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Photo- and bio-reactivity patterns of dissolved
organic matter from biomass and soil leachates
and surface waters in a subtropical wetland
Meilian Chen, Rudolf Jaff
e
*
Southeast Environmental Research Center and Department of Chemistry and Biochemistry,
Florida International University, Miami, FL 33199, USA
article info
Article history:
Received 10 December 2013
Received in revised form
26 March 2014
Accepted 29 March 2014
Available online 22 May 2014
Keywords:
DOM
Photo-reactivity
Bio-reactivity
EEM-PARAFAC
Fluorescence
Wetland
abstract
Dissolved organic carbon (DOC) measurements and optical properties were applied to
assess the photo- and bio-reactivity of dissolved organic matter (DOM) from different
sources, including biomass leaching, soil leaching and surface waters in a subtropical
wetland ecosystem. Samples were exposed to light and/or dark incubated through
controlled laboratory experiments. Changes in DOC, ultraviolet (UV-Vis) visible absor-
bance, and excitation-emission matrix (EEM) fluorescence combined with parallel factor
analysis (PARAFAC) were performed to assess sample degradation. Degradation experi-
ments showed that while significant amounts of DOC were consumed during bio-
incubation for biomass leachates, a higher degree of bio-recalcitrance for soil leachate
and particularly surface waters was displayed. Photo- and bio-humification trans-
formations were suggested for sawgrass, mangrove, and seagrass leachates, as compared
to substantial photo-degradation and very little to almost no change after bio-incubation
for the other samples. During photo-degradation in most cases the EEM-PARAFAC com-
ponents displayed photo-decay as compared to a few cases which featured photo-
production. In contrast during bio-incubation most EEM-PARAFAC components proved to
be mostly bio-refractory although some increases and decreases in abundance were also
observed. Furthermore, the sequential photo- followed by bio-degradation showed, with
some exceptions, a “priming effect”of light exposure on the bio-degradation of DOM, and
the combination of these two processes resulted in a DOM composition more similar to
that of the natural surface water for the different sub-environments. In addition, for
leachate samples there was a general enrichment of one of the EEM-PARAFAC humic-like
component (Ex/Em: <260(305)/416 nm) during photo-degradation and an enrichment of a
microbial humc-like component (Ex/Em: <260(325)/406 nm and of a tryptophan-like
component (Ex/Em: 300/342 nm) during the bio-degradation process. This study exem-
plifies the effectiveness of optical property and EEM-PARAFAC in the assessment of DOM
reactivity and highlights the importance of the coupling of photo- and bio-degradation
processes in DOM degradation.
©2014 Elsevier Ltd. All rights reserved.
*Corresponding author. Southeast Environmental Research Center, Florida International University, 11200 SW 8th Street, Miami, FL
33199, USA. Tel.: þ1 305 348 2456; fax: þ1 305 348 4096.
E-mail address: jaffer@fiu.edu (R. Jaff
e).
Available online at www.sciencedirect.com
ScienceDirect
journal homepage: www.elsevier.com/locate/watres
water research 61 (2014) 181e190
http://dx.doi.org/10.1016/j.watres.2014.03.075
0043-1354/©2014 Elsevier Ltd. All rights reserved.
1. Introduction
DOM plays diverse important ecological and environmental
roles both on local ecosystem scales and globally. Locally, it
serves as energy source for heterotrophic bacteria and thus
fuels the microbial loop, and acts as a natural sunlight
attenuator and pH buffer for aquatic ecosystems. Further-
more, as a carrier for organic and inorganic xenobiotics, DOM
can impact the transport and fate of environmentally signifi-
cant pollutants. Globally, as one of the largest and mobile
carbon pools on Earth, the biogeochemical cycling of DOM is
intricately associated with nutrient and element cycling and
as well climate change.
DOM is ultimately produced from biomass like higher
plants and phytoplankton or leached from soils and sedi-
ments. Dominant biomass and soils are believed to be main
sources of DOM in the Florida coastal Everglades FCE (Maie
et al., 2006; Yamashita et al., 2010). Previous studies reported
that biomass leaching was fast for senescent leaves (Maie
et al., 2006; Scully et al., 2004) and the incorporation of
biomass leachates along the Everglades water flow path has
clearly been demonstrated (Yamashita et al., 2010).
Better understanding the degradation and transformation
processes of DOM has been a topic of particular interest to
both ecologists and biogeochemists for decades. So far, photo-
and bio-degradation are regarded as two major processes to
transform and mineralize DOM (Obernosterer and Benner,
2004; Benner and Kaiser 2011), and heterotrophic bacteria
have a proven ability to readily uptake labile DOM molecules
such as free amino acids and carbohydrates. While the role of
microorganisms in DOM degradation has long been recog-
nized, the importance of solar irradiation has also been indi-
cated in the literature through a wealth of reports (Moran and
Zepp, 2000; Scully et al., 2004; Ortega-Retuerta et al., 2010;
Shank et al., 2010; Benner and Kaiser 2011;Helms et al., 2013;
Lu et al., 2013; Rossel et al., 2013; Lønborg et al., 2013). As such,
solar radiation has been reported to result in photo-
mineralization, photo-ammonification, photo-bleaching of
chromophoric DOM (CDOM), or in photo-humification, photo-
production of new DOM, and photo-formation of reactive
oxygen species (Moran and Zepp, 2000; Shank et al., 2010). The
photo-reactivity of DOM seems highly related to the degree of
aromaticity (Helms et al., 2013) and long-term photo bleaching
was reported to mineralize DOM nearly to completion
(V€
ah€
atalo and Wetzel, 2008). It is not surprising then, that in
clear and shallow wetland waters, photochemical trans-
formations of DOM can be significant (Cawley et al., 2012). In
addition, while DOM leached from wetlands vegetation was
found to be photo-reactive (Scully et al., 2004), only limited
bioavailability of DOM has been reported in wetlands such as
the Everglades (Qualls and Richardson, 2003) and attributed in
part to nutrient limitations. DOM bioavailability was also
observed to be controlled by source, where protein-like ma-
terials were degraded much more efficiently compared to
polyphenols (Scully et al., 2004). It is likely however, that
photo- and bio-degradation processes of DOM act in tandem,
and indeed photo-exposure can increase, decrease, or have no
net effect on biodegradability of DOM depending on DOM
source (Moran and Zepp, 2000). There is however, a general
consensus that sunlight has a “priming effect”for bacterial
uptake on old, aromatic, terrestrial-derived DOM, while the
opposite effect has been observed on fresh, non-aromatic,
algae-derived DOM (Moran and Covert, 2003).
Fluorescence spectroscopy has been successfully used to
characterize DOM sources and opens new windows to study
DOM dynamics in aquatic ecosystems (Jaff
e et al., 2008; Helms
et al., 2013). By decomposing the EEM data statistically into
different fluorescent components, EEM-PARAFAC has been
successfully applied in DOM studies (Stedmon al., 2003; Cory
and McKnight, 2005; Miller et al., 2009; Jaff
e et al., 2012),
including photo-degradation and bio-degradation studies
(Cawley et al., 2012; Lu et al., 2013; Fellman et al., 2009; Cory
and Kaplan, 2012).
The Florida coastal Everglades (FCE) is one of the largest
wetlands in the world and is undergoing a historic restoration
which aims to restore the quality, quantity, timing, and dis-
tribution of water flow (http://www.evergladesplan.org/).
DOM in this phosphorus-limited oligotrophic subtropical
wetland primarily originates from local vegetation and soil
OM oxidation. The majority of nitrogen (N) and phosphorous
(P) in the Everglades are in organic forms and therefore asso-
ciated with the DOM (Boyer et al., 1997). Therefore, the pro-
cesses controlling DOM transformation and turnover in the
FCE are critical in driving local nutrient cycling. Considering
the complexity of DOM sources in wetland ecosystems and its
ecological importance, a better understanding of the degra-
dation mechanisms controlling its transformation and ulti-
mate fate are needed. The objectives of this study were to
characterize DOM leached from biomass and soils, and to
assess photo- and bio-degradation processes comparatively to
surface water DOM from across the Everglades using optical
properties.
2. Sampling and experimental methods
2.1. Sampling
Biomass, soils and surface waters from this study were
collected at stations used in the on-going, long-tern ecological
research program of the Florida coastal Everglades (FCE-LTER;
http://fcelter.fiu.edu/). Fig. 1 displays the locations of the
sampling sites, which represent diverse sub-environments of
this ecosystem (freshwater marsh-peat eSRS3; freshwater
marsh-marl eTS/Ph2; fringe mangrove eSRS5; seagrass-
dominated coastal bay eTS/Ph11). Senescent plant mate-
rials from Cladium jamaicense (sawgrass), Elocharis cellulosa
(spikerush), Rhizophora mangle (red mangrove), fresh shoots of
Thalassia testudinidum (seagrass), floating periphyton, top soils
(including SRS peat soil, TS/Ph marl soil, mangrove mud, and
Florida Bay calcitic marl sediment), and surface water samples
from different sub-environments were collected in the FCE.
Sawgrass (from TS/Ph2) and spikerush (from SRS3) samples
were above-water senescent leaves. Floating periphyton at
site TS/Ph2 was sampled, which was entangled with Utricu-
laria. Mangrove samples consisted of yellow leaves hand-
picked from trees at site SRS5. Seagrass samples consisted of
whole shoots from site TS/Ph11. The five biomass materials
water research 61 (2014) 181e190182
and corresponding surface water samples were obtained in
July 2007 (wet season) and the four soil samples and corre-
sponding surface water samples were collected in February
2008 (dry season). Surface water from an agricultural drainage
canal (C111) was also sampled in February 2008. A suite of
Everglades surface water samples used for comparison pur-
pose (untreated) were collected from 14 FCE-LTER stations in
August 2007 and January 2008 (Fig. 1). Biomass samples were
hand-picked using pre-cleaned powder-free latex gloves
(Fisher). Top soils/sediments were collected with a pre-
cleaned small stainless steel shovel. Before collecting the
surface water samples, the bottles were rinsed three times
with surface water. The samples were placed into pre-cleaned
zip lock bags (Lab Safety Supply), pre-combusted glass jars (I-
CHEM), and pre-cleaned acid-washed brown high-density
polyethylene bottles (Nalgene) for biomass, soil, and surface
water, respectively.
All samples were stored on ice immediately after collec-
tion, transported to the laboratory within 12 h, and stored
under refrigeration until further processing. Surface water
was sequentially filtered through pre-combusted 0.7
m
m GF/F
to remove larger particles, followed by 0.22
m
m filters (Dura-
pore, Millipore) to remove bacteria and minimize microbial
activity during photo-degradation, and then stored refriger-
ated for further use within 3 days. Plant samples were rinsed
with Milli-Q water upon arrival in the lab. Biomass and soil
samples were stored refrigerated until leaching experiments
were performed (within 2 days) to ensure proper preservation
and minimize microbial growth.
2.2. Leaching experiments
Since ionic strength is an important factor affecting DOM
leaching and reactivity (Grebel et al., 2009), natural surface
water from each corresponding site was used for the leaching
experiments of local soils and biomass to better mimic the
biogeochemical conditions of the natural environment. The
background DOM of these water samples was reduced using a
pre-cleaned activated carbon filtration with Carbon-cap 150
(Whatman) cartridges. The activated carbon column was
extensively pre-washed with a large amount of 0.01 M hydro-
chloric acid (certified A.C.S. plus, Fisher) and then followed by
large amounts of Mill-Q water, until the background DOC from
the column effluent was below 0.02 mg C/L. Total phosphorous
(TP), total nitrogen (TN), and EEM were also monitored to
ensure no potential contamination or fertilization of the in-
cubations took place. Three dead volumes of column effluent
were discarded before collecting the water for leaching pur-
poses. The DOM-reduced water was sequentially filtered with
0.7
m
m GF/F and 0.22
m
m sterile filters. The carbon column was
replaced after every 8 L of natural water to ensure maximum
absorption efficiency. After completion of the activated carbon
filtration process to reduce background DOM, the remaining
background DOC from the original surface water was <5%
(DOC
remaining in background
/DOC
initially in leachate sample
100%),
except for the remaining DOC background in soil leachate from
TS/Ph 2 site which was ~10%.
All substrates were leached for a 24 h period. Biomass and
soil (about 20 g/L, 50 g/L, and 380 g/L wet weight for higher
plants, periphyton, and soil, respectively) were submerged
into pre-combusted 4 L Erlenmeyer flasks containing 2 L of
DOM-reduced reference water. Sawgrass and spikerush leaves
were cut into small pieces (~3 inches) to facilitate leaching.
Flasks were covered and wrapped with aluminum foil to avoid
photo-degradation during the leaching process. The flasks
were placed onto a shaker for soil leaching to enhance
desorption and diffusion while they were static for the
biomass. The room temperature during the 1-day leaching
Fig. 1 eSampling site locations for the Florida coastal Everglades (FCE).
water research 61 (2014) 181e190 183
experiments was about 21 C. Chemicals commonly used to
control bacterial activity were not added during the leaching
experiments considering the relative short leaching time as
well as the potential interfering effects on DOM optical prop-
erties. The leachates were sequentially filtered through
0.7
m
m GF/F and 0.22
m
m sterile filters and further diluted with
background DOM-reduced surface water to the average
UVeVis absorbance of typical FCE surface water
(~0.2e0.5 cm
1
at
l
¼280 nm) to minimize differences in op-
tical density. The obtained leached DOM samples were split
into 2 aliquots, and the photo- and bio-degradation experi-
ments started on the same day.
2.3. Photo-degradation experiments
A bench top SunTest XLS (ATLAS) solar simulator was used for
the photo-degradation experiments. The SunTest emits a full-
spectrum between 300 and 800 nm via a xenon lamp. The light
intensity was measured with a spectrophotometer (Ocean
Optics) on a daily basis and the lamp performance was found
to be stable throughout the time of the experiments. The
SunTest intensity was 765 W/m
2
equivalent to mid-day sun-
light on a sunny summer day in South Florida. One day of
SunTest treatment was estimated as equivalent to about a 4-
day sunlight dose in the FCE environment during summer.
For the photochemical exposure experiments, the leachates
were placed into 150 ml pre-combusted beakers and covered
with pre-combusted quartz covers. Dark controls were wrap-
ped with layers of aluminum foil. Sample-containing beakers
together with the blanks were placed into the water bath of
the exposure chamber of the SunTest. The water bath was
circulated pumping cool water to stabilize the water bath
temperature (~23e26 C) during the 7-day irradiation. Samples
were neither poisoned nor filtered during the 7-day irradiation
period to minimize potential interference or contamination.
Considering the samples were sterile filtered before the ex-
periments (0.22
m
m) and were subjected to very intense UV
irradiation, the bacterial activity during the photo-exposure
experiments is expected to be limited. Triplicate aliquots
were removed from the SunTest chamber on day 0 and day 7.
Samples were filtered through GF/F and 0.22
m
m filters for
consistency, and measurements of optical properties were
performed.
2.4. Bio-degradation experiments
Original unfiltered surface water from each corresponding site
was added as inoculum at a ratio of 1:100, in addition to final
concentrations of 10
m
MNH
4
NO
3
and 1
m
MKH
2
PO
4
to avoid N
and P limitations. After complete mixing, leachates were split
into pre-combusted glass jars, which were wrapped in
aluminum foil and then incubated in the dark for up to 28
days. Jars were opened every other day to aerate, and large
headspaces were left to ensure enough oxygen for the bio-
incubations. Triplicate aliquots were removed on days 0 and
28. Likewise, samples were filtered through GF/F and 0.22
m
m
filters prior to DOC, UVeVis, and EEM measurements for
consistency with the other samples.
2.5. Sequential photo- and bio-degradation experiments
Triplicate samples removed after the 7-day photo-incubation
in the SunTest were filtered with 0.22
m
m sterile filters,
spiked with an inoculum and 10
m
MNH
4
NO
3
and 1
m
MKH
2
PO
4
as described above. After complete mixing, the samples were
placed into pre-combusted glass jars, which were wrapped in
aluminum foil and then incubated in the dark for up to 28
days. Sample treatment and analysis were the same as
described above.
2.6. Analytical measurements
DOC, UVeVis and EEM fluorescence spectra were determined
for each sample. The UVeVis absorbance was measured using
a Varian Cary 50 bio spectrophotometer with a 1 cm quartz
cuvette scanning from 240 nm to 800 nm. The EEMs were
measured using a Horiba JovinYvon SPEX Fluoromax-3 spec-
trofluorometer equipped with a 150 W continuous output Xe
arc lamp. Slits were set at 5.7 nm for excitation and 2 nm for
emission. Forty-four emissions scans were acquired at exci-
tation wavelength (
l
ex
) between 240 and 455 nm at 5 nm steps.
The emission wavelengths were scanned from
l
ex
þ10 nm to
l
ex
þ250 nm (i.e., between 250 and 705 nm) in 2 nm steps
(Yamashita et al., 2010). Fluorescence signals were acquired in
signal over reference ratio mode (S/R) to eliminate potential
fluctuations of the Xe lamp. While all samples were subjected
to inner filtering effect correction, and leachates were diluted
to minimize differences in optical density, some bias between
leachates and surface water samples may still exist. More
detailed information of post-acquisition steps for correction
and standardization can be found elsewhere (Chen et al.,
2010). DOC concentrations were measured using the high-
temperature catalytic combustion method with a Shimadzu
TOC-V total organic carbon analyzer.
2.7. PARAFAC model fitting and residue components
validation
PARAFAC is a statistical tool based on an alternating least
square algorithm. It can statistically decompose EEMs into
fluorescent groups (Stedmon et al., 2003). PARAFAC modeling
can be applied either by creating and validating the model
using the complete dataset of EEMs (e.g., Stedmon et al., 2003)
or by fitting the EEMs to an already established PARAFAC
model (e.g., Cawley et al., 2012). Here all biomass and soil
leachates and surface water EEMs were fitted to existing FCE
surface water PARAFAC model (Chen et al., 2010; Yamashita
et al., 2010). The PARAFAC component spectral characteris-
tics and split-half validation data can be seen in Chen et al.
(2010). The analysis was carried out in MATLAB 7.0.4 (Math-
works, Natick, MA) with the DOMFluor toolbox (Stedmon and
Bro, 2008). Residual components (R) obtained by subtracting
the modeled EEMs from the original EEMs, were selected as
additional PARAFAC components for biomass and soil leach-
ates if they featured actual spectroscopic patterns and if their
fluorescence signal was at least three times the background
noise. The % average noise residue was estimated from the
suite of FCE surface water samples presented here.
water research 61 (2014) 181e190184
3. Results and discussion
3.1. Production of DOM from biomass and soil leaching
and their optical characteristics
The data for the incubation experiments is presented in
Tables 1 and S1. Clearly, the leaching of DOM was efficient for
biomass (about 20 g/L for higher plants and 50 g/L for wet
periphyton for leaching experiments), with DOC values be-
tween 19.5 and 56.2 mg C/L after background DOC subtraction.
The % soil OM was determined to be 88%, 59%, 12%, and 3% for
SRS3, SRS5, TS/Ph2, and TS/Ph11 soils, respectively (see http://
fce.lternet.edu/). Soil leachates (about 380 g/L wet weight) had
DOC values ranging from 1.3 to 11.7 mg C/L. Due to variations
in soil type, water content, particle size distribution, fine root
content, etc, quantitative aspects of the soil leaching experi-
ments are not discussed in a comparative way. A comparison
of SUVA values, an indicator of aromaticity, before degrada-
tion showed that biomass leachates had the lowest values,
ranging between 0.1 and 1.7 m
1
L/mg C, as compared to soil
leachates which had the highest values ranging between 1.1
and 4.4 m
1
L/mg C (Table 1). The fresh seagrass leachate had
the lowest SUVA value at 0.1 m
1
L/mg C.
The EEM data obtained from the biomass and soil leachates
and surface water degradation studies were fitted to a previ-
ously established 8-component PARAFAC model for FCE sur-
face water (Chen et al., 2010; Yamashita et al., 2010). This
approach facilitates the comparison among surface water,
biomass leachate, and soil leachate to an already existing
database (Yamashita et al., 2010). The 8 components were
assigned as six humic-like and two protein-like based on the
comparison of spectral characteristics with previous studies
(Cory and McKnight, 2005; Stedmon et al., 2003). While the
humic-like components C1 (Ex/Em: <260(345)/462), C5 (Ex/Em:
<260(405)/500), and C7 (Ex/Em: 275/326) were found to be
ubiquitous to all the samples from different sources, the
remaining five components were absent in some source ma-
terials (S1). The most noticeable absence was that of C2 (Ex/
Em: <260/454), which is a humic-like component believed to
be a photoproduct and/or photo-refractory (Chen et al., 2010;
Cawley et al., 2012), and has also been found in high abun-
dance in DOM derived from agricultural soils (Yamashita
et al., 2010). C2 was absent in all the initial biomass leach-
ates and also initial marl soil leachates, suggesting that it is
not present in biomass sources or marl soils. The next
noticeable absence is that of C3 (Ex/Em: <260(305)/416), which
is also a humic-like component, and was not found in initial
mangrove, periphyton, and marl soil leachates. C4 (Ex/Em:
<260(305)/376), a microbial humic-like component, was ab-
sent in the initial seagrass and mangrove leachates, while the
protein-like C8 (Ex/Em: <300/342) was absent in initial saw-
grass, seagrass, and mangrove leachates. In addition to these
eight components of the existing FCE PARAFAC model, three
residual PARAFAC components were identified and validated
as described above for biomass leachates and another three
for soil leachates. The excitation/emission wavelengths of all
the components are summarized in S1. Since the residue-
based PARAFAC disappeared quickly during incubations and
were not observed in surface water samples, they are
Table 1 eSummary of DOC, a (254 nm), and SUVA values (mean ±SD, n¼3).
Items DOC
a
(mg C/L) DOC
b
a(254 nm) (m
1
) SUVA
c
(m
1
L/mg C)
Time Before After
photo-
After
bio-
Background Before After
photo-
After
bio-
After
sequential
Before After
photo-
After
bio-
Biomass:
sawgrass 56.2 ±2.0 56.2 ±2.6 32.3 ±1.3 0.8 ±0 161 ±0 217 ±1 207 ±32 210 ±14 1.2 ±0 3.9 ±0.1 2.8 ±0.3
spikerush 38.3 ±1.2 38.2 ±1.6 15.3 ±0.5 0.6 ±0.1 112 ±090±4 107 ±192±3 1.3 ±0.2 2.4 ±0.1 3.0 ±0.1
periphyton 19.5 ±0.7 19.4 ±0.8 3.0 ±0.2 0.2 ±020±112±116±011±1 0.4 ±0 0.7 ±0 2.3 ±0
mangrove 38.5 ±0.2 38.5 ±0.3 10.8 ±0.4 0.4 ±0 147 ±7 234 ±12 158 ±2 183 ±29 1.7 ±0 6.1 ±0.1 5.6 ±0
seagrass 19.6 ±0.8 19.5 ±0.6 5.2 ±0.2 0.3 ±02±055±658±349±4 0.1 ±0 2.8 ±0.1 4.8 ±0
Soils:
SRS3 2.8 ±0.2 2.8 ±0.1 2.7 ±0.1 0.1 ±021±012±020±012±0 3.2 ±0 4.7 ±0 3.2 ±0
SRS5 3.3 ±0.1 3.3 ±0.1 3.2 ±0.2 0.1 ±032±020±031±019±0 4.2 ±0 6.1 ±0 4.2 ±0
TS/Ph2 1.3 ±0.1 1.2 ±0 1.2 ±0.1 0.1 ±013±17±012±26±0 4.4 ±0.1 5.9 ±0 4.3 ±0
TS/Ph11 11.7 ±0.1 11.6 ±0.2 9.0 ±0.2 0.3 ±030±116±222±214±0 1.1 ±0 1.4 ±0.1 1.0 ±0
Surface water:
SRS3 22.1 ±0.3 22.0 ±0.2 21.3 ±0.1 142 ±387±8 135 ±285±6 2.8 ±0 3.9 ±0.1 2.7 ±0
SRS5 12.1 ±0.1 12.0 ±0.2 11.3 ±0.3 106 ±173±7 105 ±172±7 3.8 ±0 6.1 ±0.1 4.2 ±0.1
TS/Ph2 6.8 ±0 6.8 ±0 6.7 ±0.1 47 ±130±440±023±0 3.0 ±0 4.4 ±0.1 2.6 ±0
TS/Ph11 3.9 ±0.1 3.8 ±0 3.8 ±022±114±022±012±2 2.4 ±0.1 3.6 ±0 2.5 ±0
C111 3.7 ±0.1 3.6 ±0.2 N/A 18 ±113±118
±012±1 2.2 ±0.1 3.7 ±0.1 N/A
N/A: not measured. DOC values after sequential experiments were not measured due to insufficient sample volume. “before”means before
degradation.
a
DOC: dissolved organic carbon (after background DOC subtraction).
b
a(254 nm): absorption coefficient ¼UV absorbance at 254 nm times 2.303, and then divided by pathlength.
c
SUVA: the UV absorption at 254 nm in inverse meters normalized to DOC in mg C/L.
water research 61 (2014) 181e190 185
suggested to be highly labile and do not accumulate in the
natural environment. Therefore, the focus of discussion will
be mostly placed on the 8 commonly observed and environ-
mentally more stable PARAFAC components.
A comparison of relative contributions from different
sources to the EEM-PARAFAC component pool is shown in S2.
Soil leachates contributed the highest level of humic-like
component C1 as compared to biomass leachate which pro-
duced the highest protein-like component C7. Surface water
samples, which are exposed to light and thus are photo-
degraded, had the highest relative abundance of C2.
3.2. Variations in DOC, a(254 nm), and EEM-PARAFAC
components during photo-, bio-, and sequential degradation
During bio-degradation (Table 1) the DOC values decreased
substantially for biomass leachates (by 42.5%, 60.1%, 84.6%,
71.9%, and 73.5% for sawgrass, spikerush, periphyton,
mangrove, and seagrass, respectively) as compared to a much
smaller decrease for soil leachates (by 3.5%, 3.0%, 7.7%, and
23.1% for SRS3, SRS5, TS/Ph2, and TS/Ph11 soils, respectively)
and surface water (by 3.6%, 6.6%, 1.5%, and 2.6% for SRS3,
SRS5, TS/Ph2, and TS/Ph11, respectively). In a recent report,
the total biodegradable DOM in a moorland ranged from 5.0 to
19% during an up to 41 days bio-degradation study (Stutter
et al., 2013). The higher biodegradability of the biomass
leachate compared to the soil leachate and surface water is
not surprising considering the higher abundance of bio-labile
materials such as carbohydrates and proteins in the fresh
biomass leachates (Maie et al., 2006). The result here are
consistent with previous Everglades DOC bio-degradation
studies which reported that less than 10% of surface water
DOC was lost after 6 months of incubation in the laboratory
(Qualls and Richardson, 2003) suggesting a generally re-
fractory character for the DOC. As for the photo-degradation,
the DOC values did not significantly change during irradiation
for the samples studied suggesting that no significant photo-
mineralization occurred under the present photo-irradiation
conditions. Similar observations have been reported by
others (Lonborg et al., 2013; Vione et al., 2009).
The absorption coefficient a(254 nm) can reflect the
character and abundance of CDOM in a wetland system.
Spikerush and periphyton leachates showed decreases in a
(254 nm) during photo- (by 19% and 37%), bio- (by 4% and
22%), and sequential degradation (by 18% and 47%). However,
sawgrass, mangrove, and seagrass leachates displayed sig-
nificant increases in absorbance during all the degradation
experiments (Table 1). While the chemistry determining
these changes remains to be determined, the decrease in
CDOM is likely related to the photolysis of aromatic struc-
tures, while the increment in CDOM for other samples could
be related to the generation of higher molecular weight, more
highly aromatic structures in the DOM (see SUVA values
below; Scully et al., 2004). Alternatively, low molecular
weight aromatics such as polyphenols from the decomposi-
tion of tannins (in the case of mangrove DOM) or through
photo-humification reactions of lower molecular weight
biomolecules leached from the biomass materials
could enhance the CDOM signal. Both photo-humification
(Ortega-Retuerta et al., 2010) and the biosynthesis of
macromolecules (Ogawa et al., 2001) have previously been
shown to influence DOM composition.
The a(254 nm) for soil leachates was determined as
ranging from 13 to 32 m
-1
, and significantly decreased by 37%,
37%, 46%, and 48% during photo-degradation for soils from
SRS3, SRS5, TS/Ph2, and TS/Ph11, respectively. In contrast, the
values of a(254 nm) decreased to a much lesser extent during
biodegradation, by 3%, 2%, 10%, and 28%, respectively. How-
ever, the decrease after sequential degradation was 38%, 38%,
53%, and 54% respectively, and only slightly higher than
photo-degradation alone for TSPh2 and TSPh11. The a
(254 nm) for surface waters was between 18 and 142 m
-1
and
generally showed similar decreasing trends to the soil leach-
ates (Table 1). The a(254 nm) decreased by 37%, 31%, 36%, 37%,
and 28% during photo-degradation, by 5%, 1%, 16%, 2%, and
2% during bio-degradation, and by 40%, 32%, 51%, 45%, 36%
during sequential degradation for SRS3, SRS5, TS/Ph2,TS/
Ph11, and C111 surface water respectively.
SUVA, the specific UV absorbance, an indicator of DOM
aromaticity (Weishaar et al., 2003), was as expected lowest for
the biomass leachates, higher for surface water DOM and
highest for soil leachates. In most instances, including some
of the biomass leachates, SUVA values increased after photo-
degradation, most likely due to photo-humification reactions.
No significant changes were observed for the bio-degradation
experiments for soil leachates and surface waters, but sig-
nificant increases in values were observed for the bio-
degraded biomass leachates, including periphyton. This sug-
gests that biomass leachates undergo both photo- and bio-
condensation/humification reactions resulting in CDOM
(Ortega-Retuerta et al., 2010; Ogawa et al., 2001).
S1 shows the data for the PARAFAC component abun-
dances prior and after the incubation experiments. During
light exposure, PARAFAC components displayed in most cases
photo-decay. However, in a few cases, components C2, C3,
and C4 were photo-produced. C2 was photo-produced from
periphyton leachate, TS/Ph2 and TS/Ph11 soil leachates sam-
ples, and its abundance increased for SRS5 soil leachate and
SRS3 and TS/Ph2 surface water samples during photo-
exposure. While C3 was photo-produced from mangrove
leachate samples, C4 was photo-produced from mangrove
and seagrass leachates samples. While photo-production of
C3 and C4 concurred with an increase of absorption coefficient
a(254 nm), photo-production of C2 did not show such a trend.
This is reasonable considering that photo-generation can be
coupled with other processes such as photo-bleaching. In
contrast, and in agreement with the DOC and CDOM data,
during bio-incubation most EEM-PARAFAC components
proved to be mostly bio-refractory except for some of the
components from biomass leachate. Interestingly, most
PARAFAC components showed higher levels (in many in-
stances significantly higher) for the sequential photo- plus
bio-degradation compared with only photo-degradation.
Since the microbial generation of CDOM and ‘humic-like’
DOM even from plankton-derived OM has been previously
reported (Romera-Castillo et al., 2010; Guillemette et al., 2012),
this observation was not surprising. Although bio-degradation
results in this study showed that PARAFAC components from
soil leaching and surface waters were in most instances not
significantly affected by this process, exposure to sunlight has
water research 61 (2014) 181e190186
been reported to stimulate subsequent bacterial processing of
DOC (Kragh et al., 2008). Thus photochemical ‘priming’
enhanced the capacity of microbial generation of fluorescent
DOM (FDOM) ranging in characteristics from humic-like to
protein-like. However, generally, the overall abundance of
PARAFAC components after sequential degradation was still
much lower than after bio-degradation alone.
Since differences and similarities in the photo-, bio- and
sequential degradation dataset using EEM-PARAFAC are
highly variable and hard to visualize, principal component
analysis (PCA) was utilized and applied to the entire dataset
with the objective to better assess general trends in the
degradation processes. The relative abundance of EEM-
PARAFAC data for the biomass leachates, soil leachates, and
surface water samples, including the residual PARAFAC
components (R), were loaded into a principal components
analysis (see Fig. 2aed). Typical, natural, untreated, surface
water PARAFAC distributions for both freshwater (Everglades)
and marine (Florida Bay) FCE environments were also
included as reference end-members. The PCA for biomass
leachate degradation (Fig. 2aeb), showed that principal
component 1 (PC1) explained 46.8% of the variance, whereas
principal component 2 (PC2), accounted for a further 20.8% of
the variance. For soil leachate degradation (Fig. 2ced), PC1
explained 50.6% of the variance, whereas PC2, accounted for a
further 20.6% of the variance. All graphs showed a common-
ality in that the original, untreated samples featured more
negative PC1 values compared to degraded samples which in
all cases were located on a more positive PC1 axis. Natural
surface water samples from the Everglades were most positive
on the PC1 axis. In most cases, the photo-irradiated samples
featured higher PC2 values compared to the original samples,
while the biodegraded samples had PC2 values that were less
positive or negative. Most commonly, the combined photo-
plus bio-degradation samples had highest PC1 values and PC2
values somewhere in between those for the photo- or the bio-
degraded samples. In fact, the composition of samples sub-
jected to the combination of photo- plus bio-degradation were
located most closely to the Everglades surface water DOM
suggesting that DOM leached from biomass or soils is trans-
formed by these combined processes into DOM of similar
PARAFAC composition to Everglades surface waters. This was
typified by the seagrass leachate clustering closely with Flor-
ida Bay surface water after sequential degradation. Periphyton
behaved differently from the other biomass samples, which is
not surprising considering it is an assemblage of algae, mi-
crobes, and detritus rather than vascular plants, and pre-
sented the least significant photo- or bio-degradation changes
in its FDOM based on the PCA plot.
The PARAFAC component loadings controlling PC1 (Fig. 2b
and d) were similar for the biomass and soil leachates in that
most of the humic-like components, including C1, C2, C4 and
C5, showed the most positive PC1 values, in contrast to the
presumably more labile components such as the residues
R1-3p and R1-3s and the protein-like C7 which featured more
negative PC1 values. Components C3, C6 and C8 showed
Fig. 2 ePCA plot of EEM-PARAFAC data from biomass leachate (a and b) and soil leachate (c and d) degradation experiments
as compared to natural surface water. Samples from different origin and subject to photo-, bio- and sequential degradation
based are indicated. Arrows indicate the degradation trends; C: EEM-PARAFAC components; R: EEM-PARAFAC residues; (a
and c) are score plots; (b and d) are loading plots.
water research 61 (2014) 181e190 187
intermediate PC1 values in comparison to the other two
groups. With respect to PC2, PARAFAC components for the
biomass and soil leachates showed positive PC2 values for the
humic-like component C3 and negative values for the humic-
like component C6 and protein-like component C8. This
consistent trend suggests that C3 may in part be a photo-
degradation intermediate leading to C2, while C6 and C8
may be microbial-derived components that are indicative of
bio-degradation and microbial processing of DOM. Protein-
like components have previously been used as proxies for
Bioavailable DOM (BDOM) where soil leachates were found to
be more reactive than surface waters (Fellman et al., 2009). In
a recent report on fluorescence characterization of DOM
molecular size fractions from plant litter leachates, tyrosine-
like fluorescence (C7) was found to be enriched in the
smallest molecular size group, while tryptophan-like fluo-
rescence (C8) with the largest molecular size group (Cuss and
Gu
eguen, 2012). These results seem consistent with the
higher reactive nature of C7 based on the size-reactivity
concept. A study in a temperate Piedmont stream also re-
ported that only 13% of the tryptophan-like FDOM was labile,
14% was semi-labile, and 73% was recalcitrant, while
tyrosine-like FDOM was 100% biodegraded and the majority
(44e69%) was classified as labile (Cory and Kaplan, 2012). As
such, the general trend observed for biomass and soil leach-
ates suggests that PC1 is indicative of microbial and poten-
tially labile components (negative PC1) vs. higher plant and
soil derived, more refractory (positive PC1) components of the
DOM. In contrast, PC2 seems specifically indicative of the
photo- (positive PC2) vs. bio-degradation (negative PC2) state
of the DOM.
Sequential photo- and bio- degradation proved to be the
most effective process in changing the soil and biomass
leachates fluorescence characteristics into EEMs more similar
to those of the natural surface water DOM. This is consistent
with previous reports that sunlight has “priming effect”for
bacterial uptake of terrestrial-derived DOM (Benner and
Biddanda, 1998; Qualls and Richardson, 2003; Moran and
Covert, 2003). Interestingly, a recent Fourier-transform ion
cyclotron resonance mass spectrometry (FT-ICR-MS) study
also showed that the combined action of sunlight and mi-
crobes was more powerful in transforming the molecular
composition of wetland plant leached DOM into that similar to
deep sea DOM (Rossel et al., 2013). However, in general terms,
the differences in photo-degradation and combined photo-
plus bio-degradation for DOC and CDOM for the FCE samples
were relatively minor, suggesting that photo-degradation
seems to be the critical mechanism in the processing of
DOM in the Everglades. Having said this, the photo- and bio-
degradation of surface water samples from the Everglades
was minimal compared to the degradation of the leachates,
suggesting that under normal conditions, surface water DOM
is already quite recalcitrant in this environment. Although
compositional changes along north to south transects in the
Everglades have shown that DOM sources change from more
soil-derived to increasingly biomass derived (Yamashita et al.,
2010), this fact also implies that the degradation of soil and
biomass leachates must be efficient and fast on time scales
similar to or less than the average water residence time in the
system (ca. two months). A quantitative assessment of DOM
reactivity comparing kinetic data with water residence times
may prove this hypothesis.
A PCA plot for all the samples, including degraded and
fresh leachate samples, degraded Everglades surface water
samples, and natural Everglades surface water samples, is
shown in Fig. 3aeb. PC1, indicative of the a refractory (positive
PC1) vs. labile (negative PC1) character of the DOM, explained
38.0% of the variance, whereas PC2, indicative of photo-
(positive PC2) vs. bio-degradation (negative PC2) state of the
DOM, accounted for a further 20.5% of the variance. This graph
depicts the relatively refractory nature of the surface water as
most of the degraded surface water samples overlapped with
the natural Everglades surface water samples. Biomass
leachate samples showed the most reactive nature (largest PC
change) with some overlapping with soil leachate samples.
After graphing the values of the PC1 and PC2 for all the
leachates and degraded surface water samples (Fig. 3ced), the
general trend observed for PC1 was that all the surface water
samples featured positive PC1 values and almost all of the
leachates samples (with 3 exceptions) had negative PC1
values. Furthermore, the changes of PC1 values after photo-,
bio-, and sequential degradations were generally much more
obvious for the leachates compared to the surface water
samples, where the latter showed only very subtle changes.
Substantial changes of PC1 values (except for sawgrass
leachate) were observed especially after sequential degrada-
tions. As for PC2, although the values and changes were more
variable, the observed general trend was that of more positive
values after photo-degradation and more negative values
after bio-degradation for leachate samples. Again, the
changes in the PC2 values for surface water DOM after
degradation were smaller compared to those for the leachate
samples.
4. Conclusions and environmental
implications
DOC measurements and optical properties were applied to
assess the photo- and bio-reactivity of DOM from different
sources, including biomass leaching, soil leaching and sur-
face waters in a subtropical wetland ecosystem. While clear
patterns of FDOM photo- and bio-reactivity were observed,
where biomass leachates showed the highest bio-reactivity
(DOC decrease), followed by soil leachates, surface waters
were nearly bio-refractory even after a 28-day period of
inoculated dark incubation. This biorefractory nature is un-
related to nutrient limitations since incubations were per-
formed after nutrient supplementation. While the more
aromatic soil leachates and surface water DOM experienced
decrease and almost no change of a(254 nm) during photo-
and bio-degradation respectively, the initially less aromatic
DOM (lower SUVA) leached from biomass seemed to undergo
humification during both photo-degradation and bio-
incubation. In addition, changes in FDOM characteristics
showed higher changes for the sequential photo- plus bio-
degradation compared with only photo-degradation.
Although FDOM represents only a relatively small fraction
of the total DOC in aquatic systems, using EEM-PARAFAC to
trace the processing of DOM during degradation proved to be
water research 61 (2014) 181e190188
quite successful. The graphical representation of all experi-
mental PARAFAC data in the PCA plots showed that indeed
specific DOM components behaved differently under photo-
or bio-degradation conditions, and these trends were fairly
consistent throughout the diverse sample set. Here, PC1 was
identified as a proxy for source and overall lability of the
samples (negative PC1 more labile/fresh; positive PC1 as more
refractory/processed), while PC2 was identified as a proxy for
either a more photo- (more positive) or more bio- (more
negative) degradation state. Some PARAFAC components
were identified as potential intermediates for photo- and bio-
degradation respectively, and as such could be applied as
reactivity proxies for this system and potentially elsewhere
(Cawley et al., 2012). The combined photo- plus bio-
degradation resulted in a DOM composition most similar to
that of the natural surface water DOM. The data presented
herein are potentially transferable to other similar subtropi-
cal wetlands regarding the assessment of DOM reactivity.
Such studies are critical for better understanding the poten-
tial effects of climate change, land use, water management,
and other anthropogenic activities on carbon cycling in
aquatic ecosystems. While the overall reactivity of surface
water DOM in this aquatic system was determined to be low,
as compared to biomass and soil leachate, a quantitative
assessment (i.e. degradation rate constants), is needed to
determine the potential effects of Everglades restoration,
such as increased water delivery resulting in reduced resi-
dence time for water and associated solutes, on the biogeo-
chemical cycling of DOM.
Acknowledgments
This study was funded by National Science Foundation (NSF)
through the FCE-LTER program (DBI-062049). Additional sup-
port was provided through the George Barley Chair to RJ. MC
thanks the Department of Chemistry and Biochemistry and
NSF for research and teaching assistantships respectively.
The authors also thank N. Maie and Y. Yamashita for their
technical assistance and helpful discussions during the per-
formance of this work. This is SERC contribution #674.
Appendix A. Supplementary data
Supplementary data related to this article can be found at
http://dx.doi.org/10.1016/j.watres.2014.03.075.
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