UNCORRECTED PROOF

Diatom-based TP and TN inference models and

indices for monitoring nutrient enrichment

of New Jersey streams

Karin C. Ponader a,

*, Donald F. Charles a

, Thomas J. Belton b

aPatrick Center for Environmental Research, The Academy of Natural Sciences,

1900 Benjamin Franklin Parkway, Philadelphia, PA 19103, USA bNew Jersey Department of Environmental Protection, Science and Research,

401 East State Street, PO Box 409, Trenton, NJ 08625, USA

Received 22 May 2005; received in revised form 1 October 2005; accepted 24 October 2005

Abstract

We evaluated the potential for using diatoms to assess and monitor nutrient enrichment in New Jersey streams and rivers, and

propose inference models and indices for regulatory purposes. We assessed the relationship between benthic diatom and water

chemistry samples (n = 101) collected from 45 sites in 3 ecoregions: Northern Piedmont, Northeastern Highlands and Ridge and

Valley. Diatom assemblages were dominated by pollution-tolerant taxa. Multivariate analysis showed that nutrient concentrations explained significant proportions of the variation in diatom species composition. Weighted-averaging partial least square

(WA-PLS) total phosphorus (TP) and total nitrogen (TN) inference models (n = 91) showed good predictive ability (TP model:

r2

apparent 1⁄4 0:87; r2boot 1⁄4 0:72; RMSEboot = 0.23 log10 mg L1 TP; TN model: r2apparent 1⁄4 0:88; r2boot 1⁄4 0:58; RMSEboot = 0.23

log10 mg L1 TN). Diatom TP and TN indices were created to simplify presentation of results for the general public by rescaling

the inferred TP and TN values from 0 to 100. The obtained index scores were assigned to nutrient impairment categories for

regulatory assessment purposes.

# 2005 Elsevier Ltd. All rights reserved.

Keywords: Diatoms; Nutrients; Inference model; Index; Streams; USA

1. Introduction

New Jersey (NJ) is the most densely populated state

in the US, with 2937 persons km2 (US Census

Bureau, 2002). Rivers and streams receive high

nutrient loadings coming from a variety of urban,

residential and agricultural sources (USGS, 2001).

Almost half of the 2179 river miles recently assessed

This article is also available online at:

www.elsevier.com/locate/ecolind

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* Corresponding author. Tel.: +1 215 405 5077;

fax: +1 215 299 1079.

E-mail addresses: ponader@acnatsci.org (K.C. Ponader),

charles@acnatsci.org (D.F. Charles),

Thomas.Belton@dep.state.nj.us (T.J. Belton).

1470-160X/$ – see front matter # 2005 Elsevier Ltd. All rights reserved.

doi:10.1016/j.ecolind.2005.10.003

ECOIND 205 1–15
UNCORRECTED PROOF

by the NJ Department of Environmental Protection

(NJDEP) did not meet standards of less than

0.1 mg L1 TP (NJDEP, 2003). These nutrient loadings and resulting algal growth can render waters

unsuitable for NJs’ designated uses, such as potable

water supply, recreational use, fisheries, and aquatic

life (NJDEP, 2001).

Monitoring of nutrient levels in rivers and

streams is problematic because of periodic and

diffuse input from non-point sources (Cattaneo and

Prairie, 1995; Che ́telat and Pick, 2001). Benthic

diatom species composition responds directly to

nutrients (Pan and Lowe, 1994; Pan et al., 1996;

Stevenson and Pan, 1999), and can be a more stable

indicator of trophic state than measurements of

nutrient concentrations or algal biomass (USEPA,

2000b).

The goal of this study was to develop robust

diatom-based tools to monitor and assess nitrogen and

phosphorus concentrations and periphyton responses

in NJ streams and rivers. It was therefore important

that these tools: (a) accurately characterize nutrient

concentrations, (b) be consistent with NJ nutrient criteria categories, (c) be practical for routine monitoring, and (d) could be used in conjunction with other NJ

bioindicators.

We chose weighted averaging (WA) inference

modeling (ter Braak and Juggins, 1993) as our

approach because it is the most accurate method for

quantifying species response to nutrients (Hall and

Smol, 1992; Dixit and Smol, 1994; Reavie et al.,

1995; Pan et al., 1996; Winter and Duthie, 2000;

Bennion et al., 2001; Bradshaw and Anderson, 2001;

Soininen and Niemela ̈, 2002). We used the nutrient

concentrations inferred from the models in two ways.

First, we used the inferred values as estimates of the

nutrient concentrations prevailing at the site during

the time the algal assemblages were developing.

Second, we rescaled the inferred concentrations from

0 to 100 to create diatom TP and TN indices to

provide a management tool that is easily interpretable and partitioned into categories of protection

for critical designated uses under the Clean Water

Act (U.S. Code, 2002), and that are similar to

existing Trophic Diatom Indices (TDI’s) (Descy and

Coste, 1990; Schiefele and Schreiner, 1991; Kelly

and Whitton, 1995; Coring et al., 1999; Rott et al.,

2003).

2. Methods

2.1. Study sites

This study was conducted in northern and central NJ,

USA, in the Northern Piedmont, the Northeastern

Highlands and the Ridge and Valley ecoregions

(Omernik, 1987, 1995) (Fig. 1). Land-use in the

Piedmont is primarily urban and agriculture, whereas in

the Highlands and the Ridge and Valley it is predominantly forest and agriculture (USEPA, 2000d).

Because of the approach we used to develop algal

indicators of anthropogenic nutrient enrichment, it

was important to select study sites with relatively

similar natural environmental conditions (e.g. geology, geomorphology), but with a wide range of

nutrient concentrations. We based our selection of

sites on chemistry data available from the NJDEP and

US Geological Survey (USGS) monitoring networks.

All selected sites were part of an Ambient Biomonitoring Network (AMNET) (NJDEP, 2000), and

reflected a range of known biological impairments

based on macroinvertebrate data collected between

1992 and 1999 (NJDEP, 1993b, 1999). All sites

selected were first to sixth order wadeable rivers and

streams. We sampled 45 AMNET sites: 12 in the

Highlands, 5 in the Ridge and Valley and 28 in the

Piedmont (Fig. 1). Physical characteristics, water

chemistry and other site information are summarized

in Table 1. Detailed site information, including site

name and location, and physical and water chemistry

data are available at http://www.diatom.acnatsci.org/

autecology/.

2.2. Field and laboratory procedures

Samples of epilithic diatoms were collected from

August through October 2000–2002. The sampling

area at each site was divided into three sections, so that

within-site variability could be assessed. For each

section, we estimated the percent of each substrate

type (boulder, cobble, gravel, sand, silt, bedrock) to

the nearest percent using habitat assessment protocols

adapted from Barbour et al. (1999). We categorized

flow velocity as slow, moderate, or fast and measured

light conditions (percent open canopy cover) using a

spherical densiometer. One quantitative composite

biomass sample was collected from three rocks per

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UNCORRECTED PROOF

section for measurement of chlorophyll a (chl a) and

ash-free dry mass (AFDM). Rock surfaces were

sampled using the ‘top-rock scrape method’ (Moulton

et al., 2002).

Water chemistry samples were collected at the time

of algal sampling. One unfiltered and one filtered

sample (250 mL each) were taken for nutrient

analysis. Filtering was done using a plastic syringe

with an attached filtration device using a 0.7 mm

Whatman glass microfiber filter. In addition, one

unfiltered 500 mL sample was collected for analysis of

major anions and cations. Laboratory analysis of

dissolved nitrate + nitrite (NO3 + NO2, here referred

to as NO3-N), total Kjeldahl nitrogen (TKN),

dissolved ammonia (NH3-N), orthophosphate (O-P),

total phosphorus (TP), chloride (Cl), total alkalinity,

total hardness, and conductivity was performed by the

Patrick Center for Environmental Research (PCER)

Geochemistry Section under the direction of Dr. D.

Velinsky following USGS methods (Fishman, 1993).

Total nitrogen (TN) was calculated from a combination of TKN and NO3 + NO2. During 2000 and 2001,

samples for TKN analysis were collected by NJDEP

and USGS. During 2003, TKN samples were collected

at the time of algal sampling and analyzed by PCER.

Chl a and AFDM samples were analyzed by PCER

using standard methods (APHA et al., 1995) and

USEPA method 445 (USEPA, 1992). Additional water

chemistry data for water quality monitoring stations

were provided by NJDEP and USGS staff. In 2002, we

measured conductivity and pH in the field using a

portable meter (Oakton pH/CON 10 Multiparameter

Meter). All other chemistry and land-use data (percent

land-use type per watershed) were provided by the

NJDEP. We used water chemistry data for samples that

were collected closest to the time of algal sampling.

Diatom samples were collected from natural rock

substrates using techniques consistent with those used

in the USGS NAWQA program (Moulton et al., 2002)

and those recommended by the USEPA (Barbour et al.,

1999). A composite diatom sample was created by

randomly selecting four to five rocks of ca. 5 cm

diameter from mid-stream. Diatoms were removed

from the rocks by scraping and brushing. Diatoms were

cleaned, permanently mounted on microscope slides,

and counted following USGS NAWQA protocols

K.C. Ponader et al. / Ecological Indicators xxx (2005) xxx–xxx 3

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Fig. 1. Geographic location of study sites and ecoregions.
UNCORRECTED PROOF

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Table 1Water chemistry parameters and physical site characteristics of all 45 sampling sites by ecoregionParameter Northern Piedmont Highlands Ridge and Valley All ec

Mean Median Minimum Maximum Mean Median Minimum Maximum Mean Median Minimum Maximum Mean 

NO3-N (mg L1) 1719 1307 232 5525 1013 441 22 7992 178 198 10 361 1524 

TN (mg L1) 2241 1856 362 6200 1272 818 247 8547 416 538 170 577 1987 

O-P (mg L1) 98 60 4 663 28 13 2 146 5 4 2 10 81 

TP (mg L1) 127 75 12 731 41 32 6 177 15 14 8 29 107 

NH3-N (mg L1) 47 24 3 177 12 5 2 52 4 4 2 7 39 

Chl a (mg m2) 124 85 3 1115 60 70 7 132 73 42 15 248 111 

AFDM (g m2) 25 17 4 153 10 9 4 17 14 7 5 39 22 

Hard (mg L1) 102 90 45 280 100 107 25 240 157 136 46 306 105 

Alk (mg L1) 69 64 20 186 66 64 7 182 109 90 30 236 70 

DO (mg L1) 7.5 8.0 2.8 11.3 8.5 9.0 4.9 11.3 7.4 8.8 2.8 10.0 7.

pH 7.6 7.6 6.5 9.0 7.6 7.7 6.8 8.2 7.9 7.8 7.7 8.1 7.Cond (mS cm1) 341 285 93 983 253 244 51 663 369 291 121 741 327 

Cl (mg L1) 50.0 42.7 2.8 240.0 27.5 21.4 3.6 69.0 21.5 17.3 1.4 55.8 44.

BASIN (km2) 190 62 1 1962 45 31 5 190 139 76 10 417 163 

URB (%) 47 45 6 92 22 21 4 60 8 8 2 14 41 AG (%) 14 15 0 56 12 11 0 21 13 11 0 29 14 FOR (%) 29 23 3 93 51 51 20 76 64 70 30 86 34 Open (%) 48 31 1 100 27 20 1 77 53 53 29 75 45 BRBD (%) 9 5 0 51 28 13 0 72 36 30 10 60 13 CBGR (%) 60 65 0 95 48 50 28 70 46 50 30 60 57 SDSTCL (%) 31 27 0 100 23 27 0 55 18 20 10 30 29 AvgW (m) 14.1 9.0 2.8 50.0 6.8 8.0 2.0 10.0 6.8 8.0 2.0 12.5 12.SecL (m) 47.8 40.0 5.0 125.0 30.6 30.0 15.0 40.0 26.0 20.0 20.0 40.0 43.Flow (estimate) 1.9 2.0 1.0 3.0 1.8 2.0 1.0 2.5 1.7 1.5 1.5 2.0 1.NO3-N: dissolved nitrate + nitrite; TN: total nitrogen; O–P: orthophosphate; TP: total phosphorus; NH3-N: dissolved ammonia; Chl a: chlorophyll a; AF

hardness; Alk: total alkalinity; DO: dissolved oxygen; Cond: conductivity; Cl: chloride; BASIN: basin size; URB: % urban land-use; AG: % agriculture 

Open: % open canopy cover; BRBD: % bedrock and boulder; CBGR: % cobble and gravel; SDSTCL: % sand silt and clay; AvgW: average stream widtestimate, mean per section (1 = slow, 2 = medium, 3 = fast).
UNCORRECTED PROOF

(Charles et al., 2002). Per slide, 600 valves were

identified to lowest taxonomic level and their relative

abundance was recorded.

2.3. Numerical analysis

2.3.1. Dataset and data transformations

The complete dataset used in the analysis contained

101 samples from 45 sites: 45 samples were

‘independent’ samples, e.g. they were collected from

section 1 of the respective sites; another 45 samples

were ‘multiple samples’, e.g. samples collected from

sections 2 and 3 of the respective sites. Eleven samples

were ‘repeat samples’, e.g. samples from sites that had

been sampled once in 2000, and which were revisited

during the summers of 2001 and/or 2002 to provide a

2–3-year record of species composition for comparison. Multiple and repeat samples were collected to

capture a bigger range of within-site as well as amongyear variability. Of the total 101 samples, 5 were

collected in the Ridge and Valley, 15 were collected in

the Highlands and 81 were collected in the Piedmont

physiographic provinces. The sites for which multiple

and repeat samples were taken were distributed evenly

along the measured TP and TN gradients and therefore

the inclusion of these additional samples did not skew

the dataset to either end of the nutrient range. Diatom

taxa were included in the analysis if their abundance

was 0.5% in at least two samples. All relative

abundance data were square-root transformed. The

water chemistry and physical parameters included in

the ordinations are listed in Table 1. Three substrate

categories were formed by combining the percentages

of two or three substrate types: % bedrock and boulder

(BRBD), % cobble and gravel (CBGR), and % sand

silt and clay (SDSTCL). Environmental variables

were transformed to reduce skewed distributions: all

water chemistry variables, except pH, were log10-

transformed; all data expressed in percentages (landuse, substrate, and open canopy cover) were squareroot transformed.

2.3.2. Ordinations

Principal components analysis (PCA) was performed to detect major gradients and principal

patterns of variation within the environmental variables (ter Braak and Prentice, 1988). The environmental variables were centered and standardized. In

the same PCA, ‘‘outliers’’ or ‘‘rogues’’ were defined as

samples with extreme sample scores on any of the first

four axes of the PCA of the environmental data (Birks

et al., 1990). Extreme sample scores were defined as

scores falling above the 95% confidence interval of all

sample score means (Winter and Duthie, 2000).

Detrended correspondence analysis (DCA) with

detrending by 26 segments and down-weighting of

rare taxa was used to examine patterns in the diatom

data, and to determine the maximum amount of

variation within the species composition data (ter

Braak, 1995). We used the gradient length of the main

DCA ordination axes to determine whether linear or

unimodal techniques were to be applied for modeling

the relationship between diatoms and environmental

variables (ter Braak and Prentice, 1988). Furthermore,

DCA was used to determine outliers, e.g. samples that

showed extreme sample scores on any of the first four

axes of the DCA of the species data (Birks et al.,

1990). Outliers were screened using the same criteria

as applied in PCA. All samples determined as outliers

by PCA and DCA were excluded from all subsequent

ordinations and from the development of calibration

models.

A series of correspondence analyses (CA) and

canonical correspondence analyses (CCAs) with

down-weighting of rare taxa was performed, in order

to determine the variables that independently

explained a significant amount of variation in diatom

species composition (ter Braak, 1995). First, a CA was

run with passive environmental variables to determine

the strength of the correlations among all 24

environmental variables and all 131 species, as well

as to identify variables that were intercorrelated, based

on weighted correlations and variance inflation factors

(VIFs). Variables with VIFs > 5 indicated strong colinearity among environmental variables and were

removed from all subsequent analyses. The data from

the remaining 13 variables and all sites were included

in a CCA analysis with forward selection in order to

identify the minimal number of variables that

explained the largest amount of variation in the

diatom species data. Unrestricted Monte Carlo

permutation tests (999 permutations) were used to

assess the statistical significance of each forward

selected variable (P < 0.05). As a last step, to assess

the strength of the relationship between diatom

species composition and the forward selected variK.C. Ponader et al. / Ecological Indicators xxx (2005) xxx–xxx 5

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UNCORRECTED PROOF

ables, we ran CCAs constrained to one variable at a

time. A high ratio between the first (constrained)

eigenvalue and the second (unconstrained) eigenvalue

(e.g. l1/l2 > 0.5) indicated strong influence of these

variables on diatom species composition, and justified

development of inference models (Bigler and Hall,

2002; Winter and Duthie, 2000). All ordinations were

produced using Canoco for Windows, version 4.5 (ter

Braak and Sˇmilauer, 2002).

2.3.3. Species optima and tolerances, and

inference models

Weighted averaging (WA) regression and calibration techniques were used to calculate diatom species

optima and tolerances as well as to develop and test

diatom inference models for TP and TN. The models

were developed using the program C2

, version 1.3

(Juggins, 2003). Several models were constructed,

based on WA inverse and classical deshrinking (Birks

et al., 1990) and on weighted-averaging partial least

square regression (WA-PLS) (ter Braak and Juggins,

1993). Model error estimation was performed by

bootstrapping with 1000 cycles (Birks, 1995). We

selected the best inference model (e.g. the ‘minimal

adequate model’ sensu Birks, 1998), based on a

combination of the following characteristics: (a) the

highest prediction accuracy, e.g. the lowest root mean

square error of prediction (RMSEP) based on crossvalidation; (b) the highest coefficient of determination

(r

2

) between observed and inferred values, and (c) the

minimum number of WA-PLS components (Birks,

1995).

2.3.4. Index development

We created diatom TP and TN indices based on

values inferred for each diatom sample using the twoand three-component WA-PLS models (n = 91). Index

values were calculated by multiplying the inferred

nutrient values (log10) obtained for each sample by a

constant that converted them to a scale from 0 to 100

(TP index = 33.33  inferred log10 TP); TN index =

50  (inferred log10 TN-2). The 0–100 scales correspond to log10 TP from 1 to 1000 mg L1 and log10 TN

from 100 to 10,000 mg L1

. The ranges for both

parameters were set slightly above the maximum and

below the minimum values in our dataset.

We divided the 0–100 scale into trophic state

categories ranging from 1 (low) to 3 (high) for TN, and

from 1 (low) to 4 (very high) for TP, with respect to the

range of values in the calibration dataset. The category

boundaries are those suggested by Wetzel (2001) for

trophic classification of temperate streams, based on

the biomass–nutrient relationships established in

Dodds et al. (1998). The four TP categories established were (1) less than 0.025 mg L1 TP, (2) 0.025–

0.075 mg L1 TP, (3) 0.075–0.1 mg L1 TP, and (4)

above 0.1 mg L1 TP. These correspond to the

following log10 TP values: (1) less than 1.4, (2)

1.4–1.9, (3) 1.9–2.0, and (4) above 2.0, and to the

index scores of (1) 0–47, (2) 47–63, (3) 63–67, and (4)

above 67. The lower boundary of category 4

(0.1 mg L1 TP) is also the nutrient criteria for NJ.

Because the TN gradient was shorter, we established

only three categories for TN, from 1 (low) to 3 (high).

The final TN categories established were (1) less than

0.7 mg L1 TN, (2) 0.7–1.5 mg L1 TN, and (3) above

1.5 mg L1 TN. These correspond to the following

log10 TN values: (1) less than 2.8, (2) 2.8–3.2, and (3)

above 3.2, and to the index scores of (1) less than 42,

(2) 42–59, and (3) above 59.

3. Results

3.1. Diatom species composition

In all, 399 diatom taxa were identified. Of these,

131 taxa were included in the data analysis. The 10

most abundant species, determined by high abundances and high effective numbers of occurrences

based on Hill’s N2 (Hill, 1973), had high TP and TN

WA-optima (Table 2). Pollution-tolerant taxa dominated.

3.2. Environmental gradients and species

distributions

We identified two major gradients in the environmental data, based on PCA. The first PCA axis (30.6%

variance explained) reflected a chemical gradient

driven by all inorganic nutrients measured, % urban

land-use, conductivity and chloride. The second axis

(16.7% variance explained) revealed a physical

gradient influenced by average river width and basin

size, % open canopy cover, and % sand, silt, and clay

substrates. In the DCA of diatom data, the eigenvalues

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K.C. Ponader et al. / Ecological Indicators xxx (2005) xxx–xxx 7

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Table 2

Species apparent optima and tolerances estimated by WA

Taxon name Hill’s N2 TP (mg L1

) TN (mg L1)

Optima Tolerance Optima Tolerance

Achnanthidium rivulare Potapova and Ponader 27 25.26 2.62 956.27 2.72

Navicula antonii Lange-Bertalot 15 30.19 2.12 919.77 1.75

Cocconeis placentula var. euglypta (Ehrenberg) Cleve 14 30.51 2.49 834.31 2.41

Nitzschia fonticola Grunow 16 30.86 2.64 915.29 2.28

A. minutissimum (Ku ̈tzing) Czarnecki 48 32.54 2.62 1016.46 2.47

Encyonema minutum (Hilse) Mann 25 35.05 2.33 1040.66 2.13

N. cryptocephala Ku ̈tzing 21 37.25 2.52 1241.02 2.65

N. cryptotenella Lange-Bertalot 44 38.84 2.35 1155.82 2.16

Cymbella tumida (Bre ́bisson ex Ku ̈tzing) Van Heurck 13 39.79 2.60 890.14 2.13

C. pediculus Ehrenberg 26 40.30 2.72 1097.97 2.67

N. decussis Østrup 18 42.34 2.23 1073.77 2.21

Diatoma vulgaris Bory 23 43.49 2.17 1216.35 1.85

Nitzschia archibaldii Lange-Bertalot 16 43.53 2.91 1152.74 2.16

Nitzschia linearis (Agardh ex W. Smith) W. Smith 11 43.53 2.25 926.96 2.25

Amphora inariensis Krammer 11 45.14 2.41 1255.19 2.22

N. tripunctata (O.F. Mu ̈ller) Bory 27 45.84 2.50 1266.25 2.09

Reimeria sinuata (Gregory) Kociolek and Stoermer 49 47.31 2.24 1279.38 2.04

Gomphonema minutum (Agardh) Agardh 24 47.99 2.39 1322.18 1.98

Synedra ulna Ehrenberg 27 48.13 2.13 1263.80 1.73

N. capitatoradiata Germain 25 50.34 2.17 1274.71 1.91

Nitzschia liebethruthii Rabenhorst 35 51.73 2.50 1498.79 2.12

Nitzschia recta Hantz. ex Rabenhorst 17 51.97 2.62 1980.48 2.22

Planothidium lanceolatum (Bre ́bisson ex Ku ̈tzing) L.-B. 49 52.22 2.14 1472.86 2.09

G. parvulum (Ku ̈tzing) Ku ̈tzing 53 52.68 2.32 1331.37 2.04

Amphora pediculus (Ku ̈tzing) Grunow 42 54.09 2.08 1448.77 1.80

Achnanthes subhudsonis var. kraeuselii Cholnoky 34 55.43 3.14 1472.55 2.28

N. perminuta Grunow 27 56.02 2.11 1639.34 1.71

Melosira varians Agardh 51 58.49 2.25 1538.05 1.83

Caloneis bacillum (Grunow) Cleve 40 59.36 2.53 1554.71 2.12

N. germainii Wallace 42 60.50 2.06 1488.02 1.68

Mayamaea atomus (Ku ̈tzing) Lange-Bertalot 22 60.67 1.66 1601.95 1.46

Nitzschia dissipata (Ku ̈tzing) Grunow 31 62.56 2.42 1606.46 1.85

C. placentula var. lineata (Ehrenberg) Van Heurck 48 63.29 2.24 1602.54 1.78

P. frequentissimum (Lange-Bertalot) Lange-Bertalot 56 63.92 2.25 1641.01 2.03

Staurosirella pinnata (Ehrenberg) Williams et Round 18 64.25 3.47 1323.58 2.57

Rhoicosphenia abbreviata (Agardh) Lange-Bertalot 70 64.46 2.36 1569.24 1.96

Achnanthes conspicua Mayer 35 67.18 1.77 1842.85 1.64

N. erifuga Lange-Bertalot 10 68.02 2.22 1527.64 1.78

Frustulia vulgaris (Thwaites) De Toni 11 68.15 2.09 1438.53 1.65

N. canalis Patrick 21 69.13 2.50 2094.02 1.87

Nitzschia palea (Ku ̈tzing) W. Smith 53 69.92 2.34 1720.68 1.81

N. rostellata Ku ̈tzing 27 71.52 1.82 1779.18 1.56

N. minima Grunow 67 71.86 2.21 1626.86 1.86

N. symmetrica Patrick 46 72.39 2.21 1613.84 1.74

Gyrosigma acuminatum (Ku ̈tzing) Rabenhorst 13 72.67 2.39 1404.91 1.76

Cyclotella meneghiniana Ku ̈tzing 33 73.71 2.74 1491.94 1.89

N. gregaria Donkin 58 74.07 2.23 1764.09 1.78

N. subminuscula (Manguin) 33 74.07 1.98 1621.92 1.54

N. lanceolata (Agardh) Ehrenberg 45 74.94 2.27 1934.33 1.71

N. capitata Ehrenberg 11 75.12 2.22 1560.52 1.70

Nitzschia amphibia Grunow 60 75.42 2.06 1841.20 1.80

Fragilaria vaucheriae (Ku ̈tzing) Petersen 17 75.56 3.02 1704.39 2.05
UNCORRECTED PROOF

of the first two axes were l1 = 0.3 and l2 = 0.17,

accounting for 20.1% of the variance in the species

data (Fig. 2). The gradient lengths of the first two DCA

axes of 2.7 and 1.8 standard deviation (S.D.) units

indicated that either linear or unimodal species

response models would be effective (ter Braak and

Prentice, 1988). As the gradient length of the first

DCA axis was above two S.D., we chose to use

techniques based on assumptions of unimodal species

distribution (ter Braak, 1995).

3.3. Data screening

Ten samples were identified as outliers, based on

PCA and DCA, and were removed from CCA analysis

and inference model development. Five of these

samples were from three sites, and were determined to

be outliers because of their high sample score means

on the PCA axes. All three sites had high conductivity

values (595–983 mS cm1

) and a high percentage of

sand, silt, and clay sediments. The nutrient values for

two of these sites were high (TP: 0.69 and

0.73 mg L1

; TN: 5.4 and 6.1 mg L1). Although it

was important to keep samples with high nutrient

concentrations in the dataset, the extreme values of

conductivity were considered to be a potential

problem for the development of nutrient inference

models. Eight samples from six sites were identified as

outliers because of their high sample score means on

the DCA axes (Fig. 2). Because all six sites had low

nutrient concentrations, and it was important that sites

in the calibration dataset have a wide nutrient gradient,

we decided to exclude only the five samples from the

two sites with the lowest pHs (6.5 and 6.7) in the

dataset. In addition, one site had a high % of sand

substrate. Despite the fact that both sites represented

the low end of the phosphorus gradient, we decided to

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Table 2 (Continued )

Taxon name Hill’s N2 TP (mg L1

) TN (mg L1)

Optima Tolerance Optima Tolerance

Nitzschia inconspicua Grunow 56 75.84 1.87 1830.04 1.60

Nitzschia capitellata Hustedt 20 76.30 1.89 2194.22 1.62

A. exiguum (Grunow) Czarnecki 17 76.53 2.18 1883.52 1.80

Bacillaria paradoxa Gmelin 18 77.08 2.30 2457.93 1.75

N. ruttnerii var. capitata Hustedt 23 78.85 1.89 1901.91 1.63

N. agrestis Hustedt 13 81.02 2.15 1866.16 1.69

Staurosira construens (Ehrenberg) Williams et Round 11 82.34 2.93 1586.97 2.66

Cyclotella pseudostelligera Hustedt 21 85.75 2.50 2077.26 1.76

Sellaphora seminulum (Grunow) Mann 55 88.40 2.11 1835.86 1.72

S. pupula (Ku ̈tzing) Meresckowsky 24 88.91 2.21 1941.51 1.63

G. kobayasii Kociolek and Kingston 39 90.67 2.58 1785.54 1.80

N. ingenua Hustedt 12 91.75 1.78 1773.62 1.30

N. recens Lange-Bertalot 21 101.10 2.28 1742.53 1.56

Aulacoseira granulata (Ehrenberg) Simonsen 10 124.04 3.13 2628.27 1.80

Cyclotella atomus Hustedt 11 155.60 2.18 2070.28 1.51

Luticola goeppertiana (Bleisch) Mann 13 163.02 2.15 2564.25 1.40

Only species with effective numbers of occurrence (Hill’s N2) > 10 are shown. Species are sorted by increasing TP optima.

Fig. 2. Detrended correspondence analysis (DCA) biplot showing

samples and passive environmental variables. Horizontal

axis = DCA axis 1; vertical axis = DCA axis 2. Ecoregions in which

sites are located are indicated by specific symbols. Circles: Northern

Piedmont; squares: Highlands; diamonds: Valley and Ridge. Abbreviations used for environmental variables can be found in Table 1.
UNCORRECTED PROOF

exclude them from further model development to

avoid the influence of a strong pH and alkalinity

gradient. Data analyses were performed using the

reduced dataset of 91 samples, that excluded the 10

outlier samples.

3.4. Relationships between diatom assemblages

and environmental variables

We used CA to determine the environmental

variables explaining most of the variation in diatom

species composition, and especially the importance of

nutrients as compared with other environmental

characteristics. In a CA including all 24 environmental

variables as passive variables and diatom data for 91

sites, 21.8% of the total variance of the diatom data

was explained by CA axis 1 (l1 = 0.27) and CA axis 2

(l2 = 0.19). The species-environment correlations

were high for both, axis 1 (r = 0.90) and axis 2

(r = 0.91), accounting for 35.1% of the variance in the

species–environment relationships. This indicates a

strong correlation between the 24 environmental

variables and the 131 species included in the complete

CA. Nevertheless, strong weighted correlations and

relatively high variance inflation factors (VIFs > 5)

indicated strong co-linearity among several environmental variables. Consequently, we omitted the

following 11 variables from further analysis: O-P,

NO3-N, NH3-N, average river width, reach length, %

urban land-use, % open canopy cover, % cobble and

gravel, AFDM, hardness, and conductivity. Because

TN and NO3-N, as well as TP and O-P, were highly

correlated, only one of the two variables, respectively,

could be selected. Because TP and TN are more

important with respect to management purposes, we

chose to eliminate the variables O-P and NO3-N. CCA

with forward selection that excluded these variables,

and included only 13 variables in total, identified 5

environmental variables that significantly (P < 0.05)

explained the variance in diatom species composition.

These were TP, TN, basin size, % forested land-use,

and % bedrock and boulder. The first two axes

explained 14.7% of the total variance in the diatom

data. Species–environment correlations of CCA axes

were high for axis 1 (r = 0.88) and for axis 2 (r = 0.85)

and accounted for 72.4% of the variance in the

species–environment relationships. The correlations

between TP and TN and the first ordination axis were

strong (r = 0.72 and 0.61). In comparison with the

CA that included all 23 variables, species–environment correlations on the first two axes were only

slightly lower, but explained most of the variance in

the species–environment relationship.

CCAs constrained to one variable at a time were

run on the remaining five variables to assess the

strength of their relationship with diatom species

composition, and to determine which variables had a

sufficiently strong relationship with diatom assemblages to justify development of inference models. Of

the five variables, TP had the highest l1/l2 ratio

(0.78), capturing 7.2% of the variation in the species

data. TN had a l1/l2 ratio of 0.54, capturing 5.6% of

the variation in species data. This confirmed that

both variables had significant influence on the diatom

assemblages and that development of inference models for both variables was justified.

3.5. Species WA-optima and tolerances

TP and TN WA species optima were calculated

using the reduced dataset (n = 91). The results are

presented for the species with effective numbers

of occurrences (Hill’s N2) > 10 (Table 2). Species

apparent WA TP optima ranged from 13 to 163

mg L1

. Weighted-average TN optima ranged from

362 to 4412 mg L1

. Taxa with high and low TP optima had values comparable to those calculated in

other studies conducted in eastern North America

(Winter and Duthie, 2000; Potapova et al., 2004).

3.6. Inference models

Models to infer TP and TN (log10 mg L1

) were

developed and tested using WA and WA-PLS. The

models were run on the n = 91 calibration set, which

included 131 taxa. In order to test if the inclusion of

multiple and repeat samples biased the model

performance, we developed additional TP and TN

WA-PLS models using the dataset with independent

samples only (n = 40). This dataset was obtained after

deleting 5 of the same outliers as in the models above,

reducing the set of 45 ‘independent’ samples to 40

samples, including only 128 taxa.

Table 3 shows performance measures for all

inference models. The best TP model was a twocomponent WA-PLS model using the n = 91 dataset,

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UNCORRECTED PROOF

showing the highest r2

boot (0.72) and the lowest

RMSEboot (0.23 log10 mg L1 TP). For TN, a threecomponent model using the n = 91 dataset showed best

performance, with an r2

boot of 0.58 and an RMSEboot of

0.23 log10 mg L1 TN. The TP gradient included in this

model was 6–458 mg L1 and the TN range was 170–

8547 mg L1

. When comparing the WA-PLS models

(n = 91), the difference between apparent and bootstrapped r

2 was less (0.87–0.72) for TP than for TN

(0.88–0.58) (Table 3 and Figs. 3A and 4A). The

residuals plot (Fig. 3B) shows that the TP model using

WA-PLS (n = 91) tends to underestimate TP values

above 100 mg L1

. A similar trend is shown in the TN

model using WA-PLS (n = 91), where the difference

between bootstrapped versus inferred values is higher

above 3000 mg L1 TN (Fig. 4A and B). Generally,

Figs. 3 and 4 show that the TN data have a less even

distribution along the nutrient gradient than the TP data.

The performance of the best TP and TN WA-PLS

models using the independent samples only (n = 40) are

slightly lower (TP model: r2

boot 1⁄4 0:69; RMSEboot = 0.28 log10 mg L1 TP; TN model: r2

boot 1⁄4 0:50;

RMSEboot = 0.27 log10 mg L1 TN) than compared

with the best models (n = 91) that include multiple and

replicate samples (Table 3). Nevertheless, because the

model performance does not significantly decrease

when excluding the multiple and replicate samples, we

believe that inclusion of these samples in the dataset

provides models that are more robust. It includes spatial

and temporal variability of the diatom species composition at these sites, which increases the reliability of

the inferred value, when the model is applied to similar

sites.

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Fig. 3. Plots of WA-PLS inference model showing (A) diatominferred log10 TP values vs. observed log10 TP values, and (B)

residuals vs. observed log10 TP values. The diagonal line is a 1:1

line. The solid line shows a LOESS scatter plot smoother (span =

0.45). The diatom TP index score is shown for comparison.

Table 3

Performance of weighted-averaging (WA) and weighted-averaging partial least squares (WA-PLS) regression and calibration models developed

for the n = 91 sample datasets including multiple and repeat samples and the n = 40 sample datasets including independent samples only

n Method Variable r

2 ðr2

bootÞ RMSE (RMSEPboot) Mean biasboot Max biasboot

91 WA-inv.desh. TP 0.70 (0.61) 0.22 (0.26) 0.009 0.57

91 WA-inv.desh. TN 0.63 (0.47) 0.19 (0.24) 0.010 0.66

91 WA-class.desh. TP 0.70 (0.61) 0.26 (0.28) 0.012 0.47

91 WA-class.desh. TN 0.63 (0.48) 0.24 (0.26) 0.014 0.64

91 WA-PLS-2 components TP 0.87 (0.72) 0.15 (0.23) 0.010 0.45

91 WA-PLS-3 components TN 0.88 (0.58) 0.11 (0.23) 0.015 0.66

40 WA-PLS-3 components TP 0.96 (0.69) 0.08 (0.28) 0.030 0.65

40 WA-PLS-1 component TN 0.71 (0.50) 0.19 (0.27) 0.015 0.81

All TP and TN units are in log10 mg L1

. WA-inv.desh.: WA-inverse deshrinking; WA-class.desh.: WA-classical deshrinking. The r

2 and RMSE

in parentheses were derived from bootstrapping; n = number of samples included in model. Models with best performance are highlighted in

bold.
UNCORRECTED PROOF

3.7. Diatom TP and TN indices

We evaluated how accurately the index values,

calculated using bootstrapped inferred nutrient concentrations, could be assigned to the nutrient

categories based on the TP and TN values measured

at each site (Fig. 5A and B). The Diatom TP Index

correctly assigned the majority (63%) of the samples.

The TP categories 1 and 3 had the highest proportion

of samples correctly assigned, whereas categories 2

and 4 had the highest proportion of samples assigned

to a neighboring category. The Diatom TN Index

correctly assigned the majority (68%) of the samples

to the correct categories, and placed 32% of the

samples into neighboring categories. The Diatom TN

index correctly assigned the highest proportion of

samples to TN category 1. The indices showed least

accuracy placing samples in the TP categories 2 and 4

and in the TN categories 2 and 3, which are the

categories with the highest number of samples.

4. Discussion

4.1. Environmental factors influencing diatom

assemblages

Multivariate analyses showed that nutrient concentrations (TP and TN) are important in explaining

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Fig. 5. Plots of inferred bootstrapped versus observed TP and TN

and the assignment of the two diatom index scores: (A) the diatom

TP index and (B) the diatom TN index to the corresponding nutrient

categories. Shaded rectangles correspond to the water chemistry

categories established for TP and TN. TP categories are: (1) <0.025

mg L1 TP, (2) 0.025–0.075 mg L1 TP, (3) 0.075–0.1 mg L1 TP,

(4) >0.1 mg L1 TP; TN categories are: (1) <0.7 mg L1 TN, (2)

0.7–1.5 mg L1 TN, (3) >1.5 mg L1 TN.

Fig. 4. Plots of WA-PLS inference model showing (A) diatominferred log10 TN values vs. observed log10 TN values, and (B)

residuals vs. observed log10 TN values. The diagonal line is a 1:1

line. The solid line shows a LOESS scatter plot smoother

(span = 0.45). The diatom TN index score is shown for comparison.
UNCORRECTED PROOF

variation in diatom assemblage composition, more so

than in other regional studies (Pan et al., 1996; Winter

and Duthie, 2000). Our analysis shows that TP and TN

were strongly correlated with the first CCA axis

(r = 0.72 and 0.61). The results of CCAs constrained to TN and TP, respectively, showed a strong

relationship between TN, TP, and diatom assemblages

with high l1/l2 ratios of 0.78 for TP and 0.54 for TN.

Other regional studies present lower l1/l2 ratios for

TP (0.21) and TN (0.32) (Winter and Duthie, 2000),

ratios comparable to the ones presented in this study

are recorded for TP (l1/l2 = 0.81) from Finland

(Soininen and Niemela ̈, 2002). This high ratio reflects

the relatively even distribution of nutrient concentrations measured in the streams in this study, and the

narrow range in other factors usually influencing

diatom distributions such as pH and conductivity.

4.2. Diatoms as nutrient indicators-inference

model performance

Compared to similar studies conducted using onetime chemistry measurements to develop benthic

diatom nutrient indicators for streams in eastern North

America (Pan et al., 1996; Winter and Duthie, 2000;

Potapova et al., 2004), the TP and TN WA-PLS models

presented here (n = 91) show high apparent and

bootstrapped r

2 and low apparent and bootstrapped

RMSE.

We took two steps to maximize the performance of

our models. First, site selection was restricted to three

ecoregions in New Jersey with a rather limited range in

water chemistry and geology. Sites were selected

based on knowledge of their chemistry to avoid wide

ranges of pH, alkalinity and conductivity that might

complicate development of nutrient inference models.

Second, we tried to capture a wide range of concentration along the nutrient gradients. This dataset

included a TP range of 6–458 mg L1 (n = 91), and the

range in TN was 170–8547 mg L1

. The nutrient

gradients included in all models were relatively large,

and the model error is low when compared to other

studies based on comparable or shorter nutrient

gradients (Pan et al., 1996; Potapova et al., 2004).

The dataset presented in Winter and Duthie (2000)

included a shorter TP gradient (5–215 mg L1

) and

a wide TN gradient, with a range of relatively

higher values (1000–16,290 mg L1

), but the models

show higher RMSEs when using one-time chemistry

measurements.

There are other factors that might affect the

performances of our models. These include the

influence of temporal variability in nutrient concentrations in streams (Cattaneo and Prairie, 1995; Pan

et al., 1996) and the indirect influence of nutrients on

diatom species composition through increasing competition with non-diatom species (e.g. Cladophora sp.)

(Winter and Duthie, 2000).

Despite the relatively wide ranges in TP and TN

captured in our models, model performance could

certainly be improved by adding samples to the

dataset. Fig. 4A and B shows that fewer numbers of

samples at the low and the high total phosphorus range

cause the TP model to be less reliable towards both

ends of the gradient. This ‘‘edge-effect’’ (Hall and

Smol, 1999) occurs with models developed on

samples with an uneven distribution along the TP

gradient (Reavie et al., 1995; Soininen and Niemela ̈,

2002), and causes the WA estimates of species optima

to be biased. Despite the rather wide range of TP

included in our dataset, the majority of samples fell

below concentrations of 100 mg L1 TP. This might

explain why the model is less reliable at concentrations of >100 mg L1

. Another explanation may be

that above higher concentrations of TP (e.g. concentrations of >100 mg L1

) smaller changes occur in

the diatom species composition and that therefore the

species response to higher TP concentrations is

weaker (Reavie et al., 1995; Pan et al., 1996).

Addition of more samples toward the ends of the

gradients might help improve the TP model performance, although significant improvement is more

likely at the lower nutrient range of the model; our

results show that diatom assemblages respond less

strongly to TP values over 100 mg L1

.

A similar trend is shown for the TN model, where

model performance decreases at the higher end of the

gradient (Fig. 4A and B). The average TN in the

Piedmont (2241 mg L1

) and the combined dataset

(1987 mg L1

) (Table 1) is low compared to the

average TN (4940 mg L1

) in the EPA Northern

Piedmont aggregate nutrient ecoregion (USEPA,

2000c), or compared to similar studies where nitrogen

inference models have been developed (Winter and

Duthie, 2000). The model therefore is built mainly

on sites with moderate nitrogen concentrations, and

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underrepresents sites with low and high nitrogen. This

explains the decrease of the TN model performance,

especially at the higher end of the gradient. To improve the TN model performance, it would be

important to add samples to both ends of the nitrogen

gradient, but especially to the higher end.

In general, our TP and TN inference models have

relatively high predictive power and showed reliable

results (e.g. inferred nutrient concentrations) when

tested using bootstrapping, and especially when

compared to existing nutrient models. When applying

the models to samples in any of the three ecoregions of

NJ, it is important to select sites with water chemistry

values that are within the ranges included in the model

calibration set, especially for pH, conductivity,

hardness and alkalinity (Table 1).

4.3. Use of diatom inference models and trophic

indices for management purposes

Our results show that diatom WA-PLS models can

reliably infer late summer nutrient concentrations in

NJ rivers. The good performance of the models

presented here confirms the conclusion of another

study conducted in the Northern Piedmont ecoregion,

that models based on WA-equations provide the best

predictions when used to model regional diatom–

nutrient responses (Potapova et al., 2004).

The diatom TP and TN indices, which are rescaled

inferred nutrient values, are meant to provide a

management tool for regulators. Because of their

presentation on a scale from 1 to 100, these indices

have the practical advantage of being more comparable than inferred nutrient values to indices commonly

used by the state, e.g. the Fish IBI (Karr, 1981), the

macroinvertebrate index and the habitat assessment

index (NJDEP, 1993a, 2000). They are similar to

commonly used diatom indices, such as trophic

diatom indices (TDIs), with the main difference that

the indices presented here are based on WA-PLS

inference models. When tested, the diatom TP and TN

indices assigned the samples to the correct water

chemistry category in the majority of cases. Because

the boundaries between the categories assigned are

based on correlations between nutrients and high algal

biomass (Dodds et al., 1998; Wetzel, 2001), the

categories not only reflect nutrient concentrations

but at the same time are estimates of impairment

associated with high biomass (i.e. rendering the waters

unsuitable for designated uses). The assignment of

samples to nutrient categories may therefore provide

NJDEP with a practical tool for regulatory assessments of eutrphication.

In addition to the categories, there are other useful

reference points for evaluating inferred nutrient

concentrations, which could serve as a basis for

alternative category boundaries. Several studies

conducted on large- and small-scale datasets on

streams in the US, including high proportions of

eastern US streams, confirm that levels of TP above

0.05 mg L1 and TN above 0.47 mg L1 lead to mean

benthic chl a values of >50 mg m2 (Dodds and

Welch, 2000; Dodds et al., 2002). We showed

previously that TP values above 0.05 mg L1 and

NO3-N values above 0.2 mg L1 can lead to nuisance

algal biomass (chl a > 100 mg m2

) (Ponader and

Charles, 2003). The establishment of category

boundaries could be refined through more extensive

exploration of relationships between nutrients and

biomass. Therefore, both the inference models and the

diatom TP and TN indices present reliable and useful

tools for monitoring nutrients and can be applied in a

regulatory context. Further improvement of the

inference models could be achieved by adding more

samples to the model calibration dataset, especially at

the both ends of the nutrient gradients.

Uncited references

USEPA (1998, 2000a).

Acknowledgements

This project was supported by the New Jersey

Department of Environmental Protection, Division of

Science, Research and Technology, Trenton, NJ. We

thank Diane Winter, Mike Hoffmann and Erin Hagan

for their work in the field and in the laboratory and

Kathleen Sprouffske for her help with data management. Marina Potapova and Eduardo Morales helped

with taxonomic identification. We also thank the

anonymous reviewers for their comments on the

manuscript.

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Works Association (AWWA), Water Environment Federation

(WEF), 1995. Standard Methods for the Examination of Water

and Wastewater, 19th ed. American Public Health Association,

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