ABSTRACT

MARTIN, CAITRIN ANNE. Delineating Sources and Estimating Cadmium Bioaccumulation and Susceptibility Differences among Aquatic Insects. (Under the direction of David B. Buchwalter). Human activities such as mining and smelting have altered the distribution of trace metals in

the environment, resulting in trace metal pollution of aquatic ecosystems worldwide. Trace

metal contamination has been shown to adversely affect benthic macroinvertebrate

community structure by influencing species diversity, density and abundance. Aquatic

macroinvertebrates play integral roles in stream ecosystem function and are widely used in

biomonitoring programs as indicators of stream integrity. Relative to marine and other lentic

invertebrates, lotic insects are grossly underrepresented in ecotoxicological studies. A major

challenge in ecotoxicology lies in generating data under experimental conditions that are

relevant to understanding contaminant effects in nature. Biodynamic modeling is one

laboratory-based approach for understanding trace metal bioaccumulation in a number of

aquatic invertebrates. This approach combines species-specific physiological traits associated

with metal bioaccumulation from both aqueous and dietary sources to make predictions of

metal- and species-specific bioaccumulation that compare well relative to field-based

measurements. This thesis represents the first application of biodynamic modeling to

stoneflies, a dominant insect group in North American streams. A second goal of these

comparative studies was to determine how much variability, or conversely, predictability

there is in some of the physiological processes related to bioaccumulation. Because insects as

a group are so species rich, a comprehensive testing regime is unfeasible. These comparative

studies were designed to determine if closely related taxa share physiological characteristics

such as cadmium uptake rate from solution, elimination rate, and subcellular

compartmentalization of accumulated metal. Biodynamic models were generated for seven

predatory stonefly (Plecoptera) species representing the families Perlidae (5) and Perlodidae

(2). Each taxon was exposed to cadmium independently via diet and via solution. Species

varied approximately 2.6 fold in predicted steady state Cd concentrations. It has become

increasingly evident that diet plays a major role in the accumulation of metals in aquatic

invertebrates and the importance of diet is consistent in stoneflies. Diet was the predominant

source of accumulated Cd in five of the seven species and averaged 53.2 ± 9.6 % and 90.2 ±

3.7% of the total Cd accumulated in perlids and perlodids, respectively. These results are

relevant to water quality criteria that are often based on toxicity tests involving aqueous

exposures only. Differences in Cd bioaccumulation between the two families were largely

driven by differences in dissolved uptake rates, which were considerably slower in perlodids

than in perlids. Cadmium uptake rates also varied within families, suggesting that family

level generalizations about metal accumulation may be erroneous. We further examined the

subcellular compartmentalization of cadmium accumulated from dissolved and dietary

exposures. Predicted steady state concentrations were modified to only consider Cd

accumulated in potentially susceptible subcellular compartments. These values ranged 5.3

fold. This variability is discussed within a phylogenetic context as well as its implications for

bioassessment.

Delineating Sources and Estimating Cadmium Bioaccumulation

and Susceptibility Differences Among Aquatic Insects

by

CAITRIN ANNE MARTIN

A thesis submitted to the Graduate Faculty of

North Carolina State University

in partial fulfillment of the requirements for the degree of

Master of Science

ENVIROMENTAL TOXICOLOGY

Raleigh, NC

2007

Approved by:

_____________________________

DAVID B. BUCHWALTER Chair of Advisory Committee

_____________________________ _____________________________

GERALD A. LEBLANC DAMIAN SHEA

ii

DEDICATION

I dedicate this work to my loving family, without whose constant support and guidance, I

would be lost. Mom, Dad, Aaron and Adrianne, life has offered me so many blessings and

our family is and always will be cherished above all.

“There is no need for temples, no need for complicated philosophy. Our own brain, our own

heart is our temple: the philosophy is kindness.”

-The Dalai Lama

iii

BIOGRAPHY

I, Caitrin Anne Martin was born in Rochester, New York to Marc Martin and Anna

Jurczynski-Martin. At an early age, my siblings and I were ingrained with a profound respect

for family and for the environment, as well as a strong work ethic fostered by our infamous

household chore charts. When I was ten years old, my family relocated to Wilmington, North

Carolina where I miraculously made it through grade school and high school without having

broken a single bone, and having caused only one relatively minor accident involving our

beloved old GMC minivan. In 2004, I graduated with a B.S. in Chemistry from the

University of North Carolina at Wilmington. One year later, after contemplating my purpose

in life, I applied and was accepted into the masters program in the Department of

Environmental and Molecular Toxicology here at North Carolina State University. Almost

two years later, I am again contemplating my life’s purpose. I began this program with only a

vague notion of the type of career I might be suited for, and although I am still uncertain of

the ideal career, my experiences in this program have directed me towards consideration of

the fields of biodiversity conservation, natural resources management/preservation, and

ecosystem restoration.

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ACKNOWLEDGMENTS

I would like to acknowledge Dr. Lingtian Xie for his willing and invaluable assistance in the

lab and on the tennis court.

I would like to give special thanks to Brian, who was so influential in my decision to

continue my education and for being my inspiration, my therapist, my stats consultant and

my best friend.

I would also like to acknowledge the members of my committee, Dr. Jerry LeBlanc and Dr.

Damian Shea for their input and encouragement on this project. Lastly, I would like to

express my gratitude to my advisor and committee chair, Dr. David Buchwalter who has

been a patient and dedicated teacher, an outstanding mentor and a constant source of

fortitude throughout many moments of self-doubt.

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TABLE OF CONTENTS

Page LIST OF FIGURES ........................................................................................................... vi LIST OF TABLES............................................................................................................ vii INTRODUCTION ...............................................................................................................1 1. Aquatic insects ...........................................................................................................1 1.1 Ecological importance of insects .............................................................................2 2. Regulatory context .....................................................................................................2 2.1 Limitations of geochemical analyses in assessing ecological condition .................4 3. The use of insects as indicators of stream condition..................................................6 3.1 Historic developments in the use of invertebrates as ecological indicators ............6 4. Current bioassessment approaches: Multimetric and Multivariate ..........................7 4.1. Bioassessment limitations ..................................................................................... 8 5. Single species toxicity bioassays ...............................................................................9 6. Trace metals in aquatic ecosystems ....................................................................... 11 6.1 Metal bioavailability in aquatic systems ............................................................. 11 6.2 Paradigms for modeling metal bioavailability, toxicity & bioaccumulation ....... 12 6.3 Species-specific differences in metal bioaccumulation/susceptibility ................. 15 6.4 Metal partitioning within the organism ................................................................16 6.5 Metallothioneins ................................................................................................. 17 6.6 Cadmium ............................................................................................................. 20 7. Research goals and hypothesis ............................................................................... 21 Literature Cited ......................................................................................................... 23 CHAPTER 1. CADMIUM ECOPHYSIOLOGY IN SEVEN STONEFLY (PLECOPTERA) SPECIES: DELINEATING SOURCES AND ESTIMATING SUSCEPTIBILITY .......30 Introduction ............................................................................................................... 32 Materials and Methods ............................................................................................... 35 Results ....................................................................................................................... 40 Discussion ..................................................................................................................44 Literature Cited ......................................................................................................... 49

CHAPTER 2. DIETARY CADMIUM DYNAMICS: INVESTIGATIONS OF EXPOSURE HISTORY AND INSECT DIGESTIVE PHYSIOLOGY ............................................... 63 1. The effect of metal exposure history on cadmium assimilation and elimination dynamics ........................................................................................ 63 2. Dual labeling technique to quantify gut retention time and Cadmium assimilation from diet ......................................................................... 66 3. Cadmium transport, absorption and elimination .................................................... 68 4. Calcium and cadmium interaction in insect haemolymph ..................................... 71 Literature Cited ......................................................................................................... 73 CONCLUSION ...........................................................................................................80

vi

LIST OF FIGURES Page INTRODUCTION

Figure 1. Domain nature of metallothionein. ....................................................................19 CHAPTER 1. Figure 1. Comparison of efflux rate constants (ke) derived from aqueous and dietary exposures. ......................................................................................................................... 58 Figure 2. Dissolved cadmium accumulation in selected taxa. ........................................ 59 Figure 3. The relative importance of dietary vs. aqueous exposure routes and total modeled cadmium accumulation. ..................................................................................... 60 Figure 4. Predicted ranges of cadmium steady state concentrations (µg g-1) wet weight in potentially susceptible subcellular compartments. ........................................................ 61 Figure 5. Relationship between total and physiologically available cadmium bioaccumulation. ................................................................................................62 CHAPTER 2. Figure 1. Cadmium elimination rates of pre-exposed larvae and naïve larvae ................ 75 Figure 2. Calcium content of Lumbriculus variegatus cultured in very hard, moderately hard and very soft ASTM reconstituted water ................................................................. 76 Figure 3. 109Cadmium activity (cpm) per mg (wet weight) in L. variegates incubated in very hard and very soft exposure media ......................................................................... 77 Figure 4. Image of a stonefly GI tract................................................................................78 Figure 5. Proportional Cd loss in Pteronarcys dorsata larvae injected with verapamil and Cd, Cd only, or Cd and Ca treatments ..................................................................... 79

vii

LIST OF TABLES

Page

Table 1. Taxa, source location, mass and numbers of individuals used in bioaccumulation experiments ............................................................................................55 Table 2. Biokinetic model parameters for seven stonefly species. ................................. 56 Table 3. Subcellular cadmium compartmentalization in seven plecopteran species ....... 57

1

INTRODUCTION 1. Aquatic insects

Insects are ubiquitous in aquatic habitats. Well over 6,500 species of aquatic insects

have been described in North America alone (1), with new species being described every

year. They can be found in habitats ranging from thermal springs to sub-arctic aquatic

environments, and from the splash zones of rocky intertidal areas to mountain springs.

However, it is in freshwater lotic systems in particular where insects thrive most. Aquatic

insects can typically comprise 70-90% of the invertebrate species pool in temperate streams

(2), and in terms of both diversity and biomass, are predominant members of stream

ecosystems. Not surprisingly, insects perform critical roles in stream ecosystem function.

They are key players both in terms of processing organic materials and serving as a conduit

of energy from primary producers to higher trophic levels as key food resources to many fish

species and birds (3). Because insects are so dominant in lotic environments, they are the

most widely used group of organisms in assessing the ecological condition of streams

worldwide (4-7).

Despite the ecological importance of insects, and their widespread use to evaluate the

ecological conditions of freshwater environments, the physiology of this important group has

received surprisingly little attention (8,9). This thesis focuses on the physiology of the trace

metal cadmium (Cd), a common pollutant in aquatic systems throughout the world. This

introduction is intended to provide background information on aquatic insects and to briefly

describe their use as ecological indicators. Furthermore, I summarize historic developments

in the fields of trace metal bioavailability and bioaccumulation, and provide material relevant

to the experimental work described in the following chapters.

2

1.1 Ecological importance of insects

Macroinvertebrates are of inestimable importance in maintaining stream ecosystem

function. Being at the intermediate trophic level of lotic food webs, macroinvertebrates

facilitate energy flow and stream productivity from the top-down and from the bottom-up.

Macroinvertebrate functional groups perform a variety of roles in critical stream ecosystem

processes such as nutrient cycling, decomposition and translocation of materials (3). Major

functional feeding groups include grazers, shredders, gatherers, scrapers, filterers and

predators (10). These functional groups have evolved through the development of

morphological and behavioral mechanisms of food acquisition. In addition to performing

integral stream ecosystem processes, insects constitute an important part of the diet of many

bird species as well as many game and commercial fish species such as salmonids (1). (Many

fly fisherman have a particular fondness for these insects in the interest of imitating fish

foods.) So, aquatic insects are also important organisms, culturally and economically, and

their numerous functions emphasize the importance of their conservation.

2. Regulatory context

Since the industrial revolution, it has been evident that human activities have

adversely effected aquatic environments. In 1972, The Federal Water Pollution Control Act

was passed as a consequence of increased public awareness and concern for controlling

pollution. The objective of the Act was to “restore and maintain the chemical, physical, and

biological integrity of the Nation’s waters.” In 1977, the Act was amended and renamed the

Clean Water Act (CWA) which established the basic regulations for discharging pollutants

into U.S. waters in the interest of surface water quality protection. The CWA gives states the

responsibility to set and enforce water quality standards in order to protect water quality,

3

promote public health and protect the chemical, physical and biological integrity of aquatic

ecosystems.

CWA section 303(c), the water quality standards (WQS) program consists of three

major components; 1.) designated uses, 2.) water quality criteria and 3.) antidegradation

policy. The designated uses (DUs) of a waterbody are uses that are determined should be

attained in the waterbody. Typically, DUs include swimming and other recreation;

commercial, subsistence and recreational fishing; supporting communities of aquatic life and

supplying water for drinking, irrigation, landscaping and industrial purposes. The second

component, water quality criteria (WQC), (based on toxicity testing, see below) are limits on

chemical concentrations that may be present in a waterbody, or descriptions of the conditions

in a waterbody necessary to support the DUs. These descriptions can be expressed as

concentrations of pollutants, temperature, pH, turbidity units, toxicity units, or other

quantitative measures. The third component of state WQS, antidegradation policies establish

a set of rules for addressing proposed activities that could impair the condition of high

quality waters.

If a waterbody does not meet established water quality standards, strategies for

meeting the standards are developed. One such strategy involves the development of a total

maximum daily load (TMDL) program. A TMDL is a sum of the allowable loads of a single

pollutant from all contributing point and nonpoint sources. The calculated load must allow

the waterbody to meet water quality standards, and ensure that the waterbody attains the

State’s designated uses. Under section 303(d); states, territories, and authorized tribes are

required to develop lists of impaired waters. Specifically, reporting requirements entail that

by April 1 of all even numbered years, “states must submit to the EPA a list of impaired and

4

threatened waters still requiring total maximum daily loads (TMDLs); identification of the

impairing pollutant(s); and priority ranking of these waters, including waters targeted for

TMDL development within the next two years (11).” Streams can be listed on the 303(d) list

of impaired waters for a number of physical and chemical impairments. It was not until 1991

that failure to meet biological criteria (see biocriteria below) based on impaired biological

community structure could result in a waterbody being listed on the 303(d) list.

2.1. Limitations of geochemical analyses in the evaluation of ecological status

Historically, monitoring activities associated with the CWA focused on determining

whether contaminants in a given waterbody, exceeded chemical water quality criteria.

However, it has been established that the ecological condition of freshwater systems is not

adequately explained by water/sediment chemistry measurements alone. Physico-chemical

measurements may fall short in assessing stream condition for reasons such as: a.) water

samples may be collected at times when contaminants are not present, for instance if a pulse

of contaminant moves through a stream. b.) a contaminant may not be recognized as an

analyte of interest or detection limits are inadequate to detect a particular analyte (e.g. dioxin)

c.) few frameworks exist for evaluating the toxicity of mixtures (12). d.) analysis of

environmental media alone presents limitations with respect to predicting the bioavailability

of contaminants and finally, e.) water chemistry analyses can conclude that a stream is in

good condition while the physical habitat of the stream could be unsuitable to support biota.

Because biological communities integrate the cumulative impact of environmental

stressors, the direct measurement of resident biota gives information that toxicity tests and

rapidly changing water chemistry measurements fail to provide. With this realization,

resource managers began utilizing the resident biota to determine if water bodies meet

5

designated aquatic life uses (ALUs), a use designation in which the waterbody provides

suitable habitat for survival and reproduction of native fish, shellfish, and other benthic

invertebrates (13). Biological assessments (bioassessments) evaluate the biological condition

of an aquatic ecosystem using biological surveys (biosurveys) and other direct measurements

of aquatic life including fish, invertebrates (primarily insects) and periphyton. These

assessments involve relatively inexpensive, though labor intensive sampling, processing, and

analyzing a representative portion of the resident aquatic community. Bioassessment data

have been instrumental in setting ecological integrity goals for use in water quality

management.

Chemical water quality criteria can be unsuccessful in “restoring and maintaining the

ecological integrity of a waterbody” for reasons described above. The concept of regulating

water quality through the use of biological critera (biocriteria) has been in existence for over

a century, but has been slow to gain support in North America. Biological criteria

(biocriteria) are regulatory-based biological measurements used to describe the qualities that

must be present in an ecosystem to support a given biological condition (14). These criteria

are either numeric or narrative descriptions derived from integrated bioassessment measures

(indices) founded on the community composition, diversity and functional organization of

unimpaired or minimally impaired reference aquatic communities. Bioassessment data are

compared with the biocriteria established for a given waterbody to determine if the

ecosystem is supporting established ALUs. Biocriteria fulfill a critical function lacking in

traditional chemical water quality criteria and can be incorporated into water quality

standards for achieving ecological integrity goals, and in identifying and reporting

biologically impaired streams under section 303(d).

6

3. The use of insects as indicators of stream condition

Aquatic macroinvertebrates are the most frequently used organisms in the routine

monitoring of biological communities (15-17). Benthic insect assemblages are diverse and

ubiquitous, offering a wide range of responses to contaminants and other stressors.

Furthermore, macroinvertebrates are suitable indicators of local conditions; their relatively

sedentary nature allows site-specific analysis of stressor effects. Many benthic invertebrates

have life cycles of one year or more, facilitating longer term monitoring. Moreover, sampling

and sorting is relatively simple and inexpensive, having minimal effect on resident biota.

Additionally, most aquatic insects are easily identified to family level, however, of continued

debate is the appropriate level of taxonomic resolution in biomonitoring (18,19).

3.1. Historic developments in the use of invertebrates as ecological indicators

Two of the inaugural programs utilizing aquatic macroinvertebrates as indicators of

stream condition include the Saprobian system and the Hilsenhoff Biotic Index. The

Saprobian system originated in 1908 and represents the first use of a biotic index. The system

is based on the concept that certain species and communities of organisms behave as

“biological indicators of organic pollution” (20). Streams are given scores based on

saprobity, or the degree of pollution (by sewage) in a stream which results in decreased

dissolved oxygen and subsequent impact on resident biota. The Hilsenhoff Biotic Index

(HBI) was developed in 1979 for use in water quality assessment and monitoring programs

(21). Hilsenhoff developed a numeric index (0-10) to classify the degree of organic pollution

using aquatic macroinvertebrates, which effectively serve as indicators of environmental

perturbations. The HBI is succeeded by the Index of Biological Integrity (IBI) which

7

integrates multiple metrics (at least 7) to determine the biological condition of a test site (22).

Multimetric approaches are discussed in greater detail below.

4. Current bioassessment approaches: Multimetric and Multivariate

Currently, the most widely used approach for water quality assessements involves

multimetric rapid assessment techniques such as the USEPA’s Rapid Bioassessment

Protocols (RBP) which use several indices to assess biotic condition (17). Metrics, which are

influenced by human impact, are measures that represent aspects of the structure and/or

function of a biological assemblage. Benthic metrics include taxa richness,

tolerance/intolerance measures, feeding measures or trophic dynamics, and habitat measures

(17). Multimetric indices are combined metrics intended to represent a range of assemblage

responses to anthropogenic impact to determine threshold levels of impairment (16).

Although these methods are relatively inexpensive and technically straightforward, they fail

to provide mechanistic explanations for stream impairment.

Alternatively, large datasets can be managed with multivariate or predictive modeling

approaches based on statistical techniques using comparisons between patterns expected at a

reference site compared to the patterns observed at a test site (16). The River Invertebrate

Prediction and Classification System (RIVPACS) multivariate predictive model is used to

compare the biota at test sites with the biological conditions at reference sites (23). The

degree of impairment is measured by the ratio of the number of observed taxa to the number

of expected taxa (O/E ratio). A highly impacted ecosystem would have an observed-to-

expected ratio of much less than 1.0. The utility in multivariate approaches lies in their

ability to provide a numeric score based on the probability of sampling a particular species

from a given site. Unfortunately, although multivariate approaches possess potential

8

diagnosic capabilities, they are not currently used to determine direct cause of stream

impairment.

4.1. Bioassessment limitations

Current and historic bioassessment approaches rely on comparisons between the

resident benthic communities in sites of interest relative to local or regional reference

(control) ecosystems. This comparative method is known as the reference condition approach

(RCA). The regional reference condition describes the condition that is “representative of a

group of minimally disturbed sites organized by physical, chemical and biological

characteristics (16).” Alarmingly, true reference sites may not exist for some regions because

anthropogenic activities have adversely impacted many lotic systems (24) to such an extent

that assessors are required to decide on an acceptable level of disturbance to represent

reference conditions.

Another limitation in the use of aquatic insects in bioassessment is the relatively poor

understanding of stressor specific and taxon specific responses to common environmental

stressors. A common practice in the use of insects in bioassessment involves assigning a

single tolerance value to a given taxon, the implicit assumption being that the particular

taxon is equally susceptible to all environmental stressors. This assumption is seldom tested,

and is likely inappropriate. Furthermore, because insects are so species rich, the current use

of these organisms in bioassessment often involves coarse generalizations about species

sensitivity at broad taxonomic levels. These generalizations are largely based on anecdotal

evidence and descriptive observations, and are likely erroneous. A common bioassessment

endpoint for instance, is the percentage of insects belonging to three major orders:

Ephemeroptera, Plecoptera, Trichoptera (EPT), representing approximately 2,465 species in

9

North America alone (1). EPT taxa account for greater than 50% of all North American

aquatic insect species. These orders are commonly thought to be the most sensitive taxa

inhabiting lotic ecosystems (17), although families, genera and species within these insect

orders can exhibit significantly different susceptibilities to a given stressor. Although these

measures provide adequate cursory evaluations of stream integrity, a greater understanding of

how and why closely related species are differentially susceptible to contaminants will likely

lead to more tenable use of these organisms in bioassessment.

5. Single species toxicity bioassays

Simple and inexpensive toxicity tests are often used to estimate the hazard associated

with a particular environmental contaminant or stressor (25). These bioassays determine

standardized endpoints such as the median lethal concentration (LC50) and the no observed

effect concentration (NOEC) and are the primary means of estimating acute toxicity of

contaminants to aquatic organisms (26). Toxicity tests can provide valuable information on

the direct effects of contaminants on survival, growth, reproduction, behavior, physiology,

etc. However, these tests generally involve short term exposures to high concentrations via

the aqueous route only. These exposure regimes lack environmental realism, as organisms in

nature are often exposed to low levels of contaminants over longer durations. Furthermore, it

is likely that neglecting a major exposure route could result in underestimating the

organism’s response to a particular contaminant.

As previously mentioned, toxicity testing is used in establishing chemical water

quality criteria. The assumption in this approach is that if the most “sensitive” species are

selected for these bioassays, the resulting water quality criteria will be sufficiently protective

for all other species. Species sensitivity distributions (SSDs) should ideally correspond to the

10

communities that water quality criteria are meant to protect (27,28). However, benthic

macroinvertebrates are grossly underrepresented in toxicity databases, relative to their

abundance in lotic stream communities (29).

Possibly the largest challenge in the field of ecotoxicology is in extrapolating

laboratory-derived data to field situations (30). Because laboratory-based mechanistic

investigations are carried out under standardized conditions, it is questionable whether such

tests accurately predict biological effects in nature (31). An issue of continued debate is

whether single species toxicity tests can adequately predict responses at higher levels of

biological organization (32). The single species approach neglects species interactions and

fails to predict changes as a result of contaminant exposure in predation, community

function, ecosystem energy flow, nutrient cycling and other indirect effects (30). These

limitations demonstrate the importance of going beyond laboratory bioassays and considering

population- and community-level ecological effects (33). It is necessary to validate

laboratory-based predictions in natural systems via field experiments and observational

approaches (34) which are often large scaled and costly. However, field studies, although

environmentally relevant, may be unsuccessful in establishing direct causal relationships

between stressor and response, in light of the complex and interacting factors that are

inherent to natural systems (32). Consequently, alternative approaches are needed to better

understand biological responses to contaminant exposure and to resolve discrepancies

between field data and toxicity testing (35).

Perhaps the best example of the discrepancy between field and laboratory data is

illustrated by toxicity testing of trace metals in insects. Laboratory toxicity bioassays with

metals in insects have reported LC50 values that are orders of magnitude higher than

11

concentrations found in nature (27), suggesting that insects are not acutely sensitive to trace

metals relative to other faunal groups. Cadmium LC50 values for several aquatic insect

species reported in EPA’s updated ambient water quality criteria (36) are also several orders

of magnitude higher than concentrations that are known to affect insect communities in

nature (37,38). These data suggest that insects are likely more susceptible to chronic metal

exposures, making acute toxicity tests relatively uninformative for predicting insect

responses to metals in nature.

6. Trace metals in aquatic ecosystems

Metals, both essential and nonessential, naturally occur in the environment therefore

natural processes as well as anthropogenic activities are known to alter the distribution of

metals in environmental media (39). Trace metals are mobilized in the environment by

human activities such as mining, smelting and through the combustion of fossil fuels. Trace

metal contamination from current and historic mining operations is recognized as one of the

most significant environmental problems in freshwater systems in the western United States

(25). The following sections provide information on the interactions of trace metals with

aquatic biota.

6.1. Metal bioavailability in aquatic systems

Metal speciation in the external environment is key to determining trace metal

mobility, bioavailability and toxicity. Metal ions form stable complexes with a number of

organic and inorganic ligands and can also complex with dissolved organic matter present in

natural waters (40). The speciation and adsorption of cadmium onto sediments and other

organic matter is highly pH-dependent, increasing as conditions become more alkaline.

When conditions are more acidic (pH below 6-7), cadmium is desorbed from these materials.

12

The uptake of cadmium in various organisms has been shown to increase with increasing pH

of the exposure media at pH of 5.5 - 9.0 and decrease when pH is greater than 9.0 (41).

Cadmium bioavailability also decreases with increasing ionic strength. Furthermore, with

every 10ºC increase in temperature, the rate of biological processes typically double,

suggesting that temperature influences metal influx and efflux rates (42). Several studies

have shown that Cd is taken up in some organisms in a temperature dependent manner.

Moreover, interactions between and among metals can have a great influence on metal

availability to aquatic organisms (see below). Interactions include: 1.) increased uptake, 2.)

competitive inhibition of uptake and 3.) competitive displacement on metal-binding proteins

(43).

6.2. Major paradigms for modeling metal bioavailability, toxicity and bioaccumulation

Early attempts to describe the biological effects of metals on aquatic biota yielded

some inconsistencies. One could see adverse biological effects associated with certain total

metal concentrations, but in other places with similar or even higher concentrations, such

adverse effects were not apparent. This evidence has established that the total aqueous

concentration of metal is a poor predictor of bioavailability (44). Because natural waters vary

in pH, conductivity, and dissolved organic material, the bioavailability and consequently the

toxicity of trace metals also varies between sites (45) depending on the ambient water

chemistry. The free ion activity model (FIAM) is based on the influences of water chemistry

on metal speciation and assumes that the rate of metal uptake from solution is determined by

the availability of free metal ion. The model emphasizes “the universal importance of free

metal ion activities in determining the uptake, nutrition and toxicity of cationic trace metals

13

(46).” In other words, the activity of the free metal ion (Mz+) in solution correlates well with

observed biological effects of metals.

The Biotic Ligand Model (BLM) is an extension of the FIAM that was developed to

explain how metal binding at the gill was related to acute toxicity. Because dissolved metal

ions are highly hydrophilic and do not readily pass through lipid bilayers, transport across

biological membranes involves active transport across specialized regions of the body

(typically on the gill surface), known as biotic ligands (9). Biotic ligands are defined as

physiological sites of metal action and typically exist as ion channels used for the

sequestration of essential metals, however, nonessential metals are also transported via these

same ion channels. Ag and Cu are transported through sodium ion channels, while Cd and Zn

are transported through Ca ion channels and as a result, most aquatic organisms are incapable

of regulating the accumulation of nonessential metals such as cadmium. In addition to

complexation, the BLM incorporates competition of free metal ion with environmental

ligands for binding at these biotic ligands, as well as concentration of the free metal ion in

solution (47). The model assumes that toxicity results from the surface binding of metal to

biotic ligands on the gill surface, and toxicity (e.g. LC50) correlates well with laboratory-

derived gill binding constants. The BLM has been successfully used to predict acute metal

toxicity in rainbow trout, as such it is widely used in establishing site-specific water quality

criteria (48). However it is not as successful in predicting the biological effects of chronic

exposures.

Several computer based equilibrium models have been developed to predict metal

speciation/complexation. MINTEQ is an equilibrium speciation model that can be used to

calculate the equilibrium composition of dilute aqueous solutions in the laboratory or in

14

natural aquatic systems. The model is useful for calculating the ratio of dissolved species,

adsorbed species, and multiple solid phases under a variety of conditions. The windermere

humic aqueous model (WHAM) was developed to explain complexation between cationic

metals and dissolved organic material (DOM) (47). The model considers specific and

nonspecific ion binding by humic substances (49), making it ideal for describing the

solubility of metals given particular water chemistry conditions, and DOM content.

A third approach to understanding trace metal bioaccumulation is through

biodynamic modeling. This approach integrates laboratory derived rate constants of

accumulation and elimination from both dissolved and dietary sources as well as the species-

specific characteristics that contribute to trace metal bioaccumulation. The dynamic multi-

pathway bioaccumulation model (DYM-BAM) has been successfully used to predict trace

metal bioaccumulation in nature (50,51). This approach was developed in marine

invertebrates, though biodynamic models have been generated for a number of freshwater

invertebrates. The model assumes that toxicity occurs when influx rates exceed loss rates and

detoxification. However, the model fails to establish a direct link between bioaccumulation

and toxicity (52). Biodynamic modeling can be combined with studies of the organism’s

subcellular compartmentalization of accumulated metal to better predict the

toxicity/susceptibility of an organism to accumulated metal (see below).

The benefit of biodynamic modeling lies partly in its consideration of both aqueous

and dietary exposure routes. Historically, metal in the soluble phase has been considered the

dominant route of trace metal exposure to aquatic organisms. However, higher metal

concentrations in food items relative to water intuitively suggest that diet is potentially an

important source of metal uptake. Experimental separation of exposure pathways has allowed

15

researchers to investigate the relative importance of metal uptake from food and it is

becoming increasingly evident that diet contributes significantly to trace metal

bioaccumulation in aquatic organisms (53-57).

Metal availability from food is influenced by a number of factors. Metal uptake from

food is dependent on feeding/ingestion rate (IR) and assimilation efficiency (AE) or the

proportion of metal in the ingested food that is absorbed from the gut. Moreover, metal

bioavailability varies according to prey type. For instance, the copepod, P. coronatus was

determined to accumulate 550% more Cd from algae than from diatoms. Furthermore, metals

in synthetic foods have proven to be more available than metals in natural food items, so the

use of natural prey labeled in vivo may give more realistic estimates of metal bioavailability

in bioaccumulation studies.

6.3. Species-specific differences in metal bioaccumulation/susceptibility

Field-based observations indicate that aquatic organisms collected from the same

habitat can have widely different metal burdens (58). These differences in trace metal

bioaccumulation among taxa are largely driven by species-specific physiological processes

such as metal uptake rate from solution (45). Buchwalter et al. found considerable variation

in dissolved uptake rates of Cd and Zn in several larvae of the EPT insect orders (9).

Furthermore, Croteau et al found notable differences in cadmium uptake rates among

Chaoborus species coexisting in nature (59). Moreover, Luoma and Rainbow concluded that

bioaccumulation varies greatly among species due in part to inter- and intra- specific

differences in elimination rates (51). Other field-based observations indicate that some

organisms are extirpated from metal contaminated systems while others persist. These

16

differential susceptibilities to trace metals could be explained by species-specific

physiological traits such as detoxification capacity.

6.4. Metal partitioning within the organism

Bioaccumulation alone does not predict sensitivity to trace metals because species

cope with accumulated metal differently. Therefore, to better evaluate the potential effects of

metal exposures in nature, it is not only important to account for the contributions of both

dietary and dissolved exposures to accumulated metal body burdens, but also account for the

subsequent subcellular disposition of metals accumulated via each exposure route. The

subcellular partitioning of metal within an organism can greatly influence the potential

effects of accumulated metal (45). Aquatic organisms are capable of coping with excess trace

metals in a number of ways, by increasing elimination, or by sequestering excess metal by

storage in detoxified forms such as granules or metallothioneins and related metal binding

proteins (60,61). Rainbow et al. observed that barnacles can accumulate up to 5% of their

body weight as zinc without detriment, due to their ability to sequester excess Zn in

pyrophosphate granules (62). Sequestration via metallothionein and related metal-binding

proteins is a well-studied detoxification strategy employed by aquatic organisms (60,63,64).

Cadmium preferentially binds to sulfur bearing functional groups of metallothionein. The

extent to which accumulated metal is associated with these types of proteins can be a useful

predictive tool in determining potential harm to the organism (65). Metallothioneins are

discussed in greater detail in the following section. Characterizing an organism’s capacity to

detoxify accumulated metal provides information that can be used to make estimates about

species specific susceptibility to trace metals.

17

6.5. Metallothioneins

Introduction

Metallothioneins (MTs) are a superfamily of low molecular weight (<7000 Da)

intracellular metal-binding proteins which, in many species, play a critical role in the

detoxification of nonessential metals such as Cd2+and Hg2+, and in the regulation of

intracellular concentrations of essential metals such as Zn2+and Cu2+. In 1957, Kagi and

Vallee first purified and characterized MT as a cadmium-binding protein in equine kidney

(66).

Classification and Structure

Metallothioneins are evolutionarily conserved, as they possess a high cysteine

content and lack aromatic amino acids. However, few invertebrate MTs have been

characterized, and these can exhibit wide variation in non-cysteine amino acid residues (60).

Initially, MTs were classified according to their structural characteristics. Class I MTs consist

of polypeptides with highly conserved cysteine residue sequences and closely resemble the

equine renal MT. Mammalian MTs consist of 61 to 68 amino acids residues and the sequence

is highly conserved with respect to the position of the cysteine residues (e.g. cys-x-cys, cys-

x-y-cys, and cys-cys sequences, where x and y are non-cysteine, non-aromatic amino acids).

Class II MTs have less conserved cysteine residues and are distantly related to mammalian

MTs. Class III MTs are defined as atypical and consist of enzymatically synthesized peptides

such as phytochelatins and cadystins (67). This former classification scheme has been

replaced by a more complex system to include the increasing number of identified isoforms.

Spectroscopic analysis of MT indicates tetrahedral–thiolate clusters in which divalent

metals are bound by sulfur atoms. However, apo-MT (metal-free thionein) exists as a

18

randomly coiled polypeptide and the structure of the holoprotein is determined by

coordinated metal ions (68). Mx+-MT consists of two subunits as identified by Winge and

coworkers using two-dimensional NMR. The two cluster structure of MT has been more

recently confirmed by Stout et al. using X-ray crystallography. Figure 1 depicts the more

stable α-domain, also known as the C-terminal, which binds four divalent metals while the β-

domain, also known as the N-terminal, is capable of binding three divalent metals. The

stoichiometry of MT dictates that up to seven divalent metals or twelve monovalent metals

can form thiolate complexes with MT. MT lacks cellular specificity in binding metal ions;

however it has the greatest propensity for incorporating group Ib and IIb transition metals.

The binding affinity of different metals for MT varies accordingly: Zn2+ < Cd2+ < Cu2+ < Ag+

= Hg2+ = Bi3+, therefore Zn can be readily displaced by other metal ions.

19

Figure 1. Domain nature of metallothionein. Four and three atoms of cadmium (Cd) are

coordinated in α- and β-domains of metallothionein, respectively. Annu. Rev.

Pharmacol. Toxicol. 1999. 39:267–94

Function

Researchers have proposed that metallothioneins possess multiple physiological

functions including protection against metal toxicity, zinc homeostasis and defense against

reactive oxygen species (ROS) (67). Yet, recent work with transgenic and knockout mice

has questioned the validity of the later two functions, while confirming that MT does indeed

play an integral role in protection against cadmium toxicity. For instance, Klaassen and

coworkers observed that in the absence of MT-1 and MT-2, incidence of CdCl2-induced

lethality and hepatotoxicity increased, whereas MT overexpression proved to be protective

(69). Using the MT -/- mouse model, Cd exposure produced considerably variable toxicity,

20

leading researchers to believe that there are factors other than MT determining Cd toxicity.

Despite the important physiological functions of MT, it should be noted that MTs, due to

their redox properties, are not necessarily a sink for metals entering cells. On the contrary,

metal exchange reactions as well as nucleophilic reactions carried out by its numerous

sulfhydryl groups make MT a rather reactive protein. Consequently, there is much work to be

done in order to further elucidate the physiological functions of metallothionein.

Degradation

The half-life of M-MT is dependent on the binding affinity of thionein for different

metal ions. For instance, upon oxidation, Cu-MT forms insoluble polymers which are

biologically unavailable and are eventually eliminated via biliary secretion. In contrast,

thionein has lower affinity for Zn, making it more easily released from the protein and

rendering the ion available for cellular processes (70). Furthermore, the rate of degradation

may be influenced by differences in metal distribution between MT isoforms. It has been

determined that MT degradation can occur in lysosomal and non-lysosomal (cytosolic)

compartments.

6.6. Cadmium

Cadmium is a nonessential metal that exists mainly as cadmium sulfide (CdS) and is a

common impurity in zinc ores, thus it is often introduced into aquatic environments as a

byproduct of zinc mining (39). Most often used in nickel-cadmium batteries, Cd is also used

in coatings, plastics, platings and pigments (cadmium selenide (CdSe) is commonly known

as cadmium red). The most common oxidation state is Cd2+ however very rarely, monovalent

cadmium can be found. Among the known cadmium isotopes, 109Cd, a gamma emitting

21

radioisotope is observed to have the longest half-life (t1/2 = 462.6 days) making it a suitable

radiotracer in bioaccumulation studies.

EPA has named cadmium a priority toxic pollutant because it has been shown to act

as a reproductive and developmental toxicant and as a carcinogen (26). Cadmium has also

been shown to elicit toxicity in aquatic organisms. The freshwater criterion maximum

concentration (CMC) is defined as “the highest concentration of a pollutant in freshwater, to

which an aquatic community can be exposed briefly (acute limit) without resulting in an

unacceptable effect. EPA has estimated the CMC for cadmium as 4.3 µg L-1. The freshwater

criterion continuous concentration (CCC) is defined as “the highest concentration of a

pollutant in freshwater, to which an aquatic community can be exposed indefinitely (chronic

limit) without resulting in an unacceptable effect.” EPA has estimated the CCC for cadmium

as 2.2 µg L-1 (71).

7. Research goals and hypothesis

Observational field-based studies have indicated that trace metal contamination in

lotic systems is often associated with impaired insect community structure (72). However,

observational studies such as these cannot directly establish causal relationships between

contaminant exposure and benthic community structure (6). Therefore, laboratory studies are

crucial to better establish this linkage (73). Unfortunately, trace metal toxicity testing with

this important group of organisms has not successfully established a linkage between metal

contamination and ecological effects. This is because metal concentrations required to elicit

responses in the lab are considerably higher than those thought to impair stream communities

in nature (35). The distribution and abundance of insects within a metal contaminated stream

has been linked to metal bioaccumulation and susceptibility differences among taxa. These

22

differences appear to be mediated by the species-specific physiological processes related to

metal bioaccumulation and detoxification.

A major objective of this research was to determine if species-specific physiological

processes such as rate of metal uptake from solution, rate of metal elimination and metal

detoxification capacity can be used in concert to estimate susceptibility differences among

closely related stonefly species. This combination of insect physiological traits is a new

approach that has potential as a tool for determining species- and stressor-specific tolerance

values for potential use in bioassessment. A secondary objective of this research was to

determine if dietary Cd exposure is an important route of uptake in predatory stoneflies.

These findings have implications for the establishment of water quality criteria which rely on

toxicity bioassays incorporating the aqueous exposure route only.

23

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48. Niyogi, S.; Wood, C. M. Biotic Ligand Model, a Flexible Tool for Developing Site-Specific Water Qualtiy Guidelines for Metals. Environ. Sci. Technol. 2004, 38 (23), 6177-6192.

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59. Croteau, M.-N.; Hare, L.; Tessier, A. Differences in Cd Accumulation among Species of the Lake-dwelling Biomonitor Chaoborus. Can. J. Fish. Aquat. Sci. 2001, 58 1737-1746.

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30

CHAPTER 2

Cadmium Ecophysiology in Seven Stonefly (Plecoptera) Species:

Delineating Sources and Estimating Susceptibility

Caitrin A. Martin1, Samuel N. Luoma2, Daniel J. Cain2 and David B. Buchwalter1†

1. Department of Environmental and Molecular Toxicology, North Carolina State University,

Raleigh, NC 27606.

2. U.S. Geological Survey, 345 Menlo Park, MS465, California 94025

† to whom correspondences should be addressed:

Department of Environmental and Molecular Toxicology, North Carolina State University,

Campus Box 7633, Raleigh, NC 27695

E-mail: david_buchwalter@ncsu.edu

Phone: 919-513-1129

Fax: 919-515-7169

*This manuscript was submitted to Environmental Science and Technology on 22 May 2007,

was accepted on 20 July 2007 and is formatted according to the journal’s requirements.

31

Abstract

A major challenge in ecotoxicology lies in generating data under experimental conditions

that are relevant to understanding contaminant effects in nature. Biodynamic modeling

combines species-specific physiological traits to make predictions of metal bioaccumulation

that fare well when tested in the field. We generated biodynamic models for seven predatory

stonefly (Plecoptera) species representing the families Perlidae (5) and Perlodidae (2). Each

taxon was exposed to cadmium independently via diet and via solution. Species varied

approximately 2.6 fold in predicted steady state cadmium concentrations. Diet was the

predominant source of accumulated cadmium in five of the seven species and averaged 53.2

± 9.6 % and 90.2 ± 3.7% of net Cd accumulation in perlids and perlodids, respectively.

Differences in Cd bioaccumulation between the two families were largely driven by

differences in dissolved accumulation rates, which were considerably slower in perlodids

than in perlids. We further examined the subcellular compartmentalization of Cd

accumulated from independent aqueous and dietary exposures. Predicted steady state

concentrations were modified to only consider Cd accumulated in metal-sensitive subcellular

compartments. These values ranged 5.3 fold. We discuss this variability within a

phylogenetic context and its implications for bioassessment.

32

Introduction

Human activities have resulted in the contamination of trace metals in aquatic

ecosystems worldwide (1). As a result, benthic community structure, more specifically,

macroinvertebrate density and species diversity can be significantly impacted (2,3). Aquatic

insects constitute a large majority of the invertebrate species pool in most lotic systems and

play fundamental roles in stream ecosystem function (4). Accordingly, aquatic insects are

widely used in bioassessment (5,6), ecological risk assessments (7) and as biomonitors of

contaminant exposure (8,9). Aquatic insects have proven to be invaluable in evaluating the

integrity of ecosystems impacted by anthropogenic stressors such as trace metals, yet their

physiology has received surprisingly little attention.

An exciting new development in bioassessment is the incorporation of species traits

to better understand species distributions in different environments (10,11). The application

of species traits to better understand aquatic insect responses to contaminants (12) such as

pesticides (13) and trace metals (14,15) will likely require focused attention to physiological

traits and their differences among species. Understanding the distribution of physiological

traits within a phylogenetic framework is a potentially powerful means of determining how

and why species are differentially susceptible to trace metal contamination. Likewise, a

better understanding of physiological differences among taxa can possibly lead to the use of

insects to diagnose the causes of ecological impairment in stream ecosystems.

Biodynamic modeling has emerged as a powerful tool for predicting trace metal

bioaccumulation differences among species (16). The dynamic multi-pathway

bioaccumulation model (DYM-BAM) integrates laboratory derived rate constants of

accumulation and elimination. These trace metal accumulation and elimination processes can

33

be considered species-specific physiological traits. The predictive power of biodynamic

modeling stems from the use of environmentally relevant exposure conditions and the

incorporation of metal accumulation from both aqueous and dietary sources.

Bioaccumulation studies that neglect dietary exposures may not provide ecologically relevant

forecasts of metal bioaccumulation. Numerous studies have acknowledged diet as a major

exposure route that significantly contributes to metal bioaccumulation in invertebrates such

as bivalves (17,18), freshwater crustaceans (19,20) and insects (21,22).

Two paradigms currently exist for understanding metal toxicity to aquatic organisms

– the Biotic Ligand Model (BLM) and biodynamic modeling. The BLM is emerging as a

powerful tool for understanding acute metal toxicity, and has been most effective when

applied to taxa that experience toxicity as a result of metal binding to biotic ligands on the

gill surface (23). The BLM approach appears best applied to species with relatively low acute

to chronic ratios. Aquatic insects however, do not appear to be acutely sensitive to trace

metals relative to other taxa (24), yet many taxa seem to be affected by chronic metal

exposures (15). Biodynamic modeling therefore appears to be better suited for understanding

differences in insect species’ susceptibilities to trace metal exposures. The underlying

premise is that metal toxicity occurs when metal influx exceeds the organism’s ability to

effectively eliminate or detoxify excess metal (2,25). While biodynamic modeling has proven

useful in predicting metal bioaccumulation difference among species, it should be linked

with other approaches to better predict biological consequences (toxicity) associated with

metal accumulation (15). One such approach is to examine the subcellular

compartmentalization of accumulated metals (14,15,26,27). The subcellular partitioning of

metals within an organism is an additional physiological trait that can be used in combination

34

with biodynamic modeling to help explain differences in trace metal susceptibility among

taxa (14,15). This approach is a first step in attempting to link metal bioaccumulation to

metal toxicity.

Here, we generated biodynamic models to compare cadmium bioaccumulation among

7 species of predatory stoneflies (order: Plecoptera). Five species from the family Perlidae,

and two from the family Perlodidae were examined. These families are often dominant

members of stream communities throughout North America. Models allowed us to predict

bioaccumulation differences among taxa and assess the relative importance of dissolved and

dietary exposures. Furthermore, we evaluated the subcellular compartmentalization of Cd

independently from each exposure route. Thus, we were able to predict differences in both

whole body cadmium bioaccumulation in each species, and also predict Cd associated with

potentially sensitive subcellular compartments. This allowed us to estimate Cd susceptibility

differences among taxa. The distribution of these physiological traits in the context of

phylogenetic framework is discussed.

35

Materials and Methods

Sample collection and handling. Larvae were field collected from wadable riffles using a

D-frame kick-net from high quality streams in California, Montana, Colorado and North

Carolina (Table 1). Specimens were sorted and placed in Whirl Pac® bags with stream water

and substrate and were transported to the lab in coolers. In the lab, larvae were placed in

clean, acid-washed 300 mL plastic exposure chambers containing ASTM artificial soft water

(48 mg L-1 NaHCO3, 30 mg L-1 CaSO4·2H2O, 30 mg L-1 MgSO4 and 2 mg L-1 KCl) with

aeration and were acclimated to laboratory conditions at 14°C for a minimum of 48 hours.

Larvae were deprived of food during this period. Following acclimation, insects were

randomly divided into three groups. Group 1 animals were used to estimate assimilation

efficiency and depuration kinetics from dietary exposure (see below). Group 2 animals were

used to investigate the subcellular compartmentalization of dietborne cadmium (see below).

Group 3 animals were used to examine dissolved Cd uptake and depuration kinetics as well

as subcellular Cd compartmentalization. Only apparently healthy, active and intact larvae

were used in experiments.

Assimilation efficiency and depuration kinetics. To investigate dietary cadmium

accumulation kinetics, approximately 10 larvae were weighed and individually placed in

separate 50 mL Nalgene® polypropylene exposure chambers containing artificial soft water

and a 2 x 3" piece of Teflon-coated mesh as artificial substrate. Each larva was hand fed one

contaminated prey item – the freshwater oligochaete, Lumbriculus variegatus. This species

has several characteristics that make it ideally suited for use as food item ((28), Xie, personal

communication). L. variegatus were uniformly labeled with 109Cd via dissolved exposure.

The exposure medium was prepared by adding 109CdCl2 (9 ng ml-1 in 0.1 N HNO3) and

36

CdCl2 (1 μg ml-1 in 0.1 N HNO3) to achieve a final cadmium concentration of 518 ng L-1 (4.6

nM), with a specific activity of approximately 2.14 μCi L-1. NaOH was added to the aqueous

exposure to adjust the pH to ~6.85. Immediately following ingestion of the contaminated

prey, larvae were thoroughly rinsed and assayed for initial radioactivity. Then, larvae were

placed in individual exposure chambers, provided with clean prey ad libitum, and assayed for

radioactivity daily over 10 days of depuration. Water was refreshed daily to minimize

reuptake of effluxed metal. Assimilation efficiencies (AE) were estimated based on the

proportion of tracer remaining in the larvae after fecal excretion. Ingestion rates were

determined by quantifying the mass of prey ingested per mass of insect, daily for 7 days

using 6-10 individuals. The rate constants for efflux (ke) were determined by fitting data from

the depuration phase to a nonlinear least squares regression model (29).

Subcellular compartmentalization of cadmium from dietary sources. To examine the

subcellular distribution of dietborne Cd, ~10 individuals (sample size varied depending on

availability) were fed contaminated prey ad libitum for a period sufficient to produce an

activity in the insect ≥ 2500 cpm (approximately five days although feeding durations varied

based on ingestion rate). Water was refreshed daily to minimize aqueous exposure from

effluxed metal. Following the exposure period, larvae were fed clean prey ad libitum for 2

days, followed by 2 additional days during which food was withheld. This regime was

designed to minimize radioactive gut contents that could possibly confound the interpretation

of subcellular compartmentalization studies. Following the depuration period, insects were

thoroughly rinsed with deionized water, weighed, assayed for radioactivity, and then stored

in liquid nitrogen.

37

For each taxon tested, three to four replicate homogenates were used to assess

subcellular compartmentalization. Each homogenate consisted of one to three individuals

depending on body size and availability. The homogenates were separated into five

operationally-defined subcellular compartments by differential centrifugation according to

methods adapted from Wallace et al. (26). All fractions were assayed for radioactivity. We

assumed that Cd in the heat-stable protein fraction was largely bound to metallothionein-like

proteins (30,31), while the heat-denatured fraction represented a variety of larger cytosolic

proteins (32). We refer to the heat stable and heat-denatured protein fractions as

metallothionein-like protein (MTLP) and non-MTLP, respectively. Data from the subcellular

fractions were summed into operationally defined metal-sensitive and detoxified

compartments. The metal-sensitive compartment comprised the non-MTLP, microsomal and

organelle fractions, each containing sites potentially susceptible to Cd binding. The

detoxified metal compartment was Cd associated with MTLP. Cadmium in the cell debris

fraction was interpreted as being associated with biologically inert tissue such as chitin.

Aqueous exposures. Methods are described previously in Cain et al. (33) and Buchwalter et

al. (15). Exposure media was identical to that used in preparation of the prey (see above).

Approximately 15-20 larvae were incubated in this Cd solution which was refreshed after

days one and three of exposure. Following 5 days of accumulation, half of the insects were

placed in clean artificial soft water and efflux was quantified daily by in vivo gamma

counting for a depuration period of 10 days (see above). The remaining larvae were frozen

for subcellular fractionation.

Radioactivity measurement. All radioisotope activity was measured using a Wallac Wizard

1480 gamma counter. Insects were individually weighed and placed into 20 ml counting vials

38

containing ~15 ml of 15°C artificial soft water. A counting time of three minutes yielded

counting errors of < 5 %. Appropriate corrections were made for radioactive decay and

counting efficiency.

Biodynamic modeling. The dynamic multi-pathway bioaccumulation model (DYM-BAM)

is discussed in detail by Schlekat et al. (29) and Luoma et al. (16). The differential equations

describing the processes of metal influx and efflux from both dissolved and dietary exposures

can be solved as an expression of the individual’s steady state concentration as in equation 1.

u W Fss

e

(k C ) (AE IR C )Ck

× + × ×= (1)

where,

ku is the uptake rate constant from solution (L g-1day-1)

CW is the concentration of metal in solution (µg L-1)

AE is assimilation efficiency

IR is ingestion rate (g g-1day-1)

CF is the concentration in the food (µg g-1)

and ke is the efflux or elimination rate constant (proportional daily loss, day-1).

Growth dilution, kg, was not considered in these experiments. Under the model assumption

that whole organism exposure is an additive function of both exposure pathways (34), the

relative importance of dietary versus aqueous exposure routes can be quantified by separating

their respective parameters (Equations 2 and 3). Equations 2 and 3 represent the expressions

used to calculate the steady state concentrations from waterborne and dietborne exposures,

respectively.

39

u wss(water)

e (water)

(k C )Ck×

= (2)

fss(food)

e (food)

(AE IR C )Ck× ×

= (3)

Where, ke(water) and ke(food) are efflux rate constants derived from aqueous and dietary

exposures, respectively.

Statistical Analysis. SAS system for Windows (version 8.10) software was used in statistical

analysis. Experimental data were normally distributed. Homogeneity of variance was verified

using Bartlett’s test for equal variances prior to application of parametric statistics. T-tests

were performed to compare between families and between exposure routes and one-way

ANOVA was performed to examine physiological parameter differences among families and

within two tribes of the perlid family.

40

Results

Assimilation efficiency (AE) and ingestion rate (IR). Cadmium AE estimates for 7 taxa

ranged from 57 to 93% (Table 2) and averaged 81.6 ± 2.0% in Perlidae, and 69.5 ± 4.5% in

Perlodidae. Ingestion rates ranged from 0.01 to 0.12 g g-1day-1, however IRs were not

determined for Calineuria californica and Baumanella alameda. For modeling purposes, the

mean IR from all other taxa was used as a surrogate value.

Cadmium elimination kinetics. Elimination rate constants (ke) were derived independently

from dissolved and dietary experiments. Dietary ke estimates ranged from 0.065 ± 0.014 day-

1 in the perlodid Baumanella alameda to 0.168 ± 0.035 day-1 in the perlid Claassenia

sabulosa. Overall, efflux rate constants were very similar among species and between

exposure routes. For example, the average ke(food) (0.084 ± 0.009 day-1) was similar to ke(water)

(0.083 ± 0.001 day-1), and there were no apparent differences in ke(food) between families

(p=0.823) (Table 2). However, ke(water) could not be estimated for the perlodids Baumanella

alameda and Isogenoides hansoni because dissolved Cd influx was so slow in these

organisms that levels of Cd loading sufficient for efflux studies were not achieved after five

days of Cd exposure (Figure 1). Within a given taxon, ke(food) and ke(water) did not differ

significantly in four of five taxa (Table 2). In Calineuria californica ke(food) and ke(water) were

significantly different (p = 0.0326), however, the difference was relatively small (~3% of

proportional daily loss). Within a given taxon, dietary and dissolved ke estimates did not

differ significantly in 4 of 5 taxa (Figure 1). ke estimates from dietary and dissolved

experiments were significantly different in Calineuria californica (p = 0.0326), however, the

difference was relatively small (~3% of proportional daily loss). For all taxa, average ke

estimates derived from dietary experiments did not differ from the average values derived

41

from dissolved exposures (p = 0.4754), supporting the use of a single ke value in biodynamic

modeling.

Dissolved cadmium uptake. There were significant differences in Cd dissolved uptake

(Figure 2) and associated rate constants (ku) (Table 2) among species (one way ANOVA, p

<0.0001). Uptake rate constants ranged 23 fold, from 0.009 ± 0.001 L g-1 d-1 in Baumanella

alameda to 0.211 ± 0.029 L g-1 d-1 in Acroneuria abnormis. Uptake rate constant estimates

were highly variable within the Perlidae, contrasting sharply with consistency in ke values

observed in this family. There was an apparent difference in Cd uptake rates between perlids

and perlodids with average ku values of 0.127 and 0.018 L g-1 d-1, respectively. While this

difference was marginally significant (p = 0.06), analyses including an additional perlid

(Doroneuria baumanni) and an additional perlodid (Skwala sp.) showed significant

differences in ku values (p = 0.04) (these taxa are not addressed in this paper because dietary

studies were not performed).

Biodynamic modeling. For each taxon, predicted ranges of Cd bioaccumulation (on a wet

weight basis, at 0.52 µg Cd L-1 in water and 1.53 µg Cd g-1 in food) were generated by

combining the minimum, median and maximum estimates of model parameters (Table 2).

For example, the lower range of predicted Cd concentration was derived using the minimum

estimated AE, IR, and ku, and the maximum estimated ke derived from each taxon. Median

Cd concentrations ranged from 1.28 µg g-1 in Paragnetina sp. to 2.17 µg g-1 in Calineuria

californica among perlids, and were 0.84 and 1.07 µg g-1 in perlodids. The average median

predicted Cd concentrations were significantly different between perlids (1.65 µg g-1) and

perlodids (0.96 µg g-1) (p = 0.004). This difference is largely accounted for by differences in

dissolved Cd accumulation between the families. The predicted median Cd concentration

42

derived from aqueous exposure differed significantly between the two families (p = 0.002).

The average dietary derived Cd concentration was 0.90 ± 0.20 µg g-1 in perlids and 0.89 ±

0.11 µg g-1 in perlodids. These were not statistically different (p = 0.49). Diet was the

predominant exposure route in five out of the seven taxa tested (Figure 3). On average, diet

was a greater contributor to Cd bioaccumulation in the perlodid family (90.2 ± 3.7%) than in

the perlids (53.2 ± 9.6 %).

Cadmium compartmentalization. There were major differences in the subcellular Cd

compartmentalization associated with exposure route (Table 3). In all seven taxa, a greater

proportion of Cd accumulated from the aqueous route was distributed in the cell debris

compartment, likely representing integument-associated Cd. Thus, the relative proportions of

Cd in both MTLP and potentially sensitive compartments were uniformly higher in dietary

exposures compared to dissolved exposures. It is difficult to determine if these relationships

would be observed in organisms chronically exposed in nature. However, an interesting trend

emerges if the modeled steady-state Cd concentrations derived from each exposure route are

plotted against the concentrations of Cd projected in metal-sensitive fractions (Figure 5).

These results suggest that dietborne Cd is potentially more harmful than dissolved Cd.

Cadmium compartmentalization varied within and among the two stonefly families.

Similar compartmentalization of Cd accumulated from diet was observed among the perlids

Claassenia sabulosa, Paragnetina sp., Calineuria californica and Hesperaperla pacifica. In

these species, only 11-21% of the total accumulated Cd was found in the detoxified

compartment while a much greater percentage (42-69%) of the Cd body burden was

partitioned into metal-sensitive compartments. Conversely, the perlid Acroneuria abnormis

successfully detoxified 32% of the total Cd body burden while a slightly larger proportion

43

(38 %) of its metal burden was recovered in potentially toxic compartments. These data

indicate that subcellular metal compartmentalization can vary considerably within members

of the same insect family. Larvae from the family Perlodidae, Baumanella alameda and

Isogenoides hansoni, partitioned their Cd burdens similarly, effectively detoxifying 37% and

32% of their total body burdens, respectively.

Combining biokinetic modeling with subcellular fractionation. Modeled steady state

concentrations based on biokinetic parameters were modified to only consider Cd

concentrations accumulated in potentially sensitive subcellular compartments (Figure 4).

These concentrations predicted varied greater than 5 fold among all seven taxa, with the

greatest variation observed within the Perlidae. Under these experimental conditions, Cd

accumulation in sensitive compartments varied from 0.25 µg g-1 in Paragnetina sp. to 1.32

µg g-1 in Calineuria californica, suggesting that generalizations about susceptibility made at

the family level are likely inappropriate for this group. There were no obvious patterns

among perlid tribes from this small sample size. The two perlodids we examined were

similar with 0.61 and 0.55 µg Cd g-1 in potentially susceptible compartments in Isogenoides

hansoni and Baumanella alameda respectively.

44

Discussion

Biodynamic modeling has provided a framework for using short term standardized

protocols in the lab to make accurate predictions of metal bioaccumulation in nature (10).

This approach is particularly useful for species that are not amenable to long term handling in

the laboratory (19). In aqueous cadmium exposures, for example, estimates of the time

required for species to achieve steady-state body burdens (relative to the ambient water

concentrations used in these studies) range from 51-110 days, whereas rate constants used to

make these predictions were generated in only 15 days. Furthermore, biodynamic modeling

allows researchers to determine the relative importance of dissolved and dietary metal

exposures and explains why species vary so greatly in their accumulation of trace metals.

The overwhelming majority of species for which biodynamic models have been

generated are marine invertebrates (29,30). We chose stoneflies for these studies because

they are dominant members of stream communities in many parts of the world. Perlids are

widespread in the northern hemisphere and can be found in Africa and South America.

Perlodids are found spanning the Holoarctic ecozone (31). Furthermore, stoneflies are a

common focus in bioassessments of freshwater ecosystems. To our knowledge, biodynamic

models incorporating the dietary route of exposure have only been generated for two lake

dwelling predatory insects – the dipteran, Chaoborus punctipennis (15) and the

megalopteran, Sialis velata (16).

Our ke estimates fall within a range observed in many invertebrates. Among aquatic

insects, ke(water) estimates range from approximately 0.14 to 0.24 day-1 in the caddisflies, H.

californica (28), H. bettini (32) and Rhyacophila sp. (19). In a number of species of

Chaoborus, ke(water) has been reported to range from 0.003 to 0.08 day-1 (33). It is often

45

assumed that elimination rate constants from dietary and aqueous exposures are comparable

and differences may be small enough to be considered negligible (10). Our results validated

this assumption in predatory stoneflies. The inter-specific variation in cadmium ku is also

consistent with other studies. For example, kus range from 0.02 – 0.55 L g-1d-1 among several

mayfly taxa and from 0.04 – 0.42 L g-1d-1 among four caddisfly taxa (Buchwalter,

unpublished data). The differences in ku were large enough to explain much of the difference

in modeled Cd accumulation among stoneflies in our studies.

Our work with stoneflies is consistent with other predatory insects, in that diet is a

major source of accumulated metal (15,16,34). One explanation for diet being such an

important contributor to Cd accumulation in these taxa is that assimilation efficiencies were

quite high which is consistent with other studies reporting AE in predatory invertebrates

(34,35). We speculate that the high AEs observed in predatory insects in general may be

associated with feeding strategy. Predators that consume a significant percentage of their

body weight in one meal and then can fast for several days may benefit from more efficient

assimilation of nutrients, proteins, lipids, and essential metals, particularly when prey

availability varies spatially and temporally. With respect to prey composition, researchers

have characterized the main component of annelid cuticle as a collagen fibril matrix (44),

which may provide a challenge to the digestive systems of animals lacking a collagenase

enzyme. We observed that these organisms use peristalsis and antiperistalsis, most likely to

scour the cuticular matrix to efficiently remove all extractable proteins and nutrients.

Also with respect to assimilation efficiency, it has been suggested that the subcellular

distribution of metal within the prey affects trace metal bioavailability, with cytosolic metal

typically being highly available for trophic transfer (36-38). Subcellular

46

compartmentalization of the prey has indicated that after five days of dissolved cadmium

accumulation, up to 70% of the Cd accumulated by L. variegatus was distributed in cytosolic

compartments (Xie and Buchwalter, unpublished data). Interestingly, during depuration of

dietborne Cd, many individuals exhibited slow, steady tracer elimination prior to observed

fecal loss. This phenomenon suggests highly efficient transport of Cd from the midgut to the

haemoceole, with a small proportion of accumulated Cd subsequently removed by

malphigian tubules during the production of urine. This process is consistent with studies in

the blowfly (39) in which calcium uptake at the midgut was unregulated even in

hypercalcemic individuals. Haemolymph ionic composition appears to be regulated by

malphigian tubules and this is likely also the case with dietary cadmium accumulation in

predatory stoneflies.

In spite of diet being the dominant exposure pathway in 5 out of 7 taxa, these

predicted dietary contributions to net Cd accumulation can be considered highly conservative

estimates. Larvae were fed a simplified diet of oligochaetes exposed to Cd for only five days

with an average Cd concentration of 1.53 µg Cd g-1 wet weight (cadmium steady state

concentration in these worms were projected to be almost ten fold higher at 15.1 µg Cd g-1).

In nature, stonefly larvae feed on a variety of prey items including mayflies which can vary

widely in their metal concentrations (9). Likewise, in predicting Cd steady state

concentrations, the aqueous contribution can be considered an overestimate, given that the

dissolved uptake experiments were performed in the absence of dissolved organic carbon,

favoring free Cd ion concentrations.

Despite its utility in accurately predicting bioaccumulation differences among

species, biodynamic modeling does not directly establish a link between bioaccumulation and

47

toxicity, but assumes that toxicity occurs when an organism accumulates metal

concentrations in excess of its elimination and detoxification capacity (40). Subcellular

compartmentalization, despite its technical limitations (41), combined with biodynamic

modeling may be a means of discriminating among taxa based on susceptibility differences

(19). For instance, net modeled Cd bioaccumulation varied 2.6 fold among taxa. But when

subcellular partitioning was considered, taxa varied approximately 5.3 fold in Cd

concentrations predicted to occur in metal-sensitive compartments.

Relevance to nature. Few field data are available to verify modeled Cd bioaccumulation for

these taxa and its toxicological significance. Relative comparisons are possible for whole

body Cd concentrations in Claassenia, Hesperoperla and Isogenoides from the Clark Fork of

the Columbia River (42). There, cadmium concentrations were similar among the three

genera ( ~1.2 µg g-1 dry weight on average), although Isogenoides fell in the lower to mid

range of the concentration distribution which would seem to agree with our relative ranking

of [Cd]ss. With respect to the distribution of these taxa along the metal contamination

gradient, Isogenoides was common throughout the river, while Claassenia and Hesperoperla

were generally found only in the less contaminated reaches. Cain et al. (2004) provided

evidence that the distributions of several species in the Clark Fork (the species here were not

considered in their study) were related to differences in their partitioning of metals (Cu, Cd,

and Zn) into metal-sensitive compartments (21).

The use of aquatic insects in bioassessment should benefit greatly from the enhanced

understanding of physiological differences among species. Here we show that species-

specific physiological processes can be used to infer Cd susceptibility differences. While

highlighting a major physiological difference between perlids and perlodids (dissolved

48

uptake rates), we found substantial variability in this trait among perlids. Resh et al. (1975)

emphasized the importance of species-level identification in the use of these insects in

evaluating ecological conditions and commented on the limitations of designating whole

taxonomic groups (typically family level or higher) as tolerant, or intolerant (43). The results

of this study suggest that physiological variation within families may limit the efficacy of

family level generalizations of tolerance. In general, a better understanding of physiological

differences among aquatic insect species can only enhance the interpretive power of the

bioassessments that utilize these important organisms. Finally, in light of the importance of

dietary exposures in metal accumulation in many aquatic invertebrates, it seems appropriate

to question whether water quality criteria generated from studies that ignore this important

exposure route are adequately protective.

Acknowledgements

The authors would like to thank Dave Penrose at NCSU for his taxonomic expertise and

William Clements at Colorado State University for providing Claassenia sabulosa. The

authors also acknowledge the editorial contributions of G.A. LeBlanc, L. Xie and three

anonymous reviewers which greatly improved the presentation of this work. C.A.M. was

supported by a departmental fellowship in the Department of Environmental and Molecular

Toxicology at NCSU.

49

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35. Ng, T. Y. T.; Wang, W.-X. Modeling of Cadmium Bioaccumulation in two Populations of the Green Mussel Perna viridis. Environ. Toxicol. Chem. 2005, 24 (9), 2299-2305.

36. An Introduction to the Aquatic Insects of North America, 3 ed.; Kendall/Hun Publishing Co.: Dubuque. 1996.

37. Timmermans, K. R.; Spijkerman, E.; Tonkes, M.; Govers, H. Cadmium and Zinc Uptake by Two Species of Aquatic Invertebrate Predators from Dietary and Aqueous Sources. Can. J. Fish. Aquat. Sci. 1992, 49 655-661.

38. Croisetiere, L.; Hare, L.; Tessier, A. A field experiment to determine the relative importance of prey and water as sources of As, Cd, Co, Cu, Pb, and Zn for the aquatic invertebrate Sialis velata. Environ. Sci. Technol. 2006, 40 (3), 873-879.

39. Wang, W.-X.; Fisher, N. S. Assimilation Efficiencies of Chemical Contaminants in Aquatic Invertebrates: A Synthesis. Environ. Toxicol. Chem. 1999, 18 (9), 2034-2045.

40. Munger, C.; Hare, L. Influences of ingestion rate and food types on cadmium accumulation by the aquatic insect Chaoborus. Can. J. Fish. Aquat. Sci. 1999, 57 (2), 327-332.

53

41. Rainbow, P. S.; Wang, W.-X. Comparative Assimilation of Cd, Cr, Se, and Zn by the barnacle Elminus Modestus from Phytoplankton and Zooplankton Diets. Mar. Ecol. Prog. Ser. 2001, 218 239-248.

42. Wang, W. X.; Fisher, N. S.; Luoma, S. N. Assimilation of Trace Elements Ingested by the Mussel Mytilus edulis: effects of algal food abundance. Mar. Ecol. Prog. Ser. 1995, 129 165-176.

43. Peckarsky, B. L.; Cowan, C. A.; Anderson, C. R. Consequences and Plasticity of the Specialized Predatory Behavior of Stream-Dwelling Stonefly Larvae. Ecology. 1994, 75 (1), 166-181.

44. Humphreys, S.; Porter, K. R. Collagenous and Other Organizations in Mature Annelid Cuticle and Epidermis. J. Morphol. 1976, 149 (1), 33-52.

45. Cheung, M.; Wang, W.-X. Influence of subcellular metal compartmentalization in different prey on the transfer of metals to a predatory gastropod. Mar. Ecol. Prog. Ser. 2005, 286 155-166.

46. Wallace, W. G.; Luoma, S. N. Subcellular compartmentalization of Cd and Zn in two bivalves. II. Significance of trophically available metal (TAM). Mar. Ecol. Prog. Ser. 2003, 257 125-137.

47. Wallace, W. G.; Lopez, G. R. Relationship Between Subcellular Cadmium Distribution in Prey and Cadmium Trophic Transfer to a Predator. Estuaries. 1996, 19 (4), 923-930.

48. Taylor, C. W. Calcium Regulation in Blowflies: Absence of a Role for Midgut. Am J Physiol Regulatory Integrative Comp Physiol. 1985, 249 209-213.

49. Evans, R. D.; Balch, G. C.; Evans, H. E.; Welbourn, P. M. Simultaneous Measurement of Uptake and Elimination of Cadmium by Caddisfly (Trichoptera: Hydropsychidae) Larvae using Stable Isotope Tracers. Environ. Toxicol. Chem. 2002, 21 (5), 1032-1039.

50. Croteau, M.-N.; Hare, L.; Tessier, A. Differences in Cd Accumulation among Species of the Lake-dwelling Biomonitor Chaoborus. Can. J. Fish. Aquat. Sci. 2001, 58 1737-1746.

54

51. Toxicity of Dietborne Metals to Aquatic Organisms, SETAC Press: 2006; p 1-303.

52. Giguere, A.; Campbell, P. G.; Hare, L.; Couture, P. Sub-cellular partitioning of cadmium, copper, nickel and zinc in indigenous yellow perch (Perca flavescens) sampled along a polymetallic gradient. Aquat. Toxicol. 2006, 77 (2), 178-189.

53. Cain, D. J.; Luoma, S. N.; Carter, J. L.; Fend, S. V. Aquatic Insects as Bioindicators of Trace Element Contamination in Cobble-Bottom Rivers and Streams. Can. J. Fish. Aquat. Sci. 1992, 49 2141-2154.

54. Resh, V. H.; Unzicker, J. D. Water Quality Monitoring and Aquatic Organisms: the Importance of Species Identification. J. Water Pollution Control Federation. 1975, 47 (1), 9-19.

55

Table 1. Taxa, source location, mass and numbers of individuals used in

bioaccumulation experiments. Masses reported here are from animals used in dietary

experiments.

n Family Genus species Sampling location

mass (mg) ± SE aq. diet

Perlidae Claassinea sabulosa 40°40’01”N, 105°13’32”W 144.3 ± 27.6 20 20

Perlidae Paragnetina sp. 36°22’47”N, 81°09’11”W 144.6 ± 15.0 14 17

Perlidae Calineuria californica 37°17’20”N, 122°04’20”W 105.1 ± 10.1 20 20

Perlidae Acroneuria abnormis 36°22’47”N, 81°09’11”W 216.5 ± 18.5 16 20

Perlidae Hesperoperla pacifica 37°17’20”N, 122°04’20”W 265.6 ± 13.3 10 14

Perlodidae Isogenoides hansoni 36°22’10”N, 80°59’30”W 108.5 ± 13.2 6 16

Perlodidae Baumanella alameda 37°04’35”N, 121°28’02”W 34.9 ± 1.6 9 19

56

Table 2. Biokinetic models for 7 stonefly species. Taxa above dashed line are from the perlid family, taxa below are from perlodid

family. Ranges of steady state cadmium concentrations were modeled based on 0.52 µg L-1 Cd in water of and 1.53 µg Cd g-1 (wet

weight) in food. Values for the model parameters are reported as mean ± SE. A range of ingestion rates are reported here. Mean IR for

all other taxa were used in modeling bioaccumulation in Calineuria and Baumenella. On average, efflux rates were similar for both

exposure routes, with greater error estimates associated with dietary exposures. However, aqueous and dietary efflux rate constants

were significantly different in Calineuria.

[Cd]ss (µg g-1) Genus species AE

(%) IR

(g g-1 day-1) ku

(L g-1 day-1) ke (aqueous)

(day-1) ke (dietary)

(day-1) min med max

Claassenia sabulosa 77 – 93 0.03 – 0.09 0.100 ± 0.020 0.082 ± 0.007 0.091 ± 0.034 0.67 1.42 3.14

Paragnetina sp. 68 – 80 0.01 – 0.06 0.164 ± 0.001 0.085 ± 0.003 0.118 ± 0.027 0.98 1.28 1.90

Calineuria californica 81 – 89 – 0.102 ± 0.012 0.081 ± 0.032 0.049 ± 0.008 0.84 2.17 6.24

Acroneuria abnormis 75 – 89 0.03 – 0.06 0.211 ± 0.029 0.087 ± 0.003 0.076 ± 0.017 1.22 1.97 5.50

Hesperoperla pacifica 74 – 90 0.04 – 0.10 0.059 ± 0.012 0.082 ± 0.020 0.078 ± 0.013 0.51 1.43 4.26

Isogenoides hansoni 57 – 73 0.05 – 0.12 0.027 ± 0.003 – 0.110 ± 0.027 0.29 0.84 2.01

Baumanella alameda 68 – 79 – 0.009 ± 0.001 – 0.065 ± 0.014 0.38 1.07 2.45

57

Table 3. Subcellular cadmium compartmentalization in seven plecopteran species.

Cadmium body burdens were subdivided into five operationally based fractions by

differential centrifugation and heat treatment. Metal associated with organelles, microsomes,

and heat labile cytosolic proteins (non-MTLP) are considered potentially toxic. The

subcellular distributions of Cd after dissolved (grey) and dietary (white) exposures are

reported as a percentage ± SE. Data are normalized to recovered radioactivity.

fractions (%) detoxified potentially susceptible inert Genus species

MTLP non MTLP organelle microsome cell debris 8.8 ± 3.6 15.5 ± 3.0 14.1 ± 5.4 6.7 ± 1.2 52.9 ± 4.0 Claassenia sabulosa 14.0 ± 2.0 38.3 ± 3.0 6.5 ± 2.6 15.2 ± 1.5 26.0 ± 3.8 31.8 ± 0.7 2.6 ± 0.2 6.9 ± 3.7 2.4 ± 0.4 56.2 ± 3.4 Paragnetina sp. 11.5 ± 2.3 24.5 ± 3.5 11.1 ± 4.0 6.4 ± 0.5 46.5 ± 2.3 3.4 ± 0.7 21.9 ± 2.4 7.9 ± 1.6 11.5 ± 0.4 55.4 ± 6.7 Calineuria californica 14.0 ± 2.5 50.3 ± 4.0 8.7 ± 0.9 9.9 ± 2.0 17.4 ± 1.7 33.9 ± 1.2 2.5 ± 0.1 4.7 ± 0.6 5.8 ± 0.3 53.2 ± 1.7

Acroneuria abnormis 31.6 ± 0.9 21.4 ± 3.3 8.2 ± 1.1 8.2 ± 0.5 30.6 ± 0.8 7.7 ± 4.8 32.7 ± 5.6 8.8 ± 0.0 8.8 ± 1.5 41.9 ± 3.9 Hesperoperla pacifica 22.3 ± 4.4 49.6 ± 6.3 3.7 ± 0.2 7.6 ± 2.1 16.8 ± 3.3 10.3 ± 0.3 48.7 ± 0.3 7.2 ± 0.3 7.9 ± 0.5 25.9 ± 0.5 Isogenoides hansoni 23.3 ± 2.5 61.3 ± 6.1 3.0 ± 1.3 3.7 ± 0.9 8.6 ± 1.2 29.4 ± 3.1 15.4 ± 2.8 10.2 ± 0.8 10.8 ± 4.3 34.3 ± 3.8 Baumanella alameda 36.8 ± 6.8 32.2 ± 8.2 8.6 ± 1.1 10.8 ± 1.4 16.6 ± 1.9

58

Claass

enia

Paragn

etina

Caline

uria

Acrone

uria

Hespe

roperl

a

Isoge

noide

s

Bauman

ella

0.00

0.05

0.10

0.15ke(aq)

ke(diet)

*

Genus

k e(d

-1)

Figure 1. Comparison of efflux rate constants (ke) derived from aqueous and dietary

exposures. On average, proportional daily loss was similar for both exposure routes, with

greater error estimates associated with dietary exposures. These findings validate the use of a

single efflux rate constant in bioaccumulation models. The asterisk denotes a significant

difference in ke between exposure routes.

59

0 1 2 3 4 50

50

100

150

200

250

300

350

Calineuria californica

Baumanella alameda

Acroneuria abnormis

Isogenoides hansoni

days

ng C

d g-1

(wet

wei

ght)

Figure 2. Short term dissolved cadmium accumulation in selected taxa. Accumulation

was generally faster in perlids (closed symbols) than in perlodids (open symbols). These taxa

were chosen to illustrate the variation in Cd uptake rates. Organisms have not reached steady

state concentrations. Deviation from zero was significant in the slopes of all four taxa.

60

Claass

enia

Paragn

etina

Caline

uria

Acrone

uria

Hespe

roperl

a

Isoge

noide

s

Bauman

ella

0.0

0.5

1.0

1.5

2.0

2.5

[Cd]ss(diet)

[Cd]ss(aq.)

Genus

[Cd]

ss(u

g/g)

Figure 3. The relative importance of dietary vs. aqueous exposure routes and total

modeled cadmium accumulation. Diet contributed between 25 – 93% of net accumulation

(based on median model parameters) and was the dominant exposure pathway in five out of

seven taxa studied.

61

1.280.550.18Baumanella alameda

1.420.610.25Isogenoides hansoni

2.540.830.29Hesperoperla pacifica

1.730.440.21Acroneuria abnormis

3.981.320.47Calineuria californica

0.490.250.13Paragnetina sp.

1.690.710.29Claassenia sabulosa

max.med.min.

[Cd]ss in metal-sensitive compartments

Perlodidae

Perlidae

Figure 4. Predicted ranges of cadmium steady state concentrations (µg g-1) wet weight

in metal-sensitive subcellular compartments. Subcellular disposition of accumulated

cadmium is used to infer susceptibility differences among taxa. The cladogram on the left

indicates taxonomic relationship (source: The Tree of Life Web Project:

http://tolweb.org/tree/phylogeny.html). Considerable variation among taxa suggests that

family level generalizations about susceptibility are unsupported.

62

0.00 0.25 0.50 0.75 1.00 1.25 1.50 1.750.00

0.25

0.50

0.75

1.00

1.25aqueousdietary

R2=0.924

R2=0.184

modeled [Cd] accumulation (µg g-1) from aqueous and dietary exposure routes

[Cd]

(µg

g-1

) par

titio

ned

into

phys

iolo

gica

lly a

vaila

ble

com

partm

ents

Figure 5. Relationship between total and physiologically available cadmium

bioaccumulation. Each species is represented by two points depicting its aqueous and

dietary accumulation. Diet derived Cd efficiently targets potentially metal-sensitive

compartments relative to the dissolved route of exposure. These data suggest that dietborne

metal is potentially more toxic than Cd accumulated from solution.

63

CHAPTER 2.

DIETARY CADMIUM DYNAMICS: INVESTIGATIONS OF EXPOSURE HISTORY

AND INSECT DIGESTIVE PHYSIOLOGY

*The following chapter includes information on studies that were either unsuccessful or

yielded results that potentially warrant further investigation.

1. The effect of metal exposure history on cadmium assimilation and elimination

dynamics

Organisms are capable of developing toxicity avoidance traits through physiological

acclimation or through the selection of resistant genotypes (1-3). Researchers have

investigated the influence of metal exposure history on physiological processes related to

metal bioaccumulation in aquatic invertebrates. Previous studies with this objective

concluded that neither the pre-exposed caddisfly, Hydropshche californica (1) nor the pre-

exposed green mussel, Perna viridis (4) exhibited efflux dynamics significantly different

from naïve organisms. Furthermore, dietary efflux dynamics of the pre-exposed freshwater

cladoceran, Daphnia magna were little affected as compared to naïve organisms (5). There is

some evidence of an increase in assimilation efficiency (AE) of Ag from diet in Perna viridis

following pre-exposure to a mixture of Cu, Zn and Ag, but the assimilation of Zn and Cu

were not influenced by exposure history (6).

We proposed that organisms inhabiting metal contaminated streams exhibit Cd

elimination kinetics and assimilation efficiencies (AE) that are distinct from those of

organisms inhabiting uncontaminated streams. The hypothesis was tested that elimination

rate constants (kes) derived in the lab would differ (increased elimination rate) such that Cd

64

bioaccumulation would be constrained in the pre-exposed population. The studies in the

previous chapter were performed using only animals with no known prior exposure history,

so we questioned how exposure history might affect the processes related to Cd accumulation

from diet in a predatory stonefly. We also examined differences in Cd AE of insects collected

from contaminated and reference sites. Due to the limited availability of this test species at

the contaminated site, dissolved Cd uptake and efflux kinetics experiments were not

performed, nor were we able to examine the compartmentalization of Cd accumulated from

aqueous or dietary routes of exposure.

Methods

Claassenia sabulosa (order: Plecoptera) were field collected in Montana from the

upper reaches of the contaminated Clark Fork of the Columbia River and from a reference

site, Rock Creek. The Clark Fork is contaminated with Cd, Zn and Cu from historic mining

operations in Butte and from an ore smelter in Anaconda, both located in the headwaters of

the Clark Fork (7). Rock Creek is a tributary to the Clark Fork and has no significant source

of metal contamination. Organisms were placed in labeled plastic containers along with mesh

substrate and river water purged with 90% pure O2, then packed in a cooler and shipped back

to the lab. In the lab, larvae were acclimated in ASTM artificial soft water for one week prior

to this experiment.

The prey, L. variegatus were uniformly labeled with 109Cd as described in the

previous chapter. Each larva was hand fed one contaminated worm, rinsed and assayed for

initial tracer activity. Larvae were provided with clean prey ad libitum and Cd elimination

was monitored daily for 10 days by in vivo gamma counting. Only the portions of the curve

65

representing the slow exchanging compartment of metal (sometimes referred to as ke2) are

shown in Figure 1.

Results and Discussion

Mean cadmium elimination rate constants, (kes) were 0.197 ± 0.034 and 0.168 ±

0.035 day-1 in larvae from Clark Fork and Rock Creek, respectively (Student’s T-test p =

0.566). These data suggest that animals with a history of metal exposure are not capable of

avoiding toxicity by eliminating dietborne metal at a faster rate than naïve organisms. No

significant differences (p =0.3306) were observed in Cd AE in insects collected from

contaminated versus reference sites (average Cd assimilation efficiencies were 79.2 ± 10.2 %

and 85.0 ± 7.8 % for pre-exposed and naïve larvae, respectively). However, the small sample

sizes

(6 individuals per sample site) due to limited availability preclude any definitive conclusions

about the influence of metal exposure history on Cd efflux dynamics and AE in this taxon.

To my knowledge, there have been no published studies reporting the effects of

exposure history on the bioaccumulation of dietborne metals in insects. However, these

results are consistent with those of studies that examined the influence of pre-exposure on

efflux dynamics observed in pre-exposed Daphnia magna following dietary Cd and Zn

exposure (5). Ideally, these comparative studies should be repeated with larger sample sizes

examining other physiological processes related to Cd bioaccumulation/susceptibility.

66

2. Dual labeling technique to quantify gut retention time and cadmium assimilation

from diet

Multiple radiotracer techniques have often been used to determine assimilation

efficiency (AE) of trace metals (8,9). This method relies on the collection and analysis of

feces for tracer content. Species specific gut retention time or gut residence time (GRT) is the

amount of time it takes for a single bolus of ingested food to pass through the GI tract. In our

Cd elimination kinetics studies, GRT a factor thought to influence AE, was difficult to

ascertain with certainty due to the relatively low tracer activity observed in feces. Dual

labeling has been an effective tool for determining GRT in invertebrates exposed to dietborne

trace metals (10,11). In this approach, the presence of an inert tracer (one that is not

absorbed/assimilated in the GI tract) in the feces is used to indicate when the initial food

bolus has passed through the organism. 51Chromium was thought to be a suitable tracer

because it has been used previously to determine AE in the freshwater snail, Lymnaea

stagnalis using enriched stable Cr isotope (10).

Methods

Nigronia sp. (Order: Megaloptera) larvae were collected from the Eno River, NC and

used to determine if 51Cr could be used as an inert tracer of GRT. L. variegatus were

incubated for five days in pH adjusted (~6.8) exposure media containing a known ratio of

51Cr to 109Cd. A protocol was established on a Perkin-Elmer 1480 automatic gamma counter

(Shelton, Connecticut, USA) for simultaneous gamma counting of these isotopes which

possess distinctly different energy ranges. Ten individuals were fed one contaminated worm

and were then rinsed and assayed for initial activities. Larvae were subsequently provided

with clean prey ad libitum. Both larvae and feces were collected and assayed daily for ten

67

days. Water was refreshed twice daily to minimize the sorption of any effluxed isotope onto

clean prey items or excreted feces.

Results and Discussion

Essentially no 51Cr was found in feces after ten days of depuration. These results

indicate that 51Cr is in fact absorbed in the GI tract and is therefore not a suitable isotope for

dual tracer studies. This study may have been complicated by the fact that these organisms

are particularly messy poopers. Whereas stoneflies likely possess a peritrophic matrix (12)

that acts as a tidy poop preparation and packaging system, conveniently enabling the ease of

feces removal from the exposure chamber. However, Cr assimilation in Nigronia sp. was

confirmed by the relatively small reduction in chromium activity observed in the larvae even

after ten days of depuration. Attempting this dual labeling technique using alternative tracers

(e.g. 60Co (8)) may prove to be worthwhile in efforts to further elucidate the digestive

physiology of these insects.

68

3. Cadmium transport, absorption and elimination

Cadmium competes with calcium for uptake via calcium transporters on chloride cells

found on the gill/body surface (13). Such interactions are poorly understood in the digestive

tract of aquatic organisms, prompting the question does calcium content of prey affect dietary

Cd dynamics? To further our understanding of Cd dynamics in the GI tract, we assessed

whether factors that control dissolved accumulation are relevant in the GI tract. The calcium

content of prey was manipulated via culture under different Ca conditions and fed to the

predatory stonefly, Acroneuria abnormis (Order: Plecoptera) to determine if calcium

competes with cadmium in the digestive tract.

Methods

Lumbriculus variegatus were cultured in very soft, moderately hard and very hard

ASTM reconstituted water (containing 12 mg/L NaHCO3, 7.5 mg/L CaSO4.2H2O, 7.5 mg/L

of MgSO4, and 0.5 mg/L of KCl (very soft), 96 mg/L NaHCO3, 60 mg/L CaSO4.2H2O, 60

mg/L MgSO4, and 4 mg/L KCl (moderately hard) and 384 mg/L NaHCO3, 240 mg/L

CaSO4.2H2O, 240 mg/L MgSO4, and 16 mg/L KCl (very hard)) for 10 days at 4oC without

aeration. Approximately 0.5 gram of worms was placed into 50 ml Nalgene® polypropylene

beakers under the very soft, moderately hard, and very hard waters. Individuals were then

removed from their respective beakers, rinsed with deionized water, blotted dry and weighed.

Individuals were then acid digested with ultrapure nitric acid (1 ml) at 80oC for 2 hours.

Calcium content was determined by inductively coupled plasma atomic emission

spectroscopy (ICP-AES). After ten days, the calcium content of Lumbriculus variegatus

incubated in very hard water differed significantly (p = 0.037) from those incubated in very

soft water (Figure 2).

69

Once it was determined that the calcium content of the prey could be successfully

manipulated, worms were again incubated in very soft and very hard water with 109Cd (0.55

µg L-1 of Cd with a final activity of 1.14 µCi L-1). pHs for the exposure media were adjusted

with 1M phosphate buffer saline (pH 6.8) to approximately 7.1. Exposure initiation times for

treatment groups varied such that upon feeding, prey was of similar Cd burdens but varying

Ca contents. (Additional worms were analyzed by ICP-AES to verify differing calcium

concentrations as in Figure 2.). L. variegatus Cd body burdens were determined by in vivo

gamma.

Acroneuria abnormis larvae were divided into two groups of ten individuals and fed

one contaminated worm incubated in either very soft or very hard Cd exposure media.

Larvae were subsequently fed clean worms ad libitum and Cd elimination was monitored

daily for 10 days by in vivo gamma counting. The rates of Cd elimination were compared

between larvae fed hypercalcemic prey (very hard exposure media) and larvae fed

hypocalcemic (very soft exposure media) prey.

Results and Discussion

Average Cd elimination rates of larvae fed hypercalcemic prey were 0.617 ± 0.298 ng

Cd g-1 hr-1, whereas average elimination rates of larvae fed hypocalcemic prey were 0.950 ±

0.461 ng Cd g-1 hr-1. Average elimination rates between treatment groups were not

significantly different due in part to the large standard errors; however these data are

compelling and warrant further investigation using larger sample sizes to minimize effects of

inter-individual variability. These results suggest that increased calcium in the ingested food

may compete with Cd in the GI tract of these organisms, thereby inhibiting Cd elimination.

Ca2+ could potentially compete with Cd2+ for absorption from the midgut, or uptake at the

70

malphigian tubules (Figure 3). Malphigian tubules are insect excretory structures and the site

of urine production (14). Previous studies examining Cd elimination in these insects indicates

steady loss of tracer for as many as six days prior to observed fecal elimination. This

observation suggests that urinary excretion could be a dominant metal elimination pathway in

these insects.

71

4. Calcium and cadmium interaction in insect haemolymph

The results of the previous study prompted us to further investigate the potential sites

of competition between cadmium and calcium in the digestive tract. We hypothesized that

this metal interaction could take place either at the midgut or at the malphigian tubules. So

we proposed to bypass the midgut and by injecting cadmium directly into the haemocele of

Pteronarcys dorsata (order: Plecoptera). This design was intended to only consider possible

competition occurring at the tubules. Treatment groups included Cd only, Cd with Ca or Cd

with verapamil, a known calcium channel blocker.

Methods

For the purpose of determining the appropriate amount of calcium to inject to alter the

haemolymph calcium content, two replicates of haemolymph were subsampled from 6

individuals using 50 µL flame-pulled capillary pipets. Pooled samples were acid digested in

ultrapure nitric acid (1 ml) at 80oC for 2 hours. Calcium content was determined by ICP-

AES. To significantly alter the total body Ca concentration, (20% increase) it was necessary

to estimate the volume of haemolymph in each individual. Buchwalter et al. determined

haemolymph volume in a closely related species, Pteronarcys californica through the

injection 14C-inulin carboxy (15). Using the standard curve from their study, we determined

haemolymph volume by each individual’s mass, and multiplied it by the Ca concentration of

the haemolymph as an estimate of total body Ca concentration.

Stock solutions were made as follows: treatment 1 (Cd only) 150 µL of 0.55 µg L-1 of

109Cd and 300 µL deionized water, treatment 2 (Cd and Ca) was prepared with 150 µL of

0.55 µg L-1 of 109Cd and 300 µL of 8 g L-1 of CaCl2 150 µL of 0.55 µg L-1 of 109Cd and 300

µL of 200 µM verapamil. All solutions were pH adjusted to approximately 7.1 with NaOH.

72

Larvae were anesthetized with CO2, blotted dry and weighed. The volume of injected

solution required to alter the total body Ca concentration by 20% was determined for each

individual then that volume of treatment solution was injected into the insect’s dorsal aorta.

Immediately following injection, larvae were placed in Petri dishes containing wet paper

towel in order to keep the gills moist but to avoid irritating the injection site. The dishes were

purged with 90% pure O2 and covered. Larvae were allowed to recover for approximately

twenty minutes and were then assayed for initial activity, and assayed again at 4, 8, and 18

hours and daily for the next 8 days.

Results and Discussion

There were no significant differences in cadmium elimination rate among larvae in

the three injection treatment groups (Figure 5) which were surprisingly slow relative to

elimination of Cd accumulated from diet or from solution. It is possible that the animals were

stressed from the injection procedure, resulting in altered Cd elimination dynamics. Another

plausible explanation for this observation is that the injected Cd may have been effectively

bound to ligands within the haemolymph and sequestered rather than bound to proteins that

facilitate Cd excretion.

73

Literature Cited

1. Cain, D. J.; Buchwalter, D. B.; Luoma, S. N. Influence of Metal Exposure History on the Bioaccumulation and Subcellular Distribution of Aqueous Cadmium in the Insect Hydropsyche californica. Environ. Toxicol. Chem. 2006, 25 (4), 1042-1049.

2. Morgan, A. J.; Kille, P.; Sturzenbaum, S. R. Microevolution and Ecotoxicology of Metals in Invertebrates. Environ. Sci. Technol. 2007, 41 1085-1096.

3. Xie, L.; Klerks, P. L. Responses to Selection for Cadmium Resistance in the Least Killifish, Heterandria Formosa. Environ. Toxicol. Chem. 2003, 22 (2), 313-320.

4. Blackmore, G.; Wang, W. X. Uptake and efflux of Cd and Zn by the green mussel Perna viridis after metal preexposure. Environ. Sci. Technol. 2002, 36 (5), 989-995.

5. Guan, R.; Wang, W.-X. Cd and Zn Uptake Kinetics in Daphnia magna in Relation to Cd Exposure History. Environ. Sci. Technol. 2004, 38 6051-6058.

6. Shi, D.; Wang, W.-X. Modification of Trace Metal Accumulation in the Green Mussel Perna viridis by Exposure to Ag, Cu, and Zn. Environ. Poll. 2004, 132 265-277.

7. Cain, D. J.; Luoma, S. N.; Wallace, W. G. Linking metal bioaccumulation of aquatic insects to their distribution patterns in a mining-impacted river. Environ. Toxicol. Chem. 2004, 23 (6), 1463-1473.

8. Weeks, J. M.; Rainbow, P. S. A Dual-labelling Technique to Measure the Relative Assimilation Efficiencies of Invertebrates Taking up Trace Metals from Food; 1990.

9. Wang, W.-X.; Fisher, N. S. Assimilation Efficiencies of Chemical Contaminants in Aquatic Invertebrates: A Synthesis. Environ. Toxicol. Chem. 1999, 18 (9), 2034-2045.

10. Croteau, M. N.; Luoma, S. N.; Pellet, B. Determining metal assimilation efficiency in aquatic invertebrates using enriched stable metal isotope tracers. Aquat. Toxicol. 2007, 83 (2), 116-125.

11. Decho, A. W.; Luoma, S. N. Time-courses in the retention of food materials in the bivalves Potamocorbula amurensis and Macoma balthica: significance to the absorption of carbon and chromium. Mar. Ecol. Prog. Ser. 1991, 78 303-314.

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12. Boonsriwong, W.; Sukontason, K.; Olson, J. K.; Vogtsberger, R. C.; Chaithong, U.; Kuntalue, B.; Ngern-Klun, R.; Upakut, S.; Sukontason, K. L. Fine structure of the alimentary canal of the larval blow fly Chrysomya megacephala (Diptera: Calliphoridae) 1. Parasitol. Res. 2007, 100 (3), 561-574.

13. Buchwalter, D. B.; Luoma, S. N. Differences in dissolved cadmium and zinc uptake among stream insects: mechanistic explanations. Environ. Sci. Technol. 2005, 39 (2), 498-504.

14. Hare, L. Aquatic Insects and Trace Metals: Bioavailability, Bioaccumulation, and Toxicity. Crit. Rev. Toxicol. 1992, 22 (5/6), 327-369.

15. Buchwalter, D. B.; Jenkins, J. J.; Curtis, L. R. Temperature Influences On Water Permeability and Chlorpyrifos Uptake in Aquatic Insects with Differing Respiratory Strategies. Environ. Toxicol. Chem. 2003, 22 (11), 2806-2812.

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5 6 7 8 9 100.0

0.1

0.2

0.3

0.4

0.5

0.6

0.7Clark ForkRock Creek

day

prop

ortio

nal d

aily

loss

Figure 1. Cadmium elimination rates of pre-exposed larvae (Clark Fork) and naïve

larvae (Rock Creek)

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very hard moderate very soft0.15

0.20

0.25

0.30

0.35

mg

Ca

g-1 w

et w

eigh

t

Figure 2. Calcium content of Lumbriculus variegatus cultured in very hard, moderately

hard and very soft ASTM reconstituted water.

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very hard very soft0

102030405060708090

cpm

mg

-1 w

et w

eigh

t

Figure 3. 109Cadmium activity (cpm) per mg (wet weight) in L. variegatus incubated in

very hard and very soft exposure media.

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Figure 4. Image of stonefly GI tract. The insect foregut functions in the storage and

partial break down of food. The midgut digests and absorbs nutrients, ions, lipids, etc.

The malphigian tubules, along with the hindgut (not depicted here, located posterior to

the tubules) manage water/ion balance and excretion.

foregut

midgut

malphigian tubules

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0 50 100 150 2000.8

0.9

1.0

Cd + verapamilCd onlyCd + Ca

hour

ln (p

ropo

rtio

nal l

oss)

Figure 5. Proportional cadmium loss in Pteronarcys dorsata larvae injected with

verapamil and Cd, Cd only, or Cd and Ca treatments.

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CONCLUSION

The objective of this thesis was to address some of the issues associated with the

presence of trace metals in stream environments. In part, this research evaluated the relative

importance of dissolved versus dietary routes of cadmium exposure in predatory stoneflies

and concludes that diet is the predominant source of accumulated cadmium in the majority of

these organisms. These results have implications for the establishment of water quality

criteria which rely on toxicity bioassays that neglect the dietary exposure route.

Additionally, this research was intended to examine the extent to which phylogenetic

position among stoneflies in the families Perlidae and Perlodidae is predictive of

physiological parameters related to cadmium accumulation and susceptibility. Physiological

processes such as the rate of dissolved Cd uptake and detoxification capacity appeared to

vary among closely related stoneflies within the perlid family. These results suggest that the

practice of designating an entire insect family as tolerant or susceptible may be unsound.

Furthermore, this work contributes significantly to a growing body of biodynamic

metal accumulation modeling literature in which insects are underrepresented. Although

biokinetic modeling provides reasonable estimates of bioaccumulation that correlate with

bioaccumulation in nature, it does not establish a direct link to toxicity. So, in order to better

understand species specific differences in metal susceptibility, bioaccumulation estimates

were modified to consider only metal partitioned into metabolically active or potentially

toxic biological compartments. This approach is becoming recognized as an initial step

towards understanding species specific susceptibility to trace metals however, more direct

toxicity endpoints (e.g. oxidative stress) should be explored to establish trace metal

susceptibility mechanisms in aquatic organisms.

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In summary, aquatic insects are an ecologically important group of organisms.

Stoneflies (Plecoptera) are an important and diverse group of organisms that are particularly

relevant due to their widespread use in ecological assessment. A major knowledge gap exists

regarding the physiology of this important insect order, but by working with native

organisms we can begin to understand how physiological processes underlie species’

responses to anthropogenic stressors. A better understanding of how species differ in terms of

these physiological processes within and among different taxonomic groups could only

augment the interpretive power of their use in ecological assessments.