Biogeosciences, 13, 2011–2028, 2016
www.biogeosciences.net/13/2011/2016/
doi:10.5194/bg-13-2011-2016
© Author(s) 2016. CC Attribution 3.0 License.
Challenges associated with modeling low-oxygen waters
in Chesapeake Bay: a multiple model comparison
Isaac D. Irby1 , Marjorie A. M. Friedrichs1 , Carl T. Friedrichs1 , Aaron J. Bever2 , Raleigh R. Hood3 , Lyon
W. J. Lanerolle4,5 , Ming Li6 , Lewis Linker7 , Malcolm E. Scully8 , Kevin Sellner9 , Jian Shen1 , Jeremy Testa6 ,
Hao Wang3 , Ping Wang10 , and Meng Xia11
1 Virginia
Institute of Marine Science, College of William & Mary, P.O. Box 1346, Gloucester Point, VA 23062, USA
QEA, LLC, 130 Battery Street, Suite 400, San Francisco, CA 94111, USA
3 Horn Point Laboratory, University of Maryland Center for Environmental Science, P.O. Box 775, Cambridge,
MD 21613, USA
4 NOAA/NOS/OCS Coast Survey Development Laboratory, 1315 East–West Highway, Silver Spring, MD 20910, USA
5 ERT Inc., 14401 Sweitzer Lane Suite 300, Laurel, MD 20707, USA
6 Chesapeake Biological Laboratory, University of Maryland Center for Environmental Science, P.O. Box 38, Solomons,
MD 20688, USA
7 US Environmental Protection Agency Chesapeake Bay Program Office, 410 Severn Avenue, Annapolis, MD 21403, USA
8 Woods Hole Oceanographic Institution, Applied Ocean Physics and Engineering Department, Woods Hole, MA 02543, USA
9 Chesapeake Research Consortium, 645 Contees Wharf Road, Edgewater, MD 21037, USA
10 VIMS/Chesapeake Bay Program Office, 410 Severn Avenue, Annapolis, MD 21403, USA
11 Department of Natural Sciences, University of Maryland Eastern Shore, MD, USA
2 Anchor
Correspondence to: Isaac D. Irby (iirby@vims.edu) and Marjorie A. M. Friedrichs (marjy@vims.edu)
Received: 2 December 2015 – Published in Biogeosciences Discuss.: 21 December 2015
Revised: 8 March 2016 – Accepted: 9 March 2016 – Published: 6 April 2016
Abstract. As three-dimensional (3-D) aquatic ecosystem
models are used more frequently for operational water quality forecasts and ecological management decisions, it is important to understand the relative strengths and limitations
of existing 3-D models of varying spatial resolution and biogeochemical complexity. To this end, 2-year simulations of
the Chesapeake Bay from eight hydrodynamic-oxygen models have been statistically compared to each other and to historical monitoring data. Results show that although models
have difficulty resolving the variables typically thought to be
the main drivers of dissolved oxygen variability (stratification, nutrients, and chlorophyll), all eight models have significant skill in reproducing the mean and seasonal variability of dissolved oxygen. In addition, models with constant net
respiration rates independent of nutrient supply and temperature reproduced observed dissolved oxygen concentrations
about as well as much more complex, nutrient-dependent
biogeochemical models. This finding has significant ramifications for short-term hypoxia forecasts in the Chesapeake
Bay, which may be possible with very simple oxygen parameterizations, in contrast to the more complex full biogeochemical models required for scenario-based forecasting.
However, models have difficulty simulating correct density
and oxygen mixed layer depths, which are important ecologically in terms of habitat compression. Observations indicate
a much stronger correlation between the depths of the top
of the pycnocline and oxycline than between their maximum
vertical gradients, highlighting the importance of the mixing
depth in defining the region of aerobic habitat in the Chesapeake Bay when low-oxygen bottom waters are present. Improvement in hypoxia simulations will thus depend more on
the ability of models to reproduce the correct mean and variability of the depth of the physically driven surface mixed
layer than the precise magnitude of the vertical density gradient.
Published by Copernicus Publications on behalf of the European Geosciences Union.
2012
I. D. Irby et al.: Multiple modeling low-oxygen waters
1 Introduction
Since the middle of the last century, anthropogenic impacts
have dramatically decreased water quality throughout the
Chesapeake Bay (Boesch et al., 2001), one of the largest estuaries in North America. Land-use change along with the industrialization and urbanization of the Chesapeake Bay watershed have caused dramatic increases in nutrient inputs to
the bay (Kemp et al., 2005), spurring additional primary production and phytoplankton abundance (Harding Jr. and Perry,
1997). Because increased primary production leads to more
organic matter throughout the water column that is eventually decomposed by bacteria, these increased nutrient inputs
to the bay have led to a corresponding decrease in dissolved
oxygen (DO) concentrations (Hagy et al., 2004). Hypoxia,
generally defined as the condition in which DO concentrations are below 2 mg L−1 , usually initiates seasonally in the
northern portion of the bay and expands southward as summer develops (Kemp et al., 2009; Testa and Kemp, 2014).
Although hypoxia in the Chesapeake Bay has likely existed since European colonization (Cooper and Brush, 1991,
1993), recent studies have highlighted an accelerated rise in
the number and spatial extent of hypoxic, as well as anoxic
(DO concentrations < 0.2 mg L−1 ), events in the bay since
the 1950s, primarily attributed to increased anthropogenic
nutrient input (Hagy et al., 2004; Kemp et al., 2005; Gilbert
et al., 2010). These impacts are likely to be exacerbated by
future climate change (Najjar et al., 2010; Meire et al., 2013;
Harding Jr. et al., 2015).
Interest in the ecological impacts of reduced DO concentrations has been elevated due to the observed proliferation of hypoxic events in the world’s coastal oceans, creating vast dead zone areas that compress suitable habitats
for many marine species (Diaz, 2001; Diaz and Rosenberg,
2008; Pierson et al., 2009). Low-DO waters can greatly
impact the abundance and health of important ecological
species, potentially resulting in suffocation and major kills of
fish, crabs, and shellfish (Breitburg, 2002; Ekau et al., 2010;
Levin et al., 2009). While the presence of DO concentrations < 2 mg L−1 have been shown to decrease the abundance of fish larvae (Keister et al., 2000), some species can
incur negative health impacts and modify their behavior at
significantly higher DO concentrations (Vaquer-Sunyer and
Duarte, 2008). DO concentrations of ∼ 4 mg L−1 have been
found to compress demersal fish habitat as fish seek out more
oxygenated waters (Buchheister et al., 2013). Zooplankton,
a crucial food source for valuable species, have also been
found to exhibit changes in distribution and predation when
subject to large volumes of low-DO water, potentially leading to further impacts along the food chain (Breitburg et
al., 1997; Pierson et al., 2009). Invertebrates have similarly
been found to alter their behavior under low-DO conditions (Riedel et al., 2014). In the Chesapeake Bay, multiple regulated fish species, such as striped bass and American
shad, require oxygen restoration targets as high as 5 mg L−1
Biogeosciences, 13, 2011–2028, 2016
Figure 1. Map of the Chesapeake Bay and its watershed.
(USEPA, 2010). The greatest impact of low DO concentrations spatially will depend on the specific living resource;
however, temporally, late spring to early fall is of most concern. As a result of the significant ecological importance of
oxygen on living resources in the bay, DO concentrations
are used as a primary indicator in assessing water quality for
Chesapeake Bay regulations (Keisman and Shenk, 2013).
Improving the health of the Chesapeake Bay has become
a priority for the Environmental Protection Agency (EPA)
along with the six states and Washington, DC that make up
the bay watershed (Fig. 1), and together they have committed to utilizing a suite of regulatory models to inform their
management decisions (USEPA, 2010). The Chesapeake Bay
Program (CBP), a regional partnership that has led and directed the restoration of the Chesapeake Bay since 1983, has
undertaken an extensive modeling effort of the bay (Cerco
and Cole, 1993; Cerco et al., 2002; Cerco and Noel, 2004,
2013). This modeling system is being used by the CBP
to estimate the aggregate effect of changes in management
practices, including land use, atmospheric deposition, animal
populations, and fertilizer and manure application. Recently,
the modeling system has been used to conduct scenario simulations to assess management actions needed to achieve desired bay water quality standards (USEPA, 2010). Ultimately
this model was used to establish a regulatory set of total maximum daily loads of nutrients and sediment delivered from
the watershed, with the goal of significantly improving water
quality throughout the bay (USEPA, 2010).
www.biogeosciences.net/13/2011/2016/
I. D. Irby et al.: Multiple modeling low-oxygen waters
Many 3-D hydrodynamic-oxygen models of varying complexity stemming from the academic research community
have also been used to simulate DO concentrations throughout the Chesapeake Bay (Scully, 2010, 2013; Hong and Shen,
2013; Feng et al., 2015; Testa et al., 2014; Y. Li et al., 2015).
Bever et al. (2013) specifically demonstrated that multiple
models of varying complexity are able to generate skillful estimates of hypoxic volume in the bay. Some of these models
are being used in the bay to simulate short-term and/or seasonal forecasts of DO conditions. Furthermore, some models
are also being used to generate scenario forecasts, or projections, that assess the impact of changes in management practices on estuarine DO concentrations, in some cases taking
into account the impacts of future changes in climate.
As ecosystem and water quality models are increasingly
used for operational forecasts as well as scenario-based
management decisions by the regulatory and academic research communities, it is important to understand the relative strengths and limitations of existing models of varying
complexity. The ability to discern which variables must be
most accurately simulated in order to adequately reproduce
the temporal and spatial variability of bay oxygen concentrations is a necessary prerequisite for fully understanding how
volumes of low-DO water are initiated and sustained within
water quality models. The utilization of multiple models can
also inform projections by providing independent confidence
bounds for management decisions. To those ends, the overarching goals of this research are to compare the relative
skill of various 3-D Chesapeake Bay models characterized
by different levels of biogeochemical complexity and spatial
resolution, to better understand factors limiting their ability
to reproduce observed DO distributions, and to suggest approaches for the continued improvement of these models.
2 Methods
2.1
Participating Chesapeake Bay models
Eight 3-D models were evaluated in this study (Table 1),
each of which includes hydrodynamic and DO components.
Among the eight models, there are four different hydrodynamic base models. Models B, C, D, F, and G utilize the
Regional Ocean Modeling System (ROMS; Shchepetkin and
McWilliams, 2005; Haidvogel et al., 2008) that employs a
structured grid with sigma layers in the vertical dimension.
Specifically, Models B, C, and F use a ROMS implementation developed for the Chesapeake Bay based on Xu et
al. (2012; ChesROMS). Model D employs a ROMS implementation for the Chesapeake Bay based on M. Li et
al. (2005), while Model G uses the ROMS-based Chesapeake
Bay Operational Forecast System (CBOFS; Lanerolle et
al., 2011). Models A, E, and H each use a different hydrodynamic base model: the Curvilinear Hydrodynamics in Three
Dimensions model (CH3D; Cerco et al., 2010), the Finitewww.biogeosciences.net/13/2011/2016/
2013
Volume Community Ocean Model (FVCOM; Jiang and Xia,
2016), and the Hydrodynamic Eutrophication Model – Environmental Fluid Dynamics Code (EFDC; Park et al., 1995;
Hong and Shen, 2012; Du and Shen, 2015), respectively.
The only model that employs a non-sigma vertical grid is
Model A and the only model utilizing an unstructured horizontal grid is Model E. While Model E contains 10 sigma
vertical layers, all of the other sigma grids use 20 layers. All
of the grids vary in terms of their horizontal resolution, with
Models A and G utilizing the highest resolution horizontal
grids.
These four hydrodynamic models are coupled to five different models used to simulate DO (Table 1). Models A, B,
C, D, and E utilize full biogeochemical models that include
various combinations of oxygen, phytoplankton, zooplankton, and multiple inorganic and organic nutrients as state
variables. Specifically, Models A and E employ a version of
the Integrated Compartment Model (ICM; Cerco et al., 2010;
Jiang et al., 2015), Model B uses the Estuarine Carbon Biogeochemistry model (ECB; Feng et al., 2015), Model C uses
the Biogeochemistry model (BGC; Brown et al., 2013), and
Model D uses the Row–Column AESOP model (RCA; Testa
et al., 2014). In terms of food web complexity the models
vary considerably: Models B and C employ a single phytoplankton group whereas Model D uses two phytoplankton
groups, Model E uses three, and Model A, the most complex
of the participating models, uses five.
In contrast to the full biogeochemical models discussed
above (Models A through E), Models F, G, and H represent
oxygen dynamics as simply as possible and therefore do not
utilize a full biogeochemical component. Rather, the models
impose a biological oxygen consumption rate that is modelspecific, but constant in both space and time. This component is referred to as a constant-respiration model (CRM).
In this model, DO is introduced to the estuary via the river
and ocean boundaries and is set to saturation at the estuarine surface. This constant-respiration oxygen parameterization (Scully, 2010) is simplistic, yet has been shown to adequately represent Chesapeake Bay oxygen dynamics (Scully,
2010, 2013; Bever et al., 2013).
The major difference in forcing between the eight model
implementations is that Models A and B use riverine input derived from watershed models, whereas Models C–
H used the measured flow from United States Geological
Survey gauging stations, extrapolated using various techniques. Model A utilized the CBP’s regulatory watershed
model (Shenk and Linker, 2013), while Model B utilized
the Dynamic Land Ecosystem Model (Yang et al., 2015a, b;
Tian et al., 2015). At the open boundary with the Atlantic
Ocean, Models B, C, D, F, G, and H utilize a sub-tidal elevation extrapolated from tidal stations on either side of the
open boundary. Model E uses the TPXO tidal model, while
Model A uses a mix of observational and model forcing
(Cerco et al., 2010). While Model B utilizes wind forcing
based on observations from the Thomas Point Light, ModBiogeosciences, 13, 2011–2028, 2016
2014
I. D. Irby et al.: Multiple modeling low-oxygen waters
Biogeochemical
complexity
Other atmospheric
forcing
Wind forcing
Sub-tidal elevation
at open boundary
River forcing
Vertical grid
Average wet-cell
resolution
Grid structure
Hydrodynamic
model-DO model
Model
Cerco et al. (2010)
High; 5 phytoplk.
groups
Multiple efforts
Multiple efforts
Multiple efforts
CBP watershed model
1.52 m
1 km
Structured
CH3D-ICM
A
Feng et al. (2015)
High; 1 phytoplk.
group
NARR
Thomas Point Light
Lewes, DE
to Duck, NC
DLEM watershed model
20 sigma
1.8 km
Structured
ChesROMS-ECB
B
Brown et al. (2013)
High; 1 phytoplk.
group
NARR
NARR
Lewes, DE
to Duck, NC
USGS data
20 sigma
1.8 km
Structured
ChesROMS-BGC
C
Testa et al. (2014)
High; 2 phytoplk.
groups
NARR
NARR
Wachapreague, VA
to Duck, NC
USGS data
20 sigma
1.89 km
Structured
ROMS-RCA
D
Jiang and Xia (2016)
High; 3 phytoplk.
groups
NARR
NARR
TPXO Tidal
Model
USGS data
10 sigma
1.26 km
Unstructured
FVCOM-ICM
E
Scully (2013)
Low; constant
respiration
NARR
NARR
Lewes, DE
to Duck, NC
USGS data
20 sigma
1.8 km
Structured
ChesROMS-CRM
F
Lanerolle et al. (2011)
Low; constant
respiration
NARR
NARR and NDBC buoys
Ocean City, MD
to Duck, NC
USGS data
20 sigma
0.565 km
Structured
CBOFS-CRM
G
Du and Shen (2015)
Low; constant
respiration
Norfolk and
Baltimore airports
NARR
Lewes, DE
to Duck, NC
USGS data
20 sigma
1.2 km
Structured
EFDC-CRM
H
Table 1. Model characteristics.
Model citation
Biogeosciences, 13, 2011–2028, 2016
els C through H use wind estimates from the North American
Regional Reanalysis (NARR).
The eight models used in this analysis have been developed for a variety of purposes. Model A is a governmental
regulatory model developed by the CBP that has been extensively calibrated specifically to examine water quality issues
in the Chesapeake Bay (Cerco and Cole, 1993; Cerco and
Noel, 2004, 2013; Cerco et al., 2010) and has been used in the
development of the 2010 Chesapeake Bay Total Maximum
Daily Load (USEPA, 2010). The National Oceanic and Atmospheric Administration employs the hydrodynamic component of Model F for operational forecasts of a variety of
physical estuarine parameters for the Chesapeake Bay (http:
//www.tidesandcurrents.noaa.gov/ofs/cbofs/cbofs.html). The
other six models are academic models used in diverse research efforts focused on the Chesapeake Bay but not necessarily specifically on DO dynamics.
Finally, a ninth model is calculated as the mean of the results from the eight models described above, and is referred
to here as Model Mean, or Model M.
2.2
Available Chesapeake Bay observations
Model simulations were compared to cruise data from the
CBP for 2004 and 2005 from 13 stations along the main
stem of the bay (Table 2, Fig. 2). The years 2004 and
2005 were selected to represent relatively wet and average years, respectively, and the 13 stations were chosen as they have been found to offer optimal estimates
of bay-wide hypoxic volume (Bever et al., 2013). Stations were sampled on up to 34 cruises over the 2 years
(Table 2), generally twice a month from April to August and once a month for the remainder of the year. Observational data can be downloaded from the CBP Water Quality Database (http://www.chesapeakebay.net/data/
downloads/cbpwaterqualitydatabase1984present). Variables
downloaded from the CBP website and used in this study
were temperature, salinity, DO, nitrate + nitrite (hereafter
abbreviated as “nitrate”), and chlorophyll a (hereafter abbreviated as “chlorophyll”). For most cruises, observations of
temperature, salinity, and DO were made at roughly 1 m intervals throughout the water column, whereas observations
of chlorophyll and nitrate were generally made only at the
surface, bottom, and sometimes one or two mid-water column locations. For further information on available water
quality observations, please see USEPA (2012). While these
observations were publicly available for model assessment
during calibration of all of the models, they represent a very
small subset of the 30 years of EPA observations across over
100 bay stations. The models compared here were calibrated
based on access to the larger data set and for conditions in
the bay in general, not specifically for the 13 stations and 2
years considered here.
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I. D. Irby et al.: Multiple modeling low-oxygen waters
Table 2. Characteristics of observation stations (from USEPA,
2012).
Station
Latitude
(◦ N)
Longitude
(◦ W)
Station depth
(m)
No. of cruises
CB3.2
CB3.3C
CB4.1C
CB4.2C
CB4.3C
CB4.4
CB5.1
CB5.2
CB5.4
CB6.2
CB6.4
CB7.1
LE2.3
39.1634
38.9951
38.8251
38.6448
38.5565
38.4132
38.3185
38.13678
37.8001
37.4868
37.2365
37.6835
38.0215
76.3063
76.3597
76.3997
76.4177
76.4347
76.3430
76.2930
76.2280
76.1747
76.1563
76.2080
75.9897
76.3477
12.1
24.3
32.3
27.2
26.9
30.3
34.1
30.6
31.1
10.5
10.2
20.9
20.1
34
34
34
34
34
34
34
34
26
30
29
27
34
Figure 2. Location of the CBP water quality monitoring stations
used in this study.
2.3
Calculation of stratification and mixed layer depth
Stratification of the density and oxygen fields was examined
to identify the maximum gradient of the pycnocline and oxycline as well as the depth of the top of the pycnocline and
oxycline. In open ocean studies, the depth of the top of stratification is commonly referred to as the mixed layer depth
(MLD), although this term is less frequently used in the estuwww.biogeosciences.net/13/2011/2016/
2015
arine literature. As the research presented here distinguishes
between the depths of the top of the pycnocline and that of
the oxycline, these will be referred to respectively as the density (ρ) mixed layer depth (MLDρ ) and the oxygen mixed
layer depth (MLDO ). Density was calculated via a classical
density formula that is also utilized by the CBP for use in
the Chesapeake Bay (Fofonoff and Millard, 1983; USEPA,
2004) and is a function of temperature and salinity.
The CBP defines the top and bottom of stratification in
order to distinguish individual designated use areas for water quality management purposes (USEPA, 2004). They suggest that the top of the pycnocline be defined as the shallowest occurrence of a density gradient of 0.1 kg m−4 or greater
as resolved by CBP profile observations, which are typically spaced at 0.5–2 m depth intervals. If density gradients
throughout the water column are less than 0.1 kg m−4 , they
define the water to be unstratified. The 0.1 kg m−4 threshold
definition is designed to identify any initiation of stratification that may serve to cut off vertical mixing from a nearly
perfectly well-mixed layer.
While the CBP definition described above delineates between designated use boundaries according to density, our
research focuses on the relationship between the pycnocline
and oxycline, requiring an alternate definition that can be
applied to both the density and oxygen distributions. In addition, the CBP definition often generates estimates for the
depth of the top of the pycnocline that are too shallow compared to the maximum depth of surface mixing (Fig. 3).
As a result, a percentage threshold criterion was developed
that identifies the bottom of the reasonably well-mixed layer,
rather than perfectly mixed layer, and is used in this analysis. The percentage threshold method defines a density or DO
profile as being stratified if a change of 10 % of the difference
between the profile’s maximum and minimum values occurs
within a single meter (Fig. 3). For example, if the maximum
DO concentration throughout the water column on an individual sampling date is 10 mg L−1 and the minimum concentration is 1 mg L−1 , stratification is defined to be present
if a difference of 0.9 mg L−1 is present within 1 m. As recommended by the CBP, the uppermost meter of the water
column is not considered (USEPA, 2004). The mixed layer
depth is therefore defined as the shallowest level (below 1 m
depth) where stratification is identified. The minimum stratification criterion utilized in this analysis requiring a profile to
pass the 10 % threshold also ensures that observations where
very little stratification exists do not bias the stratification
results while also allowing for a single criterion to be used
across multiple stratification variables.
2.4
Model skill metrics
Simulations of the Chesapeake Bay from the eight models described above were statistically compared to historical monitoring data using a variety of skill metrics including root-mean squared difference (RMSD), bias, standard
Biogeosciences, 13, 2011–2028, 2016
2016
I. D. Irby et al.: Multiple modeling low-oxygen waters
Figure 3. Density and dissolved oxygen profiles for a mid-bay station (CB4.1C) on (a) 13 January 2004 and (b) 14 June 2005, comparing
the 0.1 kg m−4 stratification definition used by the CBP (MLDCBP ) with the 10 % threshold definitions used here for density (MLDρ ) and
oxygen (MLDO ).
deviation, and correlation coefficient. These metrics are illustrated on Taylor and target diagrams (Taylor, 2001; Hofmann et al., 2008; Jolliff et al., 2009), which offer a compact way of assessing model skill by displaying a number
of different skill metrics. Target diagrams illustrate the bias
and total RMSD of model output, which Taylor diagrams
do not. Taylor diagrams include quantitative information on
the standard deviations and correlations between the model
output and the observations, which target diagrams do not.
Both diagrams, however, represent unbiased RMSD, sometimes called “centered-pattern RMSD”. On target diagrams,
a model symbol above the horizontal axis overestimates the
mean of the observations and a model symbol to the right
of the vertical axis overestimates the variability of the observations. (See Hofmann et al. (2008) and Jolliff et al. (2009)
for a more detailed description of these diagrams.) On Taylor diagrams, a model symbol lying on the horizontal axis
exactly correlates to the observations and a model symbol
further from the origin than the observation symbol overestimates the standard deviation of the observations. (See Taylor (2001) for a more detailed description of these diagrams.)
Taylor and target diagrams presented here are normalized
to the standard deviation of the observations, allowing multiple variables be represented on the same plot. This also conveniently allows the unit circle on a target diagram to represent the skill of a model defined as the mean of the observations. In effect, this means that if a model falls within the
unit circle, it exhibits a skill that is greater than the skill obtained if one were to simply use the mean of the observations.
The Taylor and target plots are either temporal (displaying
model skill at a single station over the study period) or spatial (displaying model skill during a single month over the en-
Biogeosciences, 13, 2011–2028, 2016
tire set of study stations). In addition, summary diagrams are
presented which combine both temporal (examining the seasonal changes at each individual station) and spatial (examining differences across the bay during an individual month)
variability.
Model skill was assessed using the hourly model output
(daily for CH3D-ICM chlorophyll and nitrate) that was nearest in time to that of the observation and from the grid cell
that encompassed the observation location. For months with
two observations, each observation was individually matched
to the model output and the skill statistics from those comparisons were averaged for that month. The native horizontal resolution and bathymetry of the individual model grids
was preserved in the comparison so as not to bias the analysis through varying interpolation methodologies. For stratification variables, the models and observations were interpolated to a 1 m vertical grid that extended only as deep as
the individual models’ bathymetry or deepest observation in
order to preserve the differences in bathymetric grids while
allowing for a direct comparison of the observations to the
models. Model–data comparisons at the bottom of the water column were not necessarily based on the same depths,
since in many cases the modeled bathymetry was shallower
(or at times, deeper) than the deepest data point at a given station. In order to avoid issues with extrapolation and/or grid
stretching, data at the bottom of the water column were always compared with model estimates from the deepest grid
cell provided by each particular model. Model–data comparisons for stratification and mixed layer depths only included
stations and times for which stratification was defined to exist
in both the observed and simulated fields.
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I. D. Irby et al.: Multiple modeling low-oxygen waters
2017
3 Results
An analysis of model skill of the combined temporal and
spatial variability of DO at the surface and bottom of the
water column, as well as at the observed MLDO , indicates
that all models, regardless of biogeochemical complexity or
spatial resolution, exhibit a high degree of skill in reproducing observed DO (Fig. 4). Specifically, all models produce
DO concentrations at the surface and bottom that have a normalized total RMSD less than 1. The same is true for nearly
all models for DO at the observed MLDO . However, most
models underestimate observed DO both at the surface and
at the MLDO (Fig. 4a). The correlation between the observed
and modeled DO is relatively constant with depth (Fig. 4b),
though on average slightly higher at the bottom (0.85) than
at the surface (0.80). Further, on average, the models simulate DO at the surface and bottom better than they do at
the MLDO . No statistical difference exists between the skill
of models that utilize a full biogeochemical component and
those that utilize the simple constant-respiration oxygen parameterization. Based on an analysis of variance (ANOVA)
comparing the full biogeochemical models to the CRM models, the two model types do not perform differently in terms
of their ability to reproduce the combined temporal and spatial variability of bottom DO as measured by total RMSD
(p = 0.48). Overall, Model M (the mean of the eight models) consistently performs better than any individual model
across all depths examined (Fig. 4).
The monthly temporal variability of bottom DO at each
station over the 2 years studied is resolved similarly well by
all of the models (Fig. 5a), but the models have difficulty simulating spatial DO variability during each month (Fig. 5b).
Due to the stations chosen for this analysis (Fig. 2), the spatial variability being examined here is essentially the north
to south variability. Most models exhibit a latitudinal gradient with respect to their skill in reproducing the temporal
variability of bottom DO, with models overestimating DO
at the more northern stations (Fig. 5a). Some models differ in their ability to reproduce summer (May–September)
DO concentrations and winter (October–April) DO concentrations (Fig. 5b). Models B, F, and G all distinctively overestimate mean DO in the summer compared to the winter.
In contrast, Models A and C perform similarly well in both
seasons (Fig. 5b). In addition, all three constant respiration
models, as well as Models D and E, substantially underestimate DO at several stations in the winter.
All eight models generally resolve the pycnocline and oxycline with similar skill (Fig. 6). All models consistently underestimate the mean and standard deviation of the maximum
strength of stratification within the pycnocline and oxycline,
defined herein as the maximum vertical gradients of density
and oxygen (Fig. 6a). All models, except for Model A (see
Sect. 4.2), also underestimate the mixed layer depth, regardless of whether it is computed in terms of density or oxygen. (Note that these model symbols in Fig. 6a are located
www.biogeosciences.net/13/2011/2016/
Figure 4. Normalized summary (a) target and (b) Taylor diagrams
illustrating model skill of dissolved oxygen at the surface, MLDO ,
and bottom for 13 Chesapeake Bay stations in 2004–2005. The “x”
represents the skill of a model that perfectly reproduces the observations. The dotted, dashed-dot, and dashed lines on the Taylor diagram represent lines of constant standard deviation, correlation coefficient, and unbiased RMSD, respectively.
above the y axis despite this negative bias in MLD because
the vertical coordinate system is oriented upwards.) Thus, the
models are producing stratification that is both weaker than
observed and higher (shallower) in the water column. The
correlation coefficient for these metrics is low, ranging 0.1–
0.6, and indicates that all models are missing the majority of
variability associated with the magnitude and location of the
pycnocline and oxycline (Fig. 6b). However, there is slightly
more consistency and better correlation coefficients among
the models for the strength of stratification than the depth of
the mixed layers.
All eight models are also characterized by similar skill in
representing the temporal and spatial variability of density
stratification and MLDρ (Fig. 7). There is a latitudinal difference in skill of the models in reproducing the magnitude of
Biogeosciences, 13, 2011–2028, 2016
2018
I. D. Irby et al.: Multiple modeling low-oxygen waters
Figure 5. Normalized target diagrams for Models A–H demonstrating the (a) temporal and (b) spatial skill in resolving the variability of
bottom dissolved oxygen concentrations. In (a) the individual dots represent the 13 stations along the main stem of the Chesapeake Bay.
In (b) the dots represent the 24 months of 2004–2005 and are delineated by color: red is summer (May–September) and blue is winter
(October–April).
the pycnocline and MLDρ , with model skill generally lower
at the northern stations (Fig. 7a). Contrary to the pattern
shown for bottom DO (Fig. 5b), none of the models exhibit
a significant seasonal pattern between summer and winter in
reproducing spatial variability of dρ/dz or MLDρ (Fig. 7b).
However, Model A differentiates itself from the rest of the
models in its pattern of skill at reproducing the spatial and
temporal variability of the MLDρ (see Sect. 4.2). Temporal and spatial patterns for oxycline stratification (dO/dz)
and MLDO closely match those of dρ/dz and MLDρ (not
shown).
All eight models reproduce the variability of bottom DO
better than the variables that are generally thought of as being
the primary drivers of hypoxic conditions, including stratification (Fig. 6), salinity, chlorophyll, and nitrate (Fig. 8, Table 3). However, all models reproduce patterns in temperature across the bay and through time better than any of the
other variables in this model comparison (Fig. 8). All eight
models, as well as the Model Mean, are characterized by
very low bias in modeled temperature, and correlation coefficients of approximately 0.99; this high skill results from the
very strong and predictable seasonal temperature variability.
Biogeosciences, 13, 2011–2028, 2016
Even though the five models with full biogeochemical components (Models A, B, C, D, E) are characterized by large
differences in their mechanistic approaches to modeling nitrate and chlorophyll, they produce similar total RMSDs for
all of the variables examined at both the surface and the bottom (Table 3).
The mean of the eight models (Model M) has a higher
model skill (lower RMSD) than any individual model across
nearly every variable examined (Table 3). In addition, for
nearly all observations at all stations, the 95 % confidence
interval of all model hindcasts encapsulates the observed
bottom DO concentration (Fig. 9), even though any individual model may overestimate or underestimate observed DO.
Models generally fall into greater agreement during the summer, when DO is low, and into lesser agreement in the winter
when DO is replete. While this study does not allow for a
true interannual comparison, it is interesting that at station
CB4.1C the model ensemble closely matches the timing of
the drawdown of DO in the spring of 2004 (Fig. 9), whereas
it produces a summer rather than spring initiation of hypoxic
conditions in 2005. In addition, the model ensemble produces
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I. D. Irby et al.: Multiple modeling low-oxygen waters
2019
MLDO is stronger for more severe stratification (Table 4).
The relationship between the two mixed layer depths is biased towards the MLDO being located slightly deeper in the
water column than the MLDρ . As the cut-off criteria for the
existence of stratification becomes more stringent, the relationship becomes closer to 1 : 1.
4 Discussion
4.1
How does the skill of various hydrodynamically
based DO models compare?
In examining the eight 3-D models in this study,
there is not a statistical difference between the ability of simple and complex models to simulate the
mean and monthly variability of bottom DO; in addition, models with higher spatial resolution do not
necessarily produce better estimates of DO.
Figure 6. Normalized summary (a) target and (b) Taylor diagram
illustrating model skill of MLDρ and MLDO , max dρ/dz, and max
dO/dz at 13 Chesapeake Bay stations for 2004–2005. The “x” represents the skill of a model that perfectly reproduces the observations. Since RMSD of stratification is only computed at stations
where both the observations and model exhibit stratification, the
Model Mean is not calculable for these variables.
a premature relaxing of hypoxic conditions for both years at
this observation station.
In order to better understand the impact of stratification
on DO concentrations throughout the water column, the relationship between the observed pycnocline strength and
MLDρ were compared to the observed oxycline strength and
MLDO . Observations from 1998 to 2006 demonstrate that
while there is not a strong correlation between the strengths
of the pycnocline and oxycline, there is a very strong correlation between MLDρ and MLDO (Fig. 10). Depending on
the criteria used for defining the existence of stratification
(see Sect. 2.3), the correlation of the pycnocline and oxycline strengths range r 2 = 0.18–0.26 and the correlations of
MLDρ and MLDO range r 2 = 0.51–0.82 (Table 4). Furthermore, correlation of the relationship between the MLDρ and
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Models currently simulating hypoxia throughout Chesapeake Bay compute oxygen concentrations in essentially two
distinct ways: they either utilize a simple constant respiration
model or a full biogeochemical model. In this study, the relative skill of both types of models is compared. Specifically,
in examining results of the comparison between five biogeochemical models (A, B, C, D, E) and three simplistic constant
respiration models (F, G, H), the two groups of models performed statistically similar in their skill of reproducing bottom DO concentrations (Fig. 3, Table 3). These results support those of Bever et al. (2013) who compared three constant
respiration models with the CBP regulatory model (Model A)
and similarly found that all four of the models were equally
skillful in terms of reproducing the seasonal variability in
bottom DO throughout the bay in 2004 and 2005. Consistent with the results of Scully (2013), this result implies that
the seasonal variability of DO in the Chesapeake Bay is primarily dependent on underlying hydrodynamic mechanisms
which are nearly identical for all eight models, rather than
on aspects related to the biogeochemical cycling which vary
dramatically between models and in fact are constant in three
of the eight models. It should be noted, however, that the 2
years studied here were relatively wet years and an analysis
of dry years may offer different results.
Many previous studies have examined the costs and benefits of adding complexity to biogeochemical models. For
example, increasing biogeochemical complexity has been
found to improve skill in some biogeochemical data assimilative parameter optimization studies (Friedrichs et al., 2006,
2007; Lehmann et al., 2009; Bagniewski et al., 2011; Ward
et al., 2013; Xiao and Friedrichs, 2014). The additional parameters associated with increased complexity generally provide more parameters that are available for additional tuning and subsequent improved model–data agreement. This
is in contrast to the results of this analysis demonstrating
that increased biogeochemical complexity does not necesBiogeosciences, 13, 2011–2028, 2016
2020
I. D. Irby et al.: Multiple modeling low-oxygen waters
Figure 7. Normalized target diagrams for Models A–H demonstrating the (a) temporal and (b) spatial skill in resolving the variability of
the strength of density stratification (circles) and the depth of pycnocline initiation (diamonds). In (a) the individual dots represent the 13
stations along the main stem of the Chesapeake Bay. In (b) the dots represent the 24 months of 2004–2005 and are delineated by color: red
is summer (May–September) and blue is winter (October–April).
Table 3. Mean and standard deviation (SD) of observations and total normalized RMSD for each model. RMSD for each model except when
not applicable (N/A).
Mean ± SD of Obs.
Surface temp. (◦ C)
Bottom temp. (◦ C)
Surface salinity (PSU)
Bottom salinity (PSU)
Max. dρ/dz (kg m−4 )
MLDρ (m)
Surface DO (mg L−1 )
DO at MLDO (mg L−1 )
Bottom DO (mg L−1 )
Max. dDO/dz (mg L−1 m−1 )
MLDo (m)
Surface Chl a (mg m−3 )
Bottom Chl a (mg m−3 )
Surface nitrate (mmolN m−3 )
Bottom nitrate (mmolN m−3 )
Biogeosciences, 13, 2011–2028, 2016
17.44 ± 8.82
15.75 ± 8.02
10.92 ± 4.32
18.17 ± 3.14
∼ 1.64 ± 1.15
∼ 5.32 ± 3.99
9.74 ± 2.15
∼ 8.44 ± 2.53
4.42 ± 3.61
∼ 1.81 ± 1.12
∼ 6.62 ± 4.01
11.19 ± 9.04
9.02 ± 11.52
0.32 ± 0.33
0.12 ± 0.13
Normalized RMSD
A
B
C
D
E
F
G
H
M
0.13
0.24
0.37
0.72
1.03
1.01
0.67
0.54
0.51
1.19
1.24
0.92
0.87
0.61
1.08
0.13
0.35
0.62
0.85
1.09
1.13
0.58
0.57
0.59
1.21
1.01
1.22
1.10
0.79
1.38
0.12
0.35
0.53
0.73
1.07
1.11
0.89
0.74
0.81
1.34
1.10
1.60
1.07
1.03
1.38
0.09
0.23
0.36
1.55
1.09
1.41
0.80
0.93
0.61
1.09
1.33
1.23
1.05
0.61
0.92
0.13
0.22
0.46
1.28
1.25
1.39
1.00
0.83
0.54
1.35
1.33
0.89
1.01
0.52
1.46
0.13
0.35
0.61
0.78
1.01
1.12
0.63
0.81
0.46
1.12
1.05
N/A
N/A
N/A
N/A
0.16
0.17
0.57
1.03
1.23
1.38
0.64
0.95
0.61
1.23
1.30
N/A
N/A
N/A
N/A
0.19
0.19
0.41
0.97
1.02
1.13
0.69
1.09
0.60
1.19
1.29
N/A
N/A
N/A
N/A
0.10
0.23
0.35
0.75
N/A
N/A
0.57
0.62
0.46
N/A
N/A
1.16
0.90
0.79
0.85
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I. D. Irby et al.: Multiple modeling low-oxygen waters
2021
Table 4. Pycnocline and oxycline correlation statistics (all correlations have p values ≪ 0.01).
Stratification
threshold
percentage ( %)
Max. dρ/dz
vs.
max. dO/dz
MLDρ
vs.
MLDO
10
15
20
25
0.18
0.22
0.22
0.26
0.51
0.59
0.70
0.82
Profiles
with
stratification
1613
1303
916
575
sarily improve model–data agreement. In this case, the increase in model complexity has likely outpaced the ability
of the researchers to fully tune the model to the available
observations. However, even past studies that have invoked
formal parameter optimization methodologies, such as genetic algorithms and variational adjoint methods (Friedrichs
et al., 2007; Ward et al., 2010; Xiao and Friedrichs, 2014),
have found that under certain conditions, adding too much
complexity does not necessarily improve model skill and in
fact can decrease model skill and portability, primarily due
to artifacts resulting from overtuning. This mirrors findings
from the larger ecosystem modeling community where the
best-fit models are often those with intermediate complexity
(Fulton et al., 2003).
In this study, horizontal grid resolution differed significantly between model implementations, with the most highly
resolved grid (Model G) including more than 9 times more
grid cells than the lower resolution grids (Table 1). A certain
degree of resolution is clearly required to successfully simulate dynamic processes, and a model with 8–10 km resolution will not be able to correctly simulate the hydrodynamic
processes within the bay (Feng et al., 2015). However, an increase in horizontal grid resolution from ∼ 1.8 to ∼ 0.6 km,
which results in a run-time change of a factor of 9, or possibly of 27 if the time step is accordingly decreased by a
factor of 3, does not necessarily result in a significant improvement in simulation skill of either stratification or bottom oxygen. Although not shown here, additional sensitivity
experiments with Model G revealed that doubling the vertical resolution of this model had no significant effect on the
model’s ability to resolve the depth of stratification or the
maximum magnitude of stratification. Thus, when selecting
the optimal model resolution for a simulation, it is critical
to weigh the advantages of increased resolution with the increased time required for simulation. With a given level of
computational resources, fewer sensitivity experiments can
be conducted with a model using a more highly resolved grid.
Accurately simulating the observed spatial variability of
DO (Fig. 4b) was a greater challenge than simulating the
temporal variability of DO (Fig. 4a) for all eight models participating in this intercomparison. This is especially true in
the winter months when the vast majority of the bay is oxywww.biogeosciences.net/13/2011/2016/
Figure 8. Normalized summary (a) target and (b) Taylor diagram
illustrating model skill of bottom temperature, salinity, chlorophyll,
nitrate, and dissolved oxygen at 13 Chesapeake Bay stations for
2004–2005. The “x” represents the skill of a model that perfectly
reproduces the observations.
gen replete and the models have difficulty representing the
observed variability from station to station. The majority of
the models tend to slightly overestimate mean bottom DO in
the summer whereas multiple models (e.g., Models D, E, F,
and G) exhibit a strong negative bias during January and/or
February of 2005, primarily at stations in the middle to southern portion of the bay’s deep channel. Interestingly, increased
biological complexity and higher grid resolution do not completely resolve this issue, as this is true for models utilizing
full biogeochemical models (Models D, E) as well as those
using highly resolved model grids (Model G). This is likely
due to the ephemeral nature of the biological divers of DO.
The strong performance of the constant respiration models implies that these models may be excellent candidates
for providing short-term bottom oxygen forecasts. The high
DO skill of the CRM models primarily results from the fact
that seasonal variations in physical processes (primarily wind
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I. D. Irby et al.: Multiple modeling low-oxygen waters
Figure 9. Time series of bottom dissolved concentrations for station
CB4.1C. Red dots represent the 34 observations made during 2004–
2005. Grey lines are the individual model simulations. The dark
blue line represents the Model Mean while the cyan line represents
the 95 % confidence interval of the model simulations.
mixing and temperature) play a dominant role in controlling
the seasonal cycle of oxygen (Scully, 2013). Because the underlying hydrodynamic models all use similar physical forcing, the constant respiration models are able to simulate the
seasonal cycle of DO with similar skill as the more complex biogeochemical models. As a result, these simple models that are easier to tune and require less in the way of computational resources than full biogeochemical models, may
be efficiently used to produce short-term (on the order of
days) DO forecasts. On the contrary, the more complex full
biogeochemical models will be necessary for scenario-based
and long-term (on the order of months to years) forecasting
which requires that models respond to prescribed changes
in the biogeochemical environment, such as increased rates
of nutrient loading due to changes in land use, land cover,
and/or climate.
4.2
How does model skill of DO compare to that of the
primary drivers of DO variability?
Overall, model DO skill is greater than that of the
variables generally considered to drive DO variability, such as stratification, salinity, mixed layer
depth, chlorophyll, and nitrate; only modeled temperature has higher skill than modeled DO.
Since dissolved oxygen concentrations in the Chesapeake
Bay are controlled by physical processes (e.g., advection,
wind mixing, heating/cooling, and stratification), as well as
biological processes (e.g., photosynthesis and respiration), it
is critical to understand the skill of the models in terms of
how well they reproduce the many factors influencing oxygen concentrations. As expected, the five models containing
a specific biogeochemical model component had more difficulty simulating the observed chlorophyll and nitrate concenBiogeosciences, 13, 2011–2028, 2016
Figure 10. Scatter plots comparing observations of (a) the strengths
of stratification of the pycnocline and oxycline and (b) the oxygenand density-defined mixed layer depths. Size of the circles is proportional to the number of observations. Observations are from
1998 to 2006 at the 13 Chesapeake Bay stations shown in Fig. 2.
trations than the physical variables (temperature and salinity), both at the surface (Table 3) and the bottom (Fig. 8).
Replicating the correct location, magnitude, and timing of
phytoplankton blooms and nutrient cycling is a complex issue, and as a result, these features are generally not well
simulated in the models. While the models generally simulate the total amount of chlorophyll adequately, it is more
uniformly spatially distributed in the models rather than in
patchy blooms as in nature, leading to the underestimation of
chlorophyll variability across all models. Although all models produced a relatively high correlation between observed
and modeled temperature and salinity (Fig. 8), the correlation
coefficients for chlorophyll and nitrate were much lower. The
correlations for observed vs. modeled DO was more similar
to that of the physical variables (temperature, salinity) than
the biological variables (chlorophyll and nitrate), highlighting that the seasonal variability in bottom DO is regulated
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I. D. Irby et al.: Multiple modeling low-oxygen waters
more by physical than biological factors. This also explains
the success of the constant respiration models, which by definition contain no biological variability yet reproduce DO
variability nearly as well as the most complex biogeochemical models.
In this study, model skill was also considerably higher for
bottom oxygen than it was for the vertical gradient of stratification and mixed layer depths (Figs. 6, 8). The underestimation of the vertical gradient across all models is largely
due to the numerical diffusion that characterizes all of these
hydrodynamic models, but may also be partially due to an
underestimation of the winds or a lack of diffuse freshwater input around the bay. Even though the models all underestimated the strength of stratification (Figs. 4, 6), modeled
stratification in summer was strong enough to prevent mixing
with the relatively well-oxygenated surface waters. This result suggests, somewhat surprisingly, that simulating the correct vertical gradient of stratification is not absolutely necessary for skillful bottom DO simulations. Models need only
simulate enough stratification to effectively cut off vertical
mixing in order to develop an isolated bottom layer that can
then experience a draw down in oxygen via respiration. In addition, the models must also correctly simulate the horizontal
advection of oxygen (Scully, 2013; Y. Li et al., 2015). The
fact that bottom DO is simulated so well by the eight models
analyzed here suggests that not only is the advection of oxygen well represented in the models but also the strength of
stratification, i.e., the maximum vertical gradients of density
and oxygen, produced by these models is sufficient. Thus,
although novel and somewhat unexpected, these results are
not contradictory to previous studies demonstrating the importance stratification plays in initiating summer hypoxia in
the Chesapeake Bay (Murphy et al., 2011).
Model skill in terms of reproducing observed mixed layer
depths was likewise much lower than model skill of reproducing observed oxygen concentrations. All models, except
Model A, produced mixed layer depths (MLDO and MLDρ )
that were generally too shallow in the water column (Fig. 6a).
Note that Model A is a regulatory model that has been used
for many years by the Chesapeake Bay Program, and has thus
undergone more extensive calibration aimed at matching the
mean salinity and oxygen characteristics of the bay (Cerco
and Cole, 1993). Furthermore, Model A employs a z grid
that matches the bathymetry in trench areas better than the
sigma grids used by the other models. Although Model A
produced mixed layer depths that were generally in the correct location within the water column (Fig. 6a), they were too
variable (Fig. 6b). This variability may partly be a result of
the 1.5 m z grid employed by Model A causing large jumps
between vertical grid cells and hence resulting in overestimates of MLD variability. All other models use sigma grids
typically with more highly resolved vertical resolution at the
depth of maximum stratification.
The two variables for which the models have greatest skill
are DO and temperature (Fig. 8). This is because oxygen
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2023
Figure 11. Time series of observations at station CB4.1C from 2003
to 2006 for (a) bottom dissolved oxygen, (b) dissolved oxygen at
the MLDO , and (c) MLDO .
variability is driven primarily by seasonal variability in physical processes such as solubility and wind mixing and to a
lesser degree by variability in oxygen consumption (Scully,
2013). As a result, the models using a constant mean respiration rate produce as realistic hypoxia simulations as the biogeochemically complex models. Observations clearly show
this strong seasonal variability in bottom DO (Fig. 11a) and,
to a slightly lesser extent, clear seasonal variability in DO at
the bottom of the bottom of the oxygen mixed layer (MLDO ;
Fig. 11b). However, a seasonal cycle is not manifested in the
MLDO itself (Fig. 11c). The lack of such a strong seasonal
cycle in the observed mixed layer depths makes this a more
difficult variable for the models to simulate. As a result, the
models can relatively skillfully simulate the combined spatial
and temporal variability of DO while simultaneously missing
the MLDO .
4.3
Why is it important for DO models to simulate the
MLDO correctly?
Most of the aerobic habitat in the bay during the
summer is located above the MLDO , thus it is critical for living resource managers to use models that
accurately simulate this variable.
On average, the models miss the observed depth of the
MLDO by 3.4 m, which equates to roughly a 60 % error
in the modeled mixed layer depths. While the models have
difficulty simulating the MLDO throughout the entire year
(Figs. 6, 7b), the summer months are when the mismatch
has the greatest potential to impact the available habitat for
oxygen-dependent species. Each year during this time period
low-oxygen waters occupy nearly the entire water column
below the mixed layer. At station CB4.1C, a representative
mesohaline deep trough station, the contours of low-oxygen
(5 mg L−1 ) and hypoxic (2 mg L−1 ) waters are located just
below the MLDO from late spring until late fall (Fig. 12).
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I. D. Irby et al.: Multiple modeling low-oxygen waters
Figure 12. Time series of observations of dissolved oxygen and
MLDO contours at Station CB4.1C for 2004 and 2005.
The severe depletion of oxygen below the mixed layer compresses the habitable space at this station to roughly 10 m
(from a maximum of 32 m) during the annual low-oxygen
event.
The impact of habitat compression can be substantial, as
many bay species require DO concentrations well above the
traditional hypoxic threshold (USEPA, 2010). While not all
of the main stem stations develop hypoxic water each year,
most mesohaline stations experience a dramatic drawdown of
oxygen to levels during the summer that effectively remove
a large portion of the bay from habitable space (Murphy et
al., 2011; Schlenger et al., 2013). Studies have shown that
some species modify their behavior based on the oxycline
depth, which acts to constrict the habitable space in the water column (Prince and Goodyear, 2006; Pierson et al., 2009;
Elliott et al., 2013). Since species can be negatively impacted
by low-DO concentrations as high as 5 mg L−1 (Breitburg,
2002; Vaquer-Sunyer and Duarte, 2008; USEPA, 2010), the
location of the oxycline is not only important for habitat compression in the summer months but can also be important in
the winter months when an occasional lack of vertical mixing
can substantially decrease bottom DO concentrations. Furthermore, in order to accurately estimate hypoxic volume,
models must correctly simulate the depth of the mixed layer,
since the MLDO closely follows the depth of the 2 mg L−1
contour.
4.4
How can DO simulations in the bay be improved
for management of water quality and living
resources?
To better simulate DO conditions and summer
habitat compression due to low-DO water, simulations of the depth of the top of the pycnocline
(MLDρ ) must be improved.
Biogeosciences, 13, 2011–2028, 2016
Although the suite of models examined reproduce DO
concentrations relatively well overall (Fig. 4), the models
typically overestimate summer habitat compression by producing low DO concentrations too high in the water column
(Fig. 6). Observations from the Chesapeake Bay Program
show a strong correlation between the depths of the oxygen and density-defined mixed layers (Fig. 10b). The models analyzed here also clearly exhibit a close relationship between their skill in simulating the depths of the oxygen and
density-defined mixed layers (Fig. 6). These strong relationships between the depths of the oxygen and density-defined
mixed layers result from the fact that the pycnocline represents the physical barrier that leads to the development of
the oxycline. Therefore, the inability of the models to accurately simulate habitat compression is an artifact of their lack
of skill in simulating the depth of the density-defined mixed
layer. In contrast, the strength of density stratification is not
well correlated to the strength of oxygen stratification. This
is because a relative wide range of intensities of density stratification is still sufficient to cut off vertical mixing, leading
to the observed draw-down in bottom DO. Thus, even though
all models underestimate the strength of the pycnocline, they
still produce enough stratification to greatly reduce mixing.
The results from this paper thus indicate that to further improve DO simulations and better estimate summertime habitat compression, it is even more critical for models to accurately simulate the depth of the top of the pycnocline than to
accurately simulate the absolute strength of the pycnocline.
4.5
What is the utility of the multi-model ensemble and
Model Mean?
The multi-model ensemble approach allows for the
development of a model mean, which taken as its
own model, is the most skilled model when examining the combined suite of variables analyzed in
this study.
The model skill assessment presented here demonstrates
that the average of all eight models, or five models in the case
of chlorophyll and nitrate, does better than any individual
model if looking across the suite of variables analyzed. This
finding is similar to that of other studies that examined the
value of the Model Mean from a multi-model ensemble (e.g.,
Gneiting and Raftery, 2005; Hagedorn et al., 2005). While
the concept of using a multi-model ensemble has been most
extensively employed by atmospheric, climatic, and global
circulation modelers, such as the Intergovernmental Panel on
Climate Change (e.g., Collins et al., 2013), the tool’s utility
for aquatic ecosystem modeling is gaining traction (Meier
et al., 2012; Trolle et al., 2014; Janssen et al., 2015). As
models are increasingly used in regulatory decisions regarding aquatic ecosystems, a cohort of similarly skilled models can be used to help inform a set of confidence bounds
around an environmental forecast. Due to the restrictions
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I. D. Irby et al.: Multiple modeling low-oxygen waters
placed on models used in regulatory actions, utilization of
a multi-model ensemble may not be realistic for all environmental and resource managers; however, multiple models can be integrated into the decision-making process even
when the ultimate decision must be based on a single model.
For example, a confidence interval plot could help identify
where regulatory model output might be acting out of sync
with other skilled water quality models of the same system,
thereby informing managers of the potential shortfalls associated with the regulatory model. Furthermore, if the models
tend to be predicting similar DO concentrations, a cohort of
models could enhance the confidence in regulatory decisions
based on a single regulatory model (Friedrichs et al., 2012;
Weller et al., 2013). Comparing multiple models can also
help inform how to better improve models in the future, as
this study has aimed to do.
2025
This study also helps to demonstrate how multiple community models from governmental agencies and academic institutions may be used together to provide a model mean and
a set of confidence bounds for regulatory model results that
could be used to inform management decisions.
Data Availability
Observations used in this analysis can be downloaded from
the Chesapeake Bay Program’s Water Quality Database
website at (http://www.chesapeakebay.net/data/downloads/
cbpwaterqualitydatabase1984present). Model output for the
individual stations examined in this analysis can be obtained by contacting Marjorie Friedrichs (marjy@vims.edu)
or downloaded from the Coastal & Ocean Modeling Testbed
– Estuarine Hypoxia THREDDS server (http://comt.sura.
org/thredds/catalog/comt2/cb_hypoxia/catalog.html).
5 Conclusions
All models analyzed here exhibited a high degree of skill in
simulating dissolved oxygen concentrations within the main
stem of the Chesapeake Bay in 2 years corresponding to relatively wet and average years. Their high skill results from
the fact that physical processes (e.g., solubility, wind-mixing,
and advection) exert a first order influence on the seasonal
cycle of oxygen. As a result, the models’ ability to reproduce
dissolved oxygen concentrations is independent of the complexity of the biogeochemical parameterizations: the simplest constant respiration models were found to reproduce
observed oxygen concentrations as well as the most biologically complex models. Essentially, all models are equally capable of respiring most of the available oxygen in the lower
water column during summer.
This study also suggests that for use as management tools
for water quality and living resources, it is more critical for
these models to adequately resolve the depth of the mixed
layer than the absolute strength of stratification (as long as
modeled stratification is strong enough to limit vertical mixing). This is critical because observations show that during
warmer months, oxygen-depleted water fills the water column to where stratification limits further mixing, which effectively cuts off waters below the mixed layer for use by the
majority of the Chesapeake Bay’s most recognized and valued living resources. These results furthermore suggest that
modelers should focus their efforts on improving the hydrodynamics of their models in an effort to improve simulations
of mixed layer depth dynamics and variability.
These findings have significant ramifications for shortterm bottom DO forecasts, which may be successful with
very simple oxygen parameterizations embedded in hydrodynamic models. In contrast, scenario-based water quality forecasts are likely to benefit from more complex models, which
must adequately reproduce the longer-term response of the
oxygen field to changes in nutrient and organic matter loads.
www.biogeosciences.net/13/2011/2016/
Acknowledgements. This work was supported by the NOAA IOOS
program as part of the Coastal Ocean Modeling Testbed. We thank
Yun Li and Younjoo Lee for help with the ROMS-RCA simulations
used in this analysis and Ray Najjar for his insights and comments.
This is VIMS contribution 3520 and UMCES contribution 5130.
Edited by: K. Fennel
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