, including

Pacific Grove area. Primarily rockfish and lingcod catches. Summaries available, with

location data combined, as Administrative Reports, Calif. Dept. Fish and Game, Monterey.

Carmel Bay

Crane, N.L., and K. Lashkari. 1996. Sound production of gray whales, Eschrichtius robustus,

along their migration route: A new approach to signal analysis. Journal of the Acoustical

Society of America. 100(3):1878-1886.

Cripe, D. 1998. Occurrence of xanthic grass rockfish, Sebastes rastrelliger. California Fish and

Game. 84(2):100-101.

Dingler, J.R. 1981. Stability of a Very Coarse-Grained Beach at Carmel, California. Marine

Geology. 44(3/4):241-252.

Dingler, J.R., Anima, R.J., and D.W. Hirschaut. 1982. Underwater geologic research around the

head of Carmel Submarine Canyon, Carmel Bay, California. Proceedings Of The Sixth

Meeting Of The United States - Japan Cooperative Program In Natural Resources (Ujnr).

Panel On Diving Physiology And Technology, Menlo Park And Catalina Island, California,

October 20-26, 1981, Pp. 79-89.

Dingler, J.R. and R.J. Anima. 1989. Subaqueous grain flows at the head of Carmel Submarine

Canyon, California. Journal of Sedimentary Petrology. 59(2):280-286.

Gibson, M.E. 1983. Carmel Bay. Marine oasis in the cold Pacific. Sea Frontiers. 29(3):130-

138.
112

Graham, M.H. 1996. Effect of high irradiance on recruitment of the giant kelp Macrocystis

(Phaeophyta) in shallow water. Journal of Phycology. 32(6):903-906.

Hallacher, L.E. 1977. On feeding behavior of the basking shark, Cetorhinus maximus. Environ.

Biol. Fish. 2(3):297-298.

Hallacher, L.E. 1984. Relocation of original territories by displaced black-and-yellow rockfish,

Sebastes chrysomelas, from Carmel Bay, California. California Fish and Game. 70(3):158-

162.

Hallacher, L.E. and D.A. Roberts. 1985. Differential utilization of space and food by the

inshore rockfishes (Scorpaenidae:Sebastes) of Carmel Bay, California. Environmental

Biology of Fishes. 12(2):91-110.

Hoelzer, G. A. 1988. Juvenile movement patterns in a territorial scorpaenid fish before and

during settlement. Marine Ecology Progress Series 45:193-195.

Kenner, M.C. and M.T. Lares. 1991. Size at first reproduction of the sea urchin

Strongylocentrotus purpuratus in a Central California kelp forest. Marine Ecology Progress

Series. 76(3):303-306.

Konar, B. and M.S. Foster. 1992. Distribution and recruitment of subtidal geniculate coralline

algae. Journal of Phycology. 28(3):273-280.

Reed, D.C. and M.S. Foster. 1984. The effects of canopy shading on algal recruitment and

growth in a giant kelp forest. Ecology. 65(3):937-948.

Schlining, K.L. and J.D. Spratt. 1999. Assessment of the Carmel Bay spot prawn (Pandalus

platyceros Brandt 1851) resource and trap fishery adjacent to an ecological reserve in Central

California. pp. 751-762 in Von Vaupel Klein, J.C. and F.R. Schram (ed.) The biodiversity

crisis and crustacea: Proceedings of the Fourth International Crustacean Congress,

Amsterdam, The Netherlands, July 20-24, 1998. A.A. Balkema, Rotterdam/Brookfield.

Singer, M.M. Food habits of juvenile rockfishes (Sebastes) in a Central California kelp forest.

Fishery Bulletin. 83(4):531-542.

VanWagenen, R.F., Foster, M.S., and F. Burns. 1981. Sea Otter Predation on Birds Near

Monterey, California. Journal of Mammalogy. 62(2):433-434.

Watanabe, J.M. and C. Harrold. 1991. Destructive grazing by sea urchins Strongylocentrotus

spp. in a Central California kelp forest: Potential roles of recruitment, depth, and predation.

Marine Ecology Progress Series. 71(2):125-141.

Wedi, S.E. and D.F. Dunn. 1983. Gametogenesis and reproductive periodicity of the subtidal

sea anemone Urticina lofotensis (Coelenterata: Actiniaria) in California. Biological Bulletin,

Marine Biological Laboratory, Woods Hole. 165(2):458-472.
113

Unpublished references related to Carmel Bay Ecological Reserve:

Ammann, A. J. 2001. Evaluation of a standard monitoring unit for the recruitment of fish in

Central California, M.A. Thesis, University of California, Santa Cruz. 92 pp.

Carr, M. H. 1983 . Spatial and Temporal Patterns of Recruitment of Young-of-the-Year

Rockfishes (Genus Sebastes) into a Central California Kelp Forest, San Francisco State

University. M.A. Thesis. 104 pp.

Lea, R.N. 1978. California Dept. of Fish and Game Cruise Report 78-KN-12. Central California

Marine Sportfish Survey DJ F25R 6 pp. (hook-and-line surveys)

Lea, R.N. 1979. California Dept. of Fish and Game Cruise Report 79-X-3. Central California

Marine Sportfish Survey DJ F25R 7 pp. (hook-and-line surveys)

Lea, R.N. 1979. California Dept. of Fish and Game Cruise Report 79-A-9. Central California

Marine Sportfish Survey DJ F25R 6 pp. (hook-and-line surveys)

Lea, R.N. and F. Henry. 1980. California Dept. of Fish and Game Cruise Report 80-X-5.

Central California Marine Sportfish Survey DJ F25R 4 pp. (hook-and-line surveys)

Lea, R.N. D. VenTresca, and R. McAllister. 1982. California Dept. of Fish and Game Cruise

Report 82-KB-10. Central California Marine Sportfish Survey. 7 pp. (hook-and-line surveys)

Malone, C. 1994. Temporal comparison of the intertidal biota of the Landels-Hill Big Creek

Reserve and spatial comparison of the reserve with three other Central Californian sites:

Oystercatcher Point, Carmel Point, and Natural Bridges. Senior Thesis, University of

California, Santa Cruz, 89 pp.

Miller, D. J., J. J. Geibel, and J.L. Houk. 1974. Results of the 1972 skindiver assessment

survey. Pismo Beach to Oregon. Calif. Dept. Fish and Game, Marine Resources Technical

Report No. 23: 61 pp.

Schlining, K. L. 1999. The spot prawn (Pandalus platyceros Brandt 1851) resource in Carmel

Submarine Canyon, California: aspects of fisheries and habitat associations. M. S. Thesis,

Moss Landing Marine Laboratories, Moss Landing, California.

Singer, M. M. 1982. Food Habits and Activity Patterns of Juvenile Rockfishes (Sebastes) in a

Central California Kelp Forest. M.A. Thesis, San Jose State University, San Jose, California,

75 pp.

VenTresca, D.A. 1961-1963, 1965, 1968, 1980-1984, 1986-2001. Summary reports of Central

California Council of Divers Free-diving competition spearfish meets in Carmel Bay.

California Dept. Fish and Game, Monterey.
114

Wilson-Vandenberg, D. 1987-1998. Unpublished, confidential site-specific location data of

catch and catch-per-unit effort from onboard observations of Commercial Passenger Fishing

Vessels (CPFV- recreational fishing) from Point Año Nuevo to Purisima Point, including

Carmel Bay area. Primarily rockfish and lingcod catches. Summaries available, with

location data combined, as Administrative Reports, Calif. Dept. Fish and Game, Monterey.

Point Lobos Ecological Reserve

Baldridge, A. and L.L. Rogers. 1991. Harbor seal predation on a wolf-eel. California Fish and

Game. 77(4):210-211.

Cronin, B.T. and R.B. Kidd. 1998. Heterogeneity and lithotype distribution in ancient deep-sea

canyons: Point Lobos deep-sea canyon as a reservoir analogue. Sedimentary Geology 115(1-

4):315-349.

Drury, A. 1970. Point Lobos State Reserve. Dept. of Parks and Recreation, State of California.

Gingras M.L., VenTresca D.A., Donnellan M.D., and J.L. Fisher. 1998. First observations of

vermilion rockfish courtship are from a harvest refuge. California Fish & Game. 84(4):176-

179.

Hanggi, E.B. and R.J. Schusterman. 1994. Underwater acoustic displays and individual

variation in male harbour seals, Phoca vitulina. Animal Behaviour. 48(6):1275-1283.

Lander M.E., Gulland, F.M., and R.L. DeLong. 2000. Satellite tracking a rehabilitated

Guadalupe fur seal (Arctocephalus townsendi). Aquatic Mammals. 26(2):137-142.

Long, D.J. 1992. Confirmation of the northern range of the snubnose sculpin (Orthonopias

triacis). California Fish and Game. 78(4):160-162.

Nichols, D.R. 1974. A marine survey of the north shore of the Point Lobos State Reserve. Beta

Research Oceanographic Laboratories.

Osborn, D.A. 2000. Cnidarian "parasites" on Solmissus incisa, a narcomedusan. Scientia

Marina. 64(Supplement 1):157-163.

Paddack, M.J. and J.A. Estes. 2000. Kelp forest fish populations in marine reserves and

adjacent exploited areas of Central California. Ecological Applications 10(3):855-870.

Riedman, M.L. and J.A. Estes. 1988. Predation on seabirds by sea otters. CAN. J. ZOOL.

66(6):1396-1402.

VenTresca, D.A., M.L. Gingras, M. Donnellan, J. Fisher, B. Brady, H. King, C. King, and D.

Osorio. (in preparation). Density and length estimates of some kelp forest fishes within and

adjacent to Central California marine protected areas. CalCOFI Reports 2002.
115

Unpublished references related to Point Lobos Ecological Reserve:

Castleton, M. R. 2000. Depth and substrate preference of preadult cabezon (Scorpaenichthys

marmoratus) in Point Lobos Marine Reserve. Capstone Project Paper, faculty of Earth

Systems Science and Policy, Center for Science, Technology, and Information Resources,

California State University, Monterey Bay. 28 pp.

DeMartini, J.D. and W.J. Barry. 1977. A subtidal resources inventory for Point Lobos State

Reserve. Report submitted to Dept. Parks and Recreation as a resource element of the

General Plan. 52 p.

Lea, R.N. 1978. California Dept. of Fish and Game Cruise Report 78-KN-12. Central California

Marine Sportfish Survey DJ F25R 6 pp. (hook-and-line surveys)

Lea, R.N. 1979. California Dept. of Fish and Game Cruise Report 79-X-3. Central California

Marine Sportfish Survey DJ F25R 7 pp. (hook-and-line surveys)

Lea, R.N. 1979. California Dept. of Fish and Game Cruise Report 79-A-9. Central California

Marine Sportfish Survey DJ F25R 6 pp. (hook-and-line surveys)

Lea, R.N. 1982. California Dept. of Fish and Game Cruise Report 82-KB-19. Central California

Marine Sportfish Survey. 6 pp. (hook-and-line surveys)

Lea, R.N. 1993. California Dept. of Fish and Game Cruise Report 93-M-5 Leg 2. Central

California Marine Sport Fish Project Biological Investigations. 9 pp. (hook-and-line surveys)

Paddack, M.J. 1996. The influence of marine reserves upon rockfish populations in Central

California kelp forests. M.S. Thesis, University of California, Santa Cruz. 40 pp.

Reilly, P.N., D.A. VenTresca, and M.L. Gingras. 1998. Use of marine reserves to enhance

nearshore sport fish populations . Annual Job Performance Report, Federal Aid in Sport Fish

Restoration Act. Grant Agreement F-50-R-10. California Dept. of Fish and Game. Project 5,

Job 1, 7 pp.

Reilly, P.N., and D.A. VenTresca. 1999. Use of marine reserves to enhance nearshore sport fish

populations. Final Performance Report, Federal Aid in Sport Fish Restoration Act. Grant

Agreement F-50-R-11. California Dept. of Fish and Game. Project 5, Job 1, 28 pp.

Julia Pfeiffer Burns State Park

Unpublished references related to Julia Pfeiffer Burns State Park:

Burdett, K.S., A.L Wagner, and J.S. Oliver. 1990. Biological survey of subtidal marine

communities in Julia Pfeiffer Burns State Park. Moss Landing Marine Laboratories.
116

Prepared for State of California Department of Parks and Recreation, Sacramento. 25 p. plus

attachments.

Lea, R.N. 1979. California Dept. of Fish and Game Cruise Report 79-X-3. Central California

Marine Sportfish Survey DJ F25R 7 pp. (hook-and-line surveys)

Lea, R.N. 1979. California Dept. of Fish and Game Cruise Report 79-A-9. Central California

Marine Sportfish Survey DJ F25R 6 pp. (hook-and-line surveys)

Lea, R.N. D. VenTresca, and R. McAllister. 1982. California Dept. of Fish and Game Cruise

Report 82-KB-10. Central California Marine Sportfish Survey. 7 pp. (hook-and-line surveys)

Seltenrich, C.P. J.D. Martini, and J. Barry. 1980. Water quality monitoring report No. 80-4:

California Marine Waters Areas of Special Biological Significance Reconnaissance Survey

Report, Julia Pfeiffer Burns Underwater Park. State Water Resources Control Board,

Sacramento. 76 p.

Big Creek Ecological Reserve

Ferguson, A. (ed.) 1984. Intertidal plants and animals of the Landels-Hill Big Creek Reserve,

Monterey County, California” Publication No. 14, Environmental Field Program 14,

University of California, Santa Cruz. 106 pp.

Norris, R. 1985. Geology of the Landels-Hill Big Creek Reserve : Monterey County, California.

Environmental Field Program, University of California, Santa Cruz, 1985.

Paddack, M.J. and J.A. Estes. 2000. Kelp forest fish populations in marine reserves and

adjacent exploited areas of Central California. Ecological Applications. 10(3):855-870.

Pomeroy, C. 1999. Social considerations for marine resource management: Evidence from Big

Creek Ecological Reserve. Reports of California Cooperative Oceanic Fisheries

Investigations. 40:118-127.

Pomeroy, C. and J. Beck. 1998. Cooperative management of the state's marine ecological

reserves: Preliminary evidence from Big Creek. Taking a Look at California's Ocean

Resources: An Agenda for the Future, ASCE, Reston, VA (USA). 1:105-116.

VenTresca, D.A., M.L. Gingras, M. Donnelan, J. Fisher, B. Brady, H. King, C. King, and D.

Osorio. (in preparation). Density and length estimates of some kelp forest fishes within and

adjacent to Central California marine protected areas. CalCOFI Reports 2002.

VenTresca, D.A., Gingras, M.L., Ugoretz, J., Voss, A., Blair, S., Plant, J., Hornady, R., and C.

Yoshiyama. 1998. The potential of marine reserves to enhance fisheries. Taking a Look at

California's Ocean Resources: An Agenda for the Future, ASCE, Reston, VA (USA). 1:400-

411.
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Yoklavich, M. G. Cailliet, R.N. Lea, H.G. Greene, R. Starr, J. DeMarignac, and J. Field. In

preparation. Deepwater habitat and fish resources associated with a marine reserve:

implications for fisheries. CalCOFI Reports (research in and adjacent to Big Creek

Ecological Reserve).

Yoklavich, M., Starr, R., Steger, J., Greene, H.G., Schwing, F., and C. Malzone. 1997. Mapping

benthic habitats and ocean currents in the vicinity of Central California's Big Creek

Ecological Reserve. National Marine Fisheries Service. 1997.

Unpublished references related to Big Creek Ecological Reserve:

Gingras, M.L. 1997. California Dept. of Fish and Game Cruise Report 97-M-8. Central

California Marine Sport Fish Project. 5 pp. (scuba surveys)

Gingras, M.L. 1998. California Dept. of Fish and Game Cruise Report 98-M-4. Central

California Marine Sport Fish Project. 6 pp. (scuba surveys)

Gingras, M.L. 1998. California Dept. of Fish and Game Cruise Report 98-M-5. Central

California Marine Sport Fish Project. 3 pp. (scuba surveys)

Houk , J.L. 1994. California Dept. of Fish and Game Cruise Report 94-M-12. Central California

Marine Sport Fish Project. 6 pp. (scuba surveys)

Lea, R.N. D. VenTresca, and R. McAllister. 1982. California Dept. of Fish and Game Cruise

Report 82-KB-10. Central California Marine Sportfish Survey. 7 pp. (hook-and-line surveys)

Lea, R.N. 1982. California Dept. of Fish and Game Cruise Report 82-KB-19. Central California

Marine Sportfish Survey. 6 pp. (hook-and-line surveys)

Lea, R.N. 1993. California Dept. of Fish and Game Cruise Report 93-M-5 Leg 2. Central

California Marine Sport Fish Project Biological Investigations. 9 pp. (hook-and-line surveys)

Lea, R.N. and P.N. Reilly. 1999. Biological studies utilizing research submersibles. Final

Performance Report, Federal Aid in Sport Fish Restoration Act. Grant Agreement F-50-R-11.

California Dept. of Fish and Game. Project 24, Job 3, 3 pp.

Malone, C. 1994. Temporal comparison of the intertidal biota of the Landels-Hill Big Creek

Reserve and spatial comparison of the reserve with three other Central Californian sites:

Oystercatcher Point, Carmel Point, and Natural Bridges. Senior Thesis, University of

California, Santa Cruz, 89 pp.

Paddack, M.J. 1996. The influence of marine reserves upon rockfish populations in Central

California kelp forests. M.S. Thesis, University of California, Santa Cruz. 40 pp.
118

Pattison, C.A. 1995. California Dept. of Fish and Game Cruise Report 95-M-11. Central

California Marine Sport Fish Project. 10 pp. (scuba surveys)

Pomeroy, C. 1996. An evaluative study of cooperative data collected at Big Creek. : Project

summary, University of California, California Sea Grant College: 18 pp.

Reilly, P.N., D.A. VenTresca, and J.L. Houk. 1994. Determination of the feasibility of using

marine reserves to enhance nearshore sport fish populations, using non-destructive, long-term

sampling methodologies. Annual Job Performance Report, Federal Aid in Sport Fish

Restoration Act. Grant Agreement F-50-R-6. California Dept. of Fish and Game. Project 2,

Study 1, Job 1, 18 pp.

Reilly, P.N., D.A. VenTresca, and C.A. Pattison. 1995. Use of marine reserves to enhance

nearshore sport fish populations. Annual Job Performance Report, Federal Aid in Sport Fish

Restoration Act. Grant Agreement F-50-R-7. California Dept. of Fish and Game. Project 5,

Job 1, 8 pp.

Reilly, P.N., D.A. VenTresca, and C.A. Pattison. 1996. Use of marine reserves to enhance

nearshore sport fish populations. Annual Job Performance Report, Federal Aid in Sport Fish

Restoration Act. Grant Agreement F-50-R-8. California Dept. of Fish and Game. Project 5,

Job 1, 12 pp.

Reilly, P.N., D.A. VenTresca, and M.L. Gingras. 1997. Use of marine reserves to enhance

nearshore sport fish populations. Annual Job Performance Report, Federal Aid in Sport Fish

Restoration Act. Grant Agreement F-50-R-9. California Dept. of Fish and Game. Project 5,

Job 1, 6 pp.

Reilly, P.N., D.A. VenTresca, and M.L. Gingras. 1998. Use of marine reserves to enhance

nearshore sport fish populations. Annual Job Performance Report, Federal Aid in Sport Fish

Restoration Act. Grant Agreement F-50-R-10. California Dept. of Fish and Game. Project 5,

Job 1, 7 pp.

Reilly, P.N., and D.A. VenTresca. 1999. Use of marine reserves to enhance nearshore sport fish

populations. Final Performance Report, Federal Aid in Sport Fish Restoration Act. Grant

Agreement F-50-R-11. California Dept. of Fish and Game. Project 5, Job 1, 28 pp.

Reilly, P.N., D.A. VenTresca, and D.A. Osorio. 2000. Determine the feasibility of using marine

reserves for enhancing nearshore fish populations. Annual Performance Report, Federal Aid

in Sport Fish Restoration Act. Grant Agreement F-50-R-12. California Dept. of Fish and

Game. Project 13, Job 2, 8 pp.

Smiley, J. 2000. Big Sur hook and line fishing survey, 1991-1999. (area adjacent to Big Creek

Ecological Reserve). 5 pp.
119

VenTresca, D., J. Fisher, M. Donnellan, and B.C. Brady. 1999. California dept. Fish and Game

Cruise Report 99-M-8 and 99-M-9. Central California Marine Sport Fish Project. 6 pp.

(scuba surveys)

VenTresca, D.A., M.L. Gingras, M. Donnellan, J. Fisher, and B. Brady. Kelp forest fish density

estimates from a priori habitat-stratified sampling: implications for efficient stock

assessment. Unpublished report, Calif. Dept. Fish and Game, Monterey.

Wilson, C.E.. 1996. California Dept. of Fish and Game Cruise Report 96-M-5. Central

California Marine Sport Fish Project. 11 pp. (scuba and hook-and-line surveys)

Yoklavich, M., R. Starr, J. Steger, H.G. Greene, F. Schwing, and C. Malzone. 1997. Mapping

benthic habitats and ocean currents in the vicinity of Central California's Big Creek

Ecological Reserve. NOAA Technical Memorandum NMFS, NOAA-TM-NMFS-SWFSC245. 52 pp.

Atascadero Beach Pismo Clam Preserve

Pattison, Christine. 2001. Pismo Clam. In: California’s Living Marine Resources: A Status

Report. Ed. by William S. Leet, Christopher M. Dewees, Richard Klingbeil and Eric J.

Larson. University of California Agriculture and Natural Resources Publication SG 01-11

Unpublished references related to Atascadero Beach Pismo Clam Preserve:

Pattison, C.P. California Dept. Fish and Game, Morro Bay. Annual surveys for Pismo clam

relative abundance and size frequency, 1985-2000.

Morro Beach Pismo Clam Reserve

Pattison, Christine. 2001. Pismo Clam. In: California’s Living Marine Resources: A Status

Report. Ed. by William S. Leet, Christopher M. Dewees, Richard Klingbeil and Eric J.

Larson. University of California Agriculture and Natural Resources Publication SG 01-11

Unpublished references related to Morro Beach Pismo Clam Reserve:

Pattison, C.P. California Dept. Fish and Game, Morro Bay. Annual surveys for Pismo clam

relative abundance and size frequency, 1985-2000.
120

Pismo Invertebrate Preserve

Pattison, Christine. 2001. Pismo Clam. In: California’s Living Marine Resources: A Status

Report. Ed. by William S. Leet, Christopher M. Dewees, Richard Klingbeil and Eric J.

Larson. University of California Agriculture and Natural Resources Publication SG 01-11.

Unpublished references related to Pismo Invertebrate Preserve:

Pattison, C.P. California Dept. Fish and Game, Morro Bay. Annual surveys for Pismo clam

relative abundance and size frequency, 1977-2000.

Pismo-Oceano Pismo Clam Preserve

Pattison, Christine. 2001. Pismo Clam. In: California’s Living Marine Resources: A Status

Report. Ed. by William S. Leet, Christopher M. Dewees, Richard Klingbeil and Eric J.

Larson. University of California Agriculture and Natural Resources Publication SG 01-11.

Unpublished references related to Pismo-Oceano Pismo Clam Preserve:

Pattison, C.P. California Dept. Fish and Game, Morro Bay. Annual surveys for Pismo clam

relative abundance and size frequency, 1985-2000.

Vandenberg MRPA Ecological Reserve

Watson, W., Charter, R.L., Moser, H.G., Vetter, R.D., Ambrose, D.A., Charter, S.R., Robertson,

L.L., Sandknop, E.M., Lynn, E.A., and J. Stannard. 1999. Fine-scale distributions of

planktonic fish eggs in the vicinities of Big Sycamore Canyon and Vandenberg Ecological

Reserves, and Anacapa and San Miguel islands, California. Reports of California

Cooperative Oceanic Fisheries Investigations [Calcofi Rep.], vol 40.

Unpublished references related to Vandenberg MRPA Ecological Reserve:

Friedman, C. S., P.L. Haaker, and I. Taniguchi. 2001. Density dependent recruitment of black

abalone and resistance to withering syndrome at the Vandenberg Ecological Reserve. Paper

presented at CalCOFI Symposium, La Jolla, California, November 2001.

Elkhorn Slough

ABA Consultants. 1987. Elkhorn slough wetlands management plan: draft report prepared for

the Monterey County Planning Department and the California Coastal Conservancy.
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ABA Consultants. 1989. Elkhorn Slough wetland management plan. Prepared for California

State Coastal Conservancy & Monterey County Planning Department.

Alberte, R.S., Suba, G.K., Procaccini, G., Zimmerman, R.C., and S.R. Fain. 1994. Assessment

of genetic diversity of seagrass populations using DNA fingerprinting: Implications for

population stability and management. Proceedings of the National Academy of Sciences,

USA. 91(3):1049-1053.

Antrim, B.S. 1981. Habitat and food resource utilization of three species of embiotocids in

Elkhorn Slough, Monterey Bay, California.

Arp, A.J., Hansen, B.M., and D. Julian. 1992. Burrow environment and coelomic fluid

characteristics of the echiuran worm Urechis caupo from populations at three sites in

Northern California. Marine biology. 113(4):613-623.

Barry, J.P. 1983. Utilization of shallow marsh habitats by fishes in Elkhorn Slough, California.

M.S. Thesis, Moss Landing Marine Laboratories.

Barry, J.P., Yoklavich, M.M., Cailliet, G.M., Ambrose, D.A., and B.S. Antrim. 1996. Trophic

ecology of the dominant fishes in Elkhorn Slough, California, 1974-1980. Estuaries.

19(1):115-138.

Beck, N.G. and K.W. Bruland. 2000. Diel biogeochemical cycling in a hyperventilating shallow

estuarine environment. Estuaries. 23(2):177-187.

Berkeley, E. and C. Berkeley. 1935. Some Notes on the Polychaetous Annelids of Elkhorn

Slough, Monterey Bay, California. American Midland Naturalist. 16(5):766-775.

Browning, B.M. 1972. The natural resources of Elkhorn Slough; their present and future use.

CA. Dept. of Fish and Game, State of California, 1972.

Byers, J. 1999. The distribution of an introduced mollusc and its role in the long-term demise of

a native confamilial species. Biological Invasions, 1, 339-352.

Clark, P. 1973. Aspects of the life history of Tresus nuttalli in Elkhorn Slough.

Clark, P., Nybakken, J., and L. Laurent. 1975. Aspects of the life history of Tresus nuttallii in

Elkhorn Slough. California Fish and Game. 61(4):215-227.

Cooper, J.W. 1979. Ecological aspects of Tubularia crocea (Agassiz, 1862) and its nudibranch

predators in Elkhorn Slough, California.

Crampton, T.A. 1994. Long-term effects of Moss Landing Harbor on the wetlands of Elkhorn

Slough. M.S. Thesis, Moss Landing Marine Laboratories.

De Vogelaere, A. Elkhorn Slough Bibliography, 3rd edition.
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Dickert, T.G. and A.E. Tuttle. 1980. Elkhorn Slough watershed : linking the cumulative impacts

of watershed development. Institute of Urban and Regional Development, University of

California, Berkeley.

Dickert, T.G. and A.E. Tuttle. 1985. Cumulative impact assessment in environmental planning.

A coastal wetland watershed example. Environmental Impact Assessment Review. 5: 28 pp.

Dickert, T.G. and R.B. Olshansky. 1986. Evaluating erosion susceptibility for land-use planning

in coastal watersheds. COAST. ZONE MANAGE. J. 13(3-4):309-333.

Feinholz, D.M. 1998. Abundance, distribution, and behavior of the southern sea otter (Enhydra

lutris nereis) in a California estuary. Aquatic Mammals. 24(2):105-115.

Gobalet, K.W. 1990. Prehistoric status of freshwater fishes of the Pajaro-Salinas River system

of California. Copeia. 1990(3):680-685.

Gobalet, K.W. and T.L. Jones. 1995. Prehistoric Native American fisheries of the Central

California coast. Transactions of the American Fisheries Society. 124(6):813-823.

Golet, D.S., and B.B. Ward. 2001. Vertical distribution of denitrification potential, denitrifying

bacteria, and benzoate utilization in intertidal microbial mat communities. Microbial

Ecology. 42(1):22-34.

Harvey, J.T., Helm, R.C., and G.V. Morejohn. 1995. Food habits of harbor seals inhabiting

Elkhorn Slough, California. California Fish and Game. 81(1):1-9.

Grosholz, E.D. and G.M. Ruiz. 1995. Spread and potential impact of the recently introduced

European green crab, Carcinus maenas, in Central California. Marine Biology, 122, 239-247.

Harris Jr., R.L. 1994. Application of NOAA's CoastWatch change analysis project for wetland

and upland change detection in the Elkhorn Slough watershed. M.S. Thesis, Moss Landing

Marine Laboratories.

Hsueh, P-W. 1991. Seasonal occurrence and abundance of brachyuran larvae in a coastal

embayment of Central California. Journal of Crustacean Biology. 11(4):546-552.

Hulberg,L.W. and J.S. Oliver. 1980. Caging manipulations in marine soft-bottom communities:

importance of animal interactions or sedimentary habitat modifications. Can. J. Fish. Aquat.

Sci. 37(7):1130-1139.

International Shellfish Enterprises, Inc. 1978. Elkhorn Slough Shellfish and Water Quality

Study January-June 1978.
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King, J., Mayer, M., Oliver, J., Ramer, B., Small, D., and Silberstein, M. 1986. NOAA National

Ocean Service. Monitoring Ecological Changes in the New South Marsh on Elkhorn Slough

National Estuarine Sanctuary.

Kukowski, G.E. 1972. A checklist of the fishes of the Monterey Bay area including Elkhorn

Slough, the San Lorenzo, Pajaro and Salinas Rivers. Moss Landing Marine Laboratories.

Kukowski, G.E. 1972. Southern range extension for the yellowfin goby, Acanthogobius

flavimanus (Temminck and Schlegel). California Fish and Game. 58(4):326-327.

Kvitek, R.G., Fukayama, A.K., Anderson, B.S., and B.K. Grimm. 1988. Sea otter foraging on

deep-burrowing bivalves in a California coastal lagoon. Marine Biology. 98(2):157-167.

Laurent, L.L. 1971. The spawning cycle and juvenile growth rate of the gaper clam, Tresus

nutalli, of Elkhorn Slough, California. M.S. Thesis, Moss Landing Marine Laboratories.

Levitt, J.M. 1989. The effect of sulfide on the anaerobic metabolism in two congeneric species

of mudflat clams from Elkhorn Slough.

Lindquist, D.C. 1998. The effects of erosion on the trophic ecology of fishes in Elkhorn Slough,

California.

Marelli, D.C. 1981. New records for Caprellidae in California, and notes on a morphological

variant of Caprella verrucosa Boeck, 1871. Proceedings of the Biological Society of

Washington. Washington DC. 84(3):654-662.

Martin, L.K. and G.M. Cailliet. 1988. Aspects of the reproduction of the Bat Ray, Myliobatis

californica, in Central California. Copeia. 1988(3):754-762.

Mountjoy, D.C. 1993. Farming practices survey and outreach recommendations for the Elkhorn

Slough water quality management plan; final report prepared for Association of Monterey

Bay Area Governments.

Nybakken, J.W., G. Cailliet, and W. Broenkow. 1975. A baseline study of the Moss

Landing/Elkhorn Slough environment. Moss Landing Marine Laboratories.

Nybakken, J., Cailliet, G., and W. Broenkow. 1977. Ecologic and Hydrographic Studies of

Elkhorn Slough, Moss Landing Harbor and Nearshore Coastal Waters: July 1974 to June

1976.

Oliver, J.S. and C.E. Long. 1981. Polychaete larval retention, flushing, and invasion in the

Elkhorn Slough, California. Estuaries. 4(3):278.

Oliver, J.S. and M.E. Reilly. 1981. Patch formation and competition among the flowering plants

of a salt-water marsh, Elkhorn Slough, California. Estuaries. 4(3):263.
124

Oliver, J.S. and P.N. Slattery. 1981. Community complexity, succession, and resilience in

marine soft-bottom environments. Estuaries. 4(3):259.

Pace, S.D. 1978. Distribution, abundance and rates of fecundity and growth of Acartia tonsa

Dana and Acartia californiensis Trinast (Copepoda) in Elkhorn Slough, California.

Page ,G.W., Stenzel, L.E., and J.E. Kjelmyr. 1999. Overview of shorebird abundance and

distribution in wetlands of the Pacific Coast of the contiguous United States. Condor.

101(3):461-471.

Raugh, A.A. 1976. Feeding habits of silversides (family Atherinidae) in Elkhorn Slough,

Monterey Bay, California.

Rein, F.A. 1999. An economic analysis of vegetative buffer strip implementation. Case study:

Elkhorn Slough, Monterey Bay, California. Coastal Management. 27(4):377-390.

Rice, D.W., Seltenrich, C.P., Spies, R.B. and M.L. Keller. 1993. Seasonal and annual

distribution of organic contaminants in marine sediments from Elkhorn Slough, Moss

Landing Harbor and nearshore Monterey Bay, California. Environmental Pollution.

82(1):79-91.

Rice, D.W., Seltenrich, C.P., Keller, M.L., Spies, R.B., and J.S. Felton. 1994. Mixed-function

oxidase-specific activity in wild and caged speckled sanddabs Citharichthys stigmaeus in

Elkhorn Slough, Moss Landing Harbor and nearshore Monterey Bay, California.

Environmental Pollution. 84(2):179-188.

Schew, W.A., Collins, C.T., and T.E. Harvey. 1994. Growth and breeding biology of Caspian

terns (Sterna caspia) in two coastal California environments. Colonial Waterbirds.

17(2):153-159.

Schwartz, D.L., Mullins, H.T., and D.F. Belknap. 1986. Holocene geologic history of a

transform margin estuary: Elkhorn Slough, Central California. Estuarine, Coastal and Shelf

Science. 22(3):285-302.

Silberstein, M. and E. Campbell. 1989. Elkhorn Slough. Monterey Bay Aquarium.

Sliger, M.C. 1982. Distribution and microhabitat selection of Hemigrapsus oregonensis (Dana)

and Pachygrapsus crassipes Randall in Elkhorn Slough, Monterey County, California.

Smith, R.E. 1974. The hydrography of Elkhorn Slough, a shallow California coastal

embayment. M.S. Thesis, Moss Landing Marine Laboratories.

Talent, L.G. 1973. Albinism in embryo gray smoothhound sharks, Mustelus californicus, from

Elkhorn Slough, Monterey Bay, California. Copeia. 1973(3):595-597.
125

Talent, L. 1973. The seasonal abundance and food of elasmobranchs occuring in Elkhorn

Slough, Monterey Bay, California. M.S. Thesis, Moss Landing Marine Laboratories.

Talent, L.G. 1982. Food habits of the gray smoothhound, Mustelus californicus, the brown

smoothhound, Mustelus henlei, the shovelnose guitarfish, Rhinobatos productus, and the bat

ray, Myliobatis californica, in Elk Horn Slough, California. California Fish and Game

68(4):224-234.

Talent, L.G. 1985. The occurrence, seasonal distribution, and reproductive condition of

elasmobranch fishes in Elkhorn Slough, California. California Fish and Game. 71(4):210-

219.

Van Coops, J.F. and J.A. Yap. 1998. Developing integrated access to contemporary aerial

photography and water quality-related data for a portion of the Elkhorn Slough and vicinity,

North Monterey County. Taking a Look at California's Ocean Resources: An Agenda for the

Future, ASCE, Reston, VA (USA). 1:370-381.

Wang, J.C.S. 1981. Taxonomy of the early life stages of fishes: fishes of the Sacramento-San

Joaquin estuary and Moss Landing Harbor-Elkhorn Slough, California / prepared for Pacific

Gas and Electric Co. by Johnson C.S. Wang; illustrations by Mark B. Adams.

Wasson, K., Zabin, C.J., Bedinger L., Diaz, M.C., and J.S. Pearse. 2001. Biological invasions

of estuaries without international shipping: The importance of intraregional transport.

Biological Conservation. 102(2):143-153.

Winstead, J.T., Middaugh, D.P., and L.A. Courtney. 1991. Ovarian mycosis in the topsmelt

Atherinops affinis. Diseases of Aquatic Organisms. 10(3):221-223.

Wong, C.R. 1989. Observations of tides and tidal currents in Elkhorn Slough. M.S. Thesis,

Moss Landing Marine Laboratories.

Yoklavich, M.M., Cailliet, G.M., Barry, J.P., Ambrose, D.A., and B.S. Antrim. 1991. Temporal

and spatial patterns in abundance and diversity of fish assemblages in Elkhorn Slough,

California. Estuaries. 14(4):465-480.

Yoklavich, M.M., Stevenson, M., and G.M. Cailliet. 1992. Seasonal and spatial patterns of

ichthyoplankton abundance in Elkhorn Slough, California. Estuarine, Coastal and Shelf

Science. 34(2):109-126.

Unpublished references related Elkhorn Slough:

Allen, J. K., 1992. Benthic invertebrates living in macroalgal mats on intertidal mudflats of

Elkhorn Slough. student paper, California. San Jose State University, San Jose, CA, 36 pp.
126

Kao, J. S. 2000. Diet, daily ration, and gastric evacuation of the leopard shark (Triakis

semifasciata), Department of Marine Science. Student paper, Moss Landing Marine

Laboratories, California State University Hayward, 96 pp.

King, J. M., M. Mayer, J. Oliver, B. Ramer, D. Small, and M. Silberstein. 1986. Monitoring

ecological changes in the New South Marsh on Elkhorn Slough National Estuarine

Sanctuary. Office of Ocean and Coastal Resource Management, Sanctuary Programs

Division, National Oceanic and Atmosphereic Administration, Washington, D.C., 72 pp.

Additional Literature of Interest:

Boersma, P.D. and J.K. Parrish. 1999. Limiting abuse: marine protected areas, a limited

solution. Ecological Economics. 31(2):287-304.

Brown, K., Adger, W.N., Tompkins, E., Bacon, P., Shim, D., and K.Young. 2001. Trade-off

analysis for marine protected area management. Ecological Economics. 37(3):417-434.

Castilla, J.C. 2000. Roles of experimental marine ecology in coastal management and

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