Estimating species richness: sensitivity to sample coverage and insensitivity to spatial patterns
ABSTRACT The number of species in an area is critical to the development of evolutionary and ecological theory from mass extinctions to island biogeography. Still, the factors influencing the accuracy of estimators of species richness are poorly understood. We explored these factors by simulating landscapes that varied in species richness, relative abundances, and the spatial distribution. We compared the extrapolations of nine nonparametric estimators and two species accumulation curves under three sampling intensities. Community evenness of species' abundances, sampling intensity, and the level of true species richness significantly influenced bias, precision, and accuracy of the estimations. Perhaps most surprisingly, the effects of gradient strength and spatial autocorrelation type were generally insignificant. The nonparametric estimators were substantially less biased and more precise than the species accumulation curves. Observed species richness was most biased. Community evenness, sampling intensity, and the level of true species richness influenced the performance of the nonparametric estimators indirectly via the fraction of all species found in a sample or ''sample coverage.'' For each particular level of sample coverage, a single estimator was most accurate. Choice of estimator is confounded by a priori uncertainty about the sample coverage. Accordingly, researchers can extrapolate species richness by various estimators and base the estimator choice on the mean estimated sample coverage. Alternatively, the most reliable estimator with respect to community evenness can be chosen. These predictions from our simulations are confirmed in two field studies.
 [Show abstract] [Hide abstract]
ABSTRACT: The maned wolf has been studied in nature reserves, but few researches have been carried out outside protected areas. Since only about 2 % of the Brazilian Cerrado, the maned wolf's main habitat, has been set aside as parks and reserves, determining what is happening with the species in private and disturbed areas is important for an accurate assessment of its vulnerability to extinction. Here we investigated the trophic ecology of a maned wolf population inhabiting a 1610 ha section of the Calçada Ridge, an unprotected area located in the metropolitan region of Belo Horizonte, capital of the state of Minas Gerais, Brazil. The study site is in the buffer zone (< 10 km) of two protected areas, where anthropogenic (urban areas and roads) and disturbed areas (burned fields) total a third of the study landscape. The main disturbances are mining activities, unregulated ecotourism and road proximity. Fecal samples (n= 95) collected between 2006 and 2008 revealed that the maned wolf frequently used both natural and disturbed fields. The diet was composed mostly of small mammals (9 species, 16.2 % of items and 92.6% of scats) and the plant Solanum lycocarpum (12.2% of items and 89.5% of scats), similar to what has been found in less disturbed areas. Overall diet diversity was, however, lower than has been found elsewhere, probably reflecting the poorer resource base of the study area. These results, together with recent findings from other sources, highlight the importance of buffer zones. They also suggest that the maned wolf is an ecologically flexible species that might be prone to hunt, and perhaps even survive, in disturbed areas outside protected areas.01/2012; 5:284300.  SourceAvailable from: Andrew D. BarnesAndrew D Barnes, Malte Jochum, Steffen Mumme, Noor Farikhah Haneda, Achmad Farajallah, Tri Heru Widarto, Ulrich Brose[Show abstract] [Hide abstract]
ABSTRACT: Our knowledge about landuse impacts on biodiversity and ecosystem functioning is mostly limited to single trophic levels, leaving us uncertain about wholecommunity biodiversityecosystem functioning relationships. We analyse consequences of the globally important landuse transformation from tropical forests to oil palm plantations. Species diversity, density and biomass of invertebrate communities suffer at least 45% decreases from rainforest to oil palm. Combining metabolic and foodweb theory, we calculate annual energy fluxes to model impacts of landuse intensification on multitrophic ecosystem functioning. We demonstrate a 51% reduction in energy fluxes from forest to oil palm communities. Species loss clearly explains variation in energy fluxes; however, this relationship depends on landuse systems and functional feeding guilds, whereby predators are the most heavily affected. Biodiversity decline from forest to oil palm is thus accompanied by even stronger reductions in functionality, threatening to severely limit the functional resilience of communities to cope with future global changes.Nature Communications 10/2014; 5:5351. · 10.74 Impact Factor  SourceAvailable from: Jan Beukema
Dataset: pdfspeciesrichness
Page 1
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Ecology, 84(9), 2003, pp. 2364–2377
? 2003 by the Ecological Society of America
ESTIMATING SPECIES RICHNESS: SENSITIVITY TO SAMPLE COVERAGE
AND INSENSITIVITY TO SPATIAL PATTERNS
ULRICH BROSE,1NEO D. MARTINEZ, AND RICHARD J. WILLIAMS
Romberg Tiburon Center, Department of Biology, San Francisco State University, 3152 Paradise Drive, Tiburon,
California 94920 USA
Abstract.
tionary and ecological theory from mass extinctions to island biogeography. Still, the factors
influencing the accuracy of estimators of species richness are poorly understood. We ex
plored these factors by simulating landscapes that varied in species richness, relative abun
dances, and the spatial distribution. We compared the extrapolations of nine nonparametric
estimators and two species accumulation curves under three sampling intensities. Com
munity evenness of species’ abundances, sampling intensity, and the level of true species
richness significantly influenced bias, precision, and accuracy of the estimations. Perhaps
most surprisingly, the effects of gradient strength and spatial autocorrelation type were
generally insignificant. The nonparametric estimators were substantially less biased and
more precise than the species accumulation curves. Observed species richness was most
biased. Community evenness, sampling intensity, and the level of true species richness
influenced the performance of the nonparametric estimators indirectly via the fraction of
all species found in a sample or ‘‘sample coverage.’’ For each particular level of sample
coverage, a single estimator was most accurate. Choice of estimator is confounded by a
priori uncertainty about the sample coverage. Accordingly, researchers can extrapolate
species richness by various estimators and base the estimator choice on the mean estimated
sample coverage. Alternatively, the most reliable estimator with respect to community
evenness can be chosen. These predictions from our simulations are confirmed in two field
studies.
Key words: biodiversity; estimations; extrapolations; gradients; jacknife estimators; nonpara
metric estimators; relative abundance distribution; sample coverage; spatial autocorrelation; species
accumulation curves.
The number of species in an area is critical to the development of evolu
INTRODUCTION
Studies of biodiversity initiated the study of evolu
tion and have continued to become more centrally im
portant to ecology (Gaston 1996). However, the mea
surement of biodiversity has been surrounded by a con
siderable debate (Peet 1974, May 1975, Pielou 1975,
Magurran 1988, Colwell and Coddington 1994, Gaston
1996, Gotelli and Colwell 2001, Petchey and Gaston
2002). Historically, diversity has been measured by a
bewildering range of diversity indices that often consist
of two components: the number of species and the rel
ative evenness of their abundances (Magurran 1988).
Both the compound character and the variety of dif
ferent indices make comparisons across studies diffi
cult. More recently, both components have generally
been measured separately, and much emphasis has been
put on species richness (Colwell and Coddington 1994,
Gaston 1996), which continues to be the most frequent
measure of biodiversity. Earlier, Whittaker (1972) pro
posed ? diversity to be withinhabitat species richness,
? diversity betweenhabitats species turnover, and ?
Manuscript received 11 September 2002; revised 22 Decem
ber 2002; accepted 16 January 2003. Corresponding Editor: N. J.
Gotelli.
1Email: brose@sfsu.edu
diversity landscape species richness. Measurements of
? diversity or ‘‘complementarity’’ have subsequently
been treated separately from species richness (Magur
ran 1988, Colwell and Coddington 1994, Colwell 1997,
Harte et al. 1999). In contrast to this, the separation of
? and ? diversity has been modified into a concept of
species richness on a continuous scale from local to
global (e.g., Ricklefs and Schluter 1993), and the meth
odology of measuring local or global species richness
is not considered to be qualitatively different (Colwell
and Coddington 1994, Gaston 1996). This concept of
species richness is currently the most basic and often
used parameter, not only in biodiversity studies but also
in community and trophic ecology (Gaston 1996, Mar
tinez et al. 1999, Williams and Martinez 2000). Despite
the centrality of species richness, many difficulties re
garding its measurement are not well understood nor
agreed upon. In most empirical studies, the observed
number of species, Sobs, is used as a surrogate for the
true number of species, Strue. Still, Sobstypically ex
cludes many rare species and seriously underestimates
Strue(Palmer 1990, Baltana ´s 1992, Colwell and Cod
dington 1994, Martinez et al. 1999). This bias increases
with increasing Strueand decreases with mean species
detectability, which frequently invalidates hypothesis
testing (Palmer 1990, Boulinier et al. 1998). A large
Page 2
September 2003 2365
ESTIMATING SPECIES RICHNESS
number of extrapolation methods have been developed
to reduce this bias (reviews in Bunge and Fitzpatrick
1993, Sobero ´n and Llorente 1993, Colwell and Cod
dington 1994, Colwell 1997), but their performance has
been compared in only few studies (Palmer 1990, 1991,
Baltana ´s 1992, Colwell and Coddington 1994, Walther
and Morand 1998, Longino et al. 2002). Most of these
methods can be classified as either species accumula
tion curves or nonparametric estimators. Although gen
erally less biased, the nonparametric estimators are sel
dom used in biodiversity studies, where species ac
cumulation curves are preferred (e.g., Chazdon et al.
1999).
Species accumulation curves extrapolate species
richness vs. sample size data to an asymptote of total
richness (Sobero ´n and Llorente 1993, Colwell and Cod
dington 1994). The most often used species accumu
lation curves are the exponential equation (Holdridge
et al. 1971) and the MichaelisMenten model (Mi
chaelis and Menten 1913). Their performance varies
with the underlying relative abundance distributions
(Keating and Quinn 1998), spatial aggregation of spe
cies (Baltana ´s 1992) and habitat heterogeneity (Colwell
and Coddington 1994, Lande et al. 2000).
Nonparametric estimators are sampling theoretic ex
trapolation methods that only require the number of
samples in which each species is found rather than any
parametric information about their abundance. They
were originally developed to estimate population sizes
with capture–recapture data in closed populations (Otis
et al. 1978, Burnham and Overton 1979). The technique
can be adapted to estimate species richness by substi
tuting recaptures of species in different samples for
recaptures of individuals of the same species in time
between mark and recapture. The underlying models
assume speciesspecific detection probabilities that fol
low an arbitrary distribution (e.g., model Mh, Otis et
al. 1978), which relaxes the implicit assumption of
equal detection probabilities made when using Sobs
(Boulinier et al. 1998). However, the distribution of
capture probabilities among individuals in a species’
population is likely to be more even than the distri
bution of detection probabilities among species with
strong variations in their abundances. This could re
strict legitimate use of the estimators to communities
with highly even relative abundance distributions. Fur
thermore, the estimators that are based on model Mh
have a second assumption: that the speciesspecific de
tection probabilities are constant for all sampling units,
which assumes spatially homogeneous species distri
butions. This assumption is violated in almost every
type of community due to environmental gradients and
spatial autocorrelation (Legendre 1993).
In this study, we examine the sensitivity of different
estimators’ bias and accuracy to abundance distribu
tions, strength of environmental gradients, spatial au
tocorrelation, the level of Strueand sampling intensity
in simulated landscapes. The predictions of the sen
sitivity analyses are tested in two field studies.
METHODS
Simulations
Rectangular communities were simulated using lat
tices of 100 by 50 cells (x, y), each representing 1 m2.
Strueis the number of species present in these simulated
landscapes and Sobsis the number of species found in
samples taken from these landscapes. The landscapes
contained communities that systematically differed due
to Strue, relative abundance, spatial gradients, and spatial
autocorrelation. The combination of seven levels of Strue
(25, 50, 100, 150, 200, 250, and 500 species), three
relative abundance models, three gradient strengths,
and three types of spatial autocorrelation resulted in a
total of 189 different types of communities. Monte Car
lo simulations with 100 replicates for each community
type were repeated for three sampling intensities, yield
ing a total of 56700 simulated landscapes.
In each landscape, the abundance of each species,
which also equals the number of cells occupied, was
derived from relative abundance distributions possess
ing an average of 200 individuals per species. We then
independently assigned each species i a probability of
occurrence pi(x, y) in every cell (x, y) of the lattice.
These pi(x, y) depend on the respective speciesspecific
probability functions of the gradients and the spatial
autocorrelation: pi(x, y) ? p
all individuals of a species are subjected to the same
spatial distribution of probabilities but the distribution
may be different for each species. For each species the
individuals were placed on the lattice according to pi(x,
y). First, the cells were ranked according to their pi.
Second, a number was sampled from a normal distri
bution with a mean of zero, and a standard deviation
equal to the number of empty cells divided by two, and
the chosen number was converted to a rank equal to
the absolute value of the number. Third, an individual
was assigned to the cell with the chosen rank and that
cell was no longer available for occupation by an in
dividual of the same species. The second and third steps
were repeated until the number of cells occupied
equalled the species’ abundance. This procedure was
repeated for all remaining species.
We estimated species richness under three intensities
of sampling from the 5000cell communities: low (25
cells), intermediate (200 cells), and high intensity (500
cells). We applied 12 estimators of species richness to
each sample: nine nonparametric estimators, two spe
cies accumulation curves, and Sobs. The performance of
these estimators was compared concerning bias, pre
cision, and accuracy.
(x) ? p
(x, y) such that
ii
gradauto
Relative abundance distributions
We used three stochastic relative abundance models:
the brokenstick, randomfraction, and random assort
Page 3
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ULRICH BROSE ET AL.
Ecology, Vol. 84, No. 9
FIG. 1.
with Sest? Strue. Regression details are given in Table 1. Key to abbreviations: Jsel? Jacksel; MM ? MichaelisMenten model;
Expo ? Exponential equation.
Mean estimated species richness, Sest, plotted against Strue. The solid line indicates an idealized unbiased estimator
ment models. Stochastic models generate much more
variable individual communities than deterministic
models do. We chose the stochastic models because
they represent more accurately natural communities
that are simultaneously influenced by various factors,
and they estimate the relative abundances of rare spe
cies more precisely. This allows stochastic models to
encompass a broad range of community types (Nee et
al. 1991, Tokeshi 1993, Naeem and Hawkins 1994).
The brokenstick model represents a rarely observed,
extremely even community with low sequential dom
inance and few rare species. The relative abundance of
Struespecies is determined by placing Strue?1 points
randomly on a line of unit length. Breaking the line at
those points yields Struesegments whose lengths rep
resent the relative abundances of Struespecies (Mac
Arthur 1957). Brokenstick abundance distributions
characterize some of the most even natural commu
nities observed (Tokeshi 1993). The randomfraction
model yields comparatively uneven communities with
abundance distributions similar to Preston’s (1948) de
terministic lognormal model (see Tokeshi 1993: Fig.
1). Randomly breaking a line of unit length into two
parts, and then repeating the process on randomly cho
sen parts until Struesegments are produced, generates a
randomfraction community (Tokeshi 1993). The ran
domassortment model yields more highly uneven com
munities that are analogous to those described by geo
metric series (Tokeshi 1993). The randomassortment
model assigns a random fraction of the line to the first
species, a random fraction of the remaining part of the
line to the second species, and so on until Struesegments
are produced.
For comparative purposes, we calculated community
evenness by the ratio of observed diversity to maxi
mum diversity (Magurran 1988):
?
p ln p
i
ln S
?
i
evenness ?
where piis the proportion of individuals of species i
and S is the number of species.
Gradients
Gradients such as flooding regimes on shorelines can
be expressed as functions of the geographic coordinates
x and y (Legendre 1993). We simulated landscapes with
strong and weak gradients along the x coordinate as
well as landscapes with no gradient. Gradient strength
was expressed as the turnover distance dturnat which
on average all species were exchanged once (Jongmann
et al. 1995). A strong gradient was characterized by
one full turnover along the x coordinate of the lattice
(dturn? 100 cells), a weak gradient by a quarter turnover
along the lattice (dturn? 400 cells).
The individuals of a species i on a gradient are typ
ically normally distributed with an optimum oi(the
mean) and an amplitude ai(standard deviation) (Jong
mann et al. 1995). For every species i and every cell,
we simulated these probabilities as
22
i
p
(x) ? exp[?0.5(x?o) /a ].
grad
For every species i, we randomly derived the opti
mum oifrom a uniform distribution with 0 ? oi? 100.
Given that one species turnover equals approximately
four standard deviations, the amplitude aiof every spe
cies i was randomly sampled from a normal distribution
(mean ? dturn/4, standard deviation ? dturn/8).
ii
Spatial autocorrelation
Our simulations of spatial autocorrelation (Legendre
1993) vary the occurrence probabilities associated with
each cell based on variably located autocorrelation
sources. These probabilities were independently ap
plied for every species i by locating between 1 and 10
autocorrelation sources on the lattice in the same way
as individuals, that is, either in a uniformly random
manner or according to p
igrad
source and in neighboring cells were established by
parameterizing a sill, range, and nugget. The sill is the
asymptotic occurrence probability at a certain range
from the nugget. The nugget indicates the maximum
. The probabilities at the
Page 4
September 20032367
ESTIMATING SPECIES RICHNESS
deviation from the sill at the autocorrelation source and
therefore determines the strength of autocorrelation.
Probabilities between the nugget and the sill are linear
functions of commonality dependent on the distance to
the autocorrelation source. We simulated communities
with biologically reasonable values (Mistral et al. 2000;
J. B. Wilson, personal communication) of longrange
autocorrelation (sill ? 0.27, nugget ? 0.72, and range
? 5 m) and shortrange autocorrelation (sill ? 0.27,
nugget ? 0.62, and range ? 2 m). Speciesspecific
functions were derived to determine p
tocorrelation probabilities of occurrence for every cell
(x, y):
(x, y), the au
iauto
p
(x, y)
iauto
0.033 ? 0.33d(x, y)
range ? 0.1
i
0
? c
0 ? d ? rangei
?
d ? range
i
where d(x, y) is the distance to the autocorrelation
source, rangeiis a speciesspecific range of autocor
relation, and c is the autocorrelation strength that
equals 0.35 for shortrange and 0.45 for longrange
autocorrelation. The values of rangeiwere randomly
sampled from normal distributions (shortrange auto
correlation: mean ? 2, lSD ? 1; longrange autocor
relation: mean ? 5, lSD ? 2.5). In the model without
autocorrelation we defined for all species i: p
The spatial autocorrelation patterns were indepen
dently created for every species i by (1) choosing 1–
10 autocorrelation sources from a uniform random dis
tribution, (2) placing the sources on the landscape ac
cording to the p
in the same way as individuals (see
igrad
spatial autocorrelation), and (3) calculating a new
landscape of probabilities of occurrence as pi(x, y) ?
p
(x) ? p
(x, y).
ii
grad auto
? 0.
iauto
Extrapolation of species richness
We compared estimations of two species accumu
lation curves and nine nonparametric estimators (Ap
pendix). This includes the two most frequently used
species accumulation curves, the exponential equation
(Holdridge et al. 1971) and the MichaelisMenten mod
el (Michaelis and Menten 1913). To minimize arbitrary
effects, cell sample order for accumulation curves was
randomized 100 times in each sample. The mean spe
cies numbers S(n) were calculated for each value of n
cells sampled. Following Keating and Quinn (1998),
we used least squares nonlinear regressions with the
quasiNewton algorithm to fit the MichaelisMenten
model and the exponential equation to the mean S(n)
data. The oftendiscussed difference between number
of individuals and number of samples as measurements
of sampling intensity (Chazdon et al. 1999, Gotelli and
Colwell 2001) is minimal in our simulations, because
when averaged across all species the individuals are
evenly distributed across the lattice.
The nine nonparametric estimators used include the
first to fifthorder jacknife estimators (Jack1–Jack5,
Burnham and Overton 1979), a selected and an inter
polated jacknife estimator (Jacksel, Jackint, Burnham and
Overton 1979), the incidencebased estimator (Chao2,
Chao 1987), and the incidencebased coverage esti
mator (ICE, Lee and Chao 1994). Since the original
formulations of Chao2 and ICE are undefined for zero
duplicates and when all infrequent species are uniques,
respectively, we used a corrected version of Chao2
(Colwell 1997; also see the Appendix) and replaced
ICE estimates by Chao2 estimates in the undefined cas
es. To calculate Jacksel, the jacknife estimates are se
quentially compared to the estimates of the next higher
order until the first estimate is reached that does not
significantly differ from the next (see Burnham and
Overton 1979 for the exact test statistics). Jackintis an
interpolated estimator between Jackseland the preced
ing estimator.
Estimator evaluation
We based our estimator evaluation on bias, precision,
and accuracy. Positive and negative bias indicates over
estimation and underestimation, respectively. Precision
is the closeness of repeated measurements of the same
quantity to each other (Sokal and Rohlf 1995). Accu
racy is the closeness of a measured or computed value
to its true value (Sokal and Rohlf 1995). Overall bias
and precision were calculated for each estimator in all
simulations across all community types using linear
leastsquares regressions of estimated species richness,
Sest, as a function of true species richness, Strue, forced
through an intercept of zero (method suggested by N.
Gotelli, personal communication). All data points fall
on a straight line with a slope of one for an unbiased
and perfectly precise estimator. We measured bias as
the difference between the observed slope of the re
gression and the expected slope of one: Bias ? ob
served slope ? 1. Precision was measured by the r2
values of the regressions.
We also calculated bias and inaccuracy for single
samples. In this case, bias was defined as the propor
tional deviation of Sestfrom Strue:
S
? S
Strue
esttrue
bias ?
.
We measured inaccuracy using the square proportional
deviation of Sestfrom Strue:
2
S
? S
Strue
esttrue
inaccuracy ?
.
??
For groups of data sets, we calculated mean bias and
mean inaccuracy, and imprecision was measured using
the standard deviation of the bias. Bias, imprecision
and inaccuracy were measured as percentage of Strue.
RESULTS
Overall, no single estimator consistently performed
better than all others in each situation, but the non
Page 5
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ULRICH BROSE ET AL.
Ecology, Vol. 84, No. 9
TABLE 1.
Seston Strueand correlation of the independent variables with the residuals.
General bias and precision of the estimators determined by linear regressions of
EstimatorBiasPrecisionRad†Grad‡Auto‡ Sampling†
Sobs
Chao2
Jack1
Jack2
Jack3
Jack4
Jack5
Jacksel
Jackint
ICE
MM
Expo
?0.558
?0.372
?0.457
?0.404
?0.366
?0.336
?0.309
?0.363
?0.366
?0.466
?0.474
?0.588
0.679
0.762
0.733
0.765
0.784
0.788
0.764
0.785
0.784
0.735
0.664
0.677
?0.80
?0.64
?0.83
?0.82
?0.80
?0.74
?0.62
?0.72
?0.72
?0.75
?0.87
?0.85
0.00
?0.02
0.00
?0.01
?0.01
?0.02
?0.02
?0.02
?0.02
?0.01
0.01
0.01
?0.02
?0.02
?0.02
?0.02
?0.02
?0.02
?0.01
?0.02
?0.02
?0.02
0.00
?0.01
0.39
0.50
0.33
0.31
0.32
0.34
0.35
0.43
0.44
0.42
0.17
0.25
Notes: Bias ? slope ? 1; Precision ? r2. P ? 0.001 for all regressions. Standard errors of
all biases are 0.002. Spearman rank correlation coefficients are given in the last four columns.
Key to abbreviations: ICE ? incidencebased coverage estimator, MM ? MichaelisMenten
model, Expo ? exponential equation, Rad ? relative abundance distribution, Grad ? gradient
strength, Auto ? autocorrelation type.
† All values, P ? 0.001.
‡ All values not significant.
parametric estimators as a group did generally outper
form the other estimators with respect to both bias and
precision. In contrast, several estimators consistently
performed poorly, most notably the most popular mea
sure of biodiversity, Sobs. We first present the general
performance of the estimators followed by the effects
of abundance distribution and sample coverage on this
performance. Finally, we compare results from our sim
ulated landscapes with results of similar analyses of
empirical data collected from field studies.
General performance of the estimators
We evaluated the general performance of the esti
mators using linear regressions of Sestas a function of
Strue. All estimators were much more variable in the 25
species simulations and comparatively less biased in
the500species simulations. Furthermore, Chao2
showed an extraordinary high variance in these high
richness simulations. Therefore, we avoided heteros
cedasticity and nonlinearity in evaluations of bias and
precision by excluding results from landscapes with
Strue? 25 and Strue? 500. We also excluded the Strue?
50 samples from the analyses of Jack4 and Jack5 due
to high variances. All estimators were negatively bi
ased, but only the exponential equation provided more
biased results than Sobs(Fig. 1). The MichaelisMenten
model and ICE were the most biased of the remaining
estimators. The regression details are given in Table 1.
All estimators except the MichaelisMenten model and
the exponential equation yielded more precise esti
mates than Sobs. Due to their poor general performance,
the MichaelisMenten model and the exponential equa
tion were excluded from further analyses. Bias de
creased with increasing order of the jacknife estima
tors, resulting in Jack4 and Jack5 being least biased
while Jack4 and Jackselwere most precise (Table 1).
Surprisingly, Jackintwas more biased and imprecise
than Jacksel, and was also excluded from further anal
yses. Chao2 performed intermediately with respect to
both bias and precision.
Regression residuals were significantly correlated
with relative abundance distributions and sampling in
tensity, indicating that these two factors had a sub
stantial additional impact on estimator performance
(Table 1). For most estimators, relative abundance dis
tributions were twice as highly correlated with the re
siduals as were sampling intensities. Surprisingly, gra
dient strength and autocorrelation type were not cor
related with the residuals in any of the situations an
alyzed. Therefore, we further analyzed the impact of
the relative abundance distributions on estimator per
formance in more detail.
Relative abundance distributions
The bias of all estimators was influenced by the rel
ative abundance distributions as well as by sampling
intensity and Strue(Fig. 2). At each level of sampling
intensity and Strue, the differences in mean bias between
the three relative abundance distributions groups were
highly significant for all estimators (KruskalWallis P
? 0.001). Sobswas reasonably unbiased only in very
even (brokenstick) communities under intermediate or
high sampling intensity, but Sobswas never the least
biased. In all uneven communities, all estimators were
substantially less biased than Sobs(Fig. 2). In general,
the estimators were up to 50fold more biased in the
uneven (randomfraction) communities than in the even
(brokenstick) communities (Fig. 2). The estimates in
the highly uneven (randomassortment) communities
were generally extremely biased, with mean estimates
usually well below 50% of Strue. This increase in bias
coincided with a decrease in evenness from broken
stick (mean ? 0.92, range 0.86–0.95) to random frac
Page 6
September 20032369
ESTIMATING SPECIES RICHNESS
FIG. 2.
at different levels of Strueand sampling intensity. Results are depicted for brokenstick (solid circles), random fraction (open
circles), and random assortment (solid squares) communities. Jsel? Jacksel.
Influence of the relative abundance distributions on estimator bias (mean) and imprecision (standard deviations)
tion (mean ? 0.62, range 0.25–0.87) and random as
sortment communities (mean ? 0.47, range 0.22–0.60).
Jack4 and Jack5 were generally the least biased es
timators, but they were also the least precise, as in
dicated by the large standard deviations of their mean
biases (Fig. 2). Chao2 and ICE were markedly impre
cise in some data sets (e.g., Fig. 2, random assortment
communities with Strue? 250 under high sampling).
Jack1 and Jack2 were more biased, yet substantially
more precise, than Jack4 and Jack5. Jack3 and Jacksel
performed intermediately with respect to bias and pre
cision. Jack1 and Jack2 performed best with respect to
accuracy in highly even communities, and Jack2,
Jack3, and Jack4 performed best with respect to ac
curacy in the uneven and highly uneven communities.
However, estimator ranking also depended on sampling
intensity.
Sample coverage
Strue, sampling intensity (i.e., the number of samples
per community), and the community evenness medi
ated by the relative abundance distributions influence
Sobs, and therefore the sample coverage, i.e., the pro
portion of species present in the sample. Among all
simulations, the sample coverage was strongly corre
lated with sampling intensity (Spearman rank corre
lation, r ? 0.456) and community evenness (r ? 0.836),
but correlations with Strue(r ? 0.144) were surprisingly
weak (P ? 0.001 for all three Spearman rank corre
lations). There were no significant correlations with
spatial autocorrelation type and gradient strength. Es
timator bias was significantly correlated with all in
dependent variables (Table 2). With respect to the cor
relation coefficients, sample coverage was the most im
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ULRICH BROSE ET AL.
Ecology, Vol. 84, No. 9
TABLE 2.
planatory variables: Strue, sampling intensity, community evenness, and sample coverage.
Simple Spearman rank and partial correlations among estimator bias and the ex
Estimator
Simple correlations
Strue
Sampling
intensityEvenness
Sample
coverage
Partial correlations
Strue
0.137
?0.065
?0.047
?0.032
?0.020
?0.011
0.019
0.048
Sampling
intensity Evenness
Chao2
Jack1
Jack2
Jack3
Jack4
Jack5
Jacksel
ICE
?0.033
?0.144
?0.139
?0.122
?0.090
?0.048
?0.094
?0.110
0.524
0.394
0.361
0.345
0.341
0.324
0.452
0.474
0.664
0.862
0.848
0.796
0.703
0.562
0.739
0.785
0.843
0.985
0.950
0.891
0.800
0.660
0.885
0.930
0.276
?0.243
?0.136
?0.047
0.015
0.045
0.147
0.162
?0.031
0.280
0.226
0.163
0.101
0.053
0.045
0.043
Notes: P ? 0.001 for all correlations; in partial correlations the effect of sample coverage
was removed prior to further correlation analyses.
portant independent variable and Struethe least impor
tant. After the effect of sample coverage on estimator
bias had been removed in partial correlation analyses,
the other independent variables only had a minor direct
impact on the remaining variance of estimator bias de
spite strong simple correlations (Table 2). Overall,
community evenness and sampling intensity controlled
sample coverage, which substantially influenced the
bias of the nonparametric estimators.
We analyzed the relationship between sample cov
erage and estimator performance in more detail by plot
ting estimator bias, imprecision, and inaccuracy in 1%
steps against sample coverage (Fig. 3). While bias and
inaccuracy decreased with increasing sample coverage,
imprecision was highest at intermediate levels of sam
ple coverage. All estimators reduced the bias substan
tially in comparison to Sobs, which, however, was most
precise (Fig. 3c–d). The least biased estimators, Jack3,
Jack4, and Jack5 (Fig. 3a–b), were markedly less pre
cise than the other estimators (Fig. 3c–d), whereasmore
precise estimators like Jack1, Jack2, and ICE reduced
bias less than the other estimators did except Sobs(Fig.
3a–b). Chao2 was particularly imprecise under low
sample coverage, but was among the most precise es
timators under high sample coverage (Fig. 3d). Chao2
reduced bias less under high sample coverage than the
jacknife estimators. At sample coverage ?95%, all
jacknife estimators had a positive bias. With respect to
accuracy, the estimators showed a more complex re
sponse to sample coverage: the estimators’ curves in
tersected, and for different levels of sample coverage,
different estimators were most accurate (Fig. 3e–f).
These intersections indicated that the most accurate
estimators were Jack5 up to a sample coverage of 26%,
Jack4 up to 38%, Jack3 up to 50%, Jack2 up to 74%,
Jack1 up to 96%, and Sobs? 96% (Fig. 3e–f, intersec
tions are accurate ?1%).
Field studies
To test the predictions of our simulations, we ana
lyzed data from two field studies on plant communities
that provided independent information about plant in
ventories within plots and Strueof the entire habitats.
The first study was carried out on 12 temporary wet
lands in East German farmland with a mean habitat
area of 1078 m2(Brose 2001). Systematic transect
searches (five times) were used to quantify Strue, and
nine 0.25m2plots per study site were installed in a
stratified random design. The plant communities were
characterized by a mean Strueof 30 species (range: 19–
44 species) and a mean evenness of 0.92 (range: 0.91–
0.93), which is comparable to the 50species broken
stick communities. The mean sample coverage was
75%, which is between the mean values of the simu
lated brokenstick communities sampled with low
(45%) and intermediate intensity (90%). The results
are very consistent with the simulations predicting that
(1) in brokenstick communities of 50 species, under
low to intermediate sampling intensity, the jacknife es
timators are the least biased (Figs. 2 and 4a), (2) im
precision increases with increasing order of the jacknife
estimators (Figs. 2, 3c–d, and 4a), and (3) under a
sample coverage of 75% Jack1 and Jack2 are the most
accurate estimators (Figs. 3e and 4a).
The second field study was carried out on five dry
grasslands on mountain slopes in the area of Pe ´cselyi,
?20 kilometers north of the Balaton in central Hun
gary. The mean habitat area was ?1 km2. The vege
tation was surveyed three times in 1999 and ranged
from subalpine communities to dry grasslands. On each
of the five grasslands, 40 1m2plots stratified with re
spect to altitude and soil depth were installed. Struewas
obtained by combining observed species lists with a
multiyear survey (Bauer 2000). Strue(mean: 133 spe
cies, range: 83–177 species) was bracketed by our sim
ulated Struevalues of 100 and 150 species. Their extreme
unevenness (mean ? 0.27; range: 0.26–0.27) is com
parable to our randomfraction and randomassortment
communities. Sample coverage (mean 73%) is higher
than in any of the simulated uneven communities. Ran
domly selecting 8 plots per grassland generated a less
sampled version of this data set. Species richness was
estimated in 20 replications and the mean estimates
were calculated for each estimator. Sample coverage
Page 8
September 20032371
ESTIMATING SPECIES RICHNESS
FIG. 3.
were determined for 1% ranges. Jsel? Jacksel.
Bias, imprecision, and inaccuracy of the nonparametric estimators as functions of sample coverage. All values
(mean 50%) was comparable to intermediate sampling
of our randomfraction simulations (53%, 50 species;
43%, 250 species). In accordance with our simulations,
the jacknife estimators were less biased than the other
estimators except Jack1, which was more biased than
Chao2 under intermediate sampling intensity with 8
samples (Figs. 1 and 4b–c). Most results are consistent
with our simulations predicting that (1) bias and pre
cision decreased with increasing jacknife order (Figs.
1, 2, and 4b–c), (2) under a sample coverage of 50%
Jack3 is among the most accurate estimators (Figs. 2
and 4b), and (3) under 73% sample coverage Jack2
performs most accurately (Figs. 3e and 4c). However,
Jack4 and Jack5 under intermediate sampling and Jack3
under high sampling performed more accurately than
in our simulations.
DISCUSSION
Our general results have shown that most estimators
perform significantly better than Sobs, and that the non
parametric estimators are substantially less biased and
more precise than the extrapolations of species accu
mulation curves. This is consistent with previous stud
ies (Palmer 1990, Baltana ´s 1992, Colwell and Cod
dington 1994, and with some exceptions Walther and
Morand 1998 and Longino et al. 2002). In our study,
it was shown that these results hold for a large range
of communities with differences in evenness and spa
tial patterns. In contrast to field studies, these param
eters are not only exactly known in the simulated land
scapes used in the present study, they were also sys
tematically controlled as independent variables. Based
on this approach, it was shown that species richness
estimation depends on the relative abundance distri
bution, Strue, and sampling intensity, but not on spatial
patterns.
Bias, precision and accuracy
Our results indicate that choosing estimators depends
on whether biasreduction or precision is more impor
tant. While Jack5 was the least biased and least precise
estimator, Sobswas most precise and most biased (see
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ULRICH BROSE ET AL.
Ecology, Vol. 84, No. 9
FIG. 4.
temporary wetlands, and dry grasslands in (b) a reduced data set with eight samples per grassland, and (c) the full data set
with 40 samples per grassland. Jsel? Jacksel.
Bias (mean), imprecision (standard deviation of bias), and inaccuracy (mean ? 1 SD) of the estimators on (a)
Fig. 3). In our simulations as well as in the field studies,
both estimators would have yielded problematic re
sults. Bias indicates systematic over or underestima
tion of the mean estimation when several independent
estimates are taken into account (Sokal and Rohlf
1995). Precision measures the similarity of independent
estimates, and therefore, it indicates the repeatability
of an estimate, whether or not it is biased. Biodiversity
studies typically prevent systematic evaluation of bias
and precision by restricting evaluation to one data set
rather than several independent estimates as in our
study. Accordingly, estimators used in such biodiver
sity studies not only need to be as unbiased as possible,
they also need to be reliable in terms of precision in
order to reduce arbitrary variation while conducting
similar surveys or experiments. Accordingly, choosing
Page 10
September 2003 2373
ESTIMATING SPECIES RICHNESS
FIG. 5.
of three hypothetical estimators. In panels (b)–(d) these three
estimators were used for regression analyses representing a
hypothetical relationship y ? x: (b) Estimator1 (regression:
y ? x, r2? 0.68, P ? 0.01); (c) Estimator2 (y ? 0.6x, r2?
0.93, P ? 0.001); (d) Estimator3 (y ? 0.95 x, r2? 0.95, P
? 0.001).
(a) Bias (mean) and precision (standard deviation)
the least biased estimator may be appropriate for stud
ies in which several independent data sets are available
to yield one mean estimate. This might be typical when
the concern is to estimate global biodiversity or the
species richness of an entire biome. In contrast to this,
field studies at smaller spatial scales are typically based
on single data sets, which makes precision critically
important. Our results indicate that there is a tradeoff
between biasreduction and precision of the estimators.
Following Sokal and Rohlf (1995), we feel that eval
uations should be based on accuracy defined as a com
bination of bias and precision. Similar to bias, accuracy
measures the deviation of the estimate from Strue, but
without the sign, which prevents underestimates from
averaging with overestimates to yield a very low mean
bias.
Fig. 5 illustrates this critical issue in term of three
hypothetical estimators: Estimator1 is unbiased, im
precise and inaccurate (bias ? 0, imprecision ? 0.5,
inaccuracy ? 0.23), Estimator2 is biased, precise, and
inaccurate (bias ? ?0.4, imprecision ? 0.1, inaccuracy
? 0.17), Estimator3 is slightly more biased than Es
timator1, slightly less precise than Estimator2, and
most accurate of the three estimators (bias ? ?0.05,
imprecision ? 0.15, inaccuracy ? 0.02). Although Es
timator1 is unbiased on average, its individual esti
mates are usually highly biased and over or under
estimate Struesubstantially (Fig. 5a). In contrast to this,
Estimator2 consistently underestimates Strue. Only the
accurate Estimator3 consistently provides unbiased es
timates. To see the consequences of using biased, im
precise, or accurate estimators, we then applied these
three estimators to a hypothetical study gathering data
that follows a relationship of y ? x (Fig. 5b–d). The
imprecise Estimator1 yielded a correct regression
slope, but a comparatively low coefficient of deter
mination and a lower significance level than the other
estimators (Fig. 5b). The biased Estimator2 provided
a higher r2value, but markedly underestimated the re
gression slope (Fig. 5c). Only the accurate Estimator3
yielded the approximately correct regression slope and
the highest r2value (Fig. 5d). In actual field studies
that include additional sources of variation, these ef
fects are most likely even stronger. This example dem
onstrates that basing estimator choice only on bias or
precision is potentially misleading, while accuracy
shows that the estimates are constantly close to the true
value. Our results indicate that accuracy benefits from
biasreduction under low sample coverage when all es
timators have a substantial negative bias, and precision
under high sample coverage when all estimators are
less biased.
Impact of the relative abundance distributions and
spatial patterns
Our results strongly corroborate the hypothesis that
uneven detection probabilities affect estimator perfor
mance: the relative abundance distribution was the
most important factor influencing estimator bias. This
is consistent with Boulinier et al. (1998) who concluded
that the capture–recapture model Mh(Otis et al. 1978),
which assumes heterogeneous detection probabilities
of the species, is generally the most appropriate for the
estimation of species richness. Accordingly, the non
parametric estimators based on model Mhare best suit
ed to deal with heterogeneous detection probabilities
but, as our results indicate, these estimators are still
substantially biased. This bias is most likely caused by
the fact that the nonparametric estimators have origi
nally been developed to estimate population size in
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ULRICH BROSE ET AL.
Ecology, Vol. 84, No. 9
closed capture–recapture studies. However, detection
probabilities among different species are likely to be
more uneven than the detection probabilities of differ
ent individuals of one species. Our results are also con
sistent with Heltshe and Forrester (1983) and Baltana ´s
(1992), who found that the dispersion of the lognormal
distribution in communities affects the bias of the first
order jacknife estimator. Our study has expanded this
finding to (1) all other estimators studied, (2) com
munities with high species richness, (3) communities
of a broader evenness range, and (4) communities with
strong spatial gradients and autocorrelation. Therefore,
heterogeneous detection probabilities caused by un
even relative abundance distributions pose a major
challenge for the extrapolation of species richness.
Even for the nonparametric estimators based on model
Mh, a more complex extrapolation procedure that cor
rects for heterogeneous detection probabilities is re
quired for comparing species richness and derived mea
sures (e.g., connectance, Martinez et al. 1999) between
communities that vary in evenness. Until evenness
corrected estimators are developed, our findings assist
the choice of appropriate estimators based on the even
ness in the sample. We found that the most accurate
estimators are Jack1 and Jack2 in highly even broken
stick and Jack2, Jack3, and Jack4 in the uneven random
fraction or randomassortment communities. Jack2 ap
pears to be the best overall estimator with respect to
accuracy.
To allow a detailed impact analysis of spatial pat
terns, we distinguished spatial autocorrelation from
gradients. Gradients restrict the distribution of species
below the variation permissible in random and auto
correlation communities. Therefore, the commonality
of two sampling plots on a gradient will be generally
lower than in plots lacking a gradient. Furthermore, we
differentiated the effects of the strengths of autocor
relation and its range. However, our results suggest that
spatially heterogeneous species distributions mediated
by spatial autocorrelation and gradients do not affect
estimator bias. This is in contrast to previous studies
in which spatial autocorrelation strength had a small
but significant effect on estimator bias (Baltana ´s 1992,
Chazdon et al. 1998) or diversity indices (Fager 1972).
The estimator studies, however, varied autocorrelation
strength more than the present study, and testing the
effects of observed, randomized or moderately patchy
species distributions resulted mainly in nonsignificant
differences (Chazdon et al. 1998). Furthermore, Chaz
don et al. (1998) have shown that abundancebased
estimators are more susceptible to be affected by patchy
distributions than the incidencebased estimators used
in the present study. Our results suggest that various
different spatial aggregation patterns in a realistic pa
rameter space are not likely to have a substantial impact
on estimator performance. This is consistent with an
empirical study that concluded that capture–recapture
models assuming spatial heterogeneity are rarely ap
propriate for the estimation of species richness (Bou
linier et al. 1998). Accordingly, violating the assump
tion of spatially homogeneous species distributions
does not seriously challenge the extrapolation of spe
cies richness.
Sample coverage
In our study, no single estimator performed best
across the different levels of community evenness, Strue,
and sampling intensity, and these variables should be
addressed simultaneously while evaluating estimators.
Our results suggest that community evenness, Strue, and
sampling intensity indirectly influence the nonpara
metric estimators through effects of sample coverage.
Surprisingly, there was only a minor influence of Strue,
which indicates that the nonparametric estimators are
well suited to estimate species richness in communities
with substantially different levels of Strue. Accordingly,
sample coverage mainly combines the effects of sam
pling intensity and community evenness on the esti
mators. With respect to sample coverage, studying a
more even community has the same effect as studying
an uneven community with higher sampling intensity.
Therefore, our results concerning sample coverage are
independent of the specific levels of community even
ness and sampling intensity chosen, and represent the
full range of theoretically possible data sets. Knowing
the exact sample coverage in a case study would make
the decision for a specific estimator straightforward,
but, of course, unnecessary. However, when sample
coverage is approximately known, our results indicate
which estimators are most accurate (Fig. 6). Although
the bias of each estimator was consistent for each level
of sample coverage, biascorrection based on sample
coverage is not reliable, as it would add the error of
sample coverage estimation to the error of the species
richness estimates.
Sobsis usually not the best choice to measure species
richness, and in most data sets, ICE and Chao2 were
less accurate than the jacknife estimators. The low ac
curacy of Chao2 was caused by a low precision under
low sample coverage and high bias under high sample
coverage. Surprisingly, Chao2, Jack3, Jack4, and Jack5
performed much more precisely in our general evalu
ation (Table 1) than in the more detailed evaluations
(Figs. 2 and 3). This is caused by the fact that the data
sets with their most imprecise performances were ex
cluded from the regression analyses in the general eval
uation to avoid heteroscedasticity. Therefore, the gen
eral precision values in Table 1 provide a relatively
flattering image of their precision, and the detailed re
sults shown in Figs. 2 and 3 are more meaningful.
In empirical studies, there has been some disagree
ment about the relative performance of the nonpara
metric estimators: while Palmer (1990) concluded that
Jack1 is the least biased estimator, Colwell and Cod
dington (1994) argued in favour of Chao2 and Jack2
and Walther and Morand (1998) preferred Chao2 and
Page 12
September 2003 2375
ESTIMATING SPECIES RICHNESS
FIG. 6.
coverage or the community evenness. Sobs? observed species richness. Thresholds are accurate within ?1%.
Two alternative ways to choose the most appropriate estimator based on either an estimated range of sample
Jack1. Our results indicate that this might stem from
the differences in their sample coverage. The field data
sets in the present study give an example that evalu
ations under different sampling intensity, community
evenness, and consequently different sample coverage
yield different results concerning estimator ranking. In
contrast to this, in simulated landscapes, Strue, sampling
intensity, and community evenness are not only un
questionably known, they are also controlled as inde
pendent variables. This has led to the new estimator
performance model based on sample coverage in the
present study.
All jacknife estimators were positively biased above
95% sample coverage. This has previously been re
ported in field studies by Heltshe and Forrester (1983),
Colwell and Coddington (1994), and Hellmann and
Fowler (1999), but with different magnitudes. In con
trast to this, the jacknife estimators did not overesti
mate in the studies of Palmer (1990, 1991) and Baltana ´s
(1992). Potential explanations include the following.
First, differences in the sample coverage between the
studies might be responsible. Second, particularly in
field studies, the determination of Strueis prone to un
derestimation. Such lower Struemay account for some
of the jacknife estimator’s overestimations. In our sim
ulations, however, the first and secondorder jacknife
estimators were less positively biased under high sam
ple coverage than previously reported (e.g., Hellmann
and Fowler 1999).
Estimator choice
We will subsequently present a framework for the
choice of the most appropriate nonparametric estimator
depending on sample coverage (see Fig. 6). However,
unless the sample contains the entire community, one
can only guess about the sample coverage, which
makes the estimator choice more problematic. An es
timator evaluation based on sampling intensity, i.e., the
percentage of total quadrats sampled (Hellmann and
Fowler 1999), only provides an incomplete framework,
since sampling intensity is only one of the factors in
fluencing sample coverage. This framework remains
specific for the respective levels of community even
ness, Strueand habitat area. As independent reliable in
formation about Strueis scarcely obtainable in field stud
ies, the most promising procedure to choose the ap
propriate estimator is iterative (Fig. 6, path 1). First, a
range of Strueneeds to be calculated by various esti
mators. Then, the most accurate estimator can be cho
sen based on the resulting mean estimated sample cov
erage. Applying this procedure to the field data sets
yields mean inaccuracies of 1.5% of Struefor the tem
porary wetlands and 3.5% of Strueand 9.6% of Struefor
the dry grasslands in the full and reduced data set,
respectively. While these estimates are the most ac
curate for the temporary wetlands, they are only slight
ly outperformed by Jack2 and Jack3 on the dry grass
lands. Compared to a decision based on community
evenness (Fig. 6, path 2), the decision path based on
estimated sample coverage (Fig. 6, path 1) is more
Page 13
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ULRICH BROSE ET AL.
Ecology, Vol. 84, No. 9
elaborate. For every level of sample coverage just one
estimator is most accurate. However, uncertainty of
sample coverage might force researchers to rely on path
2 if the estimated range of sample coverage is large.
In contrast to this, estimates of evenness are highly
accurate even when based on very few samples. For
example, under low sampling intensity evenness was
calculated with a mean absolute proportional deviation
of 9.8% in our simulated landscapes. Although the de
cision based on evenness is more reliable, it does not
offer conclusive advice, as estimator choice still de
pends on sampling intensity. In conclusion, we rec
ommend that researchers examine both alternatives and
rely primarily on path 1. Although this procedure surely
does not satisfy the desire for a single generally ap
plicable estimator, we do provide a method for choos
ing estimators that should generally increase accuracy
beyond current methods and certainly beyond the wide
spread measurement of species richness by Sobs. Beyond
choosing among available estimators, our in silico
methods could be used to evaluate and design new
estimators to be more highly accurate and also avoid
problems such as consistent negative bias at low sample
coverage.
CONCLUSIONS
Our analyses demonstrate that in most community
types the nonparametric estimators perform substan
tially more accurately than Sobsor the species accu
mulation curves. As specific spatial patterns and Strue
were of minor importance, our results appear spatially
scale invariant, making no differences between ? and
? diversity. Irrespective of scale, our results suggest
that species richness may be calculated accurately un
der intermediate or high sampling intensity if the most
appropriate estimator is chosen. Community evenness
and sampling intensity, however, have a strong influ
ence on estimator accuracy. With respect to these fac
tors, there was not a single generally most accurate
estimator, and an optimal decision has to be based on
knowledge about the sample coverage that can only be
estimated from the samples. There are two possibilities
to deal with this dilemma (Fig. 6). First, species rich
ness can be extrapolated by various estimators to base
the decision about the most appropriate estimator on
the estimated ranges for the sample coverage. Second,
the least sensitive estimator with respect to community
evenness can be chosen. Our study indicates a need for
nonnegatively biased estimators for low sample cov
erage and a more complex extrapolation procedure that
corrects for the very heterogeneous detection proba
bilities of species in natural communities. Until such
estimators are developed, we anticipate that our find
ings will enhance the precision and accuracy of species
richness estimations, while assisting researchers to
choose the appropriate estimator for a particular com
munity structure or sample coverage.
ACKNOWLEDGMENTS
We sincerely thank Jennifer Dunne, Robert K. Colwell,
Nicholas J. Gotelli, J. Bastow Wilson, Sa ´ndor Samu, Michael
Glemnitz, J. F. Quinn, and an anonymous reviewer for their
help and suggestions. U. Brose is supported by the German
Academy of Naturalists Leopoldina by funds of the German
Federal Ministry of Education and Science (BMBFLPD
9901/844). N. D. Martinez and R. J. Williams are supported
by an NSF Biocomplexity Incubation Grant (DEB0083929).
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APPENDIX
A comparison of formulas for the estimators of species richness is available in ESA’s Electronic Data Archive: Ecological
Archives E084057A1.
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