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Estimating Site Occupancy Rates When Detection Probabilities Are Less Than One

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  • Proteus Wildlife Research Consultants

Abstract and Figures

Nondetection of a species at a site does not imply that the species is absent unless the probability of detection is 1. We propose a model and likelihood-based method for estimating site occupancy rates when detection probabilities are 1. The model provides a flexible framework enabling covariate information to be included and allowing for missing observations. Via computer simulation, we found that the model provides good estimates of the occupancy rates, generally unbiased for moderate detection probabilities (0.3). We estimated site occupancy rates for two anuran species at 32 wetland sites in Maryland, USA, from data collected during 2000 as part of an amphibian monitoring program, Frog-watch USA. Site occupancy rates were estimated as 0.49 for American toads (Bufo amer-icanus), a 44% increase over the proportion of sites at which they were actually observed, and as 0.85 for spring peepers (Pseudacris crucifer), slightly above the observed proportion of 0.83.
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2248
Ecology,
83(8), 2002, pp. 2248–2255
q
2002 by the Ecological Society of America
ESTIMATING SITE OCCUPANCY RATES WHEN DETECTION
PROBABILITIES ARE LESS THAN ONE
D
ARRYL
I. M
AC
K
ENZIE
,
1,5
J
AMES
D. N
ICHOLS
,
2
G
IDEON
B. L
ACHMAN
,
2,6
S
AM
D
ROEGE
,
2
J. A
NDREW
R
OYLE
,
3
AND
C
ATHERINE
A. L
ANGTIMM
4
1
Department of Statistics, North Carolina State University, Raleigh, North Carolina 27695-8203 USA
2
U.S. Geological Survey, Patuxent Wildlife Research Center, 11510 American Holly Drive,
Laurel, Maryland 20708-4017 USA
3
U.S. Fish and Wildlife Service, Patuxent Wildlife Research Center, 11510 American Holly Drive,
Laurel, Maryland 20708-4017 USA
4
U.S. Geological Survey, Florida Caribbean Science Center, Southeastern Amphibian Research and Monitoring Initiative,
7920 NW 71st Street, Gainesville, Florida 32653 USA
Abstract.
Nondetection of a species at a site does not imply that the species is absent
unless the probability of detection is 1. We propose a model and likelihood-based method
for estimating site occupancy rates when detection probabilities are
,
1. The model provides
a flexible framework enabling covariate information to be included and allowing for missing
observations. Via computer simulation, we found that the model provides good estimates
of the occupancy rates, generally unbiased for moderate detection probabilities (
.
0.3). We
estimated site occupancy rates for two anuran species at 32 wetland sites in Maryland,
USA, from data collected during 2000 as part of an amphibian monitoring program, Frog-
watch USA. Site occupancy rates were estimated as 0.49 for American toads (
Bufo amer-
icanus
), a 44% increase over the proportion of sites at which they were actually observed,
and as 0.85 for spring peepers (
Pseudacris crucifer
), slightly above the observed proportion
of 0.83.
Key words: anurans; bootstrap;
Bufo americanus;
detection probability; maximum likelihood;
metapopulation; monitoring; patch occupancy;
Pseudacris crucifer;
site occupancy.
I
NTRODUCTION
We describe an approach to estimating the proportion
of sites occupied by a species of interest. We envision
a sampling method that involves multiple visits to sites
during an appropriate season during which a species
may be detectable. However, a species may go unde-
tected at these sites even when present. Sites may rep-
resent discrete habitat patches in a metapopulation dy-
namics context or sampling units (e.g., quadrats) reg-
ularly visited as part of a large-scale monitoring pro-
gram. The patterns of detection and nondetection over
the multiple visits for each site permit estimation of
detection probabilities and the parameter of interest,
proportion of sites occupied.
Our motivation for considering this problem in-
volves potential applications in (1) large-scale moni-
toring programs and (2) investigations of metapopu-
lation dynamics. Monitoring programs for animal pop-
ulations and communities have been established
throughout the world in order to meet a variety of ob-
jectives. Most programs face two important sources of
Manuscript received 29 May 2001; revised 11 October 2001;
accepted 22 October 2001.
5
Present address: Proteus Research and Consulting Ltd.,
P.O. Box 5193, Dunedin, New Zealand.
E-mail: darryl@proteus.co.nz
6
Present address: International Association of Fish and
Wildlife Agencies, 444 N. Capitol Street NW, Suite 544,
Washington, D.C., 20001 USA.
variation that must be incorporated into the design
(e.g., see Thompson 1992, Lancia et al. 1994, Thomp-
son et al. 1998, Yoccoz et al. 2001, Pollock et al. 2002).
The first source of variation is space. Many programs
seek to provide inferences about areas that are too large
to be completely surveyed. Thus, small areas must be
selected for surveying, with the selection being carried
out in a manner that permits inference to the entire area
of interest (Thompson 1992, Yoccoz et al. 2001, Pol-
lock et al. 2002).
The second source of variation important to moni-
toring program design is detectability. Few animals are
so conspicuous that they are always detected at each
survey. Instead, some sort of count statistic is obtained
(e.g., number of animals seen, heard, trapped, or oth-
erwise detected), and a method is devised to estimate
the detection probability associated with the count sta-
tistic. Virtually all of the abundance estimators de-
scribed in volumes such as Seber (1982) and Williams
et al. (
in press
) can be viewed as count statistics divided
by estimated detection probabilities. Not allowing for
detectability and solely using the count statistic as an
index to abundance is unwise. Changes in the count
may be a product of random variations or changes in
detectability, so it is impossible to make useful infer-
ence about the system under investigation.
The methods used to estimate detection probabilities
of individual animals (and hence abundance) at each
site are frequently expensive of time and effort. For
August 2002 2249
ESTIMATING SITE OCCUPANCY RATES
this reason, these estimation methods are often used in
detailed experiments or small-scale investigations, but
are not as widely used in large-scale monitoring pro-
grams. The methods proposed here to estimate the pro-
portion of sites (or more generally, the proportion of
sampled area) occupied by a species can be imple-
mented more easily and less expensively than the meth-
ods used for abundance estimation. For this reason, our
proposed method should be attractive as a basis for
large-scale monitoring programs, assuming that the
proportion of sites or area occupied is an adequate state
variable with respect to program objectives.
The second motivation for considering this estima-
tion problem involves the importance of patch occu-
pancy data to the study of metapopulation dynamics.
The proportion of patches occupied is viewed as a state
variable in various metapopulation models (e.g., Levins
1969, 1970, Lande 1987, 1988, Hanski 1992, 1994,
1997). So-called ‘‘incidence functions’’ (e.g., see Di-
amond 1975, Hanski 1992) depict the probability of
occurrence of a species in a patch, expressed as a func-
tion of patch characteristics such as area. Under the
assumption of a stationary Markov process, incidence
function data are sometimes used to estimate patch ex-
tinction and colonization probabilities (e.g., Hanski
1992, 1994, 1997, Moilanen 1999). Given the relevance
of patch occupancy data to metapopulation investiga-
tions and models, it seems important to estimate patch
occupancy probabilities properly. For most animal
sampling situations, detection of a species is indeed
indicative of ‘‘presence,’’ but nondetection of the spe-
cies is not equivalent to absence. Thus, we expect most
incidence function estimates of the proportion of patch-
es occupied to be negatively biased to some unknown
degree because species can go undetected when pre-
sent.
In this paper, we first present general sampling meth-
ods that permit estimation of the probability of site
occupancy when detection probabilities are
,
1 and
may vary as functions of site characteristics, time, or
environmental variables. We then present a statistical
model for site occupancy data and describe maximum
likelihood estimation under this model. We illustrate
use of the estimation approach with empirical data on
site occupancy by two anuran species at 32 wetland
sites in Maryland collected during 2000. Finally we
discuss extending this statistical framework to address
other issues such as colony extinction/colonization,
species co-occurrence, and allowing for heterogeneous
detection and occupancy probabilities
M
ETHODS
Notation
We use the following notation throughout this article:
c
i
, probability that a species is present at site
i
;
p
it
,
probability that a species will be detected at site
i
at
time
t
, given presence;
N
, total number of surveyed
sites;
T
, number of distinct sampling occasions;
n
t
num-
ber of sites where the species was detected at time
t
;
n.
, total number of sites at which the species was de-
tected at least once.
Our use of
p
, to signify detection probabilities, dif-
fers from its customary use in the metapopulation lit-
erature, where it is used to denote the probability of
species presence (our
c
). However, our notation is con-
sistent with the mark–recapture literature which pro-
vides the foundation of our approach.
Basic sampling situation
Here we consider situations in which surveys of spe-
cies at
N
specific sites are performed at
T
distinct oc-
casions in time. Sites are occupied by the species of
interest for the duration of the survey period, with no
new sites becoming occupied after surveying has be-
gun, and no sites abandoned before the cessation of
surveying (i.e., the sites are ‘‘closed’’ to changes in
occupancy). At each sampling occasion, investigators
use sampling methods designed to detect the species
of interest. Species are never falsely detected at a site
when absent, and a species may or may not be detected
at a site when present. Detection of the species at a
site is also assumed to be independent of detecting the
species at all other sites. The resulting data for each
site can be recorded as a vector of 1’s and 0’s denoting
detection and nondetection, respectively, for the oc-
casions on which the site was sampled. The set of such
detection histories is used to estimate the quantity of
interest, the proportion of sites occupied by the species.
General likelihood
We propose a method that parallels a closed-popu-
lation, mark–recapture model, with an additional pa-
rameter (
c
) that represents the probability of species
presence. In closed-population models, the focus is to
estimate the number of individuals never encountered
by using information garnered from those individuals
encountered at least once (e.g., see Otis et al. 1978,
Williams et al.,
in press
). In our application, sites are
analogous to individuals except that we observe the
number of sites with the history comprising
T
0’s (sites
at which the species is never detected over the
T
sam-
pling occasions); hence, the total population size of
sites is known, but the focus is to estimate the fraction
of those sites that the species actually occupies. One
could recast this problem into a more conventional
closed mark–recapture framework by only considering
those sites where the species was detected at least once.
Use of such data with closed-population, capture–re-
capture models (e.g., Otis et al. 1978) would yield es-
timates of population size that correspond to the num-
ber of sites where the species is present. However, the
following method enables additional modeling of
c
to
be investigated (such as including covariate informa-
tion).
A likelihood can be constructed using a series of
2250
DARRYL I. MACKENZIE ET AL.
Ecology, Vol. 83, No. 8
probabilistic arguments similar to those used in mark–
recapture modeling (Lebreton et al. 1992). For sites
where the species was detected on at least one sampling
occasion, the species must be present and was either
detected or not detected at each sampling occasion. For
example, the likelihood for site
i
with history 01010
would be
c
(1
2
p
)
p
(1
2
p
)
p
(1
2
p
).
ii1i2i3i4i5
However, nondetection of the species does not imply
absence. Either the species was present and was not
detected after
T
samples, or the species was not present.
For site
k
with history 00000, the likelihood is
5
c
(1
2
p
)+(1
2c
).
P
kktk
t
5
1
Assuming independence of the sites, the product of all
terms (one for each site) constructed in this manner
creates the model likelihood for the observed set of
data, which can be maximized to obtain maximum like-
lihood estimates of the parameters.
Note that, at this stage, presence and detection prob-
abilities have been defined as site specific. In practice,
such a model could not be fit to the data because the
likelihood contains too many parameters: the model
likelihood is over-parameterized. However, the model
is presented in these general terms because, in some
cases, the probabilities may be modeled as a function
of site-specific covariates, to which we shall return.
When presence and detection probabilities are con-
stant across monitoring sites, the combined model like-
lihood can be written as
T
n. n n.
2
n
tt
L
(
c
,p)
5c
p
(1
2
p
)
P
tt
[]
t
5
1
(1)
N
2
n.
T
3c
(1
2
p
)+(1
2c
).
P
t
[]
t
5
1
Using the likelihood in this form, our model could be
implemented with relative ease via spreadsheet soft-
ware with built-in function maximization routines, be-
cause only the summary statistics (
n
1
,...,
n
T
,
n.
) and
N
are required. Detection probabilities could be time
specific, or reduced forms of the model could be in-
vestigated by constraining
p
to be constant across time
or a function of environmental covariates.
We suggest that the standard error of
c
be estimated
using a nonparametric bootstrap method (Buckland and
Garthwaite 1991), rather than the asymptotic (large-
sample) estimate involving the second partial deriva-
tives of the model likelihood (Lebreton et al. 1992).
The asymptotic estimate represents a lower bound on
the value of the standard error, and may be too small
when sample sizes are small. A random bootstrap sam-
ple of
N
sites is taken (with replacement) from the
N
monitored sites. The histories of the sites in the boot-
strap sample are used to obtain a bootstrap estimate of
c
. The bootstrap procedure is repeated a large number
of times, and the estimated standard error is the sample
standard deviation of the bootstrap estimates (Manly
1997).
Extensions to the model
Covariates
.—It would be reasonable to expect that
c
may be some function of site characteristics such as
habitat type or patch size. Similarly,
p
may also vary
with certain measurable variables such as weather con-
ditions. This covariate information (X) can be easily
introduced to the model using a logistic model (Eq. 2)
for
c
and/or
p
(denote the parameter of interest as
u
and the vector of model parameters as B:
exp(XB)
u5
. (2)
1 + exp(XB)
Because
c
does not change over time during the sam-
pling (the population is closed), appropriate covariates
would be time constant and site specific, whereas cov-
ariates for detection probabilities could be time varying
and site specific (such as air or water temperature).
This is in contrast to mark–recapture models in
which time-varying individual covariates cannot be
used. In mark–recapture, a time-varying individual
covariate can only be measured on those occasions
when the individual is captured; the covariate value is
unknown otherwise. Here, time-varying, site-specific
covariates can be collected and used regardless of
whether the species is detected. It would not be pos-
sible, however, to use covariates that change over time
and cannot be measured independent of the detection
process.
If
c
is modeled as a function of covariates, the av-
erage species presence probability is
N
cˆ
O
i
i
5
1
cˆ5
. (3)
N
Missing observations
.—In some circumstances, it
may not be possible to survey all sites at all sampling
occasions. Sites may not be surveyed for a number of
reasons, from logistic difficulties in getting field per-
sonnel to all sites, to the technician’s vehicle breaking
down en route. These sampling inconsistencies can be
easily accommodated using the proposed model like-
lihood.
If sampling does not take place at site
i
at time
t
,
then that occasion contributes no information to the
model likelihood for that site. For example, consider
the history 10 11, where no sampling occurred at time
3. The likelihood for this site would be:
c
p
(1
2
p
)
pp
.
1245
Missing observations can only be accounted for in this
manner when the model likelihood is evaluated sepa-
rately for each site, rather than using the combined form
of Eq. 1.
August 2002 2251
ESTIMATING SITE OCCUPANCY RATES
F
IG
. 1. Results of the 500 simulated sets of data for
N
5
40, with no missing values. Indicated are the average value of
, ; the replication-based estimate of the true standard error of ,
SE
( ); and the average estimate of the standard error
cˆcˆcˆcˆ
obtained from 200 nonparametric bootstrap samples, , for various levels of
T
,
p
, and
c
.
SE
(
cˆ
)
S
IMULATION
S
TUDY
Simulation methods
A simulation study was undertaken to evaluate the
proposed method for estimating
c
. Data were generated
for situations in which all sites had the same probability
of species presence, and the detection probability was
constant across time and sites,
c
(·)
p
(·). The effects of
five factors were investigated: (1)
N
5
20, 40, or 60;
(2)
c5
0.5, 0.7, or 0.9; (3)
p
5
0.1, 0.3, or 0.5; (4)
T
5
2, 5, or 10; (5) probability of a missing observation
5
0.0, 0.1, or 0.2.
For each of the 243 scenarios, 500 sets of data were
simulated. For each site, a uniformly distributed, pseu-
do-random number between 0 and 1 was generated (
y
),
and if
y
#c
then the site was occupied. Further pseudo-
random numbers were generated and similarly com-
pared to
p
to determine whether the species was de-
tected at each time period, with additional random
numbers being used to establish missing observations.
The
c
(·)
p
(·) model was applied to each set of simulated
data. The resulting estimate of
c
was recorded and the
nonparametric bootstrap estimate of the standard error
was also obtained using 200 bootstrap samples.
Simulation results
Fig. 1 presents the simulation results for scenarios
where
N
5
40 with no missing values only, but these
are representative of the results in general. The full
simulation results are included in the Appendix.
Generally, this method provides reasonable estimates
of the proportion of sites occupied. When detection
probability is 0.3 or greater, the estimates of
c
are
reasonably unbiased in all scenarios considered for
T
$
5. When
T
5
2, only when detection probability is
at least 0.5 do the estimates of
c
appear to be reason-
able. For low detection probabilities, however,
c
tends
to be overestimated when the true value is 0.5 or 0.7,
but underestimated when
c
equals 0.9. A closer ex-
amination of the results reveals that, in some situations
in which detection probability is low, tends to 1.
cˆ
In most cases, the nonparametric bootstrap provides
a good estimate of the standard error for , the excep-
cˆ
tion being for situations with low detection probabil-
2252
DARRYL I. MACKENZIE ET AL.
Ecology, Vol. 83, No. 8
T
ABLE
1. Relative difference in AIC (
D
AIC), AIC model
weights (
w
i
), overall estimate of the fraction of sites oc-
cupied by each species ( ), and associated standard error
cˆ
(
SE
( )).
cˆ
Model, by species
D
AIC
w
i
cˆ
SE
()
cˆ
American toad
c
(Habitat)
p
(Temperature)
c
(·)
p
(Temperature)
c
(Habitat)
p
(·)
c
(·)
p
(·)
0.00
0.42
0.49
0.70
0.36
0.24
0.22
0.18
0.50
0.49
0.49
0.49
0.13
0.14
0.12
0.13
Spring peeper
c
(Habitat)
p
(Temperature)
c
(·)
p
(Temperature)
c
(Habitat)
p
(·)
c
(·)
p
(·)
0.00
1.72
40.49
42.18
0.85
0.15
0.00
0.00
0.84
0.85
0.84
0.85
0.07
0.07
0.07
0.07
ities. Again, this is caused by
c
estimates close to 1;
in such situations, the bootstrap estimate of the stan-
dard error is very small, which overstates the precision
of .
cˆ
In general, increasing the number of sampling oc-
casions improves both the accuracy and precision of
, although in some instances there is little gain in
cˆ
using 10 occasions rather than five. If only two occa-
sions are used, however, accuracy tends to be poor
unless detection probabilities are high, and even then
the standard error of is approximately double that of
cˆ
using five sampling occasions.
Similarly, increasing the number of sites sampled,
N
,
also improves both the accuracy and precision of .
cˆ
Not presented here are the simulation results for sce-
narios with missing observations. The proposed meth-
od appears to be robust to missing data, with the only
noticeable effect being (unsurprisingly) a loss of pre-
cision. In this study, on average, the standard error of
increased by 5% with 10% missing observations, and
cˆ
by 11% with 20% missing observations. The bootstrap
standard error estimates also increased by a similar
amount, accounting well for the loss of information.
F
IELD
S
TUDY OF
A
NURANS AT
M
ARYLAND
W
ETLANDS
Field methods and data collection
We illustrate our method by considering monitoring
data collected on American toads (
Bufo americanus
)
and spring peepers (
Pseudacris crucifer
) at 32 wetland
sites located in the Piedmont and Upper Coastal Plain
physiographic provinces surrounding Washington,
D.C., and Baltimore, Maryland, USA. Volunteers en-
rolled in the National Wildlife Federation/U.S. Geo-
logical Survey’s amphibian monitoring program,
FrogwatchUSA, visited monitoring sites between 19
February 2000 and 12 October 2000. Sites were chosen
nonrandomly by volunteers and were monitored at their
convenience. Observers collected information on the
species of frogs and toads heard calling during a 3-min
counting period taken sometime after sundown. Each
species of calling frog and toad was assigned a three-
level calling index, which, for this study, was truncated
to reflect either detection (1) or nondetection (0).
The data set was reduced by considering only the
portion of data for each species between the dates of
first and last detection exclusive. Truncating the data
in this manner ensures that species were available to
be detected throughout that portion of the monitoring
period, thus satisfying our closure assumption. Includ-
ing the dates of first and last detection in the analysis
would bias parameter estimates because the data set
was defined using these points; hence, they were ex-
cluded.
Three sites were removed after the truncation be-
cause they were never monitored during the redefined
period. Fewer than eight of the 29 sites were monitored
on any given day and the number of visits per site
varied tremendously, with a very large number of miss-
ing observations (
;
90%). Note that in the context of
this sampling, the entire sampling period included the
interval between the date at which the first wetland was
sampled and the date at which all sampling ended. A
missing observation was thus any date during this in-
terval on which a wetland was not sampled. Each time
a site was visited, air temperature was recorded. Sites
were defined as being either a distinct body of water
(pond, lake) or other habitat (swamp, marsh, wet mead-
ow). These variables were considered as potential cov-
ariates for detection and presence probabilities, re-
spectively. The data used in this analysis have been
included in the Supplement.
Results of field study
American toad
.—Daily records for the 29 sites, mon-
itored between 9 March 2000 and 30 May 2000, were
included for analysis. Sites were visited 8.9 times on
average (minimum
5
2, maximum
5
58 times), with
American toads being detected at least once at 10 lo-
cations (0.34). Three models with covariates and one
without were fit to the data (Table 1) and ranked ac-
cording to AIC (Burnham and Anderson 1998). The
four models considered have virtually identical weight,
suggesting that all models provide a similar description
of the data, despite the different structural forms.
Therefore we cannot make any conclusive statement
regarding the importance of the covariates, but there
is some suggestion that detection probabilities may in-
crease with increasing temperature and occupancy rates
may be lower for habitats consisting of a distinct body
of water. However, all models provide very similar es-
timates of the overall occupancy rate (
;
0.49), which
is 44% larger than the proportion of sites where toads
were detected at least once. The standard error for the
estimate is reasonably large and corresponds to a co-
efficient of variation of 27%.
Spring peeper
.—Daily records for the 29 sites, mon-
itored between 27 February 2000 and 30 May 2000,
were included for analysis. Sites were visited, on av-
August 2002 2253
ESTIMATING SITE OCCUPANCY RATES
erage, 9.6 times (minimum
5
2, maximum
5
66 visits),
with spring peepers being detected at least once at 24
locations (0.83). The same models as those for the
American toad were fit to the spring peeper data and
the results are also displayed in Table 1. Here the two
p
(·) models have virtually zero weight, indicating that
the
p
(Temperature) models provide a much better de-
scription of the data. We suspect that this effectis due,
partially, to a tapering off of the calling season as spring
progresses into summer. The
c
(Habitat)
p
(Temperature)
model clearly has greatest weight and suggests that
estimated occupancy rates are lower for distinct bodies
of water (0.77) than for other habitat types (1.00). This
is not unexpected, given spring peepers were actually
detected at all sites of the latter type. Regardless of
how the models ranked, however, all models provide
a similar estimate of the overall occupancy rate that is
only marginally greater than the number of sites where
spring peepers were detected at least once. This sug-
gests that detection probabilities were large enough that
spring peepers probably would be detected during the
monitoring if present.
D
ISCUSSION
The method proposed here to estimate site occupancy
rate uses a simple probabilistic argument to allow for
species detection probabilities of
,
1. As shown, it pro-
vides a flexible modeling framework for incorporating
both covariate information and missing observations.
It also lays the groundwork for some potentially ex-
citing extensions that would enable important ecolog-
ical questions to be addressed.
From the full simulation results for scenarios with
low detection probabilities, it is very easy to identify
circumstances in which one should doubt the estimates
of
c
. We advise caution if an estimate of
c
very close
to 1 is obtained when detection probabilities are low
(
,
0.15), particularly when the number of sampling oc-
casions is also small (
,
7). In such circumstances, the
level of information collected on species presence/ab-
sence is small, so it is difficult for the model to dis-
tinguish between a site where the species is genuinely
absent and a site where the species has merely not been
detected.
Our simulation results may also provide some guid-
ance on the number of visits to each site required in
order to obtain reasonable estimates of occupancy rate.
If one wishes to visit a site only twice, then it appears
that the true occupancy rate needs to be
.
0.7 and de-
tection probability (at each visit) should be
.
0.3. Even
then, however, precision of the estimate may be low.
Increasing the number of visits per site improves the
precision of the estimated occupancy rate, and the re-
sulting increase in information improves the accuracy
of the estimate when detection probabilities are low.
We stress that whenever a survey (of any type) is being
designed, some thought should be given to the likely
results and method of analysis, because these consid-
erations can provide valuable insight on the level of
sampling effort required to achieve ‘‘good’’ results.
Logistical considerations of multiple visits will prob-
ably result in some hesitancy to use this approach, but
we suggest that the expenditure of extra effort to obtain
unbiased estimates of parameters of interest generally
will be preferable to the expenditure of less effort to
obtain biased estimates. If travel time to sites is sub-
stantial, then multiple searches or samples may be con-
ducted by multiple observers, or even by a single ob-
server, at a single trip to a site, e.g., conduct two or
more 3-min amphibian calling surveys in a single night
at the same pond. If large numbers of patches must be
surveyed, then it may be reasonable to conduct multiple
visits at a subset of sites for the purpose of estimating
detection probability, and perhaps associated covariate
relationships. Then this information on detection prob-
ability, perhaps modeled as a function of site-specific
covariates, could be applied to sites visited only once.
Issues about optimal design require additional work,
but it is clear that a great deal of flexibility is possible
in approaches to sampling.
Site occupancy may well change over years or be-
tween seasons as populations change; new colonies
could be formed or colonies could become locally ex-
tinct. When sites are surveyed on more than one oc-
casion between these periods of change, for multiple
periods, the approach described here could be com-
bined with the robust design mark–recapture approach
(Pollock et al. 1990). For example, suppose that the
anuran sampling described in our examples is contin-
ued in the future, such that the same wetland sites are
surveyed multiple times each summer, for multiple
years. During the periods when sites are closed to
changes in occupancy, our approach could be used to
estimate the occupancy rate as in our example. The
change in occupancy rates over years could then be
modeled as functions of site colonization and extinction
rates, analogous with the birth and death rates in an
open-population mark–recapture study. Such Markov
models of patch occupancy dynamics will permit time-
specific estimation and modeling of patch extinction
and colonization rates that do not require the assump-
tions of
p
5
1 or process stationarity invoked in pre-
vious modeling efforts (e.g., Erwin et al. [1998] re-
quired
p
5
1; Hanski [1992, 1994] and Clark and Ro-
senzweig [1994] required both assumptions).
Often monitoring programs collect information on
the presence/absence of multiple species at the same
sites. An important biological question is whether spe-
cies co-occur independently. Does the presence/ab-
sence of species A depend upon the occupancy state
of species B? Our method of modeling species presence
could be extended in this direction, enabling such im-
portant ecological questions to be addressed. The mod-
el could be parameterized in terms of
c
AB
(in addition
to
c
A
and
c
B
): the probability that both species A and
species B are present at a site. However, the number
2254
DARRYL I. MACKENZIE ET AL.
Ecology, Vol. 83, No. 8
of parameters in the model would increase exponen-
tially with the number of species, so reasonably good
data sets might be required. For example, four addi-
tional parameters would be required to model co-oc-
currences between species A, B, and C (
c
AB
,
c
AC
,
c
BC
,
c
ABC
), but if six species were being modeled, 57 extra
parameters would need to be estimated.
Not addressed are situations in which presence and
detection probabilities are heterogeneous, varying
across sites. Some forms of heterogeneity may be ac-
counted for with covariate information such as site
characteristics or environmental conditions at the time
of sampling. On other occasions, however, the source
of heterogeneity may be unknown. We foresee that
combining our method with the mixture model ap-
proach to closed-population, mark–recapture models of
Pledger (2000) would be one solution, which enables
the problem to be contained within a likelihood frame-
work. It may also be possible to combine our method
with other closed-population, mark–recapture methods
such as the jackknife (Burnham and Overton 1978) or
coverage estimators (Chao et al. 1992). For different
sampling frameworks, where monitoring is performed
on a continuous or incidental basis rather than at dis-
crete sampling occasions, combining our methods with
the Poisson family of models (Boyce et al. 2001,
MacKenzie and Boyce 2001) may also be feasible, par-
ticularly for multiple years of data.
The three extensions to the proposed methods are
currently the focus of ongoing research on this general
topic of estimating site occupancy rates.
Software to perform the above modeling has been
included in the Supplement.
A
CKNOWLEDGMENTS
We would like to thank ChristopheBarbraud, Mike Conroy,
Ullas Karanth, Bill Kendall, Ken Pollock, John Sauer, and
Rob Swihart for useful discussions of this general estimation
problem and members of the USGS Southeastern Amphibian
Research and Monitoring Initiative for discussions on am-
phibian monitoring. Atte Moilanen provided a constructive
review of the manuscript, as did a second anonymous re-
viewer. Our thanks also go to the Frogwatch USAvolunteers
directed by Sue Muller at Howard County Department of
Parks and Recreation for collection of the data.
L
ITERATURE
C
ITED
Boyce, M. S., D. I. MacKenzie, B. F. J. Manly, M. A. Har-
oldson, and D. Moody. 2001. Negative binomial models
for abundance estimation of multiple closed populations.
Journal of Wildlife Management 65:498–509.
Buckland, S. T., and P. H. Garthwaite. 1991. Quantifying
precision of mark–recapture estimates using the bootstrap
and related methods. Biometrics 47:255–268.
Burnham, K. P., and D. R. Anderson. 1998. Model selection
and inference—a practical information-theoretic approach.
Springer-Verlag, New York, New York, USA.
Burnham, K. P., and W. S. Overton. 1978. Estimation of the
size of a closed population when capture probabilities vary
among animals. Biometrika 65:625–633.
Chao, A., S.-M. Lee, and S.-L. Jeng. 1992. Estimating pop-
ulation size for capture–recapture data when capture prob-
abilities vary by time and individual animal. Biometrics
48:201–216.
Clark, C. W., and M. L. Rosenzweig. 1994. Extinction and
colonization processes: parameter estimates from sporadic
surveys. American Naturalist 143:583–596.
Diamond, J. M. 1975. Assembly of species communities.
Pages 342–444
in
M. L. Cody and J. M. Diamond, editors.
Ecology and evolution of communities. Harvard University
Press, Cambridge, Massachusetts, USA.
Erwin, R. M., J. D. Nichols, T. B. Eyler, D. B. Stotts, and B.
R. Truitt. 1998. Modeling colony site dynamics: a case
study of Gull-billed Terns (
Sterna nilotica
) in coastal Vir-
ginia. Auk 115:970–978.
Hanski, I. 1992. Inferences from ecological incidence func-
tions. American Naturalist 139:657–662.
Hanski, I. 1994. A practical model of metapopulation dy-
namics. Journal of Animal Ecology 63:151–162.
Hanski, I. 1997. Metapopulation dynamics: from concepts
and observations to predictive models. Pages 69–91
in
I.
A. Hanski and M. E. Gilpin, editors. Metapopulation bi-
ology: ecology, genetics, and evolution. Academic Press,
New York, New York, USA.
Lancia, R. A., J. D. Nichols, and K. H. Pollock. 1994. Es-
timating the number of animals in wildlife populations.
Pages 215–253
in
T. Bookhout, editor. Research and man-
agement techniques for wildlife and habitats. The Wildlife
Society, Bethesda, Maryland, USA.
Lande, R. 1987. Extinction thresholds in demographic mod-
els of territorial populations. American Naturalist 130:624–
635.
Lande, R. 1988. Demographic models of the northern spotted
owl (
Strix occidentalis caurina
). Oecologia 75:601–607.
Lebreton, J. D., K. P. Burnham, J. Clobert, and D. R. An-
derson. 1992. Modeling survival and testing biological hy-
potheses using marked animals. A unified approach with
case studies. Ecological Monographs 62:67–118.
Levins, R. 1969. Some demographic and genetic consequenc-
es of environmental heterogeneity for biological control.
Bulletin of the Entomological Society of America 15:237–
240.
Levins, R. 1970. Extinction. Pages 77–107
in
M. Gersten-
haber, editor. Some mathematical questions in biology. Vol-
ume II. American Mathematical Society, Providence,
Rhode Island, USA.
MacKenzie, D. I., and M. S. Boyce. 2001. Estimating closed
population size using negative binomial models. Western
Black Bear Workshop 7:21–23.
Manly, B. F. J. 1997. Randomization, bootstrap and Monte
Carlo methods in biology. Second edition. Chapman and
Hall, London, UK.
Moilanen, A. 1999. Patch occupancy models of metapopu-
lation dynamics: efficient parameter estimation using im-
plicit statistical inference. Ecology 80:1031–1043.
Otis, D. L., K. P. Burnham, G. C. White, and D. R. Anderson.
1978. Statistical inference from capture data on closed an-
imal populations. Wildlife Monographs 62.
Pledger, S. 2000. Unified maximum likelihood estimates for
closed capture–recapture models using mixtures. Biomet-
rics 56:434–442.
Pollock, K. H., J. D. Nichols, C. Brownie, and J. E. Hines.
1990. Statistical inference for capture–recapture experi-
ments. Wildlife Monographs 107.
Pollock, K. H., J. D. Nichols, T. R. Simons, G. L.Farnsworth,
L. L. Bailey, and J. R. Sauer. 2002. Large scale wildlife
monitoring studies: statistical methods for design and anal-
ysis. Environmetrics 13:1–15.
Seber, G. A. F. 1982. The estimation of animal abundance
and related parameters. MacMillan Press, New York, New
York, USA.
August 2002 2255
ESTIMATING SITE OCCUPANCY RATES
Thompson, S. K. 1992. Sampling. John Wiley, New York,
New York, USA.
Thompson, W. L., G. C. White, and C. Gowan. 1998. Mon-
itoring vertebrate populations. Academic Press, San Diego,
California, USA.
Williams, B. K., J. D. Nichols, and M. J. Conroy.
In press.
Analysis and management of animal populations. Academ-
ic Press, San Diego, California, USA.
Yoccoz, N. G., J. D. Nichols, and T. Boulinier. 2001. Mon-
itoring of biological diversity in space and time; concepts,
methods and designs. Trends in Ecology andEvolution 16:
446–453.
APPENDIX
Full results of the simulation study are available in ESA’s Electronic Data Archive:
Ecological Archives
E083-041-A1.
SUPPLEMENT
Software, source code, and the sample data sets are available in ESA’s Electronic DataArchive:
Ecological Archives
E083-
041-S1.
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Preface. Basic Concepts. Sampling Designs and Related Topics. Enumeration Methods. Community Surveys. Detection of a Trend in Population Estimates. Guidelines for Planning Surveys. Fish. Amphibians and Reptiles. Birds. Mammals. Glossary of Terms. Glossary of Notation. Sampling Estimators. Common and Scientific Names of Cited Vertebrates. Subject Index.
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The practical value of a predictive metapopulation model is much affected by the amount of data required for parameter estimation. Some metapopulation models require information on population turnover events for parameterization, whereas other models, such as the incidence function model that is used in this study, can be parameterized with spatial data on patch occupancy. The latter data are more readily available. The original method of using spatial pattern data to parameterize the incidence function and other patch models has been criticized for involving potentially troublesome assumptions, such as the independence of habitat patches and constant colonization probabilities. This study describes an improved parameter estimation method that is not affected by these problems. The proposed method is based on Monte Carlo inference for implicit statistical models, and it can be adapted to any stochastic patch occupancy model of metapopulation dynamics. As an additional advantage, the new method allows the estimation of the amplitude of regional stochasticity. Tested with simulated data, the new method was found to produce substantially more accurate parameter estimates than the original method. The new approach is applied to two empirical metapopulations, the false heath fritillary butterfly in Finland and the American pika at Bodie, California.
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We discuss a simple method, based on maximum likelihood, to estimate the rates of extinction and recolonization of a species, for example, on an island. Our method applies to both regular and sporadic surveys and thus allows one to use time series with missing data. When only a few surveys have been conducted, it may be possible to lump data from several species. The shape of the likelihood surface may indicate whether the lumping is appropriate.