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Stress-induced alterations of ecosystem function: the role of acidification in lotic

metabolism and biogeochemistry

Damon T. Ely

Dissertation submitted to the faculty of the Virginia Polytechnic Institute and State University in

partial fulfillment of the requirements for the degree of

Doctor of Philosophy

In

Biological Sciences

Herbert M. Valett, Chair

James A. Burger

Walter L. Daniels

Robert H. Jones

Jackson R. Webster

31 March 2010

Blacksburg, Virginia

Keywords: respiration, nitrogen uptake, spiraling, DIN, 15

N, stream metabolism, aquatic fungi,

chlorophyll a, qCO2

Copyright © 2010, Damon T. Ely, All Rights Reserved

Stress-induced alterations of ecosystem function: the role of acidification in lotic metabolism and

biogeochemistry

Damon T. Ely

Abstract

I investigated how anthropogenic acidification influences stream metabolism and nitrogen (N)

cycling by considering the stress response of microbial compartments responsible for these

ecosystem processes. Microcosm incubations of leaf biofilms from streams of differing pH

revealed greater rates of fungal biomass-specific respiration (i.e. the stress metric qCO2) and

biomass-specific N uptake (i.e. qN) with increasing acidity. The positive relationship between

qCO2 and qN indicated alternate fates for N other than structural biomass, possibly related to

increased exoenzyme production as part of the stress response.

Whole-stream 15

N experiments and measurements of respiration and fungal standing crop

across the pH gradient resulted in similar patterns in qCO2 and qN found in microcosm

experiments, supporting qCO2 as an ecosystem-level stress indicator and providing insight

towards controls over N cycling across the pH gradient. Fungal biomass and ecosystem

respiration declined with increasing acidity while N uptake metrics were not related to pH, which

suggested qN in acid streams was sufficiently high to counteract declines in fungal abundance.

During spring, chlorophyll a standing crops were higher in more acidic streams despite

lower nutrient concentrations. However, N uptake rates and gross primary production differed

little between acid and circumneutral streams. Reduced heterotrophy in acid streams was

apparent in lower whole-stream respiration rates, less ability to process organic carbon, and little

response of N uptake to added carbon resources. Overall, acid-induced stress in streams was

found to impair decomposer activity and caused a decoupling of carbon and nitrogen cycles in

these systems.

iii

Acknowledgements

I am thankful for the guidance and friendship of my advisor Maury Valett, who allowed and

encouraged my creative freedom. His dedication to mentoring cannot be understated and has

made all the difference in my professional development. I thank Jack Webster for always being

there to help, and for his enthusiasm toward the mysteries of nature. I thank Bob Jones for his

excellence as department head and his support of all graduate students. I thank both Bob Jones

and Lee Daniels for their many helpful comments that greatly improved the quality of this work.

I thank Jim Burger for his insight, and for providing the stimulation for this research, possibly

without knowing it.

I thank the students and faculty whom compose the Stream Team, especially Fred Benfield

for providing me the opportunity to join this group, Jeb Barrett for our many stimulating

conversations, and Bobbie Niederlehner for her patience and support in the laboratory. I would

like to thank Daniel von Schiller, a proud colleague and good friend, for all of his hard work. I

also thank NSF for funding much of the research herein.

I am thankful for my wife, whose unwavering support of all my efforts, academic and

otherwise, has provided me the strength to always ‗push on‘. My son has given me joy beyond

all expectations, a factor that has contributed to the completion of this dissertation in numerous

valuable and indescribable ways. Many, many thanks are deserved for Gayle Pierce and Julia

Armstrong, who gave us their home out of the kindness of their hearts because of their strong

belief in the value of family and hard work.

iv

Attribution

A colleague and my advisor aided in the writing and research behind the chapters of this

dissertation. A brief description of their background and their contributions are included here.

Prof. H. Maurice Valett, PhD (Department of Biological Sciences, Virginia Tech) currently at

Flathead Lake Biological Station, Montana, is the primary Advisor and Committee Chair. Prof.

Valett provided considerable help in the development of the ideas herein and was the primary

editor of the dissertation. Prof. Valett is a co-author on Chapter 2, which has been published in

the journal Freshwater Biology V.55 p.1337-1348.

Daniel von Schiller, PhD (Leibniz-Institute of Freshwater Ecology and Inland Fisheries,

Müggleseedamm, Berlin, Germany) was a visiting scholar who worked closely with me to

develop and carry out the laboratory experiments in Chapter 2. Dr. von Schiller is also a co-

author on this published manuscript.

v

Table of Contents

Abstract .......................................................................................................................................... ii

Acknowledgements ...................................................................................................................... iii

Attribution .................................................................................................................................... iv

List of tables.................................................................................................................................. vi

List of figures ............................................................................................................................... vii

Chapter 1: General Introduction ................................................................................................ 1

Chapter 2: Stream acidification increases nitrogen uptake by leaf biofilms: implications at

the ecosystem scale ........................................................................................................................ 5

Abstract ..................................................................................................................................................... 5

Introduction .............................................................................................................................................. 6

Methods .................................................................................................................................................... 9

Results ..................................................................................................................................................... 14

Discussion................................................................................................................................................ 17

Literature cited........................................................................................................................................ 23

Chapter 3: Ecosystem stress: acidification influences on metabolic and biogeochemical

processes in lotic ecosystems ...................................................................................................... 34

Abstract ................................................................................................................................................... 34

Introduction ............................................................................................................................................ 35

Methods .................................................................................................................................................. 38

Results ..................................................................................................................................................... 44

Discussion................................................................................................................................................ 46

Literature Cited ....................................................................................................................................... 52

Chapter 4: Acidification-induced enhancement of benthic algal biomass: influences on

metabolism and C and N uptake in forested headwater streams ........................................... 67

Abstract ................................................................................................................................................... 67

Introduction ............................................................................................................................................ 68

Methods .................................................................................................................................................. 71

Results ..................................................................................................................................................... 77

Discussion................................................................................................................................................ 82

Literature Cited ....................................................................................................................................... 90

Chapter 5: Synthesis ................................................................................................................. 103

vi

List of tables


the ecosystem scale

Table 1. Geographical and geological characteristics of study streams in Shenandoah

National Park, VA, USA...…………………………………………………..…… 28

Table 2. Chemical parameters and leaf characteristics of the study streams.……………… 29


processes in lotic ecosystems

Table 1. Temperature and chemical characteristics of the five study streams...................... 59

Table 2. Standing crops of coarse benthic organic matter (CBOM) and fungal biomass, flow

characteristics, and N uptake parameters for the five study streams...................... 60


metabolism and C and N uptake in forested headwater streams

Table 1. Average daily (24h) temperature, mean concentrations of various dissolved solutes

and mean standing crops of chlorophyll a and epilithic biomass in the study streams

from 12 April 2009 to 11 July 2009.........................................................................96

Table 2. Stream pH and nutrients, photosynthetically active radiation (PAR) and flow

conditions of study streams during 15

N releases with (+ DOC) and without glucose

additions...................................................................................................................97

Table 3. Stream periphyton, metabolism and carbon (C) uptake metrics measured during

15

N tracer experiments..............................................................................................98

vii

List of figures


the ecosystem scale

Figure 1. Linear regressions of stream pH vs. a) Ln fungal biomass, b) microbial qCO2 in

microcosms with low (25 μg NH4-N L-1

) and high (100 μg NH4-N L-1

) nitrogen (N)

availabilities and c) respiration per unit leaf mass in microcosms with low and high

N………………………………………………………………………………….. 30

Figure 2. Linear regressions of stream pH vs. a) microbial N uptake and b) N uptake

standardized to leaf mass in microcosms with low (25 μg NH4-N L-1

) and high (100

μg NH4-N L-1

) nitrogen (N) availabilities…………………………………………31

Figure 3. Log-log relationship between microbial qCO2 and N uptake measured in

laboratory microcosms………………………………………………………...… 32

Figure 4. Conceptual model of proposed linkage between acidification as an environmental

stressor, exogenous N as a subsidy, physiological responses of microbes

influencing C and N processing, and consequences at the ecosystem level.......... 33



Figure 1. Map of Shenandoah National Park, VA, U.S.A. showing the three major bedrock

classes and the location of catchments drained by the five study streams...............61

Figure 2. Correlations between stream pH and a) fungal biomass and b) ecosystem

respiration................................................................................................................ 62

Figure 3. Correlations between stream pH and a) whole-stream qCO2 and b) qN................ 63

Figure 4. Metabolic quotients (qCO2) calculated from three separate studies of nutrient

enrichment effects on respiration and fungal biomass in stream ecosystems......... 64

Figure 5. Relationship between qCO2 and qN in nine North American streams.................. 66



Figure 1. Measurements of a) stream pH, and concentrations (± SE) of b) NH4-N, c) NO3-N,

d) soluble reactive phosphorus (SRP) and e) dissolved organic carbon (DOC) in

SNP streams over the spring study period.............................................................. 99

Figure 2. Correlations of the charge sum of base cations (microequivalents per liter) with a)

chlorophyll a, b) soluble reactive phosphorus (SRP) and c) dissolved organic

carbon (DOC) in SNP streams.............................................................................. 100

Figure 3. Standing crops of a) chlorophyll a and b) epilithic biomass, and c) autotrophic

index values in the study streams over the spring study period............................ 101

Figure 4. a) Uptake lengths (Sw), b) uptake velocities (vf) and c) areal uptake rates (U) for

NO3-N in Meadow Run (pH 5.5) and Piney River (pH 6.9) under ambient (open

bars) and DOC-enriched (solid bars) conditions................................................... 102

1

Chapter 1: General Introduction

Earth systems are currently challenged by unprecedented rates of change in myriad

environmental conditions, including changes in climate, widespread deforestation, water

abstraction, increased nutrient loading, and acidification of both marine and freshwater

ecosystems. Such shifts in the natural template for biological organization pose unique and

critical challenges to ecosystems science due to the large-scale and pervasive nature of these

perturbations. Understanding the flow of energy through ecosystems and the biogeochemical

processes that accompany energy transfers provides the fundamental basis for prediction of

ecosystem response and sustainability under strong anthropogenic pressure.

Acid deposition is a large-scale, anthropogenically-caused alteration of the chemical and

biological makeup of atmospheric, terrestrial, and aquatic systems. Since the advent of the

industrial revolution, increased emissions of acid-forming SO2 and NOx gases have altered

atmospheric chemistry, lowered the pH of precipitation, and increased the dry deposition of

acidifying particulates. The deleterious effects on high-elevation forests of the northeastern

United States and Europe raised widespread awareness of ‗acid rain‘ and led to legislation

curtailing emissions. These efforts effectively halted the decline in precipitation pH and some

terrestrial ecosystems have shown signs of recovery. However, few surface waters in

historically-affected regions have shown any reversal in acid-status due to the substantial

declines in buffering capacities of acid-impaired soils. Thus, excluding drastic remediation

efforts such as liming, acidified surface waters are expected to remain acidic for some time.

Extended declines in pH are cause for concern because the acid-base status of environmental

conditions governs the physiological processes that organize ecosystem function.

2

The majority of research efforts in acidified lakes and streams have focused on changes in

community composition with an overwhelming emphasis on fish assemblages and, to a lesser

extent, macroinvertebrates. While these investigations have provided valuable knowledge

concerning the ecology of higher trophic levels, and brought considerable public and political

attention to surface water acidification, they provide little information on the role of acidification

in ecosystem function. More specifically, a significant gap exists in our understanding of

acidification influences on microbial-mediated rates of carbon (C) flux and nutrient cycling in

freshwater ecosystems. Advancement of our understanding of these processes is especially

urgent in lotic systems, which connect geographically distant regions through the advective

transport of materials in running waters.

Coincident with watershed acidification is the increase in deposition of nitrogen (N), which

enters soils as nitric acid (HNO3) and quickly dissociates into its ionic form, nitrate (NO3-). Over

the last century, human activities have more than doubled the amount of available N (as both

NH4 and NO3) across the landscape via multiple pathways in addition to atmospheric deposition,

pushing ecosystems toward ‗N saturation‘ – where N supplies are in excess of biological

demand. The potential for downstream eutrophication and, consequently, degradation has

brought investigation of controls over watershed export of N to the forefront of research efforts

in lotic biogeochemistry. Headwater streams play a critical role in landscapes as the final point to

remove harmful N pollutants before export from the watershed. Indeed, these systems have

displayed strong potential for N removal through microbial-driven processes of assimilatory

uptake and denitrification.

Both autotrophic (algae + diatoms) and heterotrophic (bacteria + fungi) microbes are

responsible for nutrient uptake in stream ecosystems and many studies have found opposite

3

effects of acidification on the structure and function of these organisms. Autotrophic biomass

and production often increases with increasing acidity through releases from competition or

grazing pressure, while heterotrophic biomass often declines. The deleterious effects on

heterotrophs in acidified headwater streams is especially concerning as these systems are

overwhelmingly detritus-based, i.e., heterotrophic microbes play key roles in the decomposition

of organic matter and the transfer of energy to higher trophic levels. These roles are pronounced

during autumn, when forested temperate streams receive large pulses of detrital resources in the

form of fallen leaves. The respiratory activity of decomposers results in strong demand for

inorganic N. Thus, acidification is predicted to influence the cycling of both C and N, warranting

the simultaneous consideration of C and N dynamics in the assessment of acidification effects on

lotic biogeochemistry.

Here, I use the concept of ecosystem stress to address the role of acidification in stream

metabolism and N cycling. Stress is conceptualized as a decrease in metabolic efficiency due to

the increased energy requirements for survival and maintenance of biological compartments, and

the consequent reduction in the allocation of energy towards growth and reproduction. Stress

impairments are recognized by an increase in the metabolic quotient qCO2 (i.e., biomass-specific

respiration) and remains little used outside of the soil sciences. In chapter 2, I investigate the

respiratory and N uptake activities of heterotrophic leaf biofilms from streams of differing pH

during autumn. In experimental microcosms, I show how rising qCO2 reflects the gradient of

increasing acidity and is positively related to biomass-specific N uptake (i.e., qN), and discuss

the implications for whole-stream metabolism and N uptake. In chapter 3, I apply qCO2 and qN

at the ecosystem level by quantifying standing stocks of microbial biomass and rates of

metabolism and N uptake across a gradient of stream pH during autumn. I show how results at

4

the ecosystem level compare to microcosm experiments, and use results from other published

studies to display the utility of qCO2 as an indicator of ecosystem-level stress and its relationship

with qN across multiple systems. Chapter 4 addresses how stream acidity influences metabolism

and N uptake during increased autotrophic presence. I show how greater algal biomass in a

highly acidic stream influences its capacity to remove inorganic N and dissolved organic carbon

(DOC) relative to a circumneutral stream, and how additions of highly-labile DOC differentially

affect N uptake in streams of differing acid status.

5


the ecosystem scale

Damon T. Ely, Daniel von Schiller, and H. Maurice Valett

Used with permission of Freshwater Biology 55: 1337-1348.

Abstract

While anthropogenic stream acidification is known to lower species diversity and impair

decomposition, its effects on nutrient cycling remain unclear. The influence of acid-stress on

microbial physiology can have implications for carbon (C) and nitrogen (N) cycles, linking

environmental conditions to ecosystem processes. We collected leaf biofilms from streams

spanning a gradient of pH (5.1 – 6.7), related to chronic acidification, to investigate the

relationship between qCO2 (biomass-specific respiration; mg CO2-C g-1

fungal C h-1

), a known

indicator of stress, and biomass-specific N uptake (μg NH4-N mg-1

fungal biomass h-1

) at two

levels of N availability (25 and 100 μg NH4-N L-1

) in experimental microcosms.

Strong patterns of increasing qCO2 (i.e. increasing stress) and increasing microbial N uptake

were observed with a decrease in ambient (i.e. chronic) stream pH at both levels of N

availability. However, fungal biomass was lower on leaves from more acidic streams, resulting

in lower overall respiration and N uptake when rates were standardized by leaf biomass. Results

suggest that chronic acidification decreases fungal metabolic efficiency because, under acid

conditions, these organisms allocate more resources to maintenance and survival and increase

their removal of N, possibly via increased exoenzyme production. At the same time, greater N

availability enhanced N uptake without influencing CO2 production, implying increased growth

efficiency.

At the ecosystem level, reductions in growth due to chronic acidification reduce microbial

biomass and may impair decomposition and N uptake; however, in systems where N is initially

6

scarce, increased N availability may alleviate these effects. Ecosystem response to chronic

stressors may be better understood by a greater focus on microbial physiology, coupled

elemental cycling, and responses across several scales of investigation.

Introduction

Despite long-term efforts to curb emissions of atmospheric pollutants, acid deposition continues

to influence ecosystems in industrialized regions of the world (Driscoll et al., 2001; Sanderson et

al., 2006). Freshwaters are particularly sensitive to the suite of geochemical changes associated

with acidification, which can lower species diversity in both lakes and streams (Schindler et al.,

1985; Elwood & Mulholland, 1989; Hesthagen et al., 1999; Lovett et al., 2009). In this regard,

stream ecosystems have received less attention than lakes; furthermore, the influence of

acidification on ecosystem processes, such as metabolism and nutrient cycling, are less well

known than are implications for species composition and abundance.

The impairment of litter decomposition is a stream ecosystem process that has been examined

in the context of anthropogenic acidification (Chamier, 1987; Mulholland et al., 1987; Dangles et

al., 2004; Riipinen, Davy-Bowker & Dobson, 2009; Simon, Simon & Benfield, 2009). Aquatic

hyphomycetes (i.e. microfungi), which may constitute 98% of microbial decomposer (i.e. fungi +

bacteria) biomass on leaf litter (Gulis & Suberkropp, 2003), have much (up to 100 times) greater

production rates than bacteria (Findlay & Arsuffi, 1989; Newell et al., 1995; Suberkropp &

Weyers, 1996), play a key role in the breakdown of leaf detritus (Gessner & Chauvet, 1994), and

respond negatively to stream acidification (Chamier, 1987; Chamier & Tipping, 1997). Indeed,

reduced leaf decomposition rates are often attributed to microbial impairment (Mulholland et al.,

1987; Simon et al., 2009), possibly due to low pH inhibition of key leaf-degrading exoenzymes

7

such as pectin lyases (Jenkins & Suberkropp, 1995) and the interference of metals like

aluminium (Al) which adsorb to hyphal surfaces or accumulate in fungal cells (Chamier &

Tipping, 1997). The effect that these acid-induced influences have on nutrient cycling is less

well-known but should result in a reduced rate of nutrient uptake from the water column, because

reduced carbon (C) processing by heterotrophic microbes is linked stoichiometrically to reduced

nutrient demand (Hall & Tank, 2003).

The deposition of atmospheric nitrogen (N) also leads to N saturation of the catchment and

thus greater availability of N to freshwater systems (Aber, Nadelhoffer & Steudler, 1989;

Driscoll et al., 2003; Fowler et al., 2005). The uptake of inorganic N from surface waters is an

ecosystem service that is microbially-driven (Peterson et al., 2001; Bernot & Dodds, 2005), and

generally increases with N concentration when N is scarce and limits biotic processing (Bernot &

Dodds 2005). The influence of acidification on stream N cycling at both natural and, as

catchments approach N saturation, raised N concentrations is unknown. Because microbial

assemblages have large influences on energy flow and material cycling, understanding the

responses of these organisms to change in environmental conditions is necessary for interpreting

ecosystem-level patterns (Schimel, Balser & Wallenstein, 2007).

Microbes, stress and links between metabolism and N cycling

The influence of acidification on ecosystem processes may best be addressed by considering the

metabolic response of microbial assemblages in the context of physiological stress. Stress is

realized at both organismal and ecosystem levels as a sustained reduction in metabolic

efficiency, resulting from a diversion of energy from growth and reproduction toward

maintenance and survival (Odum, 1985). Recently, Schimel et al. (2007) recommended that

8

ecosystem ecologists place greater emphasis on microbial physiological responses to stress,

because shifts in resource allocation can have influences on ecosystem-level C, energy and

nutrient dynamics.

In this context, biomass-specific respiration rate can be a useful indicator of stress for

ecosystem-level investigations. Referred to as the metabolic quotient (qCO2, Anderson &

Domsch, 1993), this index of metabolic efficiency is derived from Odum‘s theory of ecosystem

succession (Odum, 1969, 1985) and reflects large-scale shifts in carbon use. Greater values

indicate increased energy expenditure for survival at the expense of growth. In forest soils, qCO2

may change with both stress and disturbance (Wardle & Ghani, 1995), where the latter refers to

unpredictable, stochastic events often resulting in mass mortality and habitat alteration (Resh et

al., 1988). Accordingly, in the absence of punctuated disturbance, qCO2 increases with stressful

conditions (e.g., historic metal contamination, low soil pH; Wardle & Ghani, 1995) and is valued

for its relative ease of measurement.

To address how chronic exposure to stream acidification may alter critical ecosystem

processes and how N subsidies may alleviate or exacerbate these influences, we used microcosm

assays to investigate the response of fungi in leaf biofilms to chronic stream pH and acute N

addition, and to build hypotheses about ecosystem-level N cycling (Benton et al., 2007). We

collected leaves from a series of streams where ambient pH varied as a result of acid deposition

and catchment lithology to assess experimentally acid stress and N subsidy. Given the proposed

linkages between metabolism and nutrient cycling described above, biofilms from more acidic

streams should display higher qCO2 values and lower biomass-specific N uptake rates, reflecting

a stress-induced reduction in the capacity to process inorganic N.

9

Methods

Study Sites

Streams were located within Shenandoah National Park (SNP), Virginia, USA, which receives

high acid deposition due to its position along the Blue Ridge Mountains downwind of the coal-

fired power plants of the Ohio River Valley (Driscoll et al., 2001). Within SNP, variation in

underlying lithology produces differing buffering capacities of soils and streams that generate a

gradient in stream pH. Over a period of 13 years, hydrogen ion concentrations in SNP streams

have remained essentially unchanged (Webb et al., 2004). Furthermore, monthly stream pH data

collected in a subset of these and other western Virginia streams indicate minimal within-year

variation (Shenandoah Watershed Study, http://swas.evsc.virginia.edu). We selected five second

to fourth order streams distributed across three bedrock classes (in order of decreasing acid

sensitivity): siliceous (quartzite and sandstone), felsic (granitic), and mafic (basaltic; Table 1).

The median pH of SNP streams rises from 5 to 7 across this gradient (Cosby et al., 2006). The

catchment of one stream, Madison Run, has both siliceous (58%) and mafic (42%) bedrock types

(James R. Webb, University of Virginia, personal communication). These streams drain heavily

forested catchments undisturbed by anthropogenic influences since creation of the park in 1936.

Leaf collection, stream sampling and chemical analyses

Large quantities (~ 500 g ash-free dry mass, AFDM) of submerged chestnut oak (Quercus prinus

L.) leaves were collected from all streams on 17 December 2006 and kept in stream water in the

dark near ambient temperatures (7°C) until the end of all experiments. Because we did not

control for leaf residence time (i.e. time within the stream), we collected large composite

samples on a single date approximately 7-8 weeks after most leaf-fall had occurred to minimize

http://swas.evsc.virginia.edu/

10

between-site differences in mean colonization time. Immediately prior to leaf collections we

measured stream pH, temperature and specific conductivity, and collected samples for the

analysis of water chemistry. Stream pH was measured with a Thermo Orion model 290A

portable meter (Orion Research Inc., Beverly, Massachusetts, USA) following calibration in the

field. Temperature and conductivity were measured using a YSI model 30 hand-held probe (YSI,

Yellow Springs, Ohio, USA). Water samples (n = 3) from each stream were filtered (Whatman

GF/F, pore size = 0.7 μm) into acid-washed 125-mL polyethylene bottles and frozen until

analysis within five days of collection. Unfiltered samples (n = 3 per stream) were collected in

acid-washed 1-L polyethylene bottles and kept at 7 °C until measurements of acid-neutralizing

capacity (ANC) were made (within 48 hours) using Gran titration (Gran 1952) to an endpoint of

pH 3.5.

Base cations (Ca2+

, Mg2+

, Na+ and K

+) in stream water were measured using a Dionex 500

Ion Chromatograph (Dionex Corporation, Sunnyvale, CA, USA) using an Ionpac CG12a cation

column following standard methods (APHA, 1998). Concentrations of nitrite-N and nitrate-N

(presented as NO3-N) and NH4-N were determined colorimetrically following the acidic diazo

method after cadmium reduction (Wood, Armstrong & Richards, 1967) and a modified form of

the automated phenate method (USEPA, 1997), respectively, using a Lachat QuikChem 8500

auto analyser (Lachat Instruments, Loveland, Colorado, USA). Soluble reactive phosphorus

(SRP) was used to represent orthophosphate (PO4-P) as determined by the molybdate-antimony

method (Murphy & Riley, 1962). Total dissolved Al was measured using inductively coupled

plasma spectrometry (Spectro CirOS VISION, Spectro Analytical Instruments Inc., Mahwah,

New Jersey, USA). Dissolved organic carbon was determined by persulphate digestion (Menzel

& Vaccaro, 1964) on a model OI Analytical 1010 Total Organic Carbon Analyser

11

(Oceanographic International, College Station, Texas, USA). Percent C and N content of leaves

from each stream (n = 3 per stream) was measured by a Flash EA 1112 elemental analyser

(Thermo Fisher Scientific, Delft, the Netherlands).

Laboratory microcosm experiments

A stock solution of ―stream water‖ was prepared from reagent-grade water with 12 mg L-1

NaHCO3, 7.5 mg L-1

CaSO4•2H2O, 7.5 mg L-1

MgSO4, and 0.5 mg L-1

KCl following APHA

(1998) and kept at 7°C. To evaluate possible influences of acute pH in the test solutions and

acute exposure to Al, two-thirds of the stock solution (pH 6.75) was titrated to pH 4.5 with 1N

H2SO4, and Al was added as AlCl3 to half of the pH 4.5 solution for a final concentration of 15

μM Al. Thus, three levels of a treatment factor labelled ―acute pH‖ were created (pH 6.75, pH

4.5 and pH 4.5 + 15 μM Al). The latter two treatments reflect the temporary conditions

associated with high discharge-related episodic acidification events observed in SNP streams

(Bulger et al., 1995; Cosby et al., 2006). We found no influence of experimentally-altered acute

pH or Al on either microbial qCO2 or N uptake rates by biofilms from any stream. Subsequently,

we pooled all data across pH treatments and eliminated ―acute pH‖ as a factor. Each morning for

five consecutive days, two test solutions of different N concentration were created for each

stream for a two-factor block design: factor 1 – Stream (five), factor 2 – N availability (25 μg

NH4-N L-1

[hereafter, ―low N‖] and 100 μg NH4-N L-1

[hereafter, ―high N‖]). To control for

experiment duration we used Day (levels = 1 – 5) as a blocking factor. Nitrogen was added as

NH4Cl. The experimental N concentrations were chosen to represent a large (four-fold)

difference in N availability between annual high (150-250 μg L-1

) and low (10 μg L-1

)

concentrations in summer and winter, respectively, for these streams (D.T. Ely, personal

12

observations). Phosphorus (10 μg L-1

) was added as Na2HPO4 to all treatment solutions to

eliminate P-limitation.

For each stream on each day, thirty leaf discs (1.5-cm diameter) were randomly chosen from

~100 freshly cut discs and five leaf discs were added to each of six 50-mL centrifuge tubes

completely filled (i.e. no headspace) with the appropriate treatment solution (three each for low

and high N solutions) and placed in the dark for 24 h at 7°C with mild shaking (100 rpm) to

ensure contact of all leaf surfaces with the treatment solution. Five blanks (solution only) were

conducted simultaneously for each treatment on each day. This design produced fifteen

replicates for each level of Stream and N availability. After 24 h, the concentration of dissolved

oxygen (DO) in each tube was immediately measured (YSI model 55 handheld meter, Yellow

Springs, Ohio, USA) following the removal of 10 mL of solution for NH4-N and NO3-N

analyses. To minimize the exposure of treatment solutions to air, we kept the time between cap

removal and DO measurement in each tube to < 10 s. Leaf discs from each microcosm were then

dried at 55°C (24 hrs), weighed, ashed at 550°C (24 hrs), and reweighed to obtain AFDM.

Fungal biomass

Five leaf discs (~40 – 60 mg AFDM; n = 5 per stream) cut from the leaves collected at each

stream were kept in 5 mL methanol (HPLC-grade) and frozen at -20°C until measurement of

ergosterol content to determine fungal biomass (Gessner & Chauvet, 1993). Ergosterol was

extracted by heating in methanol for 2 h at 65°C, purified by liquid-liquid extraction, and

quantified by high pressure liquid chromatography following the methods of Tank et al. (1998).

The mobile phase was 2 mL min-1

methanol and the stationary phase was a Nova-Pak ODS C18

reverse phase column; UV absorbance detection was set at 282 nm (Dionex DX-500, Dionex

13

Corp., Sunnyvale, CA, USA). Ergosterol mass was multiplied by a conversion factor of 182 to

estimate fungal biomass (Gessner & Chauvet, 1993) and reported as mg fungal biomass per g

leaf AFDM.

Metabolic response

We calculated CO2-C produced in each experimental unit as:

CR = t

CFRQVDODO trtctrl (1)

where CR is respiratory carbon production (mg C h-1

), DOctrl and DOtrt are the post-incubation

dissolved oxygen concentrations (moles L-1

) of the controls (n = 5) and each treatment replicate,

respectively, V is the solution volume, RQ is the respiratory quotient (0.85 moles CO2 produced

per mole O2 respired; Bott, 2006), CF is a correction factor for C mass per mole of CO2 (12011

mg C mol-1

CO2), and t is the incubation time. To quantify fungal C in each experimental unit

(assuming that fungal C was 50% of fungal biomass) stream-specific fungal biomass per unit leaf

AFDM was multiplied by leaf mass within a microcosm. The microbial metabolic quotient (i.e.

qCO2) was then calculated as the quotient of CO2-C production and fungal C and reported as mg

CO2-C g-1

fungal C h-1

. Ammonium-N uptake rates were expressed as NH4-N mass lost from

solution (mean control NH4-N – individual treatment NH4-N) per unit fungal biomass per hour.

In addition, we calculated CO2-C production and N uptake per unit leaf AFDM per hour to

compare biomass- and detrital substrate-specific rates.

Statistical analyses

Differences in stream chemistry and leaf C and N content among sites were investigated using

one-way analysis of variance (ANOVA). Two-way ANOVA (blocking for day) was used to

14

evaluate the effects of Stream and N level on qCO2 and N uptake. Data that did not meet

assumptions of normality or constant variance were log-transformed. Tukey‘s honest significant

difference (HSD) was used as a post-hoc multiple comparison test following a significant

ANOVA. We used linear regression to evaluate the relationship between ambient stream pH and

response variables, and between qCO2 and biomass-specific N uptake at either level of N

availability. Slopes and intercepts of significant regressions at both low and high N availability

were compared using a t-test for parallelism (Kleinbaum & Kupper, 1978). ANOVAs and post-

hoc tests were performed using SAS 9.1.3 (SAS Institute Inc., Cary, NC); linear regressions were

performed using SigmaStat 3.11.0 (Systat Software Inc., San Jose, CA).

Results

Stream chemistry

Water temperature (5.6 – 9.0 ºC) and specific conductivity (20.3 – 31.1 μS cm-1

) were similar

among streams while pH (5.1 – 6.7) and ANC (-10.6 – 127.2 μeq L-1

) increased across

siliceous<felsic<mafic catchments (Table 2). The exception was Madison Run, where pH was

higher (6.5) given a relatively low ANC (36 μeq L-1

; Table 2). We found close agreement

between our values of pH and long-term (12-24 y) averages for Meadow Run, Paine Run, and

Madison Run (James N. Galloway, University of Virginia, personal communication). Total Al

concentrations were below detection (< 0.003 mg L-1

) for all streams. Calcium (0.7 – 2.8 mg L-1

)

and the charge sum of base cations (106 – 252 μeq L-1

) consistently increased with ANC and pH,

again with the exception of Madison Run (Table 2). Low electrical conductivity was reflected in

low concentrations of other base cations (e.g., Mg2+

; K+; and Na

+), which differed among

streams (ANOVA, P < 0.05) but displayed no trends with ambient pH (Table 2). Stream

15

concentrations of NH4-N and SRP were low (< 5 μg L-1

) and did not differ among streams

(ANOVA, P > 0.05; Table 2). Nitrate concentrations were also low (< 5 μg L-1

) in all but the

most circumneutral stream (Rose River), where NO3-N was 37.8 μg L-1

(Table 2).

Concentrations of DOC varied from 0.8 – 1.4 mg L-1

and did not differ among streams (ANOVA,

P > 0.05).

Leaf and biofilm characteristics

Leaf N content differed by 0.5% among streams with the highest value found in the most pH

neutral stream (i.e., Rose River = pH 6.7) and the lowest value in a stream of intermediate pH

along the gradient (i.e., Lower Hazel River = pH 6.3; Table 2). Molar C:N of leaves was lowest

in Rose River (55.3) and highest in Lower Hazel (98.2). While both %N and molar C:N varied

among streams (ANOVA, P < 0.05; Table 2), differences were fairly low and mean values were

not correlated with stream pH. Fungal biomass on leaves increased along the gradient of

increasing pH, with the most circumneutral stream (39.0 mg g-1

AFDM) supporting 8.6 times

more fungal biomass than the most acidic (4.5 mg g-1

AFDM; ANOVA, P < 0.05; Fig. 1a).

Microcosm experiments

Microbial respiratory quotient (qCO2) was strongly influenced by stream of origin (two-way

ANOVA, Stream main effect P < 0.001). Microbial qCO2 declined with increasing pH of the

stream of origin at both low and high N availability (Fig. 1b) and response to pH change was

similar between N treatments (t test of slopes, P = 0.17). When the results from each N level

were combined, the relationship remained significant (n = 10, r2 = 0.75, P = 0.001). In contrast,

N availability had little effect on qCO2 with a significantly higher value found for Meadow Run

16

only (12.4 ± 0.6 vs. 18.2 ± 0.8 mg CO2-C g-1

fungal C h-1

) which resulted in a significant Stream

X N availability interaction (P < 0.001).

When measures of respiratory activity included variation in fungal standing stock (i.e. leaf

respiration normalized for leaf mass, mg CO2-C g-1

leaf AFDM h-1

), the rate increased with

stream pH in the low N treatment but showed no relationship at higher N availability (Fig. 1c).

When results from N treatments were combined, leaf biomass-specific respiration declined with

increasing acidity of the stream of origin (n = 10, r2 = 0.51, P = 0.02).

Biomass-specific rate of N uptake by fungi was closely related to stream identity (two-way

ANOVA, Stream main effect, P < 0.001). Nitrogen uptake by leaf biofilms declined with

increasing pH of the stream of origin both in low (P = 0.008) and high (P = 0.025) N treatments

(Fig. 2a). Unlike qCO2, which responded little to N manipulation, biomass-specific N uptake

increased with greater N availability (two-way ANOVA, N level main effect, P < 0.001; Fig. 2a).

Further, the increase in N uptake differed among streams (two-way ANOVA, Stream X N

interaction, P = 0.01), with biofilms from more acidic streams exhibiting greater positive

responses to increased N availability than those from more circumneutral streams (t test of

slopes, P = 0.024; Fig. 2a). On the other hand, mean N uptake normalized to leaf AFDM (μg N

g-1

leaf AFDM h-1

) increased with stream pH at low N availability, but this was only marginally

significant (P = 0.056; Fig. 2b). Leaf mass-specific N uptake was not related to stream pH at

high N availability (Fig. 2b) or when low and high N treatments were combined.

Biomass-specific N uptake was tightly and positively linked to microbial qCO2 in both low

and high N treatments (Fig. 3). Further, study streams occupied discrete regions of these trends,

and were organized by ambient pH. The proportional increase in N uptake per unit change in

qCO2 remained consistent between levels of N availability (t test of slopes, P = 0.355). At the

17

same time, greater absolute rates of N uptake occurred with higher N availability at similar qCO2

(t test of intercepts, P = 0.003).

Discussion

The geochemical differences we observed among the five study streams are consistent with

chronic stream acidification (Chamier, 1987; Likens, Driscoll & Buso, 1996; Cosby et al., 2006)

and represent environmental conditions that typically result in the physiological impairment of

aquatic organisms (Schindler et al., 1985; Bulger et al., 1995; Ormerod & Durance, 2009). With

decreases in ambient stream pH, we found reduced microbial biomass and large increases in

qCO2 similar to trends found in terrestrial soils (Anderson & Domsch, 1993). However, in

contrast to our proposed link between metabolism and N cycling, microbial biomass-specific N

uptake increased with increasing acidity. Further, increased N uptake was strongly and positively

related to qCO2 with streams occupying consistent positions along these functional axes. As

such, fungi in biofilms in acidic streams were least abundant and most metabolically inefficient

(i.e., highest qCO2), but had the highest rates of N uptake.

Influence of acidification: stress effects on ecosystem function

High concentrations of protons and inorganic Al are known to interfere with the enzyme activity,

reproduction, and other physiological processes of aquatic microfungi (Jenkins & Suberkropp,

1995; Chamier & Tipping, 1997). We were surprised to find no differences in qCO2 or N uptake

due to our acute pH and Al treatments despite the strong trends in these response variables

associated with pH of the stream of origin. We do not know if there was an interaction between

the added materials and our synthetic water solution, but assuming the intended conditions were

18

met, the response times of these organisms may have been longer than our incubation times. This

seems unlikely, however, as differences in respiration rates were observed over 6-10 hours

between leaf discs incubated in ambient circumneutral (pH 6.8) stream water and those incubated

in water from an acidic (pH 4.9) stream (Simon et al., 2009).

Another explanation may be that these microbial assemblages acclimate rapidly and persist

functionally throughout short-term exposures to low pH and high Al typical of episodic

acidification events associated with storms. During these events, which last 1-7 days, strong acid

anions are mobilized from upland soils and can substantially alter stream pH and Al

concentrations (Deviney et al., 2006; Kowalik & Ormerod, 2006). For example, total monomeric

Al in SNP streams is generally scarce during baseflow but may exceed 100 μg L-1

during high

discharge storm events (Bulger et al. 1995). These events occur regularly in SNP; between

November 1992 and July 1994, Hyer, Webb & Eshleman (1995) recorded twenty-five storm

events sufficient to initiate 41 acidic episodes in three SNP streams. Our 24-hour acute exposures

to experimentally-altered pH and [Al] conditions may have simulated episodic acidification

events which these biofilms are able to resist.

Typically, N uptake by stream biota is associated with biotic immobilization and growth.

However, understanding the fate of N uptake requires consideration of associated C dynamics

and the processes of C sequestration. Microbial acquisition of C relies on the catalytic activities

of exoenzymes (among the highest-priority molecules facilitating heterotrophic microbial C

flow) and both C and N are required for their production (Schimel & Weintraub, 2003). In acidic

streams, enzymes with low pH optima (pH 4.0 to 6.5), such as hydrolytic polysaccharidases, are

primarily responsible for leaf softening and maceration (Jenkins & Suberkropp, 1995). In

19

addition to exoenzyme production, other proximate fates for C within microorganisms are

growth (also requiring N) and respiration (Schimel et al., 2007).

If we assume C acquisition compensates for respiratory C losses, as required for steady-state

biomass, increased qCO2 requires greater C and N allocation toward exoenzyme production,

increasing N demand. Alternatively, if external C acquisition is decoupled from respiration no

change in C or N allocation toward exoenzymes should accompany increasing qCO2. Under the

latter circumstances, enhanced qCO2 would necessarily consume biomass C, require reductions

in C and N allocation toward growth, and result in lower N demand. Use of biomass C to

compensate for enhanced qCO2 is not sustainable and we observed enhanced N uptake associated

with increased qCO2. Thus, we offer the hypothesis that the positive relationship between

biomass-specific N uptake and qCO2, as observed across the pH gradient, reflects a stress-

induced physiological response because acidification results in increased C and N allocation

toward exoenzyme production, due to heightened resource requirements for processes related to

maintenance and survival.

The activities of certain exoenzymes at low pH may increase (Griffith et al., 1995) or

decrease (Jenkins & Suberkropp, 1995) in accordance with pH optima. However, Simon et al.

(2009) found greater phosphatase activity in biofilms from acid streams independent of the direct

effects of pH on enzyme activity, implying greater production of this enzyme. Resource

reallocation towards exoenzymes probably includes declines in the production rates of other

proteins, lipids, cell walls and reproductive structures in comparison to unstressed organisms

(Odum, 1985; Schimel et al., 2007). This type of metabolic compensation is consistent with the

observed decline in microbial abundance commensurate with increased acidity.

20

The potential reductions in growth and reproductive success associated with lower microbial

biomass on leaves from more chronically acidic streams have implications for ecosystem-level

processes (Fig. 4). We standardized qCO2 and rates of N uptake by leaf AFDM to address

ecosystem-scale detrital pools and include the influence of variation in microbial biomass among

streams. Respiration normalized by leaf mass generally declined with increasing acidity (i.e.

lower abundance of microbial flora per unit substrate), which has been observed elsewhere (e.g.

Groom & Hildrew, 1989) and is consistent with slower decomposition rates found in acidified

streams (Hildrew et al., 1984; Chamier, 1987; Dangles et al., 2004; Riipinen et al., 2009; Simon

et al., 2009). Similarly, leaf mass-normalized N uptake declined with increasing acidity, but only

at low N availability. Thus, the fates of C and N in acidified streams may be altered at both the

organismal- (increased CO2-C losses, greater N uptake) and ecosystem- (slower decomposition

and greater proportions of N uptake entering the exoenzyme pool) levels of organization, due to

stress influences on microbial physiology and abundance.

Influence of N availability: subsidy vs. stress

At the ecosystem level, enhanced uptake of inorganic N with increasing N concentration is

frequently observed in streams with low N, as a result of both mass-transfer into biofilms and

biotic demand (Dodds et al., 2002). Similarly, in our microcosm experiments, increased

availability of labile inorganic N was a subsidy to stream microbes that enhanced biomass-

specific rates of N uptake across all streams.

Because heterotrophic micro-consumers are not believed to be capable of luxury nutrient

uptake (i.e., they maintain a constant C:N), N uptake is linked to the manufacture of N-

containing organic molecules (Sterner & Elser, 2002). Increased N availability had almost no

21

effect on qCO2 and, considering C and N allocation pathways, we may interpret greater N uptake

without changes in CO2 production as subsidizing a shift in resource allocation towards growth

(Fisk & Fahey, 2001). In this case, sequestered N is probably incorporated into a pool of organic

molecules with a longer residence time (e.g. associated with structural biomass) as opposed to

exoenzymes, since respiratory C loss remained unchanged and need not be compensated for by

exoenzyme production. Indeed, in similar microcosm experiments, Gulis & Suberkropp (2003)

found a 72% increase in microbial production efficiency with greater N availability. Thus, for the

microbial assemblages that constitute leaf biofilms, increases in N availability appear to

positively influence both N uptake from stream water and the efficiency of N-use across the

acid-stress gradient.

The influence of increased N availability may be greater in more stressed systems that face

future increases in N loading, as suggested by the much larger increases in biomass-specific N

uptake we observed for acidified, compared to circumneutral, streams when presented with an N

subsidy. Under high N availability, N uptake normalized to leaf mass did not decline with

increasing acidity because disproportionately higher biomass-specific N uptake rates

compensated for overall lower fungal biomass in acid streams. For example, fungal biomass was

8.6 times greater in Rose River (pH 6.7) than in Meadow Run (pH 5.1) but biomass-specific N

uptake in Meadow Run was approximately 8.7 times greater under high N availability than in

Rose River. At low N availability, biomass-specific N uptake in Meadow Run was only five

times greater than in Rose River and failed to counteract the biomass deficit.

These data suggest that ecosystem response to acidification may be mediated by increased N

loading; depending on the magnitude of exogenous N supply, an increase in microbial N uptake

due to both acid-stress and N subsidy may (or may not) counterbalance a decline in microbial

22

abundance. The result may be similar ecosystem-level rates of N uptake between circumneutral

and acidified streams under high N supply, whereas N-subsidized increases in N-use efficiency

may lead to greater incorporation of N into biomass compartments (Fig. 4). Net effects in

environments with lower pH minima and greater N availability may differ from those found here

as physiological domains and uptake kinetics respond to subsidy and stress magnitudes (Kemp &

Dodds, 2002; Earl, Valett & Webster, 2006; O‘Brien et al., 2007). Consideration of longer time

scales is also important; greater production of more labile, N-containing exoenzymes with

increased stress could result in high turnover and the rapid return of inorganic N to the water

column when these enzymes degrade.

Odum, Finn & Franz (1979) recognized the interplay of various levels of organization in

determining ecosystem response to subsidy and stress; apparent subsidies at one level may be

overwhelmed by stress effects at another. Higher maintenance costs at the organismal level may

explain lower fungal biomass in detrital pools from acid-stressed streams because less energy is

routed to growth and reproduction. Thus, at a larger scale, lowered microbial abundance may

result in the net decrease of stream C and N processing, despite the positive response to N

subsidies observed for these organisms. Ecosystem-scale assessments will be required to address

how N uptake and respiration processes in acidified streams compare to reference conditions

with future enhanced N loading to streams. In addition, measurements of microbial growth rates

and exoenzyme production, as well as consideration of bacterial metabolic dynamics, are

primary research needs in future investigations of acidification effects on stream ecosystem

functions of C and N processing.

23

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28

Table 1 Geographical and geological characteristics of study streams in Shenandoah National Park, VA, USA.

Stream

Characteristics Meadow Run Paine Run Lower Hazel Madison Run Rose River

Location 38°09‘30‖N

78°48‘36‖W

38°11‘56‖N

78°47‘39‖W

38°36‘53‖N

78°15‘24‖W

38°15‘27‖N

78°46‘18‖W

38°30‘56‖N

78°21‘58‖W

Altitude (m) 483 432 303 433 335

Bedrock

Geology

Siliceous Siliceous Felsic Siliceous/Mafic Mafic

Catchment

area (ha)

900 1240 1320 2110 2370

Stream Order 2 2 3 4 3

29

Table 2 Chemical parameters and leaf characteristics of the study streams. Data are means (n = 3) from samples collected on 17 December 2006.

Means within a row that share the same superscript are not significantly different (ANOVA, Tukey‘s HSD, α = 0.05). ANC = acid neutralizing

capacity. SRP = soluble reactive phosphorus.

Parameter Meadow Run Paine Run Lower Hazel Madison Run Rose River

Temperature (°C) 6.4 5.6 9.0 6.1 7.5

Specific Conductivity (μS cm-1

) 20 23 31 24 31

pH 5.1 5.7 6.3 6.5 6.7

ANC (μeq·L-1

) -11a -2

a 93

b 36

c 127

d

Ca2+

(mg·L-1

) 0.7a 0.8

a 1.6

b 1.1

a 2.8

c

Mg2+

(mg·L-1

) 0.3a 0.4

b 0.3

a 0.5

c 0.6

d

K+ (mg·L

-1) 1.0

a 1.5

b 0.4

c 1.1

d 0.2

e

Na+ (mg·L

-1) 0.5

ac 0.5

a 1.4

b 0.6

c 1.3

d

Base Cations (μeq·L-1

) 106a 133

ac 179

b 149

c 252

d

NH4-N (μg·L-1

) 1.9a 1.7

a 4.5

a 1.5

a 2.7

a

NO3-N (μg·L-1

) 1.1a 1.3

a 1.1

a 1.3

a 37.8

b

SRP (μg·L-1

) 1.7a 1.5

a 2.9

a 3.6

a 3.2

a

DOC (mg·L-1

) 1.2a 1.4

a 1.3

a 1.1

a 0.8

a

leaf %C 53.6a 55.7

ab 56.0

b 54.4

ab 54.6

ab

leaf %N 0.9ab

0.9ab

0.7a 0.9

b 1.2

c

leaf C:N (atomic) 71.3ac

76.7a 98.2

b 71.9

ac 55.3

c

30

Figure 1. Linear regressions of stream pH vs. a) Ln fungal biomass, b) microbial qCO2 in

microcosms with low ( 25 μg NH4-N L-1

) and high ( 100 μg NH4-N L-1

) nitrogen

(N) availabilities and c) respiration per unit leaf mass in microcosms with low and high N. For

each regression, n = 5 using mean values; plotted symbols are means ± SE with n = 5 in a) and n

= 15 in b) & c).

31

Figure 2. Linear regressions of stream pH vs. a) microbial N uptake and b) N uptake

standardized to leaf mass in microcosms with low ( 25 μg NH4-N L-1

) and high (

100 μg NH4-N L-1

) nitrogen (N) availabilities. For each regression, n = 5 using mean values;

plotted symbols are means (n = 15) ± SE. Slopes are significantly different by a t-test (P =

0.024).

32

Figure 3. Log-log relationship between microbial qCO2 and N uptake measured in laboratory

microcosms. Filled symbols with the solid regression line represent data from the low N

treatment (25 μg NH4-N L-1

: y = 0.98x – 1.97, r2

= 0.83, P < 0.001, n = 75) and open symbols

with the dashed regression line are from the high N treatment (100 μg NH4-N L-1

: y = 1.04x –

1.56, r2 = 0.91, P < 0.001, n = 75). Symbols represent different ambient stream pH as follows:

pH 6.7 (▲), pH 6.5 (■), pH 6.3 (●), pH 5.7 (▼), and pH 5.1 (♦).

33

Figure 4. Conceptual model of proposed linkage between acidification as an environmental

stressor, exogenous N as a subsidy, physiological responses of microbes influencing C and N

processing, and consequences at the ecosystem level. Italics indicate hypothetical relationships.

34



Abstract

Ecosystem responses to gradual, directional global change will differ markedly from responses

to punctuated disturbance and therefore require a distinct context for assessment. In this chapter I

use a stress concept to address the role of stream acidification in metabolism and nitrogen (N)

uptake at the ecosystem level. Whole-stream measurements of fungal standing crops,

respiration, and nitrate-N uptake (using 15

N tracers) were conducted in five streams of differing

pH (5.3 – 7.0) following autumn leaf-fall. Biomass-specific rates of respiration (i.e., the stress

metric qCO2) and N uptake (i.e., qN) were compared among streams. Both fungal biomass and

respiration declined with increasing acidity, and when combined, displayed increasing qCO2,

reflecting metabolic inefficiencies associated with environmental stress. N spiraling metrics (Sw,

vf, and U) showed no relationship to pH while qN generally increased with acidity. Though

somewhat weaker and of different magnitude, these relationships are consistent with findings in

experimental microcosms (Chapter 2) and reveal changes in both ecosystem structure and

function associated with acidification stress. I further explore the sensitivity of ecosystem qCO2

to large-scale environmental change by calculating qCO2 from three published studies of nutrient

enrichment in aquatic systems; a positive influence of stress on qN was also found using data

from nine streams of broad geographic distribution. Collectively these data revealed increasing

biological impairment with rising stream acidity, provided insight towards controls over in-

stream N cycling, and supported qCO2 as a functionally-based indicator of environmental

change.

35

Introduction

One of the current challenges facing ecosystems ecologists is prediction of how long-term

directional changes in environmental conditions (e.g., climate change, increasing nitrogen (N)

availability, and ocean acidification) will affect biogeochemical cycling. These modern scenarios

of global change differ from the abrupt occurrence of discrete disturbance events (e.g., fire,

spates), which typically induce mass mortality or severely alter habitat structure (White and

Pickett 1985). Instead, trend-based changes in certain variables, such as rising temperature and

increasing nutrient availability, are subtle but extensive alterations in environmental conditions

that present a challenge to organismal homeostasis. A better understanding of ecosystem

response to these types of global change may be gained by addressing them in the context of

chronic stress and through consideration of the physiological stress response of the biological

compartments that play key roles in ecosystem processes (Schimel et al. 2007). Investigations

addressing ecosystem response to altered environmental conditions have increasingly focused on

the physiological responses of microbial assemblages to provide insight (Buckeridge and Grogan

2008; Aanderud and Richards 2009; Muhr et al. 2009). Because microbial-mediated processes

drive ecosystem-level energy and nutrient dynamics (Falkowski et al. 2008), the functional

response of these organisms to changing conditions may have direct consequences on ecosystem

processes.

Stress is recognized at the ecosystem level as a sustained reduction in metabolic efficiency

and associated alterations of biogeochemical function (Odum 1985). These inefficiencies result

from shifts in resource allocation away from growth and reproduction and towards survival and

maintenance (Odum 1985; Schimel et al. 2007). Over time, either through physiological

adjustments within species or shifts in community structure, the efficiency of resource-use within

36

the ecosystem declines, manifesting as greater biomass-specific rates of respiration (Odum

1985). In forest soils, respiration rate standardized by microbial biomass (i.e., the microbial

metabolic quotient, qCO2) is a widely used soil quality index (Bastida et al. 2008) that increases

under stressful conditions, such as heavy metal accumulation (Brookes 1995) and enhanced soil

acidity (Anderson and Domsch 1993; Wardle and Ghani 1995). Despite its foundation in

ecosystems theory and successful application in soil science, qCO2 has not been used as a

bioenergetic indicator of stress outside of soil systems.

Acidification of surface waters through atmospheric deposition of strong acid anions (sulfate

and nitrate) is an example of a global change in many aquatic systems that began with the

industrial revolution. Legislation curbing the emissions of acid-forming pollutants has halted

further acidification in the most severely affected areas; however, streams in those regions have

shown little chemical or biological recovery (Stoddard et al. 2003; Wright et al. 2005). Much

attention has been paid to the degradation of fish populations and the overall reductions in

biodiversity associated with surface water acidification (Elwood and Mulholland 1989; Lovett et

al. 2009). Fewer studies have examined the influence of acidification on stream function, the

analysis of which has been increasingly recommended to accompany traditional, structure-based

metrics (e.g., species distribution and abundance) in the development of indicators of ecosystem

integrity (Gessner and Chauvet 2002; Fellows et al. 2006a; Young et al. 2008; Sandin and

Solomini 2009).

The impairment of litter decomposition in acidified versus circumneutral streams has been

widely observed (Chamier 1987; Dangles and Chauvet 2003; Simon et al. 2009) and is often

attributed directly to microbial impairment (Allard and Moreau 1986; Griffith and Perry 1994;

Jenkins and Suberkropp 1995). The metabolic activity of heterotrophic microbes associated with

37

decaying leaf litter in temperate forest streams drives both energy flow through lotic food webs

(Wallace et al. 1997) and the cycling of nitrogen (N) and phosphorus (P) (Mulholland 1992;

Valett et al. 2008). At the watershed scale, functional linkage between metabolism and nutrient

budgets has been a mainstay of terrestrial research (Vitousek and Reiners 1975; Hedin et al.

1995). Nevertheless, despite a rich literature addressing acidification influences on aquatic biota

relatively little work has addressed how coupling of metabolic and biogeochemical processes

responds to stress.

We previously investigated the respiratory and N uptake activities of leaf biofilms across a

gradient of stream pH in laboratory microcosm experiments addressing the relationship between

acidification and microbial response (Ely et al. 2010). In that study, we observed declines in

fungal biomass (FB) and rates of respiration and N uptake per unit leaf biomass with increasing

acidity (Ely et al. 2010). When these rates were standardized by FB, however, we found

increases in qCO2 and FB-specific N uptake with greater stream acidity. The increases in qCO2

we observed with declining stream pH supported the notion of acidification as an ecosystem

stressor, and biomass-specific increases in N uptake were hypothesized to reflect stress-induced

increases in production of N-rich exoenzymes as opposed to growth-related demands.

For the current study, we sought to perform an ecosystem-level bioassay addressing the

influence of chronic acidification on stream metabolism and N uptake. Here we focus on

ecosystem-level manifestation of stress, introduce the use of biomass-specific N uptake (i.e., qN)

as a biogeochemical stress response, and emphasize the links between stress influences on

metabolic efficiency and biogeochemical ‗purification‘ (e.g., Peterson et al. 2001; Mulholland et

al. 2008) in running water ecosystems. Understanding ecosystem-level stress responses in

streams is critical to interpreting nutrient budgets and material retention at watershed (Bernhardt

38

et al. 2003; Brookshire et al. 2007) and landscape scales (Alexander et al. 2007). Further, given

the links between carbon (C) processing and nutrient demand, acid-induced impairment of

decomposition should lower N uptake and thus the potential for self-purification in these

systems; however, there are no whole-system assessments of nutrient cycling in the context of

stream acidification.

To address these goals, we performed short-term (3-5 h) 15

N releases and measured stream

metabolism during late autumn in five streams spanning a gradient of stream pH (5.3 to 7.0). We

also quantified standing stocks of coarse benthic organic matter (CBOM) and FB on leaf

biofilms to obtain whole-stream qCO2 and qN across the pH gradient.

Methods

Study Sites

The study was conducted in five second- to fourth-order streams draining catchments within

Shenandoah National Park (SNP), Virginia, USA (Fig. 1). SNP comprises ~80,000 ha of

predominantly (95%) forested land (Young et al. 2006) in the Blue Ridge Mountains and has

been an active location for acid deposition research since 1979 (http://swas.evsc.virginia.edu).

The park receives high acid deposition due to its position downwind of the coal-fired power

plants of the Ohio River Valley (Driscoll et al. 2001). Differences in stream pH arise within SNP

due to variation in underlying geology and subsequently, the capacity of catchments to neutralize

acidic inputs. Three major bedrock types are recognized in order of increasing acid sensitivity:

basaltic (mafic), granitic (felsic), and siliciclastic (quartzite and sandstone; Cosby et al. 2006;

Fig. 1). The five study streams are distributed across all three bedrock types. Long-term (11-16

y) records of pH measurements during either December (the month we conducted our study) or


39

January (when December data were not available) show a gradient in pH from 5.3 to 7.0 across

the five streams (Table 1; James N. Galloway, University of Virginia, personal communication).

Dissolved inorganic nitrogen is predominantly nitrate-N (NO3-N) and concentrations are

relatively low (<0.2 mg·L-1

), with the highest levels during early spring and lowest

concentrations (<5 μg·L-1

) during and immediately following autumn leaf-fall (October-

December). Soluble reactive phosphorus (SRP) concentration in all streams is low (<5 μg·L-1

)

year-round.

Solute release experiments

We conducted continuous additions (3 to 5h duration) of 99% 15

N-enriched KNO3 along with

chloride (Cl-, from NaCl), as a hydrologically conservative tracer, to each stream between 27

November and 14 December 2007. Prior to releases, we collected background water samples (n

= 3) at each of four transects along each reach (60-100 m) for analysis of NO3-N, ammonium-N

(NH4-N), SRP, dissolved organic carbon (DOC), Cl-, and

15N-NO3. Releases of K

15NO3 were

designed to increase the delta (δ) 15

N of stream water by 500‰ at the addition site. Background

concentrations of NO3-N were ≤ 2 μg·L-1

at all sites and our 15

N enrichments increased these

concentrations by < 2%.

When streams reached steady-state conditions as indicated by unchanging conductivity (YSI

model 30 hand-held probe, YSI, Yellow Springs, Ohio, USA) at the most downstream transect,

three replicate samples were collected at each transect and corrected for background and dilution

influences (Stream Solute Workshop 1990). Samples for the analysis of 15

NO3-N were collected

in acid-washed 1-L bottles and kept on ice until filtered (Whatman GF/F, pore size = 0.7 μm) in

the laboratory and thereafter refrigerated (~4° C) until processed. Additional samples (n = 3 per

40

transect) were collected for NO3-N and Cl- analyses, filtered (Whatman GF/F, pore size = 0.7

μm) into acid-washed 125-ml polyethylene bottles, and frozen.

Temperature and specific conductivity of stream water was continuously monitored at 5-min

intervals throughout the releases using an automated sonde (Hydrolab Model 4a, Hydrolab, Hach

Environmental, Loveland, Colorado, USA) at the farthest downstream transect. Velocity was

determined from the resulting conservative tracer curve (Bencala and Walters 1983). Discharge

at each transect was calculated by mass balance using background-corrected [Cl-] at each

transect and the known [Cl-] and delivery rate of the injectate (Mulholland et al. 2006).

Wetted widths and depths were measured at 10 m intervals along the study reach following each

release.

Sample processing and analysis

We followed the reduction-diffusion method of Sigman et al. (1997) to process 15

N-NO3

samples. Ammonium was removed by boiling samples under basic conditions (by adding MgO)

to a final volume of ~100 ml. Devarda‘s Alloy was then added to reduce nitrate to ammonia,

followed by the immediate addition of a pre-combusted, acidified (25 μL; 2.5 M KHSO4) glass-

fiber filter (Whatman GF/D) wrapped in Teflon tape. The sealed sample was kept at 60°C for 48

h and then gently shaken for 7 days at room temperature to facilitate the diffusion of NH3 into

the acidified filter. Filters were then removed, dried, encapsulated in tins, and sent to the

University of California, Davis Stable Isotope Facility for analysis of 15

N:14

N ratios on a Europa

Integra mass spectrometer (Sercon, Cheshire, UK). Nitrate concentrations in the streams were

too low to provide ample N mass for 15

N measurement so we spiked all samples with 100 μg

NO3-N; diffusions on spiked reagent-grade water were used to correct for the added 14

N.

41

Concentrations of nitrite-N and nitrate-N (presented as NO3-N) and NH4-N were determined

colorimetrically following the acidic diazo method after cadmium reduction (Wood et al. 1967)

and a modified form of the automated phenate method (USEPA 1997), respectively, using a

Lachat QuikChem 8500 auto analyser (Lachat Instruments, Loveland, Colorado, USA). Soluble

reactive phosphorus was used to represent orthophosphate (PO4-P) as determined by the

molybdate-antimony method (Murphy and Riley 1962). Dissolved organic carbon (DOC) was

determined by persulphate digestion (Menzel and Vaccaro 1964) on a model OI Analytical 1010

Total Organic Carbon Analyzer (Oceanographic International, College Station, Texas, USA).

Chloride was analyzed on a Dionex DX 500 Ion Chromatograph (Dionex, Sunnyvale, California,

USA). Base cation (Ca2+

, Mg2+

, K+ and Na

+) concentrations were measured using a Dionex 500

Ion Chromatograph (Dionex Corporation, Sunnyvale, CA, U.S.A.) following standard methods

(American Public Health Association [APHA], 1998).

Whole-system respiration

Reach-scale measurements of stream ecosystem respiration (ER) were made using open-system

single-station analyses of diel oxygen curves (Owens 1974, Young and Huryn 1996, Mulholland

et al. 2001) obtained from an automated sonde placed at the bottom of the reach. Sondes were

equilibrated (at least 1 hr) with 100% moisture-saturated air prior to calibration, placed at the

bottom of the reach before the start of the solute release, retrieved 36-48 hrs after termination of

the solute release, and immediately placed back into saturated air to correct for drift. Reaeration

rate was calculated from longitudinal declines in sulfur hexafluoride (SF6; Marzolf et al. 1994).

DO concentrations were collected every five minutes, and the rate of ER for each 5-min interval

during the night was calculated as the difference between the change in DO and the quantity of

42

atmospheric exchange over the interval. Atmospheric exchange was determined as the product of

the reaeration coefficient and the oxygen deficit (i.e., the difference between observed DO and

expected DO at saturation). Daytime ER was extrapolated from average ER during 1h pre-dawn

and 1hr post-dusk periods. Values for ER were summed for all intervals during the 24-h period

following sonde deployment and the resulting volumetric rates (g O2·m-3

·d-1

) were converted to

areal rates (g O2·m-2

·d-1

) by multiplying by average stream depth (m).

CBOM and fungal biomass

Ten samples (0.25 m2) of CBOM were collected at random locations along each study reach

following completion of solute releases. All organic matter within the sampling area was

removed by hand, sieved (8 mm), and chilled until processing. In the laboratory, samples were

dried at 55°C (24 hrs), weighed, ashed at 550°C (24 hrs), and reweighed to obtain ash-free dry

mass (AFDM).

A subset (n = 5) of the CBOM samples in each stream was chosen for analysis of ergosterol

to estimate FB. For each sample, we cut ~50 leaf discs from randomly chosen, recognizable leaf

material. Five of these leaf discs (~40-60 mg AFDM) were then kept in 5 mL methanol (HPLC-

grade) and frozen at -20°C until measurement of ergosterol content (Gessner and Chauvet 1993).

The remaining leaf discs were added back to the CBOM sample of origin for inclusion in mass

calculations. Ergosterol was extracted by heating in methanol for 2 h at 65°C, purified by liquid-

liquid extraction, and quantified by high pressure liquid chromatography following the methods

of Tank et al. (1998). The mobile phase was 2 mL min-1

methanol and the stationary phase was a

Nova-Pak ODS C18 reverse phase column; UV absorbance detection was set at 282 nm (Dionex

DX-500, Dionex Corp., Sunnyvale, CA, USA). Ergosterol mass was multiplied by a conversion

43

factor of 182 to estimate FB (Gessner and Chauvet 1993). Leaf disc mass was determined by

combustion and added to the CBOM sample of origin, and FB was reported as g FB per g leaf

AFDM.

Nutrient spiraling and qCO2 metrics

The NO3-N uptake length (Sw; m) was derived as the inverse of the slope (longitudinal uptake

rate; kL; m-1

) of the line relating ln (15

N flux) and distance downstream (m). The uptake velocity,

vf (m·s-1

), standardizes Sw to hydrologic influences:

w

fS

udv

where u is stream velocity (m·s-1

) and d is mean stream depth (m). Stream background NO3-N

concentration is combined with the uptake velocity to determine the areal rate of NO3-N uptake

(U, μg N·m-2

·s-1

):

CvU f

where C is ambient NO3-N concentration (μg·L-1

).

Ecosystem respiration (g O2·m-2

·d-1

) was converted to a rate of CO2-C production (g C·m-2

·d-

1) in each stream using a respiratory quotient of 0.85 moles of CO2 produced per mole of O2

respired (Bott 2006). Stream-specific FB (g·g-1

AFDM) was multiplied by standing stocks of

CBOM (g AFDM·m-2

) and divided in half to estimate standing stocks of fungal C (g C·m-2

)

within the reach. Carbon-based respiration was then divided by fungal C to obtain the metabolic

quotient, qCO2 (g·CO2-C·g-1

fungal C·d-1

). We also calculated qN (μg N·g FB·d-1

), a biomass-

specific rate of N uptake for each stream by dividing U by FB calculated for the reach.

Statistical analyses

44

Differences in base cations, CBOM standing crops, and background nutrient and DOC

concentrations among streams were investigated using one-way analysis of variance (ANOVA).

Tukey‘s honest significant difference (HSD) was used as a post-hoc multiple comparison test

following significant (P < 0.05) ANOVA results. Pearson correlation was used to evaluate

relationships between ambient pH and response variables. All statistical analyses were performed

using SigmaStat 3.11.0 (Systat Software Inc., San Jose, CA).

Results

Average stream temperature during the 24 h encompassing each 15

N release ranged 3.7 °C

among streams from 3.1°C in Piney River to 6.8°C in Paine Run (Table 1). Concentrations of

individual base cations generally increased with greater stream pH (except K+, Table 1) but only

Ca2+

(r = 0.89, P = 0.042, n = 5) and Na+ (r = 0.99, P = 0.001, n = 5) displayed clear positive

relationships. The charge sum of base cations was also positively related to increasing stream pH

(r = 0.95, P = 0.013, n = 5). Concentrations of NH4-N were consistently below detection (1 μg·L-

1) in all streams, and NO3-N, while detectable, was scarce (1-2 μg·L

-1, Table 1). SRP

concentrations were low (1-5 μg·L-1

) and DOC ranged 1.3 mg·L-1

, from 0.8 mg·L-1

in Meadow

Run to 2.1 mg·L-1

in both Brokenback Run and Hazel River (Table 1). There was no relationship

between stream pH and either nutrient or DOC concentrations (Pearson correlation, P > 0.05).

Standing crops of CBOM varied from 7.9 to 11.9 g AFDM·m-2

across four of the five study

streams while CBOM was more abundant in Hazel River (20.3 g AFDM·m-2

, Table 2). CBOM

stock showed no relationship to stream pH (Table 2). Fungal biomass differed 13-fold between

the most acidic (2.9 mg·g-1

AFDM) and the most circumneutral stream (37.7 mg·g-1

AFDM;

45

Table 2) and was positively correlated to stream pH (Pearson correlation of ln-transformed FB

with stream pH, r = 0.98, P = 0.004; Figure 2a).

Uptake lengths (Sw) for NO3-N were between 254 and 319 m across four of the five study

streams while Sw was longer in Hazel River (502 m, Table 2). Uptake velocities (vf) varied from

0.61 mm·min-1

(Brokenback Run) to 1.12 mm·min-1

(Hazel River, Table 2). Areal uptake rate

(U) was highest in Hazel River (3.3 mg N·m-2

·d-1

) and lowest in Paine Run (0.83 mg N·m-2

·d-1

;

Table 2). We found no relationship between any metric of N uptake (Sw, vf, and U) and stream

pH or fungal standing stocks (Pearson correlation, P > 0.05).

Unlike N uptake, ER displayed a strong positive relationship with increasing stream pH (r =

0.99, P = 0.002; Fig. 2b). Across the pH gradient, ER increased 5-fold from 1.14 g O2·m-2

·d-1

in

the most acidic stream (Meadow Run) to 5.97 g O2·m-2

·d-1

in the most circumneutral stream

(Piney River; Fig. 2b).

Whole-stream qCO2 increased with stream acidity (i.e., declined with pH; r = -0.81, P =

0.09; Fig. 3a), from 8.5 g CO2-C·g-1

fungal C·d-1

in Piney River (pH 7.0) to 31.8 and 37.6 g CO2-

C·g-1

fungal C·d-1

in Meadow Run (pH 5.3) and Paine Run (pH 5.7), respectively. While U was

not related to acidity, U standardized to areal fungal biomass (i.e., qN) increased with acidity

(i.e., declined with pH; r = -0.86, P = 0.06; Fig. 3b). Although both qCO2 and qN increased with

increasing acidity, these variables were only weakly correlated with each other (r = 0.56, P =

0.32) due to the much higher qN/qCO2 found in Meadow Run relative to the other streams.

Without this data point, the relationship between qCO2 and qN improved (r = 0.89, P = 0.11, n =

4) but was still not significant.

46

Discussion

The patterns we observed in whole-stream qCO2 and qN across the pH gradient were similar to

our previous findings in laboratory microcosms, implicating stream acidification as a

physiological stressor with consequences for ecosystem-level energy and nutrient dynamics. FB

and ER both showed strong negative relationships with increasing stream acidity and these

patterns have either been found in (in the case of FB) or inferred from (ER) studies of

acidification effects on leaf decomposition (e.g., Mulholland et al. 1987; Simon et al. 2009).

However, because FB declined exponentially with increasing acidity while the decline in ER was

arithmetic, qCO2 increased with increasing acidity. Differences in qCO2 may be the result of

either physiological impairment of resident biota (Schimel et al. 2007) or community shifts along

the pH gradient (Lovett et al. 2009). Whatever the mechanism, stream pH and potentially the

suite of geochemical alterations associated with acidification influence ecosystem-level

efficiencies of resource use. The acid-induced inefficiencies in C use observed at the whole-

stream level indicate greater amounts of energy allocation toward maintenance and survival at

the expense of growth and reproduction, patterns that characterize physiological stress (Schimel

et al. 2007).

The use of qCO2 as an index of metabolic efficiency has largely been restricted to soil

systems with no applications in aquatic ecosystems. Further, few studies investigating lotic

metabolism responses to a particular change in environmental conditions simultaneously report

standing stocks of microbial biomass. We found three such studies, all investigating

eutrophication effects on ecosystem structure and function, from which we calculated qCO2

based on their published values of respiration and microbial biomass (Fig. 4). In a five-year

continuous nutrient enrichment (N+P) experiment in a mountain stream within Coweeta

47

Hydrologic Laboratory, NC, USA, mean qCO2 over the five-year treatment period was greater

than the reference during this time suggesting that eutrophication decreased the efficiency of

carbon use in this system (Fig. 4a; Suberkropp et al. 2010). Although qCO2 in the treatment

stream was 1.2 times higher during the single pre-treatment year, greater differences between

streams were observed in all subsequent years.

Similarly, qCO2 was greatest in a hypertrophic stream in France, where trophic status was

closely related to concentrations of phosphorus (P) (Fig. 4b; Baldy et al. 2007). Interestingly, the

relationship between qCO2 and P concentration was non-linear, with an increase in qCO2

observed for the most oligotrophic stream as well, suggesting decreases in carbon-use efficiency

at the extremes of trophic status (Fig. 4b). Lastly, mean qCO2 measured on various wood and leaf

types declined when stream reaches were enriched with N and P together (Fig. 4c; Stelzer et al.

2003). Although Suberkropp et al. (2010) found the opposite pattern with N+P additions (Fig.

4a), their enrichment levels were within the hypertrophic range of P concentrations reported by

Baldy et al. (2007), while the P concentrations in the treatment streams reported by Stelzer et al.

(2003) were similar to the intermediate concentrations where qCO2 was lowest.

In all three of the above cases, qCO2 varies with treatment along a scale of values (0.03 to

0.8 g CO2-C·g-1

fungal C·d-1

) that encompasses our previous findings in microcosms (0.07 to

0.45 g CO2-C·g-1

fungal C·d-1

; Ely et al. 2010). These values are all within the range of 200 qCO2

measurements (0.009 to 0.903 g CO2-C·g-1

microbial C·d-1

) published in six different studies of

soil systems addressing the influences of myriad conditions including increased O3 and CO2

concentrations (Islam et al. 2000), various soil management practices and crop rotation systems

(Hungria et al. 2009), different forest soil types (Anderson and Joergensen 1997; Agnelli et al.

2001; Armas et al. 2007) and 22,000 years of succession (Wardle and Ghani 1995). These

48

findings suggest that qCO2 may be a valuable metric of ecosystem metabolic efficiency that

combines structure and function to allow comparisons across a broad range of ecosystem types.

We found that qN at the ecosystem level increased with acidity along with declines in

metabolic efficiency (i.e., increased qCO2). These results corroborate a tight coupling between

acidity, metabolic efficiency, and N uptake derived from 150 measures in experimental

microcosms (Ely et al. 2010). This same study illustrated that fungal biomass declined by over

and order of magnitude across the gradient; the same decline observed one year later during the

present study. As a result, Ely et al. (2010) found no relationship between N uptake standardized

by leaf mass and stream pH and hypothesized that greater biomass-specific rates (possibly due to

increases in exoenzyme production) may compensate for reductions in microbial abundance,

leading to undetectable differences in U among streams. Indeed, we found no relationship

between U, Sw, or vf and stream pH in the current study. These results are surprising given a) the

large role microbes (and more specifically following autumn leaf fall, aquatic hyphomycetes)

play in nutrient dynamics (Gulis and Suberkropp 2003; Mulholland 2004), b) the large declines

in FB we observed with increasing stream acidity and c) the stoichiometric linkage between

energy flow and nutrient demand (Hall and Tank 2003). Combined with our similar findings in

controlled microcosm experiments, the decoupling of respiratory and nitrogen uptake dynamics

we have observed here suggests alternate fates for acquired N than incorporation into biomass,

and may have implications for DON production (as labile exoenzymes), cycling, and

downstream transport (Bronk et al. 1998; Berman and Bronk 2003).

The relatively weak relationship we observed between ecosystem-level qCO2 and qN (n = 5)

compared to those observed in highly controlled microcosm experiments (n = 75 at two levels of

N availability; Ely et al. 2010) prompted us to explore this relationship further using published

49

values of ER, N uptake and microbial biomass for streams spanning multiple biomes. The Lotic

Intersite Nitrogen eXperiment (LINX) was a coordinated study investigating factors controlling

metabolism and NH4-N uptake in streams distributed across arid-temperate, humid-temperate,

and humid-tropical climates in North America (Peterson et al. 2001). In that study, no general

relationships could explain the variability observed in N uptake; however, calculations of N

demand based on metabolic rates corresponded with measured assimilative (i.e., total N uptake

minus nitrification) N uptake (Webster et al. 2003). We calculated qCO2 and qN in 9 LINX

streams with published measures of microbial (fungi + bacteria) biomass (Findlay et al. 2002),

NH4-N uptake, and ER (Webster et al. 2003). Using these data we found a strong, positive

relationship between qCO2 and qN (qN = 0.0088*qCO2 + 0.025, r2 = 0.94, P < 0.001, Figure 5)

with a slope similar to that found in microcosm experiments (0.0106, D.T. Ely unpublished data)

having comparable N availabilities (LINX streams: 1.9 – 23.4 μg NH4-N·L-1

; microcosms: 25 μg

NH4-N·L-1

). These data suggest that N flux through resident microbes increases with decreasing

metabolic efficiency and indicate stress as an influential factor in nutrient dynamics. We do not

know what factors are responsible for the differences in qCO2 among the LINX streams.

While qCO2 and qN increased with stream acidity both during microcosm (Ely et al. 2010)

and whole-stream experiments, whole-stream values were often greater by 1 to 2 orders of

magnitude, suggesting that ecosystem experiments underestimated the biotic compartments

responsible for respiratory and nutrient uptake functions (Carpenter et al. 1989). Losses of

inorganic N from the water column may be dictated by N concentration, abiotic influences of

hydrology, and sorption and biotic uptake by the microbial assemblage (algae + bacteria + fungi)

(Bernot and Dodds 2005). The use of a 15

N tracer in both this study and the LINX experiments

ensured minimal N enrichment, and the simultaneous release of a conservative tracer accounted

50

for abiotic influences (Stream Solute Workshop 1990). Gross primary production (GPP) in SNP

streams was extremely low (between 0.09 and 0.45 g O2·m-2

·d-1

, D.T. Ely, unpublished data)

suggesting little contribution of algae to whole-stream rates of either respiration or N uptake.

However, when LINX standing crops of epilithon and algae were included, qCO2 and qN in four

streams were lowered to within the set of values observed in microcosms.

We did not include bacterial biomass in estimates of qCO2 and qN because bacteria often

contribute less than 10% of microbial biomass and production during the initial stages of leaf

decomposition (Suberkropp and Weyers 1996; Findlay et al. 2002; Gulis and Suberkropp 2003;

Stelzer et al. 2003). In the LINX streams, Findlay et al. (2002) found that while fungal biomass

was much greater than bacterial biomass on leaves and wood, the opposite held for smaller size

classes (i.e., fine benthic organic matter, FBOM). We did not quantify FBOM pools, thus we do

not know how the contribution of bacteria to qCO2 and qN varied among sites. The qCO2 and qN

values we calculated from LINX data included both fungal and bacterial biomass, and while one

stream (Walker Branch, Tennessee) displayed values for these metrics within the range observed

in microcosms, all other streams had qCO2 and qN values similar to those found in SNP streams.

This study emphasizes the need to address the coupling between energy flow (i.e., C

processing) and biogeochemistry (N uptake) and uses changes in this relationship as a means to

assess the influence of ecosystem stress associated with global change. We have shown that

anthropogenic acidification can have deleterious effects on microbial biomass and associated

decomposition processes and that these translate to microbial processing of N in a manner

opposite that initially suggested by reduced microbial biomass. Indeed, the potential exists that

enhanced biomass-specific N uptake (i.e., qN) may compensate for depleted microbial biomass

without influencing ecosystem-level rates of N uptake.

51

Respiration and nutrient acquisition are expected to be linked stoichiometrically, especially

in strongly heterotrophic detritus-based systems, due to requirements for consumer growth, and

some lotic studies have found such linkages (Hall and Tank 2003, Webster et al. 2003; but see

Fellows et al. 2006b). The similarities we observed in areal N uptake among streams despite

robust declines in ecosystem respiration with increasing acidity suggest that N demand under

stressful conditions may not necessarily be related to growth requirements, or that altered

microbial assemblages in acid streams are functionally redundant, even compensatory, causing

structure to lag behind this particular function (N uptake) along a long-term trajectory of

ecosystem recovery (Kelly and Harwell 1990). Furthermore, patterns in qCO2 and qN associated

with increasing stream acidity are indicative of compensatory N uptake associated with a general

decrease in the efficiency of carbon use and point to the idea that the relationship between qCO2

and qN may serve as a potentially valuable indicator of ecosystem response to large-scale

changes in environmental conditions.

52

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59

Table 1. Temperature and chemical characteristics of the five study streams. Values in the same

row that share a similar letter are not significantly different (Tukey‘s HSD, P>0.05).

Parameter Meadow

Run

Paine Run Brokenback

Run

Hazel River Piney River

pH 5.3 5.7 6.5 6.6 7.0

Temperature (°C)* 5.2 6.8 5.2 4.7 3.1

Ca2+

(mg·L-1

) 0.7a 0.9

a 2.2

bc 2.1

b 4.7

c

Mg2+

(mg·L-1

) 0.4a 0.5

bc 0.4

ab 0.5

ab 1.3

c

K+ (mg·L

-1) 0.9

ab 1.7

a 0.4

c 0.4

bc 0.3

c

Na+ (mg·L

-1) 0.6

a 0.7

a 1.6

b 1.7

b 2.1

c

Base Cations (μeq·L-1

) 109a 157

b 226

c 230

c 438

d

NH4-N (μg·L-1

) bd1 bd bd bd bd

NO3-N (μg·L-1

) 1a 1

ab 2

b 2

bc 1

ac

SRP (μg·L-1

) 1a 2

ac 5

b 4

b 3

bc

DOC (mg·L-1

) 0.8a 0.7

a 2.1

b 2.1

b 1.3

ab

*24h average 1bd: below detection (1 μg·L

-1)

60

Table 2. Standing crops of coarse benthic organic matter (CBOM) and fungal biomass, flow

characteristics, and N uptake parameters for the five study streams.

Meadow

Run

Paine Run Brokenback

Run

Hazel

River

Piney

River

Detrital stocks

CBOM

(g AFDM·m-2

±SE)

7.9 ± 1.5 9.6 ± 3.4 11.1 ± 3.2 20.3 ± 7.0 11.9 ± 3.9

Fungal biomass

(mg·g-1

AFDM ±SE)

2.9 ± 0.2 4.6 ± 0.1 9.8 ± 2.1 11.8 ± 2.3 37.7 ± 4.7

Flow conditions

Wetted width (m) 5.2 4.9 5.9 4.2 7.0

Depth (cm) 10 25 16 18 12

Current velocity (cm·s-1

) 2.9 1.3 1.8 5.2 4.2

Discharge (L·s-1

) 15.4 16.1 33.0 40.5 34.3

N uptake

Sw (m) 254 319 287 502 305

vf (mm·min-1

) 0.67 0.62 0.61 1.12 0.99

U (mg N·m-2

·d-1

) 1.38 0.83 2.07 3.30 1.33

61

Figure 1. Map of Shenandoah National Park, VA, U.S.A. showing the three major bedrock

classes and the location of catchments drained by the five study streams. (Reproduced with

permission from Webb et al. 2004)

62

Figure 2. Correlations between stream pH and a) fungal biomass and b) ecosystem respiration.

a

b

63

Figure 3. Correlations between stream pH and a) whole-stream qCO2 and b) qN.

64

Figure 4. Metabolic quotients (qCO2) calculated from three separate studies of nutrient

enrichment effects on respiration and fungal biomass in stream ecosystems. a) qCO2 measured

on leaf discs in streamside chambers in a reference stream (white bars) and a stream

experimentally enriched with inorganic N and P (gray bars) during one year before enrichment of

the treatment stream (pre-treatment) and the following five years of continuous enrichment

(Years 1-5; mean +1SE; from Suberkropp et al. 2010). Both are headwater streams located

65

within Coweeta Hydrologic Laboratory, NC, U.S.A. Values in the Years 1-5 category are

significantly different by a paired t-test (P = 0.003, n = 5). b) qCO2 measured on alder leaf discs

in laboratory microcosms of water from six streams in southwestern France of differing trophic

status related to SRP concentrations (from Baldy et al. 2007). The relationship is best described

by a second-order polynomial regression (y = 0.000106x2 – 0.00663x + 0.463, r

2 = 0.841, P =

0.067). c) average qCO2 (+1SE, n = 7) of mean rates measured on five different types of wood

and two types of leaves following incubations in a reference reach (white bars) and downstream

reaches experimentally enriched with either inorganic N alone (gray bars) or N + P (black bars)

during periods before (pre-treatment) and after (post-treatment) enrichment (from Stelzer et al.

2003). The study stream is located in Hubbard Brook Experimental Forest, NH, U.S.A.

Measurements were conducted during mid-summer (pre-treatment) and early autumn (post-

treatment). Values were not significantly different during the pre-treatment period (one-way

ANOVA); qCO2 during the post-treatment period was significantly lower in the N+P-enriched

reach than either of the other reaches, which did not differ from each other (one-way ANOVA,

Tukey‘s post-hoc comparison).

66

Fig. 5

Figure 5. Relationship between qCO2 and qN in nine North American streams. BBNH: Bear

Brook, New Hampshire; BCNC: Ball Creek, North Carolina; ECMI: Eagle Creek, Michigan;

GCNM: Gallina Creek, New Mexico; KCKS: Kings Creek, Kansas; MCOR: Mack Creek,

Oregon; QBPR: Quebrada Bisley, Puerto Rico; SCAZ: Sycamore Creek, Arizona; WBTN:

Walker Branch, Tennessee. Data from Findlay et al. (2002) and Webster et al. (2003).

67



Abstract

While effects on consumers are generally negative, stream acidification often has positive

influences on autotrophic structure and function, which are emphasized during the spring period

of high light availability prior to leaf emergence. In this chapter I ask how acidification-enhanced

endogenous organization affects whole-stream carbon (C) and nitrogen (N) cycling. I predict

greater rates of N uptake associated with increased acidity, since autotrophs experience less

stoichiometric constraint in nutrient uptake requirements than heterotrophs, and a simultaneous

decrease in the capacity of acid streams to process organic C. Algal biomass and water chemistry

were monitored over a 3-month period (April – July 2009) in four streams of differing pH. Near

the end of this period I measured N uptake using 15

N tracers in the most acidic and most

circumneutral streams under ambient conditions and again with experimentally-added glucose as

a source of dissolved organic carbon (DOC). Measurements of metabolism (gross primary

production [GPP] and ecosystem respiration [ER]) and algal standing crops were also measured

during whole-stream experiments. Acidic streams generally had higher algal biomass than

circumneutral streams but the most acidic stream did not display any differences in N uptake

under ambient conditions and had slightly lower GPP than the most circumneutral stream.

However, the acidic stream had a much lower rate of ER, making its P:R ratio higher than the

circumneutral stream. When streams were DOC-enriched, the acidic stream showed little

response in terms of N demand, whereas the most circumneutral stream was stimulated by the

carbon addition made evident by a large decrease in uptake length and increases in uptake

velocity and areal uptake of N. Uptake metrics for C calculated from the glucose additions

showed a much lower demand for organic C in the acid stream.

68

Introduction

Knowledge of the relative degree of endogenous (energy base generated within the system) and

exogenous (imported energy base) organization of ecosystem function, and greater

understanding of how their comparative influences shift with altered environmental conditions,

may better inform our predictions of ecosystem response to global change. For instance,

continuous nutrient enrichment of an exogenously-controlled stream in North Carolina resulted

in accelerated rates of carbon (C) flow through the detritus-based food web (Cross et al. 2007;

Suberkropp et al. 2010), whereas coastal eutrophication of the Mississippi River delta, an

endogenously-controlled system, has led to large increases in C flow through autotrophic

pathways (Lohrenz et al. 1997). Recently, Valett et al. (2008) suggested that a combination of

endogenous and exogenous control is typical of most ecosystems, where endogenous

organization is the baseline and exogenous control grows in importance as systems become more

open.

Streams are considered among the most open ecosystems because of their position at the

high end of a continuum of allochthonous inputs (Leroux and Loreau 2008). Such strong

linkages exist between streams and the surrounding landscape that stream structure and function

essentially integrates landscape-level responses to environmental changes (Williamson et al.

2008). While both temporal and spatial chemical profiles of streams have provided insight into

terrestrial processes (Vitousek and Reiners 1975; Hedin et al. 1995; Brookshire et al. 2009), the

quantification of in-stream metabolism and nutrient cycling has revealed strong relationships

with seasonal environmental variation (Roberts et al. 2007) and human-induced alteration of

entire landscapes (Hall et al. 2009), linking atmospheric and terrestrial catchment properties to

local and downstream aquatic communities.

69

The openness of streams is a result of the large influx of materials across system

boundaries, and consequently, most streams are a net sink for carbon (i.e., they are heterotrophic;

Mulholland et al. 2001). In temperate regions, autumnal pulses of leaf litter are the trophic basis

of in-stream production (Wallace et al. 1997) and stimulate nitrogen (N) and phosphorous (P)

uptake through the metabolic activity of heterotrophic microbes (Roberts and Mulholland 2007).

While streams are net heterotrophic on an annual basis, gross primary production (GPP) can

exceed ecosystem respiration (ER) during one or more seasons (Bott et al. 1985), and nutrient

uptake by autotrophs can be substantial during this period (typically spring; Mulholland et al.

2006; Roberts and Mulholland 2007). Moreover, producer-driven functions of metabolism and

nutrient uptake can be sensitive to large-scale changes in environmental conditions. For example,

stimulation of GPP increased the rate of nitrate (NO3-) removal from the water column in a

forested stream of eastern Tennessee when a spring freeze led to delayed leaf emergence across

the central and southeastern United States in 2007 (Mulholland et al. 2009).

Acidification and in-stream processes

Stream acidification continues to be a widespread detriment to ecosystems in industrialized

regions of the world despite successful attempts to curb emissions of acid-forming pollutants

(Stoddard et al. 2003; Sanderson et al. 2006). While acidification lowers aquatic diversity across

all levels of biological organization (Lovett et al. 2009), opposite effects on both the standing

stocks and functional attributes of autotrophic and heterotrophic compartments of microbial

communities are observed. Fungi and bacteria are negatively impacted with increasing acidity

(Hall et al. 1980; Palumbo et al. 1987; Griffith and Perry 1994), while algal abundance is

frequently observed to be higher in acidified streams (Mulholland et al. 1986; Stokes 1986;

70

Niyogi et al. 1999) with corresponding increases in primary productivity (Mulholland et al.

1986). Reasons for positive autotrophic responses include the stimulation of acid-tolerant species

(Novis 2006), decreased competition (Niyogi et al. 1999), and release from grazing pressure

(Rosemond et al. 1993).

The impairment of microbial heterotrophs and the potential for concomitant stimulation of

autotrophs (i.e., periphyton) conceivably enhances endogenous control in acidified streams, with

consequences for energy and nutrient dynamics. The uptake of NO3-, a common pollutant of

surface waters, is often more strongly related to GPP over ER (Hall and Tank 2003; Hall et al.

2009) and benthic algae can play key roles in N immobilization from the water column (Fisher et

al. 1982; Covich et al. 2004; McKnight et al. 2004). At the same time, acid impairment of fungi

and bacteria should lead to reduced heterotrophic control over stream biogeochemistry evident as

a lower capacity to process dissolved organic carbon (DOC). Reductions in DOC processing

rates in acidified streams are of concern because DOC concentrations have been observed to rise

in these same systems over the past two decades; ironically, due to terrestrial recovery from acid

deposition (Monteith et al. 2007).

During previous investigations of whole-stream N cycling across a gradient of pH in

Shenandoah National Park, VA, we noted evident differences in algal patch types (i.e., green vs.

brown patch types, sensu Baker et al. 2009) with green patches abundant in acidic and brown

dominant in circumneutral streams - most obvious during spring. These observations, and the

findings of studies described above, led us to ask how apparent increases in endogenous control

due to acidification were influencing stream function. We quantified chlorophyll a, epilithic

biomass, and their ratio (i.e. the autotrophic index) over a 3 month period (April – July 2009) in

four streams spanning a pH gradient from ~5.5 to ~6.9 in order to characterize the degree of

71

autotrophy present across the gradient. We also measured stream metabolism (GPP and ER), and

conducted short-term (3-4 hours) releases of a 15

N-KNO3 tracer, with and without simultaneous

addition of glucose, in the most acidic and most circumneutral stream to evaluate acidification

influences on N uptake under ambient and increased DOC availabilities. In addition, C uptake

metrics were calculated during carbon enrichment experiments in both streams. Our objective

was to understand how stream acidification influences the degree of endogenous versus

exogenous control over C and N uptake in these systems.

Methods

Study sites

The study was conducted in four second- to fourth-order streams draining catchments within

Shenandoah National Park (SNP), Virginia, USA. SNP comprises ~80,000 ha of predominantly

(95%) forested land in the Blue Ridge Mountains and has been an active location for acid

deposition research since 1979 (http://swas.evsc.virginia.edu). The park receives high acid

deposition due to its position downwind of the coal-fired power plants of the Ohio River Valley

(Driscoll et al. 2001). Differences in stream pH arise within SNP due to variation in underlying

geology and subsequently, the capacity of catchments to neutralize acidic inputs (Cosby et al.

2006).

Water column and periphyton sampling

We began regular sampling of various dissolved solutes and benthic periphyton on 12 April

2009, and our last collection date was 11 July 2009. Our intended sampling interval was 14 days;

however, inclement weather and other complications resulted in occasionally shorter or longer


72

intervals between sampling and some dates where not all streams were sampled. Water

temperature in each stream was logged at 1-h intervals over the entire study period using HOBO

pendant data loggers (Onset Computer Corporation, Bourne, Massachusetts, USA). Stream

reaches (100 m length) were just below the park boundary and contained no tributaries. On each

sampling date, we measured stream pH with a Thermo Orion model 290A portable meter (Orion

Research Inc., Beverly, Massachusetts, USA) following calibration in the field, and collected

water samples (n = 3) for chemical analyses. Samples from each stream were filtered (Whatman

GF/F, pore size = 0.7 μm) into acid-washed 125-mL polyethylene bottles and frozen until

analysis of NO3-, ammonium (NH4

+), soluble reactive phosphorus (SRP), DOC and base cations

within five days of collection.

Periphyton was collected (n = 3) on each sampling date at three pre-determined, random

locations within each of the study reaches. For each sample, we scrubbed 3-5 randomly chosen

rocks with a wire brush, collected the combined periphyton slurries (total volume = 400-800 ml)

and immediately placed the sample on ice in a dark cooler. Scrubbed rocks were then covered

with aluminum foil and a known mass-area relationship was used to calculate surface area

(Bergey and Getty 2006). Periphyton samples were later frozen (within 3-12h) until analysis of

chlorophyll a and quantification of epilithic ash-free dry mass (AFDM).

15N and DOC releases

We conducted two continuous additions (3 to 4-h duration) of 99% 15

N-enriched KNO3 within a

48 to 72-h period in both the most acidic (Meadow Run, pH 5.5) and most circumneutral (Piney

River, pH 6.9) streams; the period between the first release in Meadow Run and the last release

in Piney River was 11 days. During the second N release in both streams, we also added glucose

73

(500 g·L-1

; Sigma-Aldrich, St. Louis, MO, U.S.A.) at a rate designed to increase stream [DOC]

four- to five-fold. To correct for dilution effects, a conservative tracer (Cl-, as NaCl) was injected

along with N in all experiments. Releases were designed to increase the delta (δ) 15

N of stream

water to 1000‰. These additions resulted in maximum NO3-N concentration increases of <2.0%.

Prior to each release, we collected background water samples (n = 3) for analysis of NO3-N,

Cl-, DOC, and δ

15N at each of four transects. Conductivity was monitored at the most

downstream transect (YSI model 30 hand-held probe, YSI, Yellow Springs, Ohio, U.S.A.) until

readings were stable (steady-state conditions), whereupon transects were re-sampled. Samples

for the analysis of 15

NO3-N were collected in acid-washed 1-L bottles and kept on ice until

filtered (Whatman GF/F, pore size = 0.7 μm) in the laboratory, and thereafter refrigerated (~4°

C) until processed. All other samples were filtered (0.7 μm) into acid-washed 125-ml

polyethylene bottles and frozen until analysis. We measured stream width and took multiple

readings of depth every 10 m along the study reach following each release.

Current velocity was calculated from tracer breakthrough curves (Bencala and Walters

1983) using conductivity readings recorded with an automated sonde (Hydrolab Model 4a,

Hydrolab, Hach Environmental, Loveland, Colorado, USA) placed at the farthest downstream

transect. Discharge at each transect was calculated by mass balance using background-corrected

[Cl-] at each transect and the known [Cl

-] and delivery rate of the injectate (Mulholland et al.

2006).

Sample processing and analysis

We measured chlorophyll a (chl a) and quantified epilithic AFDM to evaluate both the

abundance of stream autotrophs and their contribution to stream biofilms. Each periphyton

74

sample was subsampled for analysis of chl a and determination of AFDM. Chl a subsamples

were filtered onto Whatman GF/F glass fiber filters (0.7 μm), wrapped in aluminum foil and

stored at -20°C for at least 24 h. Filters were then placed in a 15-ml centrifuge tube filled with 10

ml HPLC-grade methanol and allowed to extract overnight in the dark at 6°C. The next day tubes

were centrifuged (1000 rpm, 5 min) and an aliquot of the methanol-extract solution was analyzed

spectrophotometrically for chl a (corrected for phaeophytin, Steinman et al. 2006). Total chl a in

each sample was divided by rock surface area to obtain mg chlorophyll a per square meter.

To calculate epilithic biomass, subsamples were filtered as above, dried for 7 days (50°C),

weighed, ashed (550°C, 3h), and re-weighed to obtain AFDM per subsample, which was then

converted to total mass of the sample and divided by rock surface area. We used the ratio of chl

a:epilithic AFDM as an autotrophic index (AI), where greater values of AI indicate greater

autotrophic influence over biofilm character (Crossey and La Point 1988).

We followed the reduction-diffusion method of Sigman et al. (1997) to process 15

N-NO3

samples. Ammonium was removed by boiling samples under basic conditions (by adding MgO)

to a final volume of ~100 ml. Devarda‘s Alloy was then added to reduce nitrate to ammonia,

followed by the immediate addition of a pre-combusted, acidified (25 μL; 2.5 M KHSO4) glass-

fiber filter (Whatman GF/D) wrapped in Teflon tape. The sealed sample was kept at 60°C for

48h and then gently shaken for 7 days at room temperature to facilitate the diffusion of NH3 out

of solution and sequestration by the acidified filter. Filters were then removed, dried,

encapsulated in tins, and sent to the University of California, Davis Stable Isotope Facility for

analysis of 15

N:14

N ratios on a Europa Integra mass spectrometer (Sercon, Cheshire, UK). One

liter samples from Meadow Run (background NO3-N concentration = 40 μg·L-1

) were spiked

75

with 50 μg NO3-N to ensure ample N mass for 15

N measurement via mass spectrometry. We also

ran diffusions on spiked samples of reagent-grade water in order to correct for the added 14

N.

Concentrations of nitrite-N and nitrate-N (presented as NO3-N) and NH4-N were determined

colorimetrically following the acidic diazo method after cadmium reduction (Wood et al. 1967)

and a modified form of the automated phenate method (USEPA 1997), respectively, using a

Lachat QuikChem 8500 auto analyser (Lachat Instruments, Loveland, Colorado, USA). Soluble

reactive phosphorus was used to represent orthophosphate (PO4-P) as determined by the

molybdate-antimony method (Murphy and Riley 1962). Dissolved organic carbon (DOC) was

determined by persulphate digestion (Menzel and Vaccaro 1964) on a model OI Analytical 1010

Total Organic Carbon Analyzer (Oceanographic International, College Station, Texas, USA).

Concentrations of chloride and base cations (Ca2+

, Mg2+

, K+ and Na

+) were measured using a

Dionex 500 Ion Chromatograph (Dionex Corporation, Sunnyvale, CA, U.S.A.) following

standard methods (American Public Health Association [APHA], 1998).

Uptake metrics

The NO3-N uptake length (Sw; m) was derived as the inverse of the slope (longitudinal uptake

rate; kL; m-1

) of the line relating ln (15

N flux) and distance downstream (m). The uptake velocity,

vf (m·s-1

), standardizes Sw to hydrologic influences:

Eq. (1)

w

fS

udv

where u is current velocity (m·s-1

) and d is mean depth (m). Stream background NO3-N

concentration is combined with vf to determine the areal rate of NO3-N uptake (U, μg N·m-2

·s-1

):

Eq. (2)

76

CvU f

where C is ambient NO3-N concentration (μg·L-1

).

Sw, vf, and U were also calculated for DOC during glucose enrichment experiments. Uptake

lengths for DOC were obtained by taking the inverse of the slope of the line relating the natural

log of background-corrected DOC:Cl- ratio with distance downstream. DOC-specific vf and U

were calculated similarly to N uptake metrics using C uptake length and background DOC

concentration in equations (1) and (2), respectively.

Whole-system metabolism

Reach-scale measurements of stream ecosystem metabolism (GPP and ER) were made using

open-system single-station analyses of diel oxygen curves (Roberts and Mulholland 2007)

obtained from an automated sonde placed at the bottom of the reach. Sondes were equilibrated

(at least 1 hr) with 100% moisture-saturated air prior to calibration, placed at the bottom of the

reach before the start of the first solute release, retrieved following the second solute release, and

immediately placed back into saturated air to correct for instrument drift. Barometric pressure

was recorded at 5-min intervals throughout the period of sonde deployment using a Kestrel 4000

Weather Tracker (Nielsen-Kellerman Co., Boothwyn, Pennsylvania, USA). Reaeration

coefficients were calculated from longitudinal declines in sulfur hexafluoride (SF6; Marzolf et al.

1994); the net atmospheric exchange for each 5-min interval was determined by multiplying the

reaeration coefficient by the oxygen deficit, which was calculated as the difference between

observed dissolved oxygen (DO) and expected DO concentrations at 100% saturation determined

using temperature and barometric pressure.

77

The rate of ER for each 5-min interval during the night was calculated as the difference

between the change in DO and the quantity of atmospheric exchange over the interval. Daytime

ER was extrapolated from average ER during 1h pre-dawn and 1hr post-dusk periods. GPP was

determined for each 5-min daytime interval as the difference between the rate of DO change

(corrected for atmospheric exchange) and ER (Roberts and Mulholland 2007). Values for GPP

and ER were summed for all intervals during a single 24-h period following sonde deployment

and the resulting volumetric rates (g O2·m-3

·d-1

) were converted to areal rates (g O2·m-2

·d-1

) by

multiplying by average stream depth (m). Photosynthetically active radiation (PAR) flux density

(mol·m-2

·d-1

) was monitored using a quantum sensor (LICOR 190SA, LICOR, Lincoln,

Nebraska, U.S.A.) at a single location in both streams while sondes were deployed.

Statistical assessment

Pearson product-moment correlation analysis was used to investigate relationships among

measures of ecosystem structure. Mean concentrations of dissolved solutes, chl a, and epilithic

biomass over the entire sampling period were compared among streams using one-way analysis

of variance (ANOVA) with α = 0.05. For all significant ANOVAs, differences among sites were

assessed using a Tukey multiple comparison test with α = 0.05. Data failing normality or equal

variance tests were compared using Kruskal-Wallis ANOVA on ranks followed by Dunn‘s

multiple comparisons. All statistical analyses were performed using SigmaStat 3.11.0 (Systat

Software Inc., San Jose, CA).

Results

Temporal patterns

78

Stream pH was nearly identical in Hughes and Piney River (pH 6.9) while Paine Run (pH 5.8)

and Meadow Run (pH 5.4) were considerably more acidic; pH showed little variation throughout

the 91-day period in all streams (Fig. 1a), with the highest coefficient of variation (i.e., standard

deviation/mean * 100%) for any stream equal to only 1.4% (found in Paine Run). Ammonium-N

and NO3-N generally increased over the sampling period (i.e., April – July) in all streams (Fig.

1b, c), but concentrations of NH4-N were low (e.g., 4 μg·L-1

in Meadow Run, Table 1) and

ranged only 2 μg·L-1

over the 91-day period. Nitrate-N was consistently highest in Piney River

where it increased from 37 μg·L-1

to 103 μg·L-1

over the sampling period (Fig. 1c).

Concentrations of NO3-N were low (< 10 μg·L-1

) and similar in all other streams until mid-May,

after which concentrations diverged into a consistent ranking with greater values in

circumneutral streams compared to acidified streams (Fig. 1c). Similar to patterns of NO3-N,

SRP concentrations in circumneutral streams were greater than in acidified streams (6-7 μg·L-1

vs. 1-2 μg·L-1

, respectively; Table 1) and increased over time (i.e., 3 to 9 μg·L-1

in Piney River

and 4 to 11 μg·L-1

in Hughes River, Fig. 1d). At the same time, SRP remained low (2 μg·L-1

) and

stable in the acid streams. In the circumneutral streams (i.e., Hughes River and Piney River),

NO3-N and SRP both increased over the study period and were highly correlated (Pearson

correlation, r = 0.72, P = 0.002, n = 16), while in the acid streams (i.e., Meadow Run and Paine

Run) these solutes were unrelated (r = -0.14, P = 0.61, n = 17, data not shown). Across all

streams, DOC concentrations were < 2 mg·L-1

, but were consistently lowest in the two acid

streams (Fig. 1e, Table 1).

Acid and circumneutral streams also differed in base cation concentrations. Concentrations

of both Ca2+

and Na+ in acid streams were less than half those in circumneutral streams (Table 1)

and were positively correlated with stream pH (Ca2+

: r = 0.94, P < 0.001; Na+: r = 0.95, P <

79

0.001). Concentrations of K+, on the other hand, were low in circumneutral streams (Piney River:

0.37 ± 0.03 mg·L-1

, Hughes River: 0.34 ± 0.02 mg·L-1

), ≥ 3-fold higher in acid streams (Meadow

Run: 0.99 ± 0.04 mg·L-1

, Paine Run: 1.49 ± 0.06 mg·L-1

), and across streams were negatively

correlated with pH (r = -0.81, P < 0.001). Differences in Mg2+

found among streams (Table 1)

did not correlate with stream pH. The charge sum of base cations (SBC) increased consistently

from 126 to 203 μeq·L-1

as streams became more basic (Table 1) and was positively correlated

with stream pH (r = 0.93, P < 0.0001). Stream pH was also significantly correlated with chl a,

NO3-N, NH4-N, SRP and DOC. Since we were not able to measure variation in pH for any given

stream, SBC presented a broader range of values within individual streams resulting in stronger

relationships between these same variables and SBC than with pH directly. SBC (μeq·L-1

) was

most closely related to chl a, SRP, and DOC (Fig. 2). Chl a increased with declining SBC (r = -

0.70, P < 0.001, Fig. 2a), while both SRP (r = 0.89, P < 0.001, Fig. 2b) and DOC (r = 0.71, P

<0.001, Fig. 2c) decreased as SBC declined within and across streams. Similar to SRP and DOC,

both NO3-N (r = 0.51, P = 0.02) and NH4-N (r = 0.61, P = 0.003) also decreased as SBC

declined.

Mean chl a standing crops across the entire study period were similar in acid streams (4.9

and 4.6 mg·m-2

in Meadow Run and Paine Run, respectively) and both acid streams had greater

chl a standing crops than either of the circumneutral streams (1.7 and 2.1 mg·m-2

in Hughes

River and Piney River, respectively), which did not differ from each other (Dunn‘s multiple

comparisons, P < 0.05; Table 1). Areal biomass of chl a was nearly identical among streams on

the first sampling date and was low (2-3 mg·m-2

) with little variation over the 91-day period in

the circumneutral streams. In acid streams, however, standing crops rose quickly after the first

sampling date and were frequently greater in acid streams for the remainder of the study (Fig.

80

3a). We observed greater fluctuations in chl a standing crop in the acid streams as illustrated by

higher CVs (2.8 and 2.9% vs. 0.9 and 1.1%, respectively) in acidic and circumneutral systems.

This was most notable as a steep drop in chl a in both acid streams (Meadow Run and Paine

Run) on 9 May, after which chl a either returned to previous levels (Meadow Run) or exceeded

them (Paine Run; Fig. 3a). Both acid streams declined again (from 7-8 mg·m-2

to ~4 mg·m-2

)

during mid-June and remained at this level for the remainder of the study.

Similar to chl a standing crops, epilithic organic matter (OM) was also identical among sites

on the first sampling date but showed large variation in all streams over the remainder of the

study (Fig. 3b). Mean standing crops of epilithon over the study period were between 2.2 and 3.2

g AFDM·m-2

and did not differ among streams (Table 1). Despite the inconsistent variation in

epilithic OM, the AI was generally higher in acid streams (Fig. 3c) and patterns in AI across all

streams reflected trends in chl a.

N and C uptake experiments

The study reach in Piney River was slightly wider and deeper than in Meadow Run with greater

stream flow during solute releases (Table 2). NH4-N was low (4-6 μg·L-1

) in both streams (Table

2). In Piney River, mean NO3-N (95 μg·L-1

) and SRP (8 μg·L-1

) during the two injections were

greater than in Meadow Run (NO3-N = 41 μg·L-1

; SRP = 1 μg·L

-1) (Table 2). Background [DOC]

was 0.8 mg·L-1

in Meadow Run, and 1.2 mg·L-1

in Piney River, and 24h average temperature

differed only ~0.5°C between streams (Table 2). All injections were performed on clear days

with no precipitation occurring between the first and last experiment.

Nitrate uptake length (Sw) determined under ambient conditions (i.e., no glucose addition)

was 856 m in Meadow Run (pH 5.5) and 1079 m in Piney River (pH 6.9) (Fig. 4a). Glucose

81

additions raised DOC concentrations at the first transect from 0.8 to 5.1 mg·L-1

in Meadow Run,

and from 1.2 to 6.4 mg·L-1

in Piney River. Under glucose-enriched conditions, Sw decreased with

glucose additions by 9.8% and 36.3% in acidified (Meadow Run) and circumneutral (Piney

River) sites, respectively (Fig. 4a). The NO3-N uptake velocity (vf) decreased 2.0% (from 0.945

to 0.926 mm·min-1

) in the acidified stream while glucose addition in the circumneutral stream

increased vf by 42.9% (from 0.863 to 1.232 mm·min-1

) (Fig. 4b). Areal uptake rate of NO3-N (U)

was 2-fold greater in the circumneutral site (Piney River, 111.8 mg·m-2

·d-1

) than in Meadow Run

(acidified stream, 54.4 mg·m-2

·d-1

) under ambient DOC availability. Glucose additions to the

acidic site (i.e., Meadow Run) had essentially no effect on U (54.7 mg·m-2

·d-1

) while in the

circumneutral stream (Piney River) U increased 58.8% (to 177.5 mg·m-2

·d-1

) in response to

glucose addition (Fig. 4c).

Based on results from glucose additions (Table 3), C uptake length in the acidic stream

(Meadow Run, 160 m) was 2.4 times longer than in the circumneutral site (Piney River, 68 m).

Further, C uptake velocity in Meadow Run (0.075 mm·s-1

) was only 36% that found in Piney

River (0.208 mm·s-1

; Table 3). Accordingly, areal rate of C uptake was almost 3.5 times lower in

the acid stream (Meadow Run, 0.061 mg C·m-2

·s-1

) compared to the circumneutral site (Piney

River, 0.21 mg C·m-2

·s-1

; Table 3).

Chl a standing crop was almost 3-fold greater in the acid stream (3.72 ± 0.26 mg·m-2

) than

in the circumneutral stream (1.28 ± 0.07 mg·m-2

; Table 3) during the experimental releases.

Mean epilithic biomass was 3137 ± 757 mg AFDM·m-2

in Meadow Run and 4626 ± 112 mg

AFDM·m-2

in Piney River, resulting in an almost 5-fold higher AI value in the acidified system

(Table 3).

82

Rates of GPP were similar between the two study streams (Table 3); GPP was slightly lower

in the acidified stream (Meadow Run, 0.308 g O2·m-2

·d-1

) than in the circumneutral stream (Piney

River, 0.405 g O2·m-2

·d-1

) despite greater chl a standing crop and higher AI in the acidified

system. These differences led to lower chlorophyll-specific production in the acid stream (0.083

g O2·mg-1

chl a·d-1

) compared to the circumneutral stream (0.316 g O2·mg-1

chl a·d-1

; Table 3).

Differences in ER were more pronounced; ER in the acid site (1.83 g O2·m-2

·d-1

) was only 31%

the rate of ER in the circumneutral site (5.81 g O2·m-2

·d-1

; Table 3). Thus, the P:R ratio was

higher in the acidified stream (0.17) than in Piney River (0.07) while both systems were

evidently heterotrophic based on P:R ratios far less than one. PAR over the period of metabolism

measurement was 10.1 mol·m-2

·d-1

in Meadow Run and 8.7 mol·m-2

·d-1

in Piney River (Table 2).

Discussion

Despite successful reduction of atmospheric pollution, surface waters draining acid-impaired

watersheds may remain acidic over long time scales (Stoddard et al. 2003). The delayed recovery

of stream pH results from reductions in the buffering capacities of acidified catchments.

Compromised buffering capacities are due to both accelerated loss of soil base cations, which

may only replenish over geologic time scales, and increased exchangeable acidity within the soil

matrix (Galloway 2001). Thus, stream acidification entails a suite of geochemical changes

beyond lowered pH, including reduced concentrations of base cations (Likens et al 1996), and

increases in aluminum (Gensemer and Playle 1999) and sulfate (Tomlinson 2003). In SNP

streams, we found a strong positive correlation between stream pH and SBC, and the latter better

related the acidification gradient to chlorophyll a standing crops, SRP and DOC than did pH

83

itself because SBC concentrations within each stream were more broadly distributed over time

than were measurements of pH.

SRP concentrations in acid streams were low and did not increase from April to July as

observed for circumneutral streams. Acid streams can be a sink for inorganic phosphorus inputs

through abiotic complexes formed with hydroxides of aluminum and iron, which are often

present in greater amounts in acidified surface waters (Norton et al. 2006). Though we did not

quantify metal concentrations, the relationship between NO3-N and SRP concentrations differed

between circumneutral (positive relationship) and acidic (no relationship) streams and may point

to stronger abiotic control over P in acid streams. Multi-year records of weekly NO3-N and SRP

concentrations in two forested reference streams in eastern Tennessee revealed identical seasonal

patterns in these two solutes largely due to biota (Mulholland and Hill 1997); thus, the absence

of this linkage in the acid SNP streams suggests different controls over P, possibly due to altered

abiotic processes of sorption and occlusion.

Similar to previous findings in other acidified streams (e.g. Mulholland et al. 1986), we

observed greater chlorophyll a biomass with increasing acidity, and these differences were

maintained despite various physical and chemical influences. Chl a was greater in the two acid

streams throughout the study period with the exception of the first date, when all streams were

equally low, and in early May when chl a biomass in both acid streams declined by 57% over the

ten day period since the previous sample collection. Chl a biomass then returned to, or exceeded,

the previously high levels by the next sampling date 14 days later.

The sudden loss and return of producer biomass in the acid streams was indicative of a

hydrologic event that may have caused sloughing (Roberts et al. 2007); subsequently, three

major storms were found to have occurred over the sampling period (indicated by the vertical

84

arrows in Fig. 3a; determined from daily discharge measurements provided by the USGS at

nearby stations, www.usgs.gov). One such storm occurred just two days prior to the date we

observed the sharp decline in producer biomass. Another biomass decline in the acid streams was

observed in mid-June, most likely related to increases in canopy cover (Mulholland et al. 2006)

since no storms occurred during this time. Despite the strong influences of high storm flow and

shading, chl a biomass either returned to previous abundance or generated greater areal

abundance in the acid streams than in the circumneutral streams. Storms or shading may have

had far less influence on periphyton assemblages in the circumneutral streams because standing

crops were low throughout the monitoring period.

Interestingly, producer biomass was greater in acid streams despite lower concentrations of

phosphorus (as SRP) in these systems. Management of phosphorus inputs to lakes has been

deemed essential to controlling harmful algal blooms (Schindler et al. 2008), and concentrations

of inorganic forms of both N and P are used extensively as criteria to characterize trophic state

(autotrophic vs. heterotrophic) and evaluate the potential for aquatic eutrophication (Dodds

2007). Thus, we would expect positive associations between chl a and SRP in otherwise

unaltered systems; the decoupling of this relationship in acidified streams may be due to the

modification of various other ecological interactions (e.g., competition, herbivory) proposed to

increase producer biomass (Rosemond et al. 1993; Niyogi et al. 1999) given the documented

influence of acidity on animal consumers (Elwood and Mulholland 1989).

The autotrophic index (AI) provided another structural indicator of increased autotrophic

presence in the acid streams. AI values were generally higher in acid streams throughout the

spring study period, meaning that autotrophic biomass constituted a greater proportion of total

periphyton biomass in these systems. Epilithic OM was highly variable within streams over the

http://www.usgs.gov/

85

study and few differences were apparent among streams. Therefore, the higher AI values in acid

streams probably reflect the generally higher chlorophyll content of epilithic assemblages and the

negative influence that acid has on heterotrophs described by others (e.g. Griffith and Perry

1994). In support of this contention for streams of the SNP, we found large declines in fungal

biomass with increasing acidity in previous investigations of these same streams during autumn

(Ely et al. 2010). Thus, we interpret these structural differences as evidence of the enhancement

of stream autotrophs during the spring period of high solar irradiance.

These findings suggest an intrinsic propensity for autotrophic stimulation in acidified

streams consistent with our expectations of enhanced endogenous organization of these systems.

Our objective was to understand how the enhancement of autotrophs influenced metabolic and

biogeochemical processes in these systems. Predominance of endogenous processes driven by

primary producers should manifest in terms of system energetic and material demands. From a

stoichiometric perspective, plasticity in C:N ratios of autotrophs means these organisms are

capable of luxury uptake of nutrients beyond that expected from C:N ratios of structural biomass

alone (Sterner and Elser 2002). Given constant biomass, increased stoichiometric flexibility of

primary uptake compartments (i.e., greater autotrophic presence) should result in altered, and

most likely accelerated, nutrient demand (Cross et al. 2005). Others have shown linkage between

autotrophic processes and material demand. GPP was found to play a much stronger role in NO3-

N uptake than ER as determined from 15

N tracer studies in 69 streams distributed across the

United States and Puerto Rico (Hall et al. 2009). Other studies have found either positive

relationships (Fellows et al. 2006; Mulholland et al. 2006; Newbold et al. 2006) or a lack of

relationship (Hoellein et al. 2007; Johnson et al. 2009) between GPP and uptake metrics for

either NO3-N or NH4-N using both enrichment and tracer studies. The study by Hall et al.

86

(2009), however, provides the strongest evidence of linkage between GPP and nutrient demand

to date due to their highly consistent approach using 15

N tracers across a broad geographic area

as part of the Lotic Intersite Nitrogen eXperiment (LINX, www.biol.vt.edu/faculty/webster/linx).

Before our additions of labile carbon (i.e., under ‗ambient‘ conditions), the uptake length of

NO3-N was ~20% shorter in the acid stream, where chl a biomass was almost 3 times higher than

in the circumneutral stream. However, the circumneutral stream had greater discharge, which

extends spiraling length (Stream Solute Workshop 1990) making comparisons of Sw between

these systems less informative. The uptake velocity, vf, corrects for differences in current

velocity and depth and is the preferred metric for comparisons of solute uptake among streams

(Ensign and Doyle 2006). We found that uptake velocity of NO3-N was only slightly higher

(9.5%) in the acid stream under ambient conditions. The similarity in vf between streams led to

nearly twice the rate of areal NO3-N uptake in the circumneutral stream reflecting increased

NO3-N availability. Thus, despite a greater autotrophic proportion of total epilithic biomass we

found little evidence for the enhanced NO3-N demand expected based on stoichiometric

arguments. This result is less surprising given the similar rates of GPP within the acidified and

circumneutral streams and the strong relationships between GPP and NO3-N demand mentioned

above (Hall et al. 2009). Furthermore, lower chlorophyll-specific productivity in the acid stream

indicates a reduction in biomass-specific growth rates and may counteract any advantages of

stoichiometric flexibility in the ability to acquire excess N. Mulholland et al. (1986) also found

consistently lower chlorophyll-specific productivity in a highly acidic stream within the Great

Smoky Mountains National Park. They suggest that reductions in grazing pressure due to lower

densities of algal-scraping mayflies may have led to self-inhibition of chlorophyll-specific

production due to increased self-shading. These findings indicate that the relative turnover rate of

87

autotrophic biomass is lower in acid streams and may explain the few differences we observed in

nutrient demand.

Our addition of a labile form of organic carbon led to dramatic increases in the nutrient

demand of the circumneutral stream while the acid stream remained relatively unaffected.

Glucose additions in the circumneutral stream resulted in 1.3- and 3.2-fold higher vf and U,

respectively, over the acid stream. Other studies have demonstrated tight coupling of C and N

cycles in highly heterotrophic streams through the addition of organic carbon resources.

Bernhardt et al. (2002) found greater rates of NH4+ and NO3

- uptake and higher CO2 production

in a forested headwater stream in New Hampshire during a 6-week addition of potassium acetate.

Similarly, acetate injections into the hyporheic zone of a New Mexico stream stimulated

microbial respiration and substantially reduced subsurface NO3-N concentrations (Baker et al.

1999). Multiple interactions among the uptake rates of dissolved forms of organic carbon and

both inorganic and organic nitrogen were observed when these solutes were experimentally

introduced in surface and subsurface waters of Hugh White Creek, a forested stream in North

Carolina with extremely low GPP (Brookshire et al. 2005).

We conducted simultaneous additions of 15

N and glucose in two streams with very different

pH to 1) illustrate the relative degree of autotrophy versus heterotrophy using solute dynamics in

addition to traditional analyses of dissolved oxygen and, 2) to address the possible acid-induced

impairment of DOC processing in systems predicted to experience future increases in DOC

loads. While both streams had P:R ratios considerably less than 1, P:R in the acid stream was 2.4

times higher than the circumneutral stream. Accordingly, C uptake length was longer and both

uptake velocity and the areal rate of C uptake were lower in the acid stream. However, Johnson

et al. (2009) found no relationship between benthic chl a biomass and uptake velocities of either

88

acetate or glycine in 18 Michigan streams. Furthermore, Newbold et al. (2009) found positive

relationships between the uptake velocities of glucose and arabinose with both GPP and ER in

New York streams, but no relationship with benthic chl a. Thus, a greater degree of autotrophy

may only be reflected in lower rates of organic C processing when heterotrophic compartments

are depleted. Though epilithic standing crops were similar between streams during our releases,

the acid stream had greater proportions of chl a in this biomass and ER was 69% lower. Along

with our past work showing that fungal biomass declined by more than an order of magnitude

across a pH gradient comparable to that addressed here, these results suggest acid impairment of

heterotrophic compartments of this system.

As terrestrial ecosystems recover from acid deposition, the rising solubility of DOC in soils

is increasing DOC concentrations in the surface waters of glaciated landscapes of North America

and Europe (Monteith et al. 2007). Our findings suggest that the extended depression of surface

water pH may continue to impair the ability of acidified streams to process greater DOC loads,

which could have consequences for downstream energetics (Wetzel 1995), drinking water

supplies (Oulehle and Hruška 2009), and regional carbon budgets (Siemens 2003). Further, the

importance of this impairment may be larger during autumn, when substantial pulses of DOC in

the water column result from the leaching of fallen leaves (Meyer et al. 1998). Without knowing

long-term DOC concentrations in our study streams, we may only speculate that the lower DOC

concentrations we observed in the two acidified streams are indicative of continued impairment

of terrestrial soils in these catchments, possibly due to their more acid-sensitive, siliciclastic

lithology (Cosby et al. 2006).

Greater endogenous organization of an ecosystem results in stronger controls on functional

processes by factors associated with thermal regime (i.e., light and heat) than resource subsidies

89

(i.e., imported organic carbon) (Valett et al. 2008). Accordingly, comparisons of both the

structural and functional attributes of autotrophic and heterotrophic compartments often reflect

the relative degree of endogenous versus exogenous control, respectively, within a system.

Previous findings of greater producer biomass and production in low pH streams suggests greater

endogenous influence in these systems and led us to ask how this enhancement affects nutrient

cycling during spring, when forested streams typically display the highest annual rates of GPP.

We found greater algal abundances and a higher P:R ratio in acid versus circumneutral

streams; however, despite a shorter uptake length of NO3-N in the most acidic stream, uptake

velocity (vf) and areal uptake (U) were both greater in the most circumneutral stream. Lower

chlorophyll-specific production found in the acid stream may have led to lower N demand than

expected based on the greater abundance of producers in this system. The acid stream also

displayed a reduced capacity to process experimentally-added labile DOC, apparently related to

the impairment of heterotrophic compartments. Our findings suggest that anthropogenic stream

acidification can alter the degree of internal versus external influence on metabolic and

biogeochemical processes in lotic systems.

90

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Table 1. Average daily (24h) temperature, mean concentrations of various dissolved solutes

and mean standing crops of chlorophyll a and epilithic biomass in the study streams from 12

April 2009 to 11 July 2009. Values in parentheses are (minimum, maximum) for temperature,

and (SE, n) for all others. Values within the same row that share a similar letter are not

significantly different by a Tukey multiple comparison test following a significant one-way

analysis of variance (α = 0.05) unless indicated otherwise. SRP = soluble reactive phosphorus.

DOC = dissolved organic carbon. SBC = sum of base cations. AFDM = ash-free dry mass.

Meadow Run Paine Run Hughes River Piney River

Average daily

temperature (°C) 13.4 (7.2, 17.2) 13.0 (7.5, 16.3) 13.8 (7.1, 17.9) 13.8 (6.7, 17.7)

NH4-N (μg·L-1

) 4a (0.2, 27) 5

b (0.4, 24) 6

b (0.4, 21) 5

b (0.2, 27)

NO3-N (μg·L-1

)* 29a (3.8, 27) 15

a (2.1, 24) 35

a (7.5, 21) 59

b (5.8, 27)

SRP (μg·L-1

)* 1a (0.1, 27) 2

a (0.2, 24) 7

b (0.6, 21) 6

b (0.3, 27)

DOC (mg·L-1

)* 0.95a (0.02, 25) 0.95

a (0.02, 24) 1.11

b (0.01, 21) 1.34

b (0.03, 25)

Ca2+

(mg·L-1

) 0.84a (0.03, 18) 0.87

a (0.05, 17) 1.84

b (0.09, 14) 1.86

b (0.10, 14)

Mg2+

(mg·L-1

) 0.39ab

(0.01, 18) 0.44b (0.02, 17) 0.34

a (0.02, 14) 0.42

b (0.02, 14)

K+ (mg·L

-1)* 0.99

ab (0.04, 18) 1.49

b (0.06, 17) 0.34

a (0.01, 14) 0.37

a (0.03, 14)

Na+ (mg·L

-1)* 0.63

ab (0.05, 18) 0.56

a (0.01, 17) 1.42

bc (0.04, 14) 1.51

c (0.05, 14)

SBC (μeq·L-1

) 126a (2, 18) 142

a (6, 17) 191

b (6, 14) 203

b (8, 14)

Chlorophyll a

(mg·m-2

) *

4.9a (0.6, 24) 4.6

a (0.6, 24) 1.7

b (0.2, 21) 2.1

b (0.2, 24)

Epilithic biomass

(g AFDM·m-2

) *

2.70a (0.21, 24) 2.23

a (0.15, 24) 2.30

a (0.21, 21) 3.19

a (0.27, 24)

*Means were compared using Dunn‘s multiple comparison following a significant Kruskal-

Wallis one-way ANOVA on ranks

97

Table 2. Stream pH and nutrients, photosynthetically active radiation (PAR) and flow conditions

of study streams during 15

N releases with (+ DOC) and without glucose additions. SRP = soluble

reactive phosphorus.

Meadow Run Meadow Run

+ DOC

Piney River Piney River

+ DOC

26 June 28 June 3 July 6 July

pH and nutrients

pHa 5.51 5.43 6.93 6.91

NO3-N (μg·L-1

± SE)b 40 ± 0.4 41 ± 0.2 90 ± 0.3 100 ± 1.5

NH4-N (μg·L-1

± SE)b 4 ± 0.6 4 ± 0.3 5 ± 0.7 6 ± 0.9

SRP (μg·L-1

± SE)b 1 ± 0.2 1 ± 0.1 7 ± 0.1 8 ± 0.2

Light and temperature

PAR (mol·m-2

·d-1

± SE)c 10.1 ± 1.1 5.5 ± 0.2 8.7 ± 0.6 8.2 ± 0.8

Mean 24h temp

°C (min, max)d

16.7

(15.9, 17.7)

16.7

(15.8, 17.6)

16.3

(15.8, 16.8)

16.2

(15.1, 17.2)

Flow conditions

Width (m)e 4.9 ± 0.2 4.9 ± 0.2 6.9 ± 0.1 6.9 ± 0.1

Depth (cm)f 16.2 ± 1.0 14.9 ± 0.8 26.6 ± 1.6 26.5 ± 1.4

Current velocity (m/min)g 5.0 4.8 3.5 3.2

Average discharge (L·s-1

)h 65.4 57.8 107.0 97.4

Lateral inflow (%·m-1

) 0.42 0.34 0.51 0.47 aIndividual readings prior to releases

bn = 3

cMeasured every 15 minutes and averaged for all daylight hours

dMeasured once per hour

en = 10

fn = 100

gDetermined from conservative tracer breakthrough curves

hCalculated from measurements at four transects using mass balance of conservative tracer

98

Table 3. Stream periphyton, metabolism and carbon (C) uptake metrics measured during 15

N

tracer experiments. Periphyton was sampled following the C enrichment experiments.

Calculations of GPP and ER were made over the 24h period immediately following sonde

deployment and did not encompass C enrichment experiments. AFDM = ash-free dry mass. GPP

= gross primary production. ER = ecosystem respiration. n = 3 for periphyton data.

Variable

Meadow Run Piney River

Periphyton

Chlorophyll a

(mg· m-2

± SE)

3.72 ± 0.26 1.28 ± 0.07

Epilithic biomass

(mg AFDM· m-2

± SE)

3137 ± 757 4626 ± 112

Autotrophic index

(chl a/epilithon ± SE)

1.34 ± 0.32 0.28 ± 0.02

Metabolism

GPP (g O2·m-2

·d-1

) 0.308 0.405

ER (g O2·m-2

·d-1

) 1.826 5.813

P:R ratio 0.17 0.07

Chlorophyll-specific production

(g O2·mg-1

chl a·d-1

)

0.083 0.316

Carbon uptake

C uptake length (m) 160 68

C uptake velocity (mm·s-1

) 0.075 0.208

C areal uptake (mg C·m-2

·s-1

) 0.061 0.210

99

Fig. 1

Figure 1. Measurements of a) stream pH, and concentrations (± SE) of b) NH4-N, c) NO3-N, d)

soluble reactive phosphorus (SRP) and e) dissolved organic carbon (DOC) in SNP streams over

the spring study period. Symbols for different streams are: (□) Meadow Run; (◊) Paine Run; (●)

Hughes River; (▲) Piney River.

100

Fig. 2

Figure 2. Correlations of the charge sum of base cations (microequivalents per liter) with a)

chlorophyll a, b) soluble reactive phosphorus (SRP) and c) dissolved organic carbon (DOC) in

SNP streams. n = 22 for all correlations. Symbols for the different streams are as described in

Fig. 1.

101

Fig. 3

Figure 3. Standing crops of a) chlorophyll a and b) epilithic biomass, and c) autotrophic index

values in the study streams over the spring study period. Symbols for the different streams are as

described in Fig. 1. The vertical arrows in (a) represent severe storm events as determined from

nearby USGS gauging stations (www.usgs.gov).

http://www.usgs.gov/

102

Fig. 4

Figure 4. a) Uptake lengths (Sw), b) uptake velocities (vf) and c) areal uptake rates (U) for NO3-N

in Meadow Run (pH 5.5) and Piney River (pH 6.9) under ambient (open bars) and DOC-

enriched (solid bars) conditions. K15

NO3 was used as a tracer to calculate uptake metrics;

glucose (C6H12O6) additions were designed to raise DOC concentrations at the head of the reach

approximately 5-fold over background levels.

103

Chapter 5: Synthesis

In the previous chapters I attempt to understand ecosystem response to a large-scale shift in

environmental conditions based on the premise that the collective physiological response of key

organisms dictates ecosystem-level behavior. This approach integrates multiple hierarchical

levels of ecosystem organization, from molecular resources to whole-system function, to detect

anthropogenic impacts and better understand the far-reaching consequences of subtle changes in

environmental conditions. I contend that this approach provides a more useful framework for

investigation of ecosystem response to the global changes presently confronting Earth systems

than disturbance response, which has typically operated on the concepts of resistance and

resilience that comprise recovery. Many of the environmental changes occurring at present are

irreversible, directional changes such as rising greenhouse gases in the atmosphere, dwindling

natural resources to support the growing human population, increasing availabilities of nutrients

across the landscape, and falling pH of the world‘s oceans. As such, the notion of discrete

disturbance events from which systems recover is not applicable, and we may better anticipate

the future functioning of ecosystems through consideration of the physiological adjustments

demanded of resident biota.

The study of heterotrophic leaf biofilm functions of respiration and N uptake verified stream

acidification as an ecosystem stressor through the observed increases in qCO2 with decreasing

stream pH. Increasing stress was surprisingly related to increases in qN, suggesting demands for

N unrelated to growth and possibly related to increased exoenzyme production due to greater

energetic needs as part of the stress response. The study also suggested that large declines in

consumer biomass may counteract qN, resulting in no observable differences in ecosystem-level

rates of N uptake.

104

The study applying qCO2 and qN metrics to whole-stream assessments of respiration and N

uptake showed similar patterns relating these metrics to the pH gradient as those found in

microcosm experiments. Large declines in microbial consumer biomass and ecosystem

respiration were found with increasing acidification and no differences were found in various N

uptake metrics as a result of the increases in qN. qCO2 also reflected gradients in nutrient

availability based on other published studies, and a strong relationship was found between qCO2

and qN among streams covering a broad geographic range, which mirrored the patterns found

under highly controlled microcosm experiments. These findings confirm qCO2 as a reliable

indicator of ecosystem stress and provide valuable insight into biomass-specific controls on N

cycling in lotic ecosystems.

Lastly, the study investigating acidification influences on metabolism and coupled C and N

cycling during spring showed how the simultaneous enhancement of autotrophic biomass and

declines in heterotrophic components of epilithon in an acidic stream resulted in a reduced

capacity to process organic carbon and a decoupling of C and N cycling. Similar to whole-

system experiments in autumn, few differences were observed in N uptake rates between an

acidic and a circumneutral stream under ambient conditions, however N uptake in the

circumneutral stream was greatly enhanced upon addition of glucose whereas no change was

observed with glucose enrichment in the acid stream. These findings suggest fundamental shifts

in the biogeochemical behavior of acidified streams that are related to the physiological

responses of primary uptake compartments.

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