EPA - 3O3 -R-95-OOf-
. j
Water Quality Functions of
Riparian Forest Buffer Systems in the
Chesapeake Bay Watershed
Prepared by the
Nutrient Subcommittee
of the
Chesapeake Bay Program
EPA 903-R-95-004
CBP/TRS 134/95
AUGUST 1995
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Water Quality Functions of
Riparian Forest Buffer Systems
in the
Chesapeake Bay Watershed
A Report of the
Nutrient Subcommittee
of the Chesapeake Bay Program
August 1995
Printed by the U.S. Environmental Protection Agency for the Cftesapeake Bay Program
Recycled/Recyclable • Printed with Vegetable Oil Based Inks on 100% Recycled Paper (50% Postconsumer) <
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Principal Authors
- RICHARD LOWRANCE ;
USDA-Agricultural Research Service, Tifton, GA
LEES. ALTER
USDA-Agricultural Research Service, Tifton, GA
J. DENIS NEWBOLD
Stroud Water Research Center, Avondale, PA
RONALD R. SCHNABEL
USDA-Agricultural Research Service,-University Park, PA
PETER M.GROFFMAN
Institute for Ecosystem Studies, Millbrook, NY .
; JUDITH M. DENVER
U.S. Geological Survey, Dover, DE
DAVID LCORRELL
Smithsonian Environmental Research Center, Edgewater, MD
J. WENDELL GILLIAM
North Carolina State University, Raleigh, NC
JAMES L ROBINSON
USDA-National Resources Conservation Service, Ft. Worth; TX
RUSSELL B. BRINSFIELD
University of Maryland, Wye Research Center, Queenstown, MD
KENNETH W.STAVER ,
University of Maryland, Wye Research Center, Queenstown, MD
--• : WILLIAM LUCAS
Integrated Land Management Consulting, Malvern, PA
ALBERT H.TODD ' :
USDA Forest Service, Chesapeake Bay Program, Annapolis, MD
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Preface
This document is a research synthesis requested
by the, Forestry Work Group of the Nutrient Sub-
committee of the Chesapeake Bay Program. In devel-
oping the outline for the report, the authors agreed to
specifically focus on the existing Riparian Forest
Buffer System (RFBS) specification developed by
USDA and being used as a starting point for federal,
state, 'and local RFBS specifications. Although the
report contains a general review of riparian forest and
grass vegetated filter- strip literature, the goal was to
use this literature to help determine the applicability
of the forest buffer system recommended by USDA.
The strategy for development ^of the document was
to bring together researchers in this field to: 1) dis-
cuss the current state of knowledge of RFBS;
2) determine how that knowledge related to the
Chesapeake Bay Watershed; and 3) reach consensus
about the functions of RFBS in the Bay watershed
based on that current state of knowledge. The con-
sensus statements are very important but they do riot
ensure specific functions will result from RFBS, in a
given field setting. Rather, they are Best Professional
Judgements of the entire authors group and represent
general agreement among the authors about the cur-
rent validity of the statements. In addition to the
authors, a large number of reviewers were asked to
. examine the report and form their own judgements
about the general conclusions. These reviewers,
acknowledged below, generally agree'with the con-
sensus statements contained in the report.
As readers of this report will see, numerous scien-
tific questions remain about the role of RFBS in all
of the physiographic and land use settings of the Bay
watershed. Yet, incomplete scientific knowledge can
not be used to avoid making informed management
judgements, especially when society has. determined
that a globally important natural resource such as
Chesapeake Bay must be restored to ecological
health. The scientists involved with the preparation
of this report have attempted to make the best judge-
,ments possible "to help guide the application of RFBS
to improve water quality in the Chesapeake Bay
Watershed and ultimately in the Bay itself.
v
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Ackn owledg e m e n ts
The authors are especially grateful to Mr. Spencer
Waller, who provided logistical .support for.our meet-;
ings and provided detailed minutes, of the discus- •
sions. In particular, Spencer transcribed about 25
hours of discussions which included our process of
reaching consensus on the- applicability of RFBS in
different physiographic settings. Without these direct
transcripts, we would not- have been able to ade-
quately recapture these statements for the report.
The authors also wish to thank the reviewers of the
report. Most of them spent a large amount of time on
their review and their .comments were very helpful in
completing the final draft. -The list of reviewers
includes many scientists and natural resource man-
agers active in the Chesapeake Bay Watershed. The .
reviewers were Andrew Dolloff (USDA-FS), Richard
Everett (U.S. Fish and Wildlife Service), Verna
Harrison (Maryland Dept. of Natural Resources),.
Thomas Jordan (Smithsonian Environmental
Research Center), Larry Lubbers (Maryland Dept. of
Natural Resources), Kent Mountford (U.S. En-
vironmental'Protection Agency), Adel S^hirmo^
hammadi (University of Maryland), George
Simmons (Virginia Polytechnic Institute & State
University), Thomas Simpson (Maryland Dept. of
Agriculture), Bernard Sweeny (Stroud Water
Research Center), and Donald Weller (Smithsonian'
Environmental Research Center).
A number of people participated in at least one of
our two meetings but are not-co-authors of the report.
Among these people who provided useful input at the
meetings were Ed_Corbett (USDA-Forest Service),
Rupert, Friday (Chesapeake Bay Foundation), Bob
Merrill (Pennsylvania Bureau of Forestry), Kent
Mountford (U.S. Environmental Proteetion Agency),
Ann Swanson (Chesapeake Bay Corfmiission), Bob
Tjaden (Univ, of Maryland), and Dave Welsch
(USDA-FS).
As with any undertaking of this sort, much of the
work was done by people who get little credit. Ms.
Olive Sides, Office Automation Assistant with
USDA-ARS, Tifton, GA helped with much of the
correspondence and arrangements for th& two meet- .
ings held to develop the report. Ms. Dalma Dickens,
Secretary with USDA-ARS, Tifton, typed numerous
versions of the report. -Mr. H. L. Batten, USDA-
ARS, Tifton, Ms. Wendy R. Pierce, USDA-NRCS,
Ft."Worth, Texas, and Mr. Anthony J. Kimmit,
USDA-NRCS, Ft. Worth, Texas prepared many of the
figures.
Funding for this, report was provided by the
Environmental Protection Agency, Chesapeake Bay
Program, and USDA Forest Service, Northeastern
Area, State & Private Forestry. '
VII
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Con
LIST OF TABLES AND FIGURES ... ........., : !.......... xl
EXECUTIVE SUMMARY ...:................. . . . xiii
SECTION I. Water Quality Functions of Riparian Forest Buffer Systems ....... ... 1
; A. Introduction .......... .-".' 1
B. Nonpoint Source Pollution Control Relative to Nutrient Load Reduction Strategies . 1
: . C. Watershed Approaches to Nonpoint Pollution Estimation and Abatement 4
D. Historical Overview of Scientific Interest in Riparian Ecosystems :. . . 5
E. Research Background for the Riparian Forest Buffer System Specification 5
F. Current Understanding of RFBS Functions '....'.,... •„.." .".'... 7
1. Zone 1 - Control of the Stream Environment, '. 7
a. Temperature and Light ; '....... 8
b. Habitat Diversity and Channel Morphology ....'. '.,.','.,* 9
c. Food Webs and Species Diversity 10
2. Zone 2—Removal of Nonpoint Source Pollutants ... . .......-,'., . 11
a. Nitrate Removal 1.,.,.. 11
, b. Plant Uptake ;.'...;........' ... 12
c. Microbial Processes ............;..........'. '...'...- 14
d. Removal of Surface-borne Pollutants ......'............:..... 15
• , • .3. Zone S^Sediment Removal and Spreading of Surface Runoff 1,6
4. Integrated Water Quality Functions of the Three-Zone Buffer System 17
SECTION II. Riparian Forest Buffer Systems in Physiographic Provinces of the
Chesapeake Bay Watersheds .,.....: 19
A. Coastal Plain 19
j • • 1. General Land Use and Hydrology ,19
a. Inner Coastal Plain , 19
. ' - . b. Well-Drained Upland :'. .........' 20
c, Poorly-Drained Upland , 20
• . .. d'. Surficial Confined . .'. . . ..: '.. . . 20.
2. Control of Nonpoint Source Pollutants .'. 22
: a. Nutrient Budgets for Riparian Forests 23
b. Removal of Nitrate from Groundwater ."., 24
c. Nutrient Removal Processes ..... . ........ 26
, d. Removal of Sediments arid Nutrients from Surface Runoff .^ ........... 28
3. Conclusions 31
IX
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B. Piedmont .:..... 31'
1. General Land Use andHydrology 31
2. Control of Nonppint Source Pollutants 32
a. Removal of Nitrate from Groundwater . .*, 32
b. Removal of Sediment and Nutrients in Surface Runoff . . 33
3. Conclusions 34
C. Valley and Ridge ..' ::. • 35
1. General Land Use and Hydrology 35
2. Control of Nonpoint Source Pollution . 35
a. Removal of Nitrate from Groundwater • 36
b. Streamflow Transport of Phosphorus 37
c. Removal of Sediment and Nutrients in Surface Runoff 38
3, RFBS in Forested Watersheds .. 1 38
4. Conclusions 39
SECTION III. Applicability of the Three Zone Riparian Buffer System 41
A. Control of the Stream Environment 41
B. Control of Nonpoint Source Pollution .....'..... . 41
1. CoastalPlain .......; .., . . 42
a. Inner Coastal 42
b. Outer Coastal Plain , • • 46
1) Well Drained Upland . 46
2) Poorly Drained Upland/Surficial Confined 47
c. Tidally Influenced 47
2. Piedmont 49
3. Valley and Ridge 49
C. Loading Rates and Nonpoint Source Pollution Control .......: . 52
D. Stream Order/Size 52
E. Establishment and-Sustainability 53
SECTION IV. Research Needs 55
LITERATURE CITED 57
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Tables and Figures
TABLES
1. Land use in physiographic regions Of Chesapeake Bay Watershed .2
2. Total nitrogen, nitrate-nitrogen, and total phosphorus budgets for riparian forest "
ecosystems in the Coastal Plain ; 23
3. Above-ground woody vegetation uptake of N and Pin Coastal Plain riparian forests . .. : 26
4. Sediment deposition in Coastal Plain riparian forests ..... ~ 28
5. Inputs, outputs, and % removals of sediment, total N, and total P from experimental
Ky 31-Fescue vegetated filter strips in Maryland Coastal Plain ......... 29
' 6. Effects, of different size buffer zones on reductions pf.sediment and nutrients from field surface runoff . 30
, 7. Removal of nitrate from groundwater—summary. .......:.. -43
8. Removal of phosphorus from.all sources—summary ......: ... 44
9. Removal of sediment and sediment-borne pollutants—summary ....:......... .^.... 45
FIGURES
1. Physiographic regions of the Chesapeake Bay Watershed ..'.....' 3
2.- Schematic of the three zone Riparian Forest Buffer System ............:.............,........ 6
3. Hydrogeomorphic regions of the Delmarva Peninsula 21
4. Nitrate concentrations in groundwater beneath riparian forests from five Coastal Plain sites . 25
5 , Conceptual model of below ground processes affecting groundwater nutrients in riparian forest ..... 27
6. Inner Coastal Plain ..'..; t.-......'. .......'-.:..- .......... 46
7. Outer Coastal Plain - Well-Drained Upland .... .... . . 46
8. Outer Coastal Plain - Poorly Drained Upland/Surficial Confined .... . 48
9. Coastal Plain-Tidal Influenced flow system . s .' 48
10. Piedmont (thin soils) ..../. ...,.' .......:............ 50
11. Piedmont (schist/gneiss bedrock) ...." 50
12. Piedmont (marble bedrock)/Valley and Ridge (limestone bedrock) . .. , '50
13. Valley and Ridge (sandstone/shale bedrock) , '. .-.,.'.... 51
14. Valley and Ridge/Appalachian (low order streams) :'... 1 .................. 51
XI
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Executive Summary
Riparian (streamside) forests are known to reduce
delivery of nonpoint source pollution to streams and
lakes in many types of watersheds. In addition, ripar-
ian forests are known to be important in controlling
the physical and chemical environment of streams
and in providing detritus and woody debris for
streams and near-shore areas of water bodies.
Riparian forests were the original native vegetation in
most streamside areas of the .Chesapeake Bay
Watershed. This report assesses the state of scientific
knowledge concerning the Water Quality Functions
of riparian ecosystems. This, assessment and specific
knowledge of riparian ecosys'tem function in physio-
graphic regions of the Chesapeake Bay Watershed
were used to make consensus Best Professional
Judgements as to the potential water quality func-
tions of Riparian Forest Buffer Systems (RFBS) in
the Bay Watershed. ,
Research conducted in naturally occurring riparian
forests and experimental and on-farm grass filter
strips has been used by the U.S. Department of
Agriculture to develop a general "Riparian Forest
Buffer System specification" for controlling nonpoint
source pollution from agriculture and improving gen-
eral water quality. The specification calls for a three
zone buffer system, with each zone having specific
purposes but also having interactions with the adja-
cent zones to provide the overall RFBS function.
Zone 1 of the RFBS is an area of permanent forest
vegetation immediately adjacent to the stream chan-
nel and encompassing at least the entire stream chan-
nel system. Zone 2 is an area of managed forest, ups-
lope' from Zone 1. .Zone 2 is managed for control of
pollutants in subsurface flow and surface runoff
through biological and chemical transformations,
storage in woody vegetation, infiltration, and sedi-,
ment deposition. Zone 3 is a grass or other herba-
ceous filter strip upslope from Zone 2. Zone 3 is
managed to provide spreading of concentrated flow
into sheet flow and to remove sediment and sediment
associated pollutants. '...,'•.'
The most general function of Riparian Forest
Buffer Systems is, to provide control of the stream
environment. These functions include modifying
stream temperature, and controlling light quantity and
quality; enhancing habitat diversity; modifying chan-.
nel morphology; and enhancing food webs and
species richness. All of these factors are important to
the ecological health of a stream and are best pro-
vided by a RFBS which includes a Zone 1 that
approximates the original native vegetation. These
functions occur along smaller streams, regardless of
physiographic region. These functions are most
important on smaller streams, although they are
important for bank and near-shore habitat on larger
streams and the shoreline of tlie Bay. RFBS
contribute to bank stability and thus minimize sedi-
ment loading due to instream bank erosion.
Depending on bank stability,-and soil .conditions in
Zone 1, management of Zone 2 for long-term rota-
tions may be necessary for sustainability of stream
environment functions of Zone 1.
The next most general function of RFBS is control
of sediment and sediment-borne pollutants carried in.
surface runoff. Properly managed RFBS should pro-
vide a high level of control of sediment and sediment
borne chemicals regardless of physiographic region.
Natural riparian forest studies indicate that forests
are particularly effective in filtering fine sediments
and promoting co-deposition of sediment as water
infiltrates. The slope of the RFBS is the main factor
limiting the effectiveness of the sediment removal
function. In all physiographic settings it is important
to convert concentrated flow to sheet flow in "order to
optimize RFBS function. Conversion to sheet flow
arid deposition of coarse sediment which could dam-
age young vegetation are the primary functions of
Zone 3-—the grass vegetated filter strip.
XIII
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XIV
The next most general function of RFBS is to con-
trol nitrate in shallow groundwater moving toward
streams. When groundwater moves in short, shallow
flow paths, such as in the Inner Coastal Plain (pri-
marily the western shore), 90% of the nitrate input
may be removed. In contrast, nitrate removal may be
minimal in areas where water moves to regional
groundwater such as in Piedmont and Valley and
Ridge areas with marble or limestone bedrock,
respectively. In these and some Outer Coastal Plain
regions, high nitrate groundwater may emerge in
stream channels and bypass most of the RFBS. In the
areas where this occurs or where high nitrate water
moves out in seepage faces, deeply rooted trees in'
Zone 1 or in seepage areas will be essential. The
degree to which nitrate (or other groundwater pollu-
tants) will be removed in the RFBS depends on the
proportion of groundwater moving in or near the bio-
logically active root zone and on the residence time
of the groundwater in these biologically'active areas.
The least general function of RFBS appears to be
control of dissolved phosphorus in surface runoff or
shallow groundwater. Control of sediment-borne P is
generally effective. In certain situations, dissolved P
can contribute a substantial amount of total P load.
Most of the soluble P is bioavailable, so the potential
impact of a unit of dissolved P on aquatic ecosystems
is greater. It appears that natural riparian forests have
very low net dissolved P retention. In managing for
increased P retention, effective fine sediment control
should be coupled with use of vegetation which can
increase P uptake into plant tissue:
Research on functions of natural, restored, and
enhanced RFBS is needed in all portions of the
Chesapeake Bay Watershed. Research should be
directed into four general areas: 1) assessment of
existing riparian forests relative to the RFBS stan-
dard; 2) assessment of potential RFBS restoration for
NPS pollution control; 3) assessment of NFS pollu-
tion control in pilot restoration and enhancement
projects; 4) determine the effects of management fac-
tors on both pollution control and control of the
stream environment. The research, because of the
need to do relatively large scale projects which last
for substantial periods of time, should be coordinated
with demonstration restoration/ enhancement pro-
jects. Some of the major research questions should
address the uncertainty associated with the functions
discussed above. Research should be directed toward
testing the hypotheses concerning which functions of
RFBS occur in specific physiographic settings and
the specific management conditions under which
these functions are likely to be enhanced. In'particu-
lar, research on the time to recovery of RFBS func-
tions and the processes which control the various
functions should be integrated into demonstration
projects.
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Water Quality Functions of
Riparian Forest Buffer Systems
A. INTRODUCTION
Riparian Forest Buffer Systems (RFBS) are
streamside ecosystems, managed for the enhancement
of water quality through control of noripoirit source
pollution (NFS) and protection of the stream environ-
ment. The use of riparian management zones is rela-
tively well established as a Best Management Practice
(BMP) for water quality improvement (in forestry
practices (Comerford et al., 1992), but has been much
less widely applied as a BMP in agricultural areas or
in urban or suburban settings. RFBS are especially
important on small streams where intense interaction
between terrestrial and aquatic ecosystems occurs.
First and second order streams comprise nearly three-
quarters of the total stream length in the United States
(Leopold et al., 1964). Fluvial activities influence the
composition of riparian plant communities along
these small streams (Gregory et al., 1991). Likewise,
terrestrial disturbances can have an immediate impact
on aquatic populations (Sweeney, 1993; Webster et
al., 1992). Small streams can be completely covered
by the canopies of streamside vegetation (Sweeney,
1992). Riparian vegetation has well-known beneficial
effects on the bank stability, biological diversity, and
water temperature's ;of streams (Karr and Schlosser,
1978). Riparian forests of mature trees (30 to 75 yrs.
old) are known to effectively reduce nonpoint pollu-
tion from agricultural fields (Lowrance et al., 1985b).
Compared to other NFS pollution control mea-
sures, RFBS can lead to longer-term changes in the
structure and function of agricultural landscapes. To
produce- long-term improvements in water quality,
RFBS must be designed with,an understanding of: 1)
the processes which remove or sequester pollutants
entering the riparian buffer system; 2) the effects of ri-
parian management practices on pollutant, retention;.
3) the effects of riparian forest buffers on aquatic
ecosystems; 4) the time to recovery after harvest of
trees, or reestablishment of riparian buffer systems;
and 5) the effects of underlying soil and geologic ma-
terials on chemical, hydrological, and biological
processes. . '
. This report examines the scientific basis for apply-
ing the existing RFBS specification as an agricultural
Best Management Practice (BMP) in the. different
physiographic provinces of the Chesapeake Bay
Watershed (CBW, -Table 1 and Figure 1). The report
briefly reviews NPS pollution problems in the Bay
Watersheds and approaches to NPS pollution control
(Sections I. B & C); the scientific foundation for the
Riparian Forest Buffer System specification (Sections
I. D & E); and the water quality functions of each of
the three zones of the RFBS (Section I. F). Included
is a review of the existing research on RFBS in dif-
ferent physiographic provinces that comprise the
CBW (Section II). Based on these results, the effec-
tiveness of RFBS for NPS pollution control is charac-
terized in different parts of the Bay watershed
(Section III). Finally, research needs are discussed in
Section IV. RFBS are one of many factors that influ-
ence water quality and stream health. A complex suite
of interrelated functions and mechanisms contribute
to water quality and physical habitat parameters of the
aquatic ecosystem. Other important factors, outside
the scope of this report, that may affect the function-
ing of RFBS, and should be considered in their de-
sign, include: the type and intensity of land use in the
watershed;- the effectiveness of stormwater manage-
ment; streambank and streambed stability; and stream
uses (recreation, water supply, etc.).
B. NONPOINT SOURCE POLLUTION
CONTROL RELATIVE TO NUTRIENT
LOAD REDUCTION STRATEGIES
Nonpoint source pollution is the major cause of
surface water impairment in the United States (Baker,
1992; Long, 1991) and has been addressed as a na-
tional priority since passage of the Clean Water Act
(CWA), Section 319, which requires "that programs
for the control of nonpoint sources of pollution be de-
1
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TABLE 1
Land use in physiographic regions of Chesapeake Bay Watershed (NCRI Chesapeake, 1982).
Physiographic
Region
Appalachian
Plateau
Valley & Ridge
Piedmont
Coastal Plain
TOTAL
% of TOTAL
Crop
, 659,700
986,200
825,500
768,600
3,240,000
22%
Forest
2,611,100
2,659,600
. 1,607,900
1,020,400
7,899,000
55%
Wetland
.............ha—- ~ --
181,400
60,500
141,300
509,300
892,500
6%'
Other
658,800
911,100
688,100
119,800
• 2,377,800
17%
TOTAL
4,111,000
4,617,400
3,262,900
'2,418,000
14,409,300
100%
%of
TOTAL
28
32
23
17
100
'
Rgure 1 shows generalized physiographic regions.
veloped and implemented." The effectiveness of the
RFBS is likely to be judged by their NFS pollution
control effectiveness.
Although assessments are incomplete and do not
include all states, estimates are that about 30% of US
waters are impaired—i.e. they do not fully support
their designated uses (USEPA, 1990a). Of impaired
waters, about two-thirds of the problems are primarily
from NFS pollution (USEPA, 1986). The nonpoint
sources of pollution vary, but agriculture is the major
contributor for rivers and lakes. Besides agriculture,
the other major contributors of NFS pollution are
urban areas, mining, atmospheric deposition, and nat-
ural origins. Nutrients and sediments are still the prin-
cipal sources of surface water impairment (USEPA,
1986; USEPA, 1990a; USEPA, 1990b). Sediments are
the most important cause of impairment for rivers,
and nutrients are the most important cause of impair-
ment for estuaries. Pesticides, metals, and priority
pollutants are identified as problems in less than 20%
of the assessed waters. The extent of contamination,
especially for pesticides, may be underestimated.
The earliest assessments of Chesapeake Bay water
quality in the 1980's identified non-point source pol-
lution as a major cause for water quality impairment
in the Bay (Correll, 1987; Chesapeake Bay Program,
1991). Reduction of NPS pollution has been a signif-
icant part of the strategy to improve water quality in
Chesapeake Bay since that time. The main problems
were identified as nutrient enrichment, high levels of
toxic substances, and excessive sediment loads.
Effective control of all these types of pollution, espe-
cially nutrients and sediments, requires a watershed
based program for NPS pollution control.
Improvement and maintenance of water quality is
the single most important component of the overall
protection and restoration plan established in the 1987
Chesapeake Bay Agreement (Chesapeake Bay
Program, 1991). One of the most ambitious goals of
the 1987 and 1992 agreements is to reduce nutrient
loadings to the Bay by 40% by the year 2000 and to
retain this level as a permanent cap on nutrient levels.
Strategies for nutrient load reduction require control
of both point and nonpoint sources of pollution.
Based on 1985 land uses and results from a Watershed
Model (Donigian etal., 1990), nonpoint sources dom-
inate both N (53% of total) and P (68% of total) loads
to the mainstem of the Bay (Chesapeake Bay Pro-
gram, 1991, L. Shu'yler, personal communication,
1995). The Watershed Model has been used to esti-
mate edge-of-stream nitrogen and phosphorus load-
ings from various land uses in the Bay watersheds.
Agriculture (including conventional cropland, conser-
vation cropland, pasture, and animal waste facilities)
accounted for 69% ,of total N and 79% of total P in
NPS pollution.
Of the entire loadings of N and P to the mainstem
of the Bay (point plus NPS), 44% of the N and 50%
of the P came from agricultural nonpoint sources (L.
Shuyler, personal communication, 1995). The
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Key
Coastal Plain'
Southern Piedmont
Northern .Piedmont
Appalachain
Ri
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Susquehanna Basin (Pennsylvania, New York) and the
Eastern Shore (Delaware, Maryland, Virginia) con-
tribute the highest NFS loads of N and P. Loads from
these two regions were dominated by agricultural
sources. In the Susquehanna, 74% of NPS N loads
were from agriculture. Nonpoint source loads of N,
from the Eastern Shore were 81% agriculturally re-
lated (Chesapeake Bay Program, 1991).
Until 1990, approaches for NPS pollution control
in the Bay watersheds were largely focused on con-
trolling upland sediment and sediment-borne pollu-
tants (Chesapeake Bay Program, 1990). These tradi-
tional approaches were a combination of source
reduction (i.e. reduce erosion rates in fields) and en-
gineered buffer systems or structural BMPs such as
ponds, sediment detention basins, terraces, grass
water ways and vegetated filter strips. In 1990, the
Chesapeake Bay Program's Nonpoint Source
Evaluation Panel recommended a systems approach
for nutrient load reduction with regional and water-
shed management strategies based on watershed
mass-balances (Chesapeake Bay Program, 1990). A
systems approach for NPS pollution reduction will in-
clude structural BMPs and source load reductions, as
well as approaches which seek to integrate the man-
agement and restoration of landscape features which
retain pollutants through a combination of ecosystem
processes. Examples of these pollutant sinks include
natural wetlands, constructed wetlands, and riparian
forest buffer systems (Fields, 1992). As pollutant
sinks increase in complexity from simple physical
structures to diverse natural ecosystems, both the im-
portance and difficulty of understanding processes
which sequester or remove pollutants also increase.
C. WATERSHED APPROACHES
TO NONPOINT POLLUTION
ESTIMATION AND ABATEMENT
Risk assessment and source reduction are new ap-
proaches for NPS pollution control (Baker, 1992). A
high percentage of total pollutant loadings in some
watersheds comes from a relatively small portion of
the watershed area because of improper management
of sources, improper siting of facilities, problematic
environmental and site conditions, or a combination
of these factors. Watershed scale risk assessment
seeks to identify and reduce loadings from areas
which contribute large amounts of NPS pollution.
'Concurrent with identification of problem areas
comes the opportunity for source reduction. Source
reduction has been responsible for some of the more
impressive successes of NPS pollution reduction, in-
cluding the reduction of loadings of lead from auto-
mobile emissions and of prgand-chlorine pesticides
(Baker, 1992). Source reduction should be linked with
watershed-scale risk assessments because the poten-
tial for source reduction may be greatest (and proba-
bly most economical) in areas which are generating
highest unit area loadings. The linkage of risk assess-
ment and source reduction will depend on interacting
factors such as type of pollutant (e.g. purchased input
vs. by-product), reason for high risk (e.g. poor man-
agement, siting of facilities, inherent regional risks),
and availability of alternative practices'and/or sites.
Even when risk assessment and source reduction
strategies lead to load reductions under average con-
ditions, a third aspect of watershed management -
maintenance and restoration of buffer systems be-
tween terrestrial and aquatic ecosystems - is neces-
sary to reduce the contributions of extreme events to
NPS pollutant loads. Under the best of conditions,
source reduction will likely leave watersheds vulner-
able to extreme events, including both weather ex-
tremes as well as economically generated extremes
(e.g. intensification of pollutidn generating produc-
tion practices). Watershed studies have demonstrated.
the importance of extreme events to water and pollu-
tant transport. Extreme events within a year dominate
annual totals and wet years within multi-year cycles
dominate long-term loadings (Jaworski et al., 1992;
Lowrance and Leonard, 1988; Magnien et al., 1992).
Control of NPS pollution from extreme events will re-
quire integrating risk assessment and source reduction
approaches with buffer systems as landscape scale
"insurance policies."
Buffer systems are also important components of
watershed NPS pollutant control efforts because of
•the limitations of other BMPs for NPS pollution con-
trol. For example, Hall (1992) monitored changes in
groundwater nitrate (NOs—N) concentrations be-
neath heavily fertilized and manured fields in
Lancaster County, PA following the implementation
of "input management" techniques. Fertilizer/manure
inputs were decreased from 39 to 67% (222 to 423 kg
ha'1) but groundwater nitrate, changed by -12 to 50%.
By the end of the. study, nitrate concentrations in
groundwater still exceeded federal drinking water
standards. Shirmohammadi et al. (1991) used the
CREAMS simulation model to evaluate the effects of
seven different BMPs on groundwater nitrate concen-
trations beneath cropping systems on the eastern
shore of Maryland. Although CREAMS does not pro-
-------
vide absolute predictions, none of the BMPs'were
predicted,to reduce groundwater nitrate concentra-
tions to less than the federal drinking water standard.
Under appropriate conditions, described in this report,
RFBS are likely;to be an important component
of NPS pollution control when in-field BMPs are in-
adequate. ,
D. HISTORICAL OVERVIEW OF
SCIENTIFIC INTEREST IN
RIPARIAN ECOSYSTEMS
Most of the knowledge of riparian ecosystem ef-
fects on water quality comes from research conducted
since,1975. Two publications in 1978 galvanized sci-
entitle and management interest in riparian ecosys-
tems. Karr and, Schlosser (1978) concluded that
stream environments are largely controlled by adja-
cent riparian ecosystems and provided ah overview of
relationships between water, resources and riparian
ecosystems (the land-water interface). Johnson and
McCormick (1978) edited the proceedings of a sym-
posium which included 55 reports on various aspects
of riparian research^ management, and policy; While
the symposium proceedings contained excellent dis-
cussions of the late 1970's state-of-knowledge con-
cerning riparian ecosystems and other types of wet-
lands (Brown et al., 1978; Wharton, and Bririson,
1978) only one paper (Mitsch, 1978) dealt specifi-
cally and quantitatively with the water quality func-
tions of a riparian ecosystem. The proceedings also
included a review of the general water quality func-
tions of wetlands (Kibby 1978) in which a number of
publications on nutrient cycling in riparian and other
wetlands were cited Only a few of the citations dealt
specifically with water quality effects of riparian
ecosystems (Kitchens et al., 1975; Lee et al., 1975;
Kuenzler et al., 1977; Richardson et al., 1978).
Although the 1978 symposium contained numerous
claims about the water quality functions of riparian
ecosystems, few data were presented.
In the late 1970's a number of research projects
began to develop a more quantitative understanding of
the role played by riparian ecosystems in controlling
NPS pollution by sediment and nutrients in agricul-
tural watersheds (Jacobs and .Gilliam, 1985b,
Lowrance efal., 1983; Peterjohn and Correll, 1984).
These studies were. primarily in .the Coastal Plain'
physiographic province of the Eastern U.S., where the,
typical land-use pattern is intensive row-crop agricul-
ture in upland areas with riparian forests along low-
order streams. These early studies shared at least two
otherlmportant characteristics: 1) a relatively shallow
aquiclude which forced most infiltrated water to move
laterally toward streams and pass through or near the
riparian forest root zone and 2) naturally regenerated
forests typical of the region rather, than'forests man-
aged specifically for water quality functions. These
studies focused on riparian processes related to nutri-
ents and sediment with little or no attention to the
fates of other pollutants or to the effects of riparian
areas on the physical or trophic status of the stream.
As interest in the nonpoint pollution control value
of riparian ecosystems increased, recognition of their
importance to the physical and trophic status of
streams also developed.' Karr and Schlosser (1978)
quantified the effects of riparian vegetation on sun-
light penetration and temperature of streams.
Research in the 1980s confirmed the importance of
large woody debris and leaf litter inputs to the habitat
and trpphic status of most small streams (Meyer and
O'Hop, 1983; Benke et al., 1985; Harmon et al.,
1986). By 1987, it was well established that woody
debris derived from riparian forests played an impor-
tant role in controlling channel morphology, the stor-
age and routing of organic matter and sediment, and
the amount and quality of fish habitat (Bisson et al.,
1987).
E. RESEARCH BACKGROUND FOR
THE RIPARIAN FOREST BUFFER
SYSTEM SPECIFICATION
. By the late 1980s, there was a clear need to syn-
thesize the existing knowledge into management rec-
ommendations for the establishment, maintenance,
and management of riparian ecosystems for a broad
range of water quality functions (Lowrance, 1991). In
1991, the United States Department of Agriculture-
Forest Service (USDA-FS) with assistance from
USDA-Agricultural. Research" Service, USDA-Sqil
Conservation Service, Stroud Water Research Center,
PA, Pennsylvania Dept. of Environmental Resources,
Maryland Dept. of Natural Resources, and U.S. Dept.
of Interior Fish and Wildlife Service developed draft
guidelines for riparian forest buffers. This effort re-
sulted in a booklet entitled "Riparian Forest Buffers -
Function and Design for Protection and Enhancement
of Water Resources" (Welsch, 1991) which specified
a riparian buffer system consisting of three zones
(Figure 2). ' - , :
Zone 1 is permanent woody vegetation immedi-
ately adjacent to the stream bank. Zone 2 is managed
forest occupying a strip^upslope from Zone 1. Zone 3
-------
ZONES
Runoff Control
ZONE 2
Managed Forest
ZONE1
Undisturbed Forest
Streambottom
FIGURE 2. Schematic of the three zone Riparian Forest Buffer System^
is an herbaceous filter strip upslope from Zone 2. The
specification applies to areas where cropland, grass-
lands, and/or pasture are adjacent to riparian areas on
a) permanent or intermittent streams, b) margins of
lakes and ponds, c) margins of wetlands, or d) mar-
gins of groundwater recharge areas such as sinkholes.
Although referred to as a riparian forest buffer, inclu-
sion of the non-forested herbaceous strip as Zone 3
suggests that a more correct name would be "Riparian
Forest Buffer System".
The primary purposes of Zone 3 of the RFBS are
to remove sediment from surface runoff and to con-
vert channelized flow to sheet flow. The primary
function of Zone 2 is to block transport of sediment
and chemicals from upland areas into the adjacent
wetland or aquatic ecosystem. Vegetation and litter in
these zones forms a mechanical barrier to sediment
transport. Plant roots take up chemicals that become
sequestered in growing biomass. Vegetation also pro-
duces organic matter that fosters chemical and bio-
logical processes that immobilize or transform pollu-
tants. Although most Zone 2 functions also' occur in
Zone 1, the primary purpose of Zone 1 is to maintain
the integrity of the stream bank and a favorable habi-
tat for aquatic organisms. Shade and litterfall pro-
vided by streamside vegetation has a direct influence
on water temperature and dissolved chemicals.
The USDA-FS report and specification were
based on a synthesis of literature existing through
1989 and on in-depth discussions with scientists and
managers working on various riparian ecosystems
(Welsch, 1991). Some of the generalizations which
guided the design of the RFBS were based on studies
of nutrient sequestering and nutrient transformations
in agricultural watersheds (Correll, 1983; Lowrance
et al., 1985; Yates and Sheridan, 1983).-These water-
shed-scale studies indicated that riparian forests were
important nutrient and sediment sinks in agricultural
watersheds, but provided little or no guidance,on how
to design an effective RFBS. Process studies in these
and other systems provided most of the original de-
sign guidance. Several studies on nitrate removal
from shallow groundwater in riparian forest buffers
found that most reduction in nitrate concentration
takes place within the first 10 to 15 m of forest
(Lowrance et al., 1984a, Peterjohn and Correll, 1984,
Jacobs and Gilliam, 1985b) and that the necessary-
width for shallow groundwater nitrate removal could
be relatively short. Although effective in reducing
sediment arid sediment borne chemical concentrations
-------
in sheet flow (Peterjohn and Correll, 1984), it was
known that channelized flow can bypass riparian
forests. To control channelized flow into a 'riparian
forest, a herbaceous strip in Zone 3 could be much,
more easily reshaped and revegetated than a forest.
Herbaceous buffers, especially grass filters, are effec-
tive at removing coarse suspended sediments and
some sediment-borne pollutants but may require fre-
quent maintenance and are not very effective at nutri-
ent removal from shallow groundwater (Dillaha et al.,
1989; Magette et al., 1987; Magette et at 1989). -
Long-term sequestering and removal of nutrients
and'other contaminants in the RFBS is the main pur-
pose of Zones 3 and 2. This can occur by 1) accumu-
lating sediment and adsorbed contaminants; 2) micro-
bial transformations (for N) and biochemical
degradation (for pesticides); and 3) incorporation of
nutrients and other chemicals into woody, biomass
and soil organic matter. At least one study of Coastal
Plain riparian forests showed substantial amounts of
nutrient sequestering in Woody biomass (Fail et al.,
1986). The RFBS specification encourages produc-
tion and harvest of woody biomass from Zone 2 to re-
move nutrients and other contaminants. Once vegeta-
tion has been removed from the stream channel,
recovery through plant succession may take long pe-
riods of time and revegetation may be dominated by
undesirable species (Sweeney, 1993). Therefore, the
need for permanent control of the stream physical and
trophic environment requires directed succession to-
ward desirable permanent vegetation in those portions
of the RFBS which directly influence the; stream
channel, in particular Zone 1.
A number of practical concerns were also con-
sidered in the RFBS specification (Welsch, 1991).
Most of the RFBS should be available, for manage-
ment to provide an economic return without sacrific-
ing water quality functions. Characteristics of soils,
hydrology, and potential vegetation should guide de-
sign and planning of effective RFBS. The RFBS
should be used in .conjunction with sound upland
management practices including nutrient manage-
ment and erosion control. In-stream woody debris're-
moval should be limited, but woody debris with po-
tential to form dams which cause inundation should ;
be removed. The dimensions of the RFBS-should de-
pend on the existing and potential NPS pollutant
loads and the minimum size for sustained support of
the aquatic environment.
F. CURRENT UNDERSTANDING
Of Rf BS FUNCTIONS
Several of studies are underway to test the effec-
tiveness ,of RFBS which correspond to or are similar
to the USDA specification. Vellidis et al. (1993) and
Sweeney (1993) describe RFBS restoration projects
in the Georgia Coastal Plain and the Pennsylvania -
Piedmont, respectively. Beare et al. (1994) describe
preliminary results from management of an existing
riparian forest which involves establishment of Zone
3 adjacent to mature riparian forest and tree harvest
treatments in Zone 2. Schultz et al. (1994) describe a
multi-species three zone buffer system for use in agri-
cultural areas.of Iowa and other parts of the Midwest.
Much of the current understanding of RFBS has been
incorporated into a Riparian Ecosystem Management
Model which simulates hydrologic and nutrient cy-
cling processes in RFBS that conform to the .USDA
specification (Altier et al., 1994; Sheridan et al.,
1993).
It is important to note that bur current understand-
ing of the functions, of the RFBS is based on studies
that have been done in areas where riparian forests
currently exist due to a combination of hydrology,
soils, cultural practices, and economics. Most of our
current knowledge of the functions of the three zones
of the RFBS specification is derived from studies in
existing riparian forests and on experimental and real-
world grass buffer systems. Although results can.be
extrapolated from these existing forests to restored
RFBS, most of the study sites are actually at some
stage of restoration, following clearing within the last
20-80 years. . \ '-^
1, Zone 1—Control of the Stream
Environment
Although reduction of NPS pollution is a -widely
recognized function of RFBS, they also contribute
significantly to other aspects of water quality and
physical habitat (Allan and Flecker, 1993; Karr,
1993). Habitat alterations, especially channel straight-
ening and removal of riparian vegetation, continue, to
impair the ecological health of streams more often
and for longer time periods than toxic chemicals
(Hughes et al., 1990). Sweeney (1992) considers loss
of riparian forests in eastern North America to be one
of the major causes of aquatic ecosystem 'degradation.
Zone 1, the permanent woody vegetation at the
stream edge, enhances ecosystem stability and helps
control the" physical, chemical, and trophic status of
-------
8
the stream. Healthy riparian vegetation in Zone 1 also
contributes to bank stability and minimizes instream
sediment loading due to bank erosion. Zone 1 also has
substantial ability to control NFS pollution through
denitrification (Ambus and Lowrance, 1991; Low-
ranee, 1992; Schnabel, 1986), sedimentation (Low-
rance et al., 1986), or direct root uptake of pollutants.
Zone 1 vegetation controls light quantity and qual-
ity, moderates temperature, stabilizes channel geome-
try, provides tree roots and woody debris for habitat,
and provides litter for detritivores (Barton et al., 1985;,
Beschta et al., 1987; Hax and Golladay, 1993; Hill
and Harvey, 1990; Karr and Schlosser, 1978;
Sweeney, 1992, 1993). To maintain the biological in-
tegrity of the aquatic ecosystem, an ideal managed '
buffer system should have patterns of vegetation, lit-
terfall, and light penetration similar to those in a nat-
ural, undisturbed riparian forest (Golladay. and
Webster, 1988; Karr, 1993; Sweeney, 1992, 1993).
However, for many locations, representative sites of
truly natural, undisturbed riparian ecosystems do not
exist. In fact, after a long history of human distur-
bance in many areas, the concept can be difficult to
define (Bren, 1993). Karr (1993) suggests that within
a homogeneous region, relatively pristine areas may
be identified as benchmarks for the evaluation of
other sites.
Riparian forest buffer functions related to protec-
tion of the stream environment will not be reviewed
for different physiographic regions because there is
general agreement among literature sources on the
need for riparian forests in the Eastern U.S. for this
purpose. The major differences among physiographic
regions appear to be in the importance of stream temp-
erature control for cold-water vs. warm-water fisheries.
a. Temperature and Light
The diel and seasonal patterns of water tempera-
ture are critical habitat features that directly and indi-
rectly affect the ability of a given stream to maintain
viable populations of most aquatic species, both plant
and animal. Considerable indirect evidence suggests
that the absence of riparian forests along many
streams and rivers in the Chesapeake drainage, partic-
ularly in agricultural areas, may have a profound ef-
fect on the current geographic distribution of many
species of macroinvertebrates and fish. Sweeney
(1992) reviewed the effects of temperature alterations
on the growth, development, and survival of stream
macroinvertebrates found in the Pennsylvania Pied-
mont. These studies showed that temperature changes
of 2-6°C usually alter key life-history characteristics
of most of the study species
In the absence of shading by a forest canopy, direct
sunlight can warm stream temperatures significantly,
especially during summer periods of low flow. For ex-
ample, maximum summer temperatures have been re-
ported to .increase 6-15°C following deforestation
(Beschta and Taylor, 1988, Lee and Samuel, 1976,
Brown and Krygier, 1970). Streams flowing through
forests will warm very rapidly as they enter defor-
ested areas, but excess heat dissipates quickly when
streams reenter the forest. Burton and Likens (1973)
demonstrated this alternate warming (by 4 to 5°C)
and cooling as a stream passed through clear-cut and
uncut strips in the Hubbard Brook Experimental
, Forest, New Hampshire. In Pennsylvania (Valley and
Ridge Province), average daily stream temperatures
that increased 11.7°C through a clearcut area, were
substantially moderated after flow through 500 m of
forest below the clearcut. The temperature reduction
was attributed primarily to inflows of cooler ground-
water (Lynch et al.," 1980). The impact of deforesta-
" tion on stream temperature varies seasonally. In the
Pennsylvania Piedmont, Sweeney (1993) found that
from April through October average daily tempera-
tures in a second-order meadow stream reach were
higher than in a comparable wooded reach, but that
the reverse was true from November through March.
Riparian forest buffers have been shown to prevent
the disruption of natural temperature patterns as well
' as to mitigate the increases in temperature following
deforestation (Brown and Krygier, 1970; Brazier and
Brown, 1973; Lee and Samuel, 1976). Brazier (and
Brown (1973) found that buffer strips of 10'm width
were as effective as a complete forest canopy in re-
ducing solar radiation reaching small streams in the
Pacific Northwest. The exact width of Zone 1 needed
for temperature control will vary from site-to-site de-
pending on a variety of factors. Brown (1974) pointed
out that streams oriented in a north-south direction are
less easily shaded than streams flowing east or west,
and that a buffer on the north side of a stream may
have little or no effect. Also, in larger streams and.
rivers, the width of the channel prevents a complete
canopy cover, so that the effect of canopy shading
may be reduced. In eastern North America, openings
in the canopy immediately above streams occur when
the channel width exceeds about 20 m in width (i.e.,
about stream order 4 or 5).- In a study of five
Minnesota Rivers, Sinokrot and Stefan (1993) in-
ferred midsummer shading of 40-60% for rivers rang-
-------
9
ing from 15-50 m in width but effectively no shading
along extremely wide rivers (e.g., the 300 m wide
Mississippi R.). Stream orientation relative to solar
angle may also affect the extent bf shading for larger
streams. Although shading on larger rivers may have
little or no effect on water temperature, shaded stream
'banks provide habitat microsites for fish and other
aquatic organisms. , \ " .
The ability of a given width of streamside forest to
maintain or restore the natural temperature character-
istics of a stream segment depends on how it. affects
the factors that control the diel arid seasonal thermal
regime of the stream. Such factors (other than shad-
ing) include: flow, channel geometry, solar radiation,
evaporative heat loss, conductive surface heat ex-
change, and, in some ca,ses, conductive heat exchange
with the streambed. Heat budget models can integrate
local meteorological data with the above factors to
predict stream and river temperatures with relatively
high precision (e.g., Edinger et al., 1968; Brown,
1969; Beschta, 1984; Theuer et al., 1984; Sinokrot
and Stefan, 1993; Edinger and Buchak, in press).
These models indicate that solar radiation is the major
factor influencing peak summer water temperatures
and confirm that shading by the streamside forest is
critical to the overall temperature regime of a stream
or river. Stefan and Sinokrot (1993) estimated that re-
moval of the forest canopy along the Straight R.,
Minnesota, would Increase average summer water
temperatures approximately 6 C.
Hewlett and Fortson (1982) measured unexpect-
edly large stream temperature fluctuations in the
Georgia Piedmdnt'on a clearcut site with a 5 to 8 m
buffer strip left on each side of a first-order stream.
After logging and wind damage, about a 50% cover .
, canopy remained over the stream. Despite the partial
buffer, as well as rapid regrowth of low vegetation
over the stream, stream temperature fluctuations for
four years following logging were much greater than
in an uncut forest. Since the measured temperatures
could not be accounted for by a stream temperature
model, the authors suggested that in addition to the ef-
fects of direct radiation on stream temperature, efflu-
ent groundwater temperatures may also have been
modified by the removal of vegetative cover.
b. Habitat Diversity and Channel Morphology
The biological diversity of streams depends on
the diversity of habitats available. Woody debris is
one of the major factors in habitat diversity. Woody
debris can benefit a stream in several ways: (1) by sta-
bilizing the stream environment through attenuation
of the erosive influence of stream flow; (2) by in-
. creasing the diversity and amount of habitat.for
.aquatic organisms; (3) by providing a source of
slowly decomposable nutrients; and (4) by forming
debris dams, it enhances the availability of nutrients
for aquatic organisms from mdre rapidly "decaying
material. ' ' ;'-.•,..-.
The quantity of woody debris in streams under
forested canopies in the Eastern United States has
been reported to range from 0.4 to 23 kg m"2, averag-
ing about 8 kg nr2 (Webster et al., in press). These
figures are undoubtedly lower than would be encoun-
tered in streams flowing through undisturbed forest.
because most eastern streams have been subjected to
extensive removal of streamside vegetation and, in
larger streams, clearing of woody debris for naviga-
tional purposes (Webster et al., in press). Quantities of
large woody debris (LWD) recommended for healthy
streams in the George Washington National Forest in
Virginia range from 34 pieces of LWD per km for
. warm water fisheries to 136 pieces/km for cold water
fisheries. Although the quantity of woody debris in
streams without forested riparian zones would be ex-
pected to be very low, there are few quantitative stud-
ies. Sweeney (1992) found that the volume of woody
debris under forested canopies in a Mid-Atlantic
Piedmont stream was 20 times greater than the vol-
ume in a comparable .meadow reach. Following re-
moval of a riparian forest, LWD present in the stream
declines through gradual decomposition, flushing
during storms, and lack of inputs. Smaller debris from
second-growth stands promotes less stability of the
aquatic habitat and tends to have a shorter residence
time in the stream.
Loss of streamside forest can lead to loss of habitat
through stream widening where no permanent vegeta-
tion replaces forest or through stream narrowing
• where forest is replaced by permanent sod. In the ab^
sence of other perennial vegetation, bank erosion and
channel .straightening can occur as unimpeded
streamflow scours the streambed and banks (Hartman
et al.; 1987; Oliver and Hinckley, 1987). The acceler-
ated streamflow velocity allowed by straight channels
promotes channel incision as erosion from the stream
• bottom exceeds sediment entering the stream, this
process can eventually lead to the development of
wide, shallow streams that support an impoverished
diversity of species (Shields et al., 1994). Bisson et al.
(1987) point out that stability of debris accumulation
is important for aquatic habitat. Because of the greater
-------
10
resistance to displacement by hydraulic forces,' LWD
is of greater benefit to stream stability. Longer mater-
ial is .relatively more important for the stability of
wider streams.
In contrast, narrowing of stream channels has also
been reported following the replacement of stream-
side forest with permanent grassland or grass sod.
Zimmerman et al. (1967) found that the narrowing of
deforested stream channels was evident for streams
up to drainage areas of 13 km2 (5 mi2) or about a third
or fourth-order stteam. Sweeney (1992), quantified
the narrowing phenomenon more explicitly in a
Pennsylvania Piedmont basin, showing that: (1) first
and second-order wooded reaches averaged about 2
times wider than their meadow counterparts of the
same order; and (2) third and fourth-order forested
reaches were about 1.7 times wider than in deforested
areas. The channel narrows in the absence of a
streamside forest because grassy vegetation, which is
normally shaded out', develops a sod that gradually
encroaches on the channel banks. For benthic
macroinvertebrates, microbes, and algae, which live
in and on the substratum, the loss in stream width
translates into a proportionate loss of habitat. The ef-
fects of channel narrowing on fish habitat are more
complex and involve the influence of woody debris on
the pool and riffle structure (as discussed below)..
Links between LWD in streams, the abundance of
fish habitat, and the populations, growth, and diver-
sity of fishes have been documented (see reviews by
Dolloff, 1994; Harmon et al., 1986; Bisson et al.,
1987). Even when selective harvesting of trees has
been allowed along streams, the removal of old
growth has caused a decline in aquatic habitat quality
due to diminished inputs of LWD (Bisson et al.,
1987). The surfaces of submerged logs and roots pro-
vide habitat that often support macroinvertebrate den-
sities far higher than on the stream bottom itself
(Rhodes and Hubert, 1991; Sweeney, 1992; Benke et
al., 1984).
Woody debris, like boulders and bedrock protru-
sions, tends to form pools in streams either by directly
damming flow, by the scouring effects of-plunge
pools downstream of fallen logs, or by forming back-
water eddies where logs divert flow laterally (Dolloff,
1994a). In undisturbed forests, LWD can account for
the majority of pool formation (Harmon et al., 1986;
Hedman, 1992). As expected, removal of woody de-
bris by deforestation typically results in loss of pool
habitat (Bilby, 1984). Although pools are spatially
contiguous with riffles, there is little or no overlap in
the species composition of the dominant macroinver-
tebrates occurring in the two habitats. The loss of
pools, therefore, translates directly into lower popula-
tions and diversity for this group. For fish, pools im-
prove habitat by providing space, cover, and a diver-
.sity of microenvironments. Greater depth and slower
velocity in pools afford protection to fish during
storms, drought, etc. (Dolloff, 1994a). The habitat
provided by LWD may also offset the destruction of
stream habitat structures such as pools, riffles, and
cascades by catastrophic storm events (Dolloff et al.,
1994b).
Debris dams of large woody material block the
transport of both sediment and smaller litter materials.
The impoundment and delayed transport of organic
material downstream enhances its utilization by
aquatic organisms. By slowing transport rates, dams
on small order streams serve as buffers against the
sudden deposition of sediment downstream Bisson et
al. (1987). The capacity of a stream to retain debris,
therefore, is an important characteristic influencing
the aquatic habitat. (Bisson et al., 1987; Meehan et al.,
1977). ' -•"...
Although it is often thought that LWD is less im-
portant on large rivers and openwater habitats, it has
been shown that woody debris derived from riparian
forests along tidal shorelines of the Bay provides an
important refuge habitat for numerous species of fish
and crustaceans (Everett and Ruiz, 1993). Shallow
water habitats with abundant LWD support greater
abundances of many- species of fish and crustaceans
than do areas with no woody debris bordered by nar-
row strips of march (Everett and Ruiz, 1993; Ruiz et
al., 1993). They hypothesize that the importance of
LWD along Bay shorelines has been increased due to
loss of habitat in submerged aquatic vegetation and
oysterbeds.
c. Food Webs and Species Diversity
The two primary sources of food energy input to
streams are litterfall (leaves, twigs, fruit seeds, etc.)
from streamside vegetation and algal production
within the stream. Total annual food energy inputs
(litter plus algal production) are similar under shaded
and" open canopies, but the presence or absence of a
tree canopy has a major influence on the balance be-,
tween litter input and primary production of algae in
the stream.
Meehan et al. (1987) noted that "streams flowing
through older, stratified forests receive the' greatest
variation in quality of food for detritus-processing
-------
organisms." In the Piedmont, streams flowing through
forested landscapes do not subsidize downstream
channels that have been deforested (even contiguous
reaches) because the large pieces of litter do not move
very far (Sweeney, 1992). This means that a stream-
side forest is needed along the entire length of a;
stream in order to assure a proper balance of food in-
puts appropriate to the food chain of native species.
Macroinvertebrate populations are affected by
changes in litter inputs.-The activity of benthic organ-
, isms may increase following: strearhside plant re-
moval. Woody material decomposes more quickly
following riparian forest removal, thereby further re-
ducing the stream's nutrient retention (Golladay and
Webster, 1988). ,
The quantity and, quality of algal production in a
stream is greatly affected by the quantity and quality
of light striking its surface. For-example, Bilby and
Bisson (1992) showed that the algal community of a
stream heavily shaded by an old growth forest was
dominated by diatoms all year, while a nearby stream
in a deforested area contained mainly filamentous
green algae in the spring and diatoms at other times.
Other studies have also shown that deforested sites
tend to be dominated by filamentous algae while di-
atoms prevail under dense canopy cover {Lowe et al.,
1986; Feminella et al., 1989): In the eastern
Piedmont, filamentous algae such as Cladophorq can
be dominant in deforested streams due primarily to
the a combination of high nutrients, high light levels,
and warm temperature. Although some macroinverte-
brates such as crayfish (Feminella and Resh, 1989)':
and waterboatmen, insects (Sweeney and Schnack,
1976) readily consume this type of algae, most her-
•bivorous species of stream macroinvertebrates have
evolved mouthparts specialized for scraping diatoms
from the surface of benthic substrates (Merritt and
Cummins, 1984) and cannot eat filamentous algae.
The influence of-differences in the' quality of algal
production on the aquatic, ecosystem is complex.
Algal grazing species generally benefit from an in-
crease'.in algal growth (Wallace and Gurtz, 1986;
Perrin et al., 1987; Bilby and Bisson, 1992; Sweeney,
1992). Because the growth efficiency of insects is
often higher on algae than,on detritus, the opening of
the canopy may increase the production of macroin-;
vertebrates in these reaches. For example, Behmer
and Hawkins, et al. (1986) found both higher biomass
and densities, for most grazer species in deforested
sites relative to forested sites. The pattern is not clear,
however, because Hawkins (1982) found higher bio-
mass but lower densities of grazers in deforested ver-
sus forested sites. Newbold et al. (1980) observed in
California^" streams that the benthic community in
logged watersheds became dominated by a few algal
feeding species. The diversity of the macroinverte-
brate community was significantly lower than in un-
logged watersheds, except where the stream was pro-
tected by a riparian buffer of 30 m or more. For buffer
strips less than 30 m in width, the Shannon diversity
was significantly correlated with buffer width.
2. Zone 2—Removal of
Nonpoint Source Pollutants
The primary function of Zone 2 is to remove, se-
quester, or transform nutrients, sediments, and other
pollutants. Because of its proximity to Zone 1, Zone 2
might also have direct impacts on the stream channel
system and contribute to Zone 1 functions. The pollu-
tant removal function of a Riparian Forest Buffer
System depends on two key factors; 1) the capability
of a particular area to intercept surface and/or ground-
'water-borne pollutants and 2) the activity of specific
pollutant removal processes. Focusing on these two
factors as regulators of buffer zone effectiveness is
useful for evaluating the importance of a particular
site as a buffer and for evaluating the three zone
RFBS specification. In the sections below we review
the major pollutant removal processes that operate in
Zone 2 and discuss how these processes interact with
pollutants 'in either surface runoff or groundwater
flow in the context of the three zone specification.,
,'" ' •, ^
a. Nitrate Removal
Nitrate removal from shallow groundwater has
been the focus of many completed and ongoing stud-
ies. At least four,separate studies at different sites in
the Gulf-Atlantic Coastal Plain Physiographic
Province have, shown that concentrations of nitrate in
shallow subsurface flow are markedly reduced after
passage through, portions of .natural riparian forest
analogous,to Zone 2 (Jacobs and Gilliam, I985a,b;
Jordan et al., 1993; Lowrance et al., 1983, 1984a;
Peterjohn and Correll, 1984). Studies in other physio- .
graphic settings have also shown nitrate removal from
shallow groundwater in areas analogous to Zone 2
(Groffman et al., 1992; Simmons et al., 1992). Most
studies with high levels of nitrate removal were in
areas with high water tables that caused shallow
groundwater to flow through or near the root zone.
The mechanisms for removal of nitrate in these
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12
study areas are thought to be a combination of deni-
trification and plant uptake. Linkages between plant
uptake and denitrification in surface soils have been
postulated as a means for maintaining high denitrifi-
cation rates in riparian ecosystems (Groffman et al.,
1992; Lowrance, 1992). In contrast, riparian systems
without substantial contact between the biologically
active soil layers and groundwater or with very rapid
groundvvater movement appear to allow passage of ni-
trate with only minor reductions in concentration and
load. Correll et al. (1994) reported both high nitrate
concentrations and high nitrate removal rates beneath
a riparian forest where very high nitrate flux and rapid .
groundwater movement through sandy aquifer mater-
ial limited nitrate removal efficiency. Staver and
Brinsfield (1990) showed that groundwater flow be-
neath the biologically active zone of a narrow riparian
buffer along a tidal embayment in Maryland resulted
in little removal of nitrate. It is also known that
groundwater discharging through sediments of tidal
creeks may have up to 20 times the nitrate concentra-
tions found in the main stem of the creeks (Reay et al.,
1992).
Phillips et al. (1993) indicated that groundwater ni-
trate might bypass narrow areas of riparian forest wet-
land and discharge into stream channels relatively un-
altered when the forest is underlain by an oxygenated
aquifer. This pattern of groundwater flow was sup-
ported by modelling of a small Coastal Plain water-
shed in Maryland (Reilly et al., 1994). Isotopic analy-
sis of groundwater and surface water in this watershed
suggested that denitrification was not affecting the ni-
trate concentrations of discharging groundwater. In
these cases where nitrate enriched water surfaces in
the stream channel, a wide RFBS would have little ef-
fect on nitrate. Deeply rooted vegetation near the
stream might have some effect.
Studies in New Zealand have shown that the ma-
jority of nitrate removal in a pasture watershed took-
place in organic riparian soils which received large
amounts of nitrate laden groundwater (Cooper, 1990).
The location of the high organic soils at the base of
hollows caused a high proportion of groundwater (37-
81%) to flow through the organic soils although they
occupied only 12% of the riparian zone. A related
study in New Zealand (Schipper et al., 1993) found
very high nitrate removal in the organic riparian soils
but streamflow was still enriched with nitrate. The au-
thors speculated that water movement through min-
eral soils was responsible for most of the nitrate trans-
port to streams. Puparian systems with intermingling
of organic and mineral soils point out the need to un-
derstand where groundwater is moving and what
types of soils it will contact, especially in seepage
areas.
b. Plant Uptake
Maintenance of active nutrient uptake by vegeta-
tion in Zone 2 should increase the potential for short-
term (non-woody biomass) or long-term (woody bio-
mass) sequestering of nutrients. Although plant water
uptake is chiefly a passive transpiration process, plant
nutrient uptake is mostly an active process, dependent
upon plant metabolic activity (Hoagland and Broyer,
1936). Most nutrients are transported into plants
against an electrochemical potential gradient (Bowl-
ing et al., 1966; Higinbotham et al., 1967). Obser-
vations of ion concentrations in plant xylem exceed-
ing external soil water concentrations by over 100
times indicate significant active uptake of P. (Russell
and Shorrucks, 1959). Transpiration tends to influ-
ence the uptake of a nutrient when the external con-
centration of that nutrient is high. Transpiration in ri-
parian forests is very high and can control, water
movement to streams (Correll and Weller, 1989;
Bosch et al.,' 1993). Kramer and Kozlowski (1979)
pointed out that transpiration increases the mass flow'
of solutes toward root surfaces.
Nutrient uptake by. flood-intolerant plants is
strongly influenced by the aeration status of the soil
(Hoagland and Broyer, 1936; Hopkins, 1956;
Hopkins et al. 1950). As low oxygen supply .decreases
root metabolism, the uptake of most nutrients de-
creases. Flood-tolerant species, such as those found in
many riparian forests, may tolerate low-oxygen con-
ditions by means of adaptive metabolic responses
(Crawford, 1982). They may also avoid root anoxia
by morphological adaptations that facilitate the avail-
ability of oxygen. Under flooded conditions,' roots
may become thicker and increase in porosity, allow-
ing an internal downward diffusion of oxygen
(Armstrong, 1968; Courts and Philipson, 1978). The
growth of adventitious roots may also allow water and
nutrient uptake from near-surface areas that are more
aerated (Kozlowski, 1984; Sena Gomes and Koz-
lowski, 1980).
Vegetation selection for restored or managed PJFBS
must consider the ability of different species to take
up and store nutrients under specific.conditions of the
site. Kozlowski and Pallardy (1984), point out that
flooding can.enhance the nutrient uptake and growth
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13
of some species. Bottomland hardwood seedlings
grow faster under saturated conditions than under
drained but Veil-watered conditions! Mote rapid in-
creases in total.dry weight and N and P uptake were
found in water tupelo (Nyssa aquatica L.) as well as
.several other species under saturated conditions
(Hosner and Leaf, 1962;,Hosner et al., 1965). Shoot
' weights of a majority of wetland and intermediate
plant species were either unaffected or increased
under flooded conditions (Justin and Armstrong,
1987).
Nutrient uptake and accretion in riparian forests
will be affected by vegetation management. Nutrient
demand by vegetation corresponds with growth rate
(Cole, 1981; McDonald et al., 1991). Loblolly pine
dominated forests in the. Gulf Coastal Plain attain
maximum rates of growth of about 8t. ha"1 yr.1 during
the first twenty years of age, for which 101 kg of N
and 9 kg of P are required each year (Nelson et al.,
1970; Switzer et al., 1979). Cole ,and Rapp (1980)
suggested a worldwide average annual N uptake rate
of 70.5 kg ha'1 for deciduous tree species and 39 kg
ha"1 for coniferous species. Temperate deciduous .
species produce 179 kg biom'ass kg-1 N uptake, and
temperate coniferous species produce 103 kg biomass
kg'1 N uptake (Cole and Rapp, -1980). However,
Miller (1984) disputes the notion that coniferous.
forests require less nutrients than broad-leaved
forests. His review of nutrient uptake studies indicates
that the ranges of measured uptake for coniferous and
broad-leaved forests overlap. •
Compared to the "natural" riparian forests studied
iii mosj existing research, managed riparian forests .
have the potential for increased accumulation of N
and P in biomass through both increased biomass pro-
duction and increased foliar nutrient contents. Trees
can respond to N subsidy by both increased growth
rates and luxury N uptake. The growth rate of forests
is commonly N limited. Cole (1986) suggested-that
high efficiency of N use by forests is an adaptation to
the N-deficient environments that they frequently in-
habit, '".''•
Often the potential N uptake rate is much higher
than observed rates. Forest growth has been found to
respond readily to N applications (Miller and Tarrant,
1983; Schmidtling, 1973). Mitchell and Chandler
(1939) found large tree-growth responses to N.fertil-
izer applications up to 400 to 600 kg ha'1. Cole (1981)
found that after fertilizing with 400 kg N ha'1 .yr? in
effluent'from a municipal sewage treatment plant for
three years, poplar (Populus riigra var. italica
Muench.) and Douglas fir (Pseudotsuga menziesii
(Mirb.) Franco) took up 213 and 78 kg N ha'1 yr1, re-
spectively. This contrasted with an uptake of 16 kg N
ha'1 yr1 by poplar and 23 kg N ha'1 yr! by Douglas fir
in unfertilized sites. Miller and Cooper (1973)
showed that trees can take up. "luxury" levels of N.
Growth responses by 36-year-old Corsican pine
(Pinus nigra var. maritima.(Ait.) Melv.) to different
levels of N fertilization showed that foliar N content
reached a maximum of 26,400 mg kg-1 after applying
the highest rate of 504 kg N ha'1 yr1 for three years.
Maximum volume growth corresponded to a foliar
content of about 20,000 mg kg'1, attained by applica-
tions of 336 kg N ha"1 yr1 for three years.
Conditions do exist where N is no longer the lim-
iting nutrient for forest growth. Long-term inputs of
nitrogen, such as may occur from atmospheric depo-
sition in the northeastern U.S., could result in N lev-
els exceeding the total combined plant and microbial
nutritional demands (Aber et al., 1989). Under these
conditions, P might become the limiting factor for
tree.growth. Unlike upland forests, P may often be the
most limiting nutrient in wetland ecosystems (Taylor
et al., 1990). Mitseh et al. (1979) found the growth of
bald cypress (Taxoditim distichum^ (L). Rich.) in a
-southern Illinois swamp to correspond well with P in-
puts from flooding. Foliar P content of loblolly pine
on wet Coastal Plain sites in South Carolina has been
observed to correlate well with growth (Wells and
Crutchfield, 1969). Analysis by Brinson (1977) of nu-
trient ratios in decaying litter from tupelo gum trees in
a North Carolina swamp forest suggested that P levels
may limit decomposition rates. If P is the limiting nu-
trient for tree growth, it should make vegetation an ef-
fective P sink.
. While several studies have found plant uptake to
be an important nutrient removal mechanism in areas
analogous to Zone '2 of riparian forest buffers (Correll
and Weller, 1989; Fail et al:, 1986; .Peterjohn and
Correll 1984; Groffman et al. 1992), several factors
may reduce the importance of plants as nutrient sinks.
Pollutants in groundwater flowing into the riparian
buffer will only be accessible to plants if the water
table is. high in the soil profile {Ehrenfeld 1987) or if
mass movement of water due to transpiration de-
mands moves water and solutes into the root zone.
Coastal Plain riparian forests have been shown to con-
trol localized downslope water transport by creating
moisture gradients which move water in unsaturated
flow from both the adjacent stream and the upland
field (Bosch et al.., 1993). Nutrients in surface runoff
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14
and in water percolating rapidly through soil macrop-
ores as "gravitational water" may not be available to
plants. Large rainfall events, that often transport a
high percentage of pollutants in the CBW (Jaworsjd et
al., 1992) often produce concentrated surface flow
and macropore-dominated percolation.
Plant sequestering of nutrients is also limited by
seasonal factors. In the temperate deciduous ecosys-
tems that dominate riparian forest buffers in the CBW,
plant uptake will decline or stop during the winter
season. A high percentage of surface and groundwater
flow occurs in the CBW during winter. There is also
concern that nutrients trapped in plant tissues can be
released back into the soil solution following litterfall
and decomposition. However, nutrients released from
decomposing plant litter may be subject to microbial,
physical or chemical,attenuation mechanisms in the
root zone of forest soils. Storage of nutrients in
woody tissue is a relatively long-term attenuation, but
still does not result in removal of pollutants from the
ecosystem unless biomass is removed. A final con-
cern about plant uptake as a nutrient removal mecha-
nism arises from the possibility that the ability of
trees in a buffer zone to sequester nutrients in woody
biomass becomes less as trees mature. The average
tree age in most riparian forest buffers in the CBW is
less than 100 years and should thus be accumulating
nutrients in woody biomass. Although net vegetation
accumulation of nutrients may reach zero, net ecosys-
tem accumulation may continue as nutrients are
stored in soil organic matter. Groffman et al (1992)
describes a nitrate-enriched riparian system with
symptoms of N saturation (Aber et al., 1989).
Nitrogen saturation is not likely to occur in RFBS be-
cause of high denitrification rates removing N from
the system.
Little is known about the types of vegetation
needed in new or reestablished RFBS. Crop tree man-
agement (the selection and release of desired trees by
removal of competing trees) will be possible in many
natural successional riparian forests. Numerous native
tree species are recommended for water quality im-
provement in crop tree management (Sykes et .al.,
1994). The trees were selected based ori their ability
as nutrient filters although little data exist on individ-
ual riparian species.
c. Microbial Processes
In addition to plant uptake, there are microbial
processes that attenuate pollutants in RFBS. These
processes include immobilization of nutrients, deni-
trification of nitrate and degradation of organic pollu-
tants. Microbes take.up or "immobilize" dissolved nu-
trients just as plants do. These immobilized nutrients
can be re-released or "mineralized" following death
and decomposition of microbial cells, just as nutrients
sequestered by plants can be released following litter-
fall. In ecosystems that are accumulating soil organic
matter, there will be a net storage of immobilized nu-
trients. Zone 2, if managed to foster soil organic mat-
ter accumulation, may thus support significant long-
term rates of nutrient storage by immobilization.
Denitrification refers to the anaerobic microbial
conversion of nitrate to N gases. •Denitrification is
- controlled by the availability of oxygen (Oa), nitrate,
and carbon (C). Although essentially an anaerobic
, process, denitrification can occur in well drained soils
because of the presence of anaerobic microsites, often
associated with decomposing organic matter frag-
ments which deplete available oxygen (Parkin, 1987).
It is likely that soil moisture gradients,in riparian
ecosystems cause a change in controlling factors
within most three zone RFBS. In parts of the RFBS
with better internal drainage and generally lower soil.
moisture conditions, denitrification may be generally
limited by their interacting factors of carbon avail-
ability and aeration status. While many wetlands are
often assumed to have high levels of denitrification
. because of high carbon soils and anaerobic condi-
tions, denitrification in many wetlands will be N lim-
ited (Groffman, 1994). In the more poorly drained or
wetland portions of a RFBS, denitrification is more
likely to be limited by nitrate availability.
Wetland soils develop high levels of organic matter
because of their slope position and hydrologic condi-
tion. Frequently inundated soils will have lower rates
of litter decomposition because the flow of carbon
from litter to microbial populations is reduced tinder
anaerobic conditions (Groffman, 1994). The interac-
tive nature of oxygen, nitrate, and carbon control of
denitrification means that more denitrification gener-
ally occurs in intermittently flooded sites than in per-
manently flooded conditions (Reddy and Patrick,
1984).
Denitrification, measured directly using the acety-
lene inhibition technique (Tiedje et al., 1989), ac-
counts for substantial nitrate loss from some riparian
ecosystems. Denitrification has been identified as the
key nitrate removal mechanism in several riparian for-
est buffer studies (Jacobs and Gilliam,' 1985b; Pinay
and Decamps, 1988; Correll and Weller, 1989;
Groffman et al., 1992; Haycock and Pinay, 1993;
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15
Jordan et al., 1993). Estimates in the range of 30 to 40
kg N ha'1 yr1 have been reported for natural riparian
forests in the U.S. (Hendrickson, 1981; Hanson et al.,
1994a, Lowrance et al., 1984b). In several studies of
denitrification in riparian ecosystems, denitrification
has been concentrated in surface soil and rates are
x generally much lower below the top 12 to 15 cm of
soil (Hendrickson, 1981; Grpffman et al., 1992;
Ambus and Lowrance, 1991; Lowrance, 1992).
Schipper et al. (1993) reported very high denitrifica-
tion in the top 30 cm of an organic riparian zone soil
in New Zealand. Denitrification rates (measured on
soil slurries made anaerobic with Argon gas) were
over 11 kg N ha'1 d'1 in this site. This is likely an over-
estimate of actual denitrification because the slurries
were made anaerobic. The denitrification rates mea-1
,sured were 1-3 orders of magnitude greater than most
estimates in the literature. Measurements of denitrifi-
cation in these organic soil zones showed that the den-
itrification was greatest at the upslope edge of the ri-
parian zone where nitrate-enriched water entered the
organic riparian soil (Cooper, 1990). These studies in-
dicated that most of the organic riparian soils in the
watershed were denitrifying at rates below their max-
imum capacity and could denitrify more if nitrate
loadings increased (Cooper, 1990; Schipper et al.,
. 1993). Denitrification is likely to be most important in
wetland soils such as would be found in Zone 1 and
some Zone 2 areas in the Chesapeake Bay watersheds
(Lowrance et al. 1984b, Peterjohn and Correll 1984,
Jacobs and Gilliam 1985b, Ambus and Lowrance
1991) but can also be significant in drier forest soils
subject to high nitrate loadings and in grass vegetated
filter, strips (GVFS) (Ambus and Lowrance, 1991;
Groffman et al., 1991). ,
While the factors regulating denitrification in sur-
face soils and aquifers are relatively well understood,
the amounts of direct denitrification of groundwa-
ter-rbprne nitrate are much less well established.
Subsurface denitrification has been observed in sev-
eral .studies (Truedell et al., 1986; Slater and Capone,
1987; Smith and Duff, 1988;,Francis et al.,:1989;
Obenhuber and Lowrance, 1991), yet other studies
have found the potential for denitrification in the sub-
surface to be low or non-existent (Parkin and
Meisinger, 1989; Ambus and Lowrance, 1991;
Groffman et al., 1992; Bradley et al., 1992; Lowrance,
1992; Yepmans et al., 1992; Starr and Gilham, 1993).
Subsurface microbial activity is usually limited by
carbon availability. In settings where the total and dis-
solved carbon contents of aquifers are low, they are
poor quality substrates for microbial growth (Lind
and Eiland, 1989; Hiscock~et al., 1991; Johnson and
Wood, 1992; McCarty and Bremner, 1992) and anaer-
obic conditions necessary for denitrification to pro-
ceed are not generated. -
Microbial attenuation of organic compounds arises
from their ability to degrade these compounds as food
sources or through non-energy yielding "cometabo-
lism" reactions. There are many different microbial
degradation mechanisms including aerobic, anaero-
bic, chemoautotrophic and heterotrophic pathways.
The wide range of environments and high diversity of
microbial metabolism in RFBS, should support many
of these mechanisms. Further research into specific
management strategies to foster a wide range of
degradation strategies is needed (Paterson and
Schnopr, 1992); t
In many cases, riparian zone'retention of ground-
water-borne pollutants may depend on a complex in-
teraction of hydrology, plant, soil and microbial fac-
tors. The potential importance of these interactions is
hypothesized based on studies where significant rates
of nitrate removal from groundwater were measured,
but the potential for denitrification in the subsurface
was low. Groffman et al. (1992) and Hanson et al.
(1994a,b) suggested that surface soil denitrification of
groundwater derived nitrate is an important route of N
removal in riparian forests. This route- depends on
plant uptake of nitrate from groundwater, decomposi-
tion and N release from plant litter, and nitrification
and denitrification of this N in surface soil. In riparian
forests where this route of N removal is important, the
nitrate removal function may depend on complex in-
teractions between hydrology, plant dynamics, and'
soil microbial processes. These interactions vary
within and between riparian forests and should be
strongly influenced by soil drainage class, vegetation.
and soil type, climate, and groundwater, quality.
Although soil denitrification should be sustainable in-
definitely under proper conditions with a supply of ni-
trate and available C, Hanson et al. (1994b) found that
long term groundwater nitrate loading led to symp-
toms of "N saturation in the surface soils of a riparian
forest buffer.,
d. Removal of Surface-borne Pollutants
Fewer studies have been published on NPS pollu-
tant removal from surface runoff in Zone 2 type
forests. The primary function of Zone 2 relative to
surface runoff is to remove sediment and sediment-
borne pollutants and tor infiltrate runoff. Daniels and
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16
Gilliam (in press) found that mature riparian forests,
analogous to Zone 2 vegetation, were effective for
sediment load reduction with removal of 50 to 80% of
inputs from upland fields. Sediment trapping in ripar-
ian forest buffer zones is facilitated by physical inter-
ception of surface runoff that causes flow to slow and
sediment particles to be deposited. Effective sediment
trapping requires that runoff be primarily sheet flow.
Channelized flow is not conducive to sediment depo-
sition and can actually cause erosion of the RFBS.
Tvvo studies on long-term sediment deposition in ri-
parian forests (Cooper et al., 1987, Lowrance et aL,
1986, Lowrance et al., 1988) indicated that long-term
deposition is substantial. In both these studies, two
main actions occur: 1) the forest edge fostered large
amounts of coarse sediment deposition within a few
meters of the field/forest boundary; 2) finer sediments
are deposited further into the forest and near the
stream. Both Cooper et al. (1987) and Lowrance et al.
(1986) found much higher depths of sediment deposi-
tion at the forest edge than near the stream. A second
peak of sediment depth was often found in Zone 1,
possibly from upstream sediment sources deposited in
overbank flows (Lowrance et al. 1986). The surface
runoff which passes through the forest edge environ-
ment is much reduced in sediment load because of
coarse sediment deposition but the fine sediment frac-
tion is enriched relative to total sediment load. These
fine sediments carry higher concentrations of labile
nutrients and adsorbed pollutants (Peterjohn and
Correll, 1984; Magette et al., 1989) which are carried
further into the riparian forest and are deposited
broadly across Zone 2.
Movement of nutrients through Zone 2 in surface
runoff will be controlled by a combination of: 1) sed-
iment deposition and erosion processes; 2) infiltration
of runoff; 3) dilution by incoming rainfaiythroughfall;
and 4) adsorption/desorption reactions with forest
floor soil and litter. Studies that separate -the effects
of these various processes are not available. Peterjohn
and Correll (1984) found large reductions in concen-
trations of sediment, ammonium-N, and ortho-P in
surface runoff which passed through about 50 m of a
mature riparian forest in the Maryland Coastal Plain,
analogous to Zone 2. Although the concentrations of
these pollutants were reduced by a factor of 3 or 4 in
most cases, the flow-length was about twice that rec-
ommended in the RFBS specification. Daniels and
Gilliam (in press) found that dissolved ortho-P loads
in surface runoff were not reduced markedly in a
Zone 2-like area of riparian forest. The studies of sur-
face runoff through riparian forests agreed on the im-
portance of eliminating channelized flow through the
riparian forest and recommended spreading flow be-
fore it reached the forest buffer. Flow spreading is rec-
ognized as primarily a Zone 3 function in the RFBS
specification. In-field practices are also critical in pre-
venting channelized flow from reaching the field
edge.
3. Zone 3—Sediment Removal and
Spreading of Surface Runoff
The primary functions of Zone. 3 are to remove
sediment and sediment associated chemicals and to
spread surface runoff entering as concentrated flow.
Functions of grass vegetated filter strips (GVFS),.
analogous to Zone 3 of the RFBS, have been evalu-
ated in a number of replicated experiments. Most of •
the available research on GVFS is applicable to eval-
uating the potential for sediment deposition in Zone 3
of the RFBS.
Several short-term experimental studies have found
that GVFS were effective for removal of sediment and
sediment-bound pollutants with trapping efficiencies
exceeding 50% if flow was shallow (< 5 cm depth)
(Young et al. 1980, Magette et al. 1987, Dillaha et al.
1989a).' Magette et al. (1989) and Dillaha et al.
(1989a.) evaluated relatively narrow filter strips (4.6 m
and 9.2 m) for control of nutrients and sediment mov-
ing from row-crop plots. Magette et al. concluded
that: 1) the performances of GVFS were highly vari-
able; 2) GVFS were more effective in removing sus-
pended solids than in removing nutrients; 3) GVFS
become less effective as more runoff events occur;
and 4) the effectiveness of GVFS decreased as the
ratio of GVFS length to source area decreased.
Dillaha et al. (1989a) reached similar conclusions.
They found that GVFS were effective immediately
after establishment, removing up to 98% of the in-
coming sediment and that removal of incoming sedi-
ment bound total N and total P was nearly as effec-
tive. Soluble N (both NO3 and NH4) and soluble P
were not removed effectively. Both Magette et al.
(1989) and Dillaha et al. (1989a) conclude that nar-
row GVFS would probably have relatively short use-
ful life spans. Dillaha et al. (1989a) reported that one
GVFS was nearly inundated with sediment during the
span of 6 rainfall simulator events. The sediment trap-
ping efficiency fell from 90% in run 1 to 5% in run 6.
GVFS were much less effective when flow was con-
centrated than when surface runoff was in shallow
sheet flow (Dillaha et al.,1989a). Properly managed:
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17
Zone 3 areas are likely to perform similarly to GVFS
in these experimental studies. Management of these
areas will likely include periodic removal of sedi-
ment, reestablishment of vegetation, and removal of
ephemeral channels: ,
Trapping efficiencies for sediment decrease at high
runoff rates because of increased depth of flow
(Barfield et al., 1979; Schwer and Clausen, 1989).
Concentrations, of total N, total P, suspended solids,"
and BOD were reduced up to 80% in feedlot runoff
passed through GVFS ranging from 92 to 262 m
(Dickey and Vanderholm, 1981). The need for rela-
tively long filter strips was confirmed in other studies
looking at runoff from chicken manure application "
areas (Bingham et al., 1980; Overcash et al, 1981).
They found that the ratio of buffer area to land appli-
cation area in order to achieve complete removal of
contaminants in water leaving the GVFS was about
1:1. Therefore a filter area would need to be as large
as the source area. This situation is often not possible
due to inadequate land for filter areas or competition
between land for GVFS and land for crop production. '
A buffer source area/length ratio of less than 1:1
would be adequate for less than complete removal.
Trapping efficiencies for sediment and nutrients
also decrease when runoff enters the GVFS in, con-
centrated flow (Dillaha et al., 1986). When this is the
case, very little of the filtration capacity of either the
GVFS or the riparian forest is used. If field practices
do not eliminate channelized flow,' it should be elimi-
'nated as near the upslope border of the RFBS as pos-
sible. The RFBS specification suggests using level-
lipped spreaders to convert concentrated flow to sheet
flow before it reaches Zone 2 (Welsh, .1991). These
spreaders, when needed, would be part of Zone 3 so
they could be managed (cleaned out and, periodically
restored) using farm equipment. Franklin et al. (1992)
reported on the use of a level spreader to spread flow
from agricultural fields before it entered a downslope
forest filter zone (FFZ). Although they did not com-
pare the FFZ with'and without spreaders for natural
.rainfall events, they did compare hydrologic response
with and without spreaders for simulated runoff
events. Without a spreader, the time to reach peak
flow at a flume below the FFZ was about 10 minutes •
and the time to reach zero flow below the FFZ after
the water supply was cutoff was only 20 min. In con-
trast, with the level spreader in place, this artificial
runoff took 45 minutes to reach a peak flow and 135
minutes to stop flowing after water was cut off
(Franklin et al., 1992). Although specific water qual-
ity, data are not available from this study with and
without'Spreaders, spreading the flow affected the
timing of the event with a smaller effect on peak,and
total flows. •
Used as part of the RFBS, GVFS should substan-
tially reduce sediment and sediment-borne pollutant
loads reaching the stream. Improperly installed GVFS
may serve to accentuate channelization problems in
the landscape, leading to erosiofi of the forested zones
of the buffer. For;example,, in an analysis bf existing
grass GVpS on 33 farms in Virginia (Dillaha et al.
1986, 1989b) found that sediment trapping was quite
poor in many cases. In hilly areas, sediment trapping
was generally low because runoff usually crossed the
GVFS as concentrated flow. Rapid (1-3 years) accu-
mulation of sediment caused several GVFS to become
vigorous sediment producers. In cases where sedi-
ment accumulation was significant, runoff flowed
parallel to the GVFS until a low point was reached
where it crossed the GVFS as concentrated flow. Due
to the uncertainties in long-term performance of
GVFS, overall buffer efficiency and sustainability
should be significantly increased by using a.combina-
tion of grass strip and forest buffer as described here.
4. Integrated Water Quality Functions of
the Three Zone Buffer System
Although no published studies of an integrated.
three zone buffer system are available, the" studies
cited above provide useful irisights into the potential
functions of each zone. Even with an integrated three
zone system, it is possible that there will be conflict-
ing objectives relative to the types of water quality
functions desired from the RFBS.
Perhaps the .most basic potential conflict relative
to NPS pollution control with RFBS is the need to si-
multaneously control at least three major transport
mechanisms of waterborne pollutants. It is likely that
control of pollutants transported in the sediment ad-
sorbedphase of surface runoff, the dissolved phase of
surface .runoff, and groundwater (dissolved phase
only) may be optimal on different sorts of RFBS with
differing soils, vegetation, and management. Riparian
forest buffers must be effective in controlling multiple
noripoint-sources of pollution.
Phillips (1989) proposed a general model of ri-
parian buffer effectiveness based on detention time of
surface and subsurface runoff and comparison to a
reference buffer of known'detention time and known
. effectiveness. Comparisons to a reference buffer with
an.assumed first order rate constant for nitrate ire-
-------
18
moval were done. The detention-based model indi-
cated that relatively flat, sandy, well-drained soils
with high infiltration capacities would be the most ef-
fective buffers for nitrate removal. This approach is
lacking relative to nitrate retention because it disre-
gards the effects of different soils on denitrification
and the unequal partitioning of nitrate between sur-
face runoff and subsurface transport paths. The deten-
tion model (Phillips, 1989) correctly concludes that
for surface-borne pollutants, increasing infiltration in
the RFBS will be an effective measure for both dis-
solved and adsorbed pollutant control. Conversely,
the sandy well-drained soils which have highest infil-
tration will likely have lowest denitrification rates and
may have rapid groundwater movement rates leading
to high rates of nitrate transport through the riparian
forest buffer. This type of situation is described by
Correll et al. (1994) for the entire riparian buffer and
by Cooper (1990) for the mineral soils in the riparian
zone.
For nitrate removal via denitrification, a riparian
ecosystem where high nitrate water moves into high
organic matter soils or subsoils is the best way to pro-
mote denitrification and the best way to permanently
remove nitrate from the soil-water-plant system. This
is illustrated both by the New Zealand riparian studies
of organic riparian soils (Cooper, 1990; Schipper et
al., 1993) and by the findings that denitrification is
highly stratified in mineral soils with most denitrifi-
cation occurring in the high organic carbon surface
soils (Ambus and Lowrance, 1992; Hanson et al,
1994a). Organic rich wetland soils can often respond
to increased nitrate loads with increased denitrifica-
tion (Groffmah, 1994). The same conditions which
are likely to promote denitrification are likely to de-
crease the amount of retention of surface-borne pollu-
tants. Wetland soils which are frequently inundated
will have little or no infiltration capacity or available
water storage capacity.
-------
Riparian Forest Buffer Systems in
Physiographic Provinces of the
Chesapeake Bay Watershed
A. COASTAL PLAIN
1. General Land Use and Hydrology
The Coastal Plain has higher proportions of both
cropland (32%) and wetland (21%) than any other
physiographic province of the Bay Watershed (Table
1). The Coastal Plain portions of the CBW are com-
prised of watersheds with low toppgraphic relief, rel-
atively high moisture infiltration capacities, well-dis-
tributed rainfall throughout the year, and unconfined
' surficial aquifers. Stre'amflow is mainly derived from
groundwater discharge from the surficial aquifer.
Direct .surface runoff in agricultural watersheds gen-
erally accounts for about 5 to 15% of streamflow
(Peterjohn and Correll, 1"984; Staver ef al., 1988). The
remainder of the precipitation either infiltrates and is
available for either groundwater recharge or evapo-
transpifation, or goes directly into surface water as
stream or detention storage.' Although this general
view of the Coastal Plain is useful, variations in soils,
topography, subsurface stratigraphy, and land use
within the Coastal Plain control the fate of NFS pol-
lutants relative to RFBS.
. The CBW Coastal Plain is often divided into Inner
and Outer Coastal Plains. The jhiner Coastal Plain is
mostly the western shore of Chesapeake Bay and the'
upper Eastern Shore. The Outer Coastal Plain is pri-
marily the lower Eastern Shore/Delmarva Peninsula. ,
Inner'Coastal Plain areas have relatively high topo-
graphic relief compared to Outer Coastal Plain sys-
tems and generally have finer textured, nutrient-rich
soils compared to the nutrient-deficient, sandy soils of
the Outer Coastal Plain (Correll et al., 1992). A more
detailed classification of the Coastal Plain was devel-
oped by the U.S. Geological Survey for the Delmarva
Peninsula (Phillips et al, 1993). This classification of
hydrogeomorphic regions was based.on qualitative
analysis of geologic and geomorphic features, soils,'
drainage patterns, and land cover (Figure 3). The up-
land, non-tidal area of the Delmarva was divided into .
Inner Coastal Plain which closely correlates with the
Inner Coastal Plain of Correll et al. (1992) and three
Outer Coastal Plain hydrogeomorphic regions: well-
drained upland, poorly drained upland, and surficial
confined region. Differences in the physical charac-
teristics of these regions result in variations in the
functions of RFBS within them. The following dis-
cussion presents the general hydrogeomorphic char-
acteristics associated with each .region.
a. Inner Coastal Plain
The Inner Coastal Plain (1C?) includes the portion
of the Coastal Plain located on the western shore of
Chesapeake Bay and the area immediately south of
the Fall Line on the Delmarva Peninsula. Tidal sec-
tions of rivers extend far into the ICP, near the Fall
Line in some cases. Watersheds in .the ICP are char-,
aeterized by well-drained soils on uplands with
poorly drained soils limited to riparian zones. Land
use is primarily agricultural on uplands and forested
in riparian zones. Topography of this region is gently
rolling with a high degree of stream incision.
The, ICP is a hydrologically complex region be-
cause sands' and gravels that comprise the surficial
aquifer are thin and overlie subcropping sands or
finer-textured confining beds of older Coastal Plain
aquifers. Stream valleys are commonly incised into
the older units. As a result of this configuration, the
surficial deposits-do not form an extensive aquifer as
they do. in other parts of the.Coastal Plain. Shallow
groundwater flow systems in the surficial sediments
commonly extend from topographic highs into the
deeper aquifer where they are close to the surface. In
addition, older-water from deeper aquifers often dis-
charges upward to streams. If the surficial aquifer
overlies a shallow confining bed, groundwater flow is
restricted to shallow depths where it comes into con-
tact with riparian zone, sediments and soils near
aquifer discharge areas.
The Rhode R. Watershed along the western shore
19
-------
20
of Maryland is representative of the hydrologic, con-
ditions common to much of the ICP. This 2286 ha wa-
tershed is 62% forest, 23% croplands, 12% pasture,
and 3% freshwater swamp (Jordan et al., 1986). The
watershed is underlain by a relatively impermeable
clay layer which forms an effective aquiclude. Most
groundwater flow to streams is in a shallow surficial
aquifer (Correll 1983). The 160 yr average rainfall is
108 cm. The long-term average precipitation by sea-
son is 28 cm, 31.4 cm, 24.5 cm, and 24.6 cm for
December to February, March to May, June ,to
August, and September to November, respectively
(Higman and Correll, 1982 cited in Peterjohn and
Correll, 1984). For the Rhode R. Watershed, slow
streamflow (baseflow or groundwater discharge) av-
eraged 29.6 cm of flow while quickflow (mostly
stormflow or surface runoff from all contributing
areas) accounted for 4.97 cm (Correll, unpublished in
Peterjohn and Correll, 1984). Studies on Rhode R. in-
dicated that 86% of all watershed discharge comes
from slow flow or groundwater discharge and 14%
from direct surface runoff. For one year of study
March, 1981 to March, 1982, Peterjohn and Correll
(1984) estimated that about half of all quickflow took
place in the Summer (June to August) and that over
half of slow flow (groundwater discharge) took place
in winter.
b. Well-Drained Upland t
Watersheds in the well drained upland (WDU,
Figure 3) are characterized by predominantly well-
drained soils on uplands and poorly drained soils on
fioodplains in stream valleys. The topography is rela-
tively flat to gently rolling and there is a high degree
of stream incision (Phillips et al., 1993). Most of the
upland area is used for agricultural crop production
with wooded areas generally confined to narrow ri-
parian zones. Sediments of the surficial aquifers are
primarily sand and gravel and range from about 6 to
12 m in the north to 24 to over 30m thick in the south
(Owens and Denny, 1979). The aquifer is unconfmed
and the depth to water ranges from 3 to 10 m beneath
topographic highs, to land surface in surface-water
discharge areas.
Groundwater flow paths range from about 1 km to
several km in length in the well-drained upland
(Shedlock et al., 1993). The longest; oldest flow paths
originate at topographic highs, extend to the base of
the aquifer, and discharge to 2nd and 3rd order
streams through the hyporheic zone (beneath the
stream channel). The water contained in them is gen-
erally less than 50 years old near aquifer discharge
areas (Dunkle et al., 1993). Shorter, younger flow
paths originate in near-stream recharge areas and are
the main source of baseflow to first-order streams.
c. Poorly-Drained Upland
Watersheds in the poorly drained uplands (PDU,
Figure 3) are characterized by interspersion of poorly
drained areas with forested land use, and moderately
well-drained and well-drained areas with agricultural
use (Shedlock et al., 1993). In the northern part, the
"region has hurhmocky topography and low relief with .
many seasonally ponded wooded depressions. In the
southern part, topography is relatively flat with broad
poorly drained forested areas that are seasonally
flooded (J. M. Denver, unpublished). Streams are
small and sluggish in the poorly drained upland and
flow through shallowly-incised valleys with low gra-
dients (Phillips et al., 1993). Riparian zones are usu-
ally forested and often contain wetlands. Some parts
of the poorly drained upland have been ditched to pro-
mote drainage of agricultural fields.
Sediments that make up the surficial aquifer in the
PDU are predominantly sands and gravels, similar to
those in the well-drained upland. The sediments range
in thickness from about 8 m in the north to more than
30 m in the south (Owens and Denny, 1979). The water
table is usually within 3 m of the land surface. This re-
gion is characterized as poorly drained because of the
combination of regionally high water table and small
degree of stream incision that results in groundwater
gradients too low to effectively drain the region, rather
than a low permeability substrate (Phillips et al., 1993).
Except for areas immediately adjacent to streams,
groundwater flow paths in the PDU range from about
100 m to about 1 km in the northern part of the region
where the aquifer is thin. In the southern part, where
the aquifer is thick, flow paths are up to several km in
length and generally originate near the regional
drainage divide. Local flow patterns vary seasonally,
however, smaller localized flow paths associated with
the depressional wetlands and intermittent streams in
the north and intermittent streams in the south are.ac-
tive during wet seasons (generally winter and spring).
A more regional flow system from topographic highs
to perennial streams is active throughout the PDU
during the drier seasons (generally summer and fall).
d. Surficial Confined
Watersheds in the surficial confined (SC, Figure 3)
-------
39
38
•76°
.75'
21
HYDROGEOMORPHIC .
REGIONS
20 MILES
0 10
0 10 .20 KILOMETERS
.EXPLANATION ,
CENTRAL UPLANDS. '
Poorly-drained upland
Well-drained upland
Surficial confined
region
Inner coastal plain
COASTAL LOWLANDS
Poorly-drained lowland
Fine—grained lowland
Coastal wetland and
beach region
FIGURE 3. Hydrogeomorphic regions of the Delmarva Peninsula. (From Phillips et al, 1993).
-------
22
region are geomorphologically similar to the southern
part of the poorly drained upland with low relief and
shallow incision of stream valleys, features that con-
tribute to the poor drainage in the region. Topo-
graphically, the area is a flat sandy plain with low
ridges that rise a few meters above the surrounding
landscape. The plain is dominated by poorly drained
soils and the ridges are dominated by well-drained
soils. Throughout the region large tracts of forest are
interspersed with agricultural fields on the plains;
there are broad forested riparian zones and swamps
around the major drainageways. With the exception of
the sandy dune ridges, agricultural land is heavily
ditched to promote soil drainage and would probably
be forested wetlands in the absence of ditching
(Phillips et.al., 1993).
The surficial aquifer is geologically heterogeneous
in the region, consisting of a major sand unit 25.to 30
m thick overlain by 0 to 13 m of complexly layered
clay, slit, and peat, which is itself overlain by 1 to 6 m
of wind-deposited sand with some peaty sand, slit,
and clay lenses at the base (Owens and Denny, 1979).
The complex of fine-grained deposits acts as a con-
fining unit between the sands of the upper and lower
units, except some areas where it is absent or entirely
composed of sand. The water table is generally less
than 3 m below land surface and occurs in the upper
sand unit. Local groundwater flow paths, in the upper
unit, are relatively shallow and generally less than 300
m long and extend from water-table highs in inter-
fluves between ditches and streams into the ditches
and streams. Regional groundwater flow paths, in the
lower units, are up to ten kilometers long and extend
from the uplands near the regional drainage divide to
major streams and rivers. Local and regional flow
paths are separated in most areas by the confining
layer, but local heads are higher than regional heads in
most places, and shallow flow paths extend into the
lower sand where confining beds are absent
(Shedlock et al., 1993). Residence time in the upper
sand is 15 years or less; in the deeper unit, groundwa-
ter residence tune is at least 40 to 50 years, except
where there is hydraulic connection with the shallow
unit (Dunkle et al., 1993).
2. Control of Nonpoint Source Pollutants
Although more studies have been done on Coastal
Plain riparian forests than in other physiographic re-
gions, a number of questions remain about the NPS
pollution control capacity of naturally occurring ri-
parian forest buffers. Other questions remain about
the NPS pollution control capacity of reestablished
and managed RFBS. The following discussion will
necessarily focus primarily on what is know about
naturally occurring riparian forest buffers and ex-
perimental GVFS. Although discussion of reestab-
lished RFBS will be limited, a number of useful con-
. elusions can be drawn from the existing Coastal Plain
information.
The studies on riparian forest buffer effects on NPS
pollutants in the Coastal Plain have tended to concen-
trate on the fate of nitrate in groundwater, with a sec-
ondary emphasis on the fates of N, P, and sediment in
surface runoff. Three areas of the Coastal Plain
(Georgia, Maryland, and North Carolina) have been
studied where gaged watersheds were used as the
basis for nutrient budget estimates of riparian forest
buffers. The studies from Maryland (Rhode R.) have
been used to develop nutrient budgets for watersheds
and riparian systems (Peterjohn and Correll, 1984;
Jordan et al., 1986; Correll and Weller, 1989; Correll
et al., 1992). The studies from Georgia (Little R.)
have been used to develop both nutrient and sediment
budgets (Lowrance et al., 1983,1984a,b, 1985; Fail et,
,al., 1986). The studies from North Carolina have been
used to develop nitrate budgets for riparian systems
(Jacobs and Gilliam, 1985a,b). Hydrologic conditions
for all of these studies were representative of ICP con-
ditions.
A second general type of study has been conducted
on the fate and/or transport of potential NPS pollu-
tants, primarily plant nutrients and sediment. These
studies have also been primarily in Maryland and
Delaware (Correll et al., 1993; Jordan et al., 1993;
Whigham et al, 1986), Georgia (Lowrance et al.,
1988; Ambus, and Lowrance, 1991; Lowrance, 1992;
Vellidis et al., 1993), and North Carolina (Cooper et
al., 1987; Cooper and Gilliam, 1987), In addition,.
there are several studies of Coastal Plain hydrology or
water quality which provide information on upland ri-
parian interactions or provide limited data on NPS re-
moval in riparian forest buffers. These are studies
'which, in general, were not designed specifically to
• examine the removal of potential NPS pollutants in ri-
parian forest buffers (Lowrance and Leonard, 1988;
Weil et al., 1990; Staver and Brinsfield, 1990).
Preliminary results on integrated grass and forest
buffers in the Coastal Plain have been published
(Parsons "et al., 1991, 1994) and detailed studies of
GVFS have been conducted in the Coastal Plain of
Maryland (Magette et al., 1989).'.
-------
23
a. Nutrient Budgets for Riparian Forests
The most direct means of determining the NFS pol-
lution control function of a riparian forest is to de-
velop annual or 'longer term mass .balances.
Developing nutrient or sediment budgets requires a
' watershed from which hydrolpgic measurements can
be made which assure that all'watershed outputs are
measured and sampled. If the riparian forest buffer is
continuous around the entire, stream system and
groundwater discharging to streams moves through ri-
parian soils and shallow sediments, the streamflow
output can be treated as the output from the riparian
forest system. The inputs to the riparian system must
be estimated from sampling of groundwater and sur-
face water inputs. The studies which have done this
for Coastal Plain riparian forests are summarized in
Table 2. Total N and total P retention have been esti-
mated in studies of Watershed-109 (WS-109) of the
Rhode R. in Maryland and the Heard Creek tributary
' of Little* R. in Georgia. Both of these Coastal Plain
systems have effective aquicludes at depths which
limit Techarge to deep groundwater and which cause
all or nearly all excess precipitation to move through
riparian systems and exit the watershed as streamflow,
Estimates of N retention were 89% of input (Rhode
R.), and 66% of input (Little R.). P retention in Rhode
R. was slightly less (80% of input) but much less in
Little R. (24% of input). Total N and P budgets for
Little R. (Table 2) did not include surface ^runoff in-
puts of N and P from the agricultural areas to the ri-
parian forest but did include all streamflow outputs of
N and P. Streamflow includes surface runoff, which
moved through the riparian forest and contributed to
stormflow. Therefore, the N and P retention (input-
output) estimates for the Little R. site, are underesti-
mates of the actual retention. Peterjohn and Correll
TABLE 2.
Total nitrogen, nitrate-nitrogen, and total phosphorus budgets for
riparian forest ecosystems in the Coastal Plain
Reference
Total N
Peterjohn
Correll, 1984 -'
Location
Rhode R.;MD
Input ,
83
Output
9
Retention* Flux Notes*
•
74 NO3, NH4, Org-N in &
. SRO, GW, R PSF, PQF.
Lowrance et al., 1983 Little R., GA
39
Nitrate-N
Correll & "
Weller, 1989 '
Rhode R., Mb 45
Lowrance et al., 1983 Little R., GA 22
Cooper et al., 1986 Beaverdam Cr., 35
.-.-._ ' , „ NC
Total-P
Peterjohn & Correll, Rhode R., MD 3.6
1984 "•'.• -;: -.':-.
Lowrance et al., 1983 Little R.. GA 5.1
13
6.4
2.1
5.1
0.7
3.9
26 N03, NH4, Org-N In GW,
P, SF. ,
38.6 NO3linGW,SF
(baseflow only).
19.9 NO3inGW,SF.
29.9 NO3 in GW, SRO, SF.
2.9 Total P in SRO, GW, R
PSF, PQF.
1.2 Total P in GW, P, SF.
+Retention = Input-Output "
*SRO = surface runoff input; GW = groundwater input; P = precipitation input; SF = streamflow output; PSF = partitioned
slowflow; PQF = partitioned quickflow
-------
24
(1984) included direct estimates of both surface
runoff and groimdwater inputs and outputs for Rhode
R, Their budget estimates were based on these direct
measurements rather than streamflow. outputs.
Streamflow outputs for Rhode R. were different than
the riparian budget output for both total N and P. This
difference has only a negligible effect on the total N
budget, but has a large effect on the total P budget. If
streamflow outputs are considered the output from the
riparian forest for the Rhode R. site, .the total N reten-
tion is still 83% of inputs, but P retention is zero.
The Little R. and Rhode R. studies were both done
in systems which are likely to maximize retention by
natural riparian forests. Although the studies report
different ranges of percent retention for N and P, re-
tention of N was generally high. Both watersheds
have percentages of agricultural land typical for the
more agricultural portions of the Coastal Plain and are
representative of potential inputs to riparian systems
in the absence of animal confinement facilities and
manure disposal systems. These natural riparian sys-
tems would appear to retain at least two-thirds of the
N inputs but perhaps as little as one-third of the P
input. • ' •
In both the Rhode R . and Little R. studies, nitrate-
in subsurface flow made up the majority of total in-
puts to the riparian forest system.. The input in
groundvvater for WS-109 of Rhode R. in the year re-
ported on in Peterjohn and Correll (1984) was 57 kg
NOa-N ha"1 yr1 based on the area of riparian forest.
This accounted for 69% of the total N input (Table 2).
Based on two more years of data for WS-109 of
Rhode R., the input averaged 45 kg N03-N ha'1 yr1
(Correll and Weller, 1989). Data from Little R.
showed that groundwater input was 22 kg NOs-N ha"
1 yr1, 56% of total N input. A third study of nitrate
budgets (Cooper et al., 1985) on a Coastal Plain wa-
tershed in North Carolina showed similar results to
the MD and GA studies. Nitrate retention rates of
85%, 86%, and 90% for the three studies (NC, MD,
GA, respectively) reflect removal of nitrate through
both denitriflcation and plant uptake. Plant uptake
(and perhaps microbial immobilization) contribute to
transformation of a predominately nitrate input to the
riparian zone into a predominately organic N output
in streamflow. Total N input to the riparian forest on
Rhode R. was 69% nitrate. Streamflow was 51% or-
ganic N (Correll et al, 1992, Correll, 1983). On the
Little R., groundwater inputs to the riparian forest
were 74% nitrate. Streamflow outputs were 18% ni-
trate and 80% organic N. A later study of the entire
Little R. watershed showed consistent trends of ni-
trate increase during stormflow, indicating that the ni-
trate removal/transformation capacity of riparian
forests is partially by-passed when water moves
through more quickly during high flows (Lowrance
and Leonard, 1988).
b. Removal of Nitrate from Groundwater
Although elemental, nutrient, chemical, and sedi-
ment budgets on a watershed scale are the most com-
plete way to evaluate the functions of riparian forest
buffers and offer the best information on potential
load reductions, a number of studies have examined
nitrate concentration changes in riparian forests/This
emphasis on nitrate is due to a number of factors in-
cluding the relatively high transport rate of nitrate
from most agricultural systems, the availability of ni-
trate for algal uptake.as a stimulus for eutrophication,
and possible impacts on downstream or shallow
groundwater drinking water supplies. Studies in at
least five separate Coastal Plain locations have exam-
ined the changes in nitrate concentrations as shallow
groundwater moves from agricultural fields through
naturally occurring riparian forests (Figure 4). Studies
in four separate locations, have shown that average an-
nual edge-of-field nitrate levels of 7 to 14 mg NOs-N
L"1 decreased to 1 mg L'1 or less in shallow ground-
water near streams. Some studies, have used chloride
concentrations and nitrate:chloride ratios to separate
the effects of dilution from the effects of biological
removal of nitrate. Decreases in chloride concentra-
tions were generally small compared to nitrate de-
creases. Chloride concentrations usually increased at
some point in the shallow groundwater system proba-
bly due to exclusion of Chloride from the transpirar
tion stream (Peterjohn and Correll, 1986; Jordan et
al., 1993; Correll et al., 1993; Lowrance, 1992).
Most studies of nitrate dynamics in riparian forests
have shown that removal of nitrate from groundwater
continued year-round. Mechanisms to explain this
have not been elucidated, although it is likely that in
some of the Southeastern Coastal Plain areas,, rela-
tively warm soils and evergreen or tardily deciduous
(broad-leaf trees that lose leaves in the spring) vege-
tation can provide biological removal of the nitrate.
Most Coastal Plain areas of the CBW have lower
groundwater and soil temperatures in the. winter and
little or no evergreen vegetation. Weil et al. (1990) ob-
served year-round reductions of groundwater nitrate
in streamside forests on tributaries of the Choptank R.
on the Eastern Shore of Maryland. Groundwater
-------
25
•'_, 26
* 24
z
'.,22
o
2 20
• z • -
2 16
PETERJOHN & CORRELL, 1984 (MD)
CORRELL ET AL, 1993 (MD)
LOWRANCE, 1992 (GA)
JORDAN ET AL, 1993 (MD)
JACOBS &. GILLIAM, 1985 (NC)
STREAM CHANNEL LOCATION
RELATIVE TO DISTANCE
SCALE (NOT N03 -N)
30 40 50
DISTANCE FROM FIELD (m)
60
70
80
FIGURE 4. Nitrate concentrations in groundwater beneath riparian forests from five Coastal Plain sites.
under riparian forests always had less than 1 mg NO3-
N L"1 while adjacent fields had concentrations of 15-
40 mg NO3-N L'1. The decreases in chloride concen-
trations were much less than the nitrate decreases.
Year-round nitrate removal has been observed, but not
explained.
At least one study has shown that in situations with
relatively high nitrate concentrations entering from an
adjacent field, substantial nitrate concentration reduc-
. tions can occur but still leave high concentrations in
shallow groundwater at the stream (Correll et al,
1993), (Figure 4). This site, on a'tributary of the
Choptank R. on the Delmarva Peninsula is located in
the Well Drained Uplands. Nitrate concentration re-
ductions were actually higher at this site than at two
other Maryland Goastal Plain sites (Peterjohn and
Correll, 1984; Jordan et al, 1993) but groundwater
concentrations near the stream were 12 to 18 mg
s-N L"1. Similar results were inferred from a study
of nitrate in regional groundwater and nitrate levels in
streamflow for the WDU hydrogeomorphic region
(Phillips et al., 1993). In .related work, Bohlke and,
Denver (in press) concluded that, the riparian forest
wetland next to the stream in the Locust Grove
Watershed in Maryland had little effect on nitrate
movement to the stream. Hydrologic data and ground-
water flow modeling, show that groundwater dis-
charges upward directly to the streambed from the ;
aquifer system, effectively bypassing the riparian
zone (Reilly et al., 1994). Baseflow concentrations of
. nitrate commonly exceeded 9 mg NOs-N L"1 in, the
stream draining this watershed, and isotopic analysis
indicated that denitrification was not significantly af-
fecting nitrate concentrations (Bohlke, and Denver, in
review).
Nitrate transport into tidal streams' is often domi-
nated by direct recharge through sediments in inter-'
tidal zones (Reay et al., 1992; Simmons et al., 1992;
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26
Stayer and Brinsfield, 1994). Approximately 80 kg
ha'1 yr1 of NOs-N was discharged to a tidal creek in
Maryland with apparently most groundwater moving
at least 2 m below the ground surface in near-stream
areas (Staver and Brinsfield, 1994). These situations
may allow little chance for nitrate removal. The direct
NOa-N discharge to tidal streams make riparian
buffers desirable (Simmons et al., 1992), if proper
management could allow direct vegetation uptake
from the groundwater.
c. Nutrient Removal Processes
Removal processes were quantified in most of the
riparian forest research on nutrient budgets and nitrate
transport. Studies in Maryland and Georgia have
made direct estimates of N and P uptake by vegetation
and storage of N and P in woody biomass. Estimates
from Watershed 109 of Rhode R. (Peterjohn and
Correll, 1984; Correll andWeller, 1989) indicated that
total vegetation uptake of N and P was 77 and 10 kg
ha'1 yr1, respectively. N and P storage in woody bio-
mass was less than total uptake (Table 3).
Extensive data on total N and P uptake and woody
storage were reported by Fail et al. (1986, 1987).
Values for P uptake and storage are similar for the
Little R. and Rhode R. studies (Table 3). Major
differences between the two studies were found for N
woody storage and N uptake. Fail et al. (1986, 1987)
reported mean storage of N in wood as 52 kg N ha'1
yr1. The range was from about 35, to 98 kg N ha'1
yr1. The net primary productivity reported by Fail et
al. and Peterjohn and Correll are similar as are leaf N
concentrations and leaf and wood P concentrations.
Wood N concentrations averaged 7900 ug g'1 in the
Little R. studies, compared to average sapwood val-
ues of about 900 ug g"1 in the Rhode R. study. Fail et
al. (1987) used branch wood samples to represent the
entire woody biomass of the tree and so overestimated
N accretion in wood. Based on a number of studies,
they pointed out that bole wood N content averaged
about 43% of branch wood N content. This correction
would bring the net wood accumulation of N down to
about 22 kg N ha"1 yr1. Total N uptake would be
about 84 kg N ha'1 yr1 if this correction is applied.
Denitrification has been shown to be an important
N removal process in Coastal Plain riparian forests ei-
ther: 1) through indirect measurement using the
acetylene inhibition technique; 2) through measure-
ment of environmental conditions which control den-
itrification (Eh, water-filled pore space, N and C
availability) and verifying that proper environmental
conditions exist; or 3) through measurement of deni-
trification potential (Ambus and Lowrance, 1991;
Lowrance et al., 1984b; Hendrickson 1981, Jacobs
TABLES.
Above-ground woody vegetation uptake of N and P in Coastal Plain riparian forests.
Reference
Nitrogen
Location
Total
Input
Woody
Storage
Phosphorus
Total
Uptake
Woody
Storage
--kg ha"1 yr1
Correll & Weller, 1989
Peterjohn & Correll, 1984
Failetal., 1986, 1987
(mean)
Fail, 1986 (maximum)
Fail, 1986 (minimum)
Rhode R., MD
Rhode R., MD
Little R., GA
Little R., GA
Little R, GA
ND*
77
114
194.4
80
12 to 20
12 -~
52
97.6
34.6
ND
10
7.5
12.6
4.5
3 to 5
1.7
3.8
6.9
1.9
' ND = not determined
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27
and Gilliam, 1985,b; Correll et al., 1994; Jordan et al,
1993, Lowrance, 1992). The general conclusion of all
these studies was that denitrification occurred in most
riparian forest soils, especially in the root zone, or
that conditions were'favorable for denitrification.
Recent work by-Bohlke and Denver (in press) indi-
cated that denitrification can also occur in sediments
beneath the influence of the riparian root zone.
Denitrification was/measured in riparian forests
of Little R., GA in conjunction with water quality and
hydrologic measurements (Hendrickson 1981). A
total of 1114 soil cores,(0 to 10 cm), were taken
monthly for a year from 6 riparian forest sites on the
Heard Cr. tributary of Little R. Summarized data from
these ;studies were'used to estimate a denitrification
rate of 31 kg N ha'1 yr1 for the top 5.0 cm of soil for
the entire riparian zone of the watershed (Lowrance et
al., 1984b). Denitrification rates under conditions of
high N and C subsidy from a swine operation ranged
up to 295 kg N ha -1 yr1 (Hendrickson, 1981). Lowest
denitrification rates (1.4 kg N ha'1 yr1) were mea-
sured in a riparian zone adjacent to an old field which
received no fertilizer application. Hendrickson (1981)
found that the active cores (those producing N2O
above background levels) ranged from 1.1% to 6.6% of
the cor.es taken, depending on the site. This study con-
firmed the potential for denitrification in surface soils
as well as the high variability to be expected in field
measurements of denitrification. Soil cores taken
to 50 cm in 10 cm increments showed that, except
near the stream channel, denitrification activity below
20 cm depth was much lower than activity in the-top
20cm. -..'.....'"
Later studies from Little Ri, GA have also shown
that denitrification potentials at the top of the water
table are measurable, but very low (Lowrance, 1992).
Nitrate which moves into upper soil layers is likely to
be reduced by denitrification. Nitrate moving through
a restored RFBS was reduced by high rates of denitri-
fication averaging 68 kg N ha'1 yr1. These high rates
were due to a relict forested wetland soil and move-
ment of high nitrate water in the root zone (Lowrance
et al., in press). In addition, nitrate which moves
through anoxia sediments in riparian zones is .also
likely to be reduced. In contrast, nitrate in groundwa-
ter which moves through generally oxic aquifer mate-
'. rial or nitrate which.does not generally come in con-
tact with the root zone soil layers is much less likely
to be denitrified.
The interaction of vegetation nitrogen uptake, or-
ganic carbon production via. litterfall and root senes-
cence, and microbial denitrification appear to be dri-
ving nitrate removal in most Coastal Plain riparian
. forests. Correll and Weller (1989) proposed a model
of belpwground processes affecting nutrients (Figure,
5) .which conceptualized the system' as being con-
trolled largely by oxidation-reduction conditions.
Organic matter from decomposing litter and roots
DIFFUSION
PRECIPITATION
Cropland, Forest
Boundary _:
Ground
Water
FIGURE 5. Conceptual model of below ground processes'affecting groundwater nutrients in riparian forest
(from Correll and Weller, 1989),
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28
serves as an energy source and oxygen is consumed
through aerobic respiration, followed by nitrate re-
duction, followed by sulfate reduction when condi-
tions become sufficiently reduced. In the presence of
carbon rich sediments or relict organic matter hori-
zons, these processes could potentially proceed with-
out forest vegetation. A similar conceptual model for
nitrate removal in Coastal Plain riparian forests was
proposed by Lowrance (1992). Stratified denitrifica-
tion potential in riparian forests of Little R. indicated
that denitrification coincided with the stratification of
N and C from litter and roots. These findings support
the hypothesis that nitrate removal in RFBS is depen-
dent on interactions in the forest ecosystem rather
than just a poorly drained soil adjacent to a stream. It
is likely that nitrate removal in all Coastal Plain forest
sites (where substantial removal has been demon-
strated) was due to these complex interactions of veg-
etation and the belowground environment. It should
be noted, however, that hydrologic conditions in
which groundwater containing nitrate passes through
or near the root zone must be present for this mecha-
nism to operate effectively. Although most of the
Coastal Plain studies of nitrate removal were in areas
with relatively flat wetland soils near the stream, re-
moval often took place in areas immediately downs-
lope from the fields on better drained soils.
d. Removal of Sediments and Nutrients
from Surface Runoff
Removal of nutrients and sediment from surface
runoff in the RFBS will be a function of both Zone 3
and Zone 2. Sediment and nutrient deposition from
surface runoff moving through a Coastal Plain ripar-
ian forest has been estimated from direct sampling of
surface runoff in the Rhode R. watershed (Peterjohn
andCorrell, 1984). Estimates of sediment deposition
have been made based on soil morphology and 137Cs
profiles in Little R., GA and in Cypress Creek, NC.
GVFS have been widely studied, with at least one de-
tailed study of fescue buffers in the Coastal Plain of
Maryland (Magette et al., 1989).
The estimated range of sediment deposition rates
in riparian forests is large and apparently somewhat
dependent on estimation technique (Table 4). Al-
though the different methods give widely divergent
numbers, in all cases sediment deposition accounted
TABLE 4.
Sediment deposition in Coastal Plain riparian forests.
Reference
Location
Sediment Deposition Notes
Mg ha"1 yr1
Peterjohn & Correll, 1984 Rhode R. (MD)
4.2
Cooper etal., 1987
Cooper etal., 1987
Lowrance, et al., 1987
Cypress Cr. (NC) 105-315*
Cypress Cr. (NC) • 35-105*
Cooper et al., 1987 Cypress Cr. (NC)
Lowrance, et al., 1986 Little River (GA)
Little River (GA)
0-35*
35-52
256-262
Annual measurements, first order ,
stream, runoff samples
137Cs measurements—forest edge
137Cs measurements^ephemeral &
Intermittent streams
137Cs measurements—floodplain
swamp
Watershed based, long term,
sediment delivery ratio, soil
morphology
Single field/forest system 137Cs
measurements
*Based on sediment depths in Cooper et al. (1987) and assumed bulk density of 1.4 g cm-3.
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29
for 80 to 90% of gross erosion from the uplands.
Relatively low overall deposition rates (4.2 Mg ha'1)
reported from direct sampling were associated with
90% reductions in sediment concentration in 19 m of
flow through a riparian forest (Peterjohn and Correll,
1984; Table 6). Sediment deposition estimates need to
be compared to the gross erosion rates from'cropland
with information on the contributing area:riparian
. area ratio. With a field: forest ratio of approximately
2:1,. the riparian forest would attenuate cropland ero-
sion rates of about 2.1 Mgha^yr1 per year (Peterjohn
and Correll, 1984). This is well below the tolerance
value for the upland soils and many fields would con-
tribute higher sediment loads from erosion. In con-
trast, a sediment deposition rate of 35 Mg ha'1 yr!.at
a 2:1 field to forest ratio would attenuate erosion from ;
cropland contributing up to 17 Mg ha'1 yr1. Very" high
sediment deposition rates (up to 315 Mg ha'1 yr1) re-
ported from 137Cs' distribution studies (Table 4) were
due to high deposition at field edge. This deposition
was mostly coarse material and did not contain large
amounts of adsorbed nutrients.
Sediment removal in GVFS in Coastal Plain areas
is very effective in relatively short distances (Table 5).
The RFBS would generally include a grass strip of a
little more than 4.6 m. If concentrated flow occurs
across the GVFS, sediment removal is much less effi-
cient. The grass strips also become less effective
when multiple rainfall events take place in a few days
or when sediment begins to accumulate and forms
berms which can lead'to channelized flow (Magette et
al., 1989). Field evaluations of GVFS indicated that
they were more effective in Coastal Plain areas of
Virginia than in steeper topography (Dillaha et al.,
1989b). Slopes in Coastal Plain areas were more uni-
form and field reconnaissance indicated that signifi-
cant portions of stormwater runoff entered the GVFS
as shallow uniform flow. These GVFS needed regular
maintenance (sediment removal and possible revege-
tation every 1 to 3 years) because of the amounts of
sediment deposition (Dillaha et al, 1989b).
Nutrient removal from surface runoff has received
very limited study (Tables 5 and 6). The 4.6 m -filter
strips used by Magette et al, in the Maryland Coastal
Plain generally did not remove total N from surface
runoff and removed only 27% of the total P load. The
9.2 m filter strips had total N and P removals of nearly
50%. Peterjohn and Correll (1984) reported concen-
tration reductions of 74% for total N and 70% for total
P in flow through 19 m of mature riparian forest in
Watershed 109 of Rhode R. (Table 6). This width of
forest would be very similar to a Zone 2 which con-
formed to the RFBS specification.
.' Data from Magette et al., (1989) .and Peterjohn
and Correll (1984) have been combined to estimate
the /effects of combined Zones 3 and 2 on sediment
and nutrients in surface runoff (Table 6). The GVFS
of Magette et al. are analogous to Zone 3 and the 19
m of mature forest from Peterjohn and Correll is anal-
ogous to Zone 2. These widths, 4.6 m and 19 m, are
almost the exact widths specified in Welsch (1991) for
/Zones 3 and 2, respectively. Applying the 89.8% sed-
TABLE5.
Inputs, outputs, and % removals of sediment (total suspended solids), total N
(particulate + dissolved), and total P (particulate + dissolved) from experimental
Ky 31-Fescue vegetated filter strips in Maryland Coastal Plain. From Magette et al., 1989.
Filter
Strip ' •"-•---
Width Total Suspended Solids
Input Output Removal*
m —-Mgha'1-— %
4.6 27.2 9.3 66
9.2 , 27.2 4.9 82
Total Nitrogen
Input Output Removal
-—kg ha-1 %
39.4 41.6 -5
39.4 20.7 , 47
Total Phosphorus
Input Output Removal
-'—kg ha'1 ~- %
32.3 23.6 27
32:3 17.4 46
*Removal (%) = (Input-Output)/Input. Negative removal is percent increase in load after movement of runoff through filter strip.
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30
TABLE 6,
Effects of different size buffer zones on reductions of sediment and nutrients from field surface runoff.
Buffer Buffer
Width Type
m
4.61 Grass
9.21 Grass
19.02-3 Forest
23.6s Grass/
Forest
28.26 Grass/
Forest
Sediment
Input. Output Reduction4
Cone. Cone.
Nitrogen
input Output Reduction4
Cone. Cone.
Phosphorus
Input Output Reduction4
Gone. Cone.
— mg L'1—
7284 2841
7284 1852
6480 661
7284 290
7284 188
61.0
74.6
89.8
96.0
97.4
--mgL-1--
14.11 13.55
14.11 10.91
27.59 7.08
14.11 3.48
4.0
22.7
74.3
75.3
--mgL'1 —
i L
11.30 8.09
11.30 8.56
\
5.03 1.51
11.30 2.43
%
28.5
24.2
70.0
78.5
14.11 2.80 80.1 11.30 2.57
77.2
Calculated from masses of total suspended solids, total N, total P, runoff depth, and plot size (22 x 55 m) from Magette et
al.(1989)
2lnput concentrations from Table 2, Peterjohn & Correll (1984). Nitrogen = Nitrate-N + exch. part, ammonium + diss. am-
monium + part, organic N + diss. organic N. Phosphorus = part. P + diss P.
3Surface runoff concentrations at 19 m into forest reported by Peterjohn & Correll (1984). N and P constituents same as
Input (footnote 2).
^Percent reduction = 100 * (lnput-Output)/lnput.
54.6 m grass buffer plus 19m of forest.
69.2 m grass buffer plus 19 m of forest. ,
iment concentration reduction found in Peterjohn and
Correll (1984) to the output from a 4.6 m grass buffer
(2841 mg L"1) yields a sediment concentration of 290
mg L'1 from the 4.6 m grass and 19 m of forest (Table
6). This is an overall reduction of 96%. Applying the
same approach to total N and total P yields an output
concentrations of 3.48 and 2.43 mg L'1, respectively.
These represent concentration reductions of 75.3%
and 78.5% for total N and total P, respectively.
Increasing the width of the grass buffer to 9.2 m
would increase sediment retention by 1.4% of input,
N retention by 4.8%, but increase P concentrations
slightly.
Although a number of experimental studies are
ongoing which link grass filters and riparian forests
for sediment and nutrient removal from surface
runoff, most have only made preliminary reports.
Parsons et al. (1994) reports sediment load reductions
of 80 to 90% of field edge loads for both 4.2 and 8.5
m,Ky-31 fescue buffer strips at a lower Coastal Plain
site in North Carolina. Cutover riparian forests (per-
haps analogous to early natural regeneration in Zone
2) showed somewhat higher sediment and total N
yields than the 8.5 m grass strips. In a study of Zone
2 management on a tributary of Little R., Georgia,,
sediment loads in surface runoff entering the stream
channel system were significantly higher from a clear
cut Zone 2 than from a mature or thinned Zone 2
(Lowrance et al., unpublished). Although these results
are preliminary they suggest the importance of the
GVFS in Zone 3 during the early regeneration of
Zone 2 after tree harvest. In a study of a reestablished
RFBS, Vellidis et al.(1993), reports consistent but rel-
atively minor reductions in PO4-P in surface runoff in
the first year after establishment of slash pine in a re-
stored riparian forest buffer system in Little R. water-
shed. ,
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31
3. Conclusions ,
For purposes of estimating riparian ecosystem
functions in other physiographic regions^ results from
Inner Coastal Plain RFBS, probably represent the
upper limits for NFS pollution control in naturally oc-
curring riparian forests equivalent to Zones 2 and 1.
Other naturally occurring Coastal Plain and non-
Coastal Plain systems are likely to be less effective
, , than Inner Coastal Plain RFBS because of groundwa-
ter flow paths that bypass the riparian zone. Although
, numerous questions remain, the understanding of
Coastal Plain riparian systems is much advanced
compared to other portions of the CBW..
The ratio of source areas to RFBS which is re-
quired for continued improvement in water quality
can probably be greatest in ICP conditions. Under op-
timum hydrologic conditions, such as the ICP, where
groundwater moves in shallow pathways through nat-
: urally occurring riparian forests, a ratio of 2:1 or 3:1
(upland to riparian) is typical. These are the types of
.systems where some of the first data linking riparian
forests and water quality were collected. However,
data on nitrate concentration reductions suggest that
much of the removal occurs within a relatively narrow
ecotone at the field edge, implying that the ratio of
field/forest can be increased; Management of upland
source areas, to reduce NFS pollutants, andof RFBS,
to increase effectiveness of removal of NFS pollu-
tants, should provide opportunities for raising the
, ratio of cropland to RFBS.
Ongoing research on managed and experimental
RFBS in the Coastal Plain suggests that restoration of
the NFS pollution control function can be rapid, espe-
cially when nitrate moves through relict wetland soils.
These studies also confirm the need to control chan-
nelized flow and to use an effective GVFS for sedi-
i ment control when Zone 2 trees are harvested.
B. PIEDMONT
1. General Land Use and Hydrology
The Piedmont Province is an upland region lying
between the Coastal Plain and the Valley and Ridge
Provinces at elevations ranging from 30 to 300 m. The
Piedmont accounts for 23% of the Chesapeake
Drainage or 32,600 km2 (NCPJ Chesapeake, 1982).
Of this area,'49% is in, woodland, 25% is used as
cropland, 4% is wetland, and 21% is in other uses.in-
cluding pastures, and suburban and urban land uses
(NCR! Chesapeake, 1982). Of the total cropland
within the Chesapeake drainage, 25% lies within the
Piedmont '.
The Piedmont is underlain primarily by metamor-
phic Precambrian and early Paleozoic rocks subject to
several episodes of folding. The majority of Piedmont
basementjmaterials are quartzites, gneisses, schists,
and marbles. These rocks were metamorphosed from
ancient sandstones, gabbros and granites, shales, and
limestones, respectively. During the Paleozoic, these
basement rocks were interspersed with igneous peg-
matite intrusions, and portions were covered by sedi-
• mentary deposits during the Triassic era. In
Pennsylvania and Maryland, the marble belts form
valleys; the gneiss,- schist, quartzite, and granites form
uplands (Hunt, 1974). Pavich et al., 1989 described
the upland residual mantle (regolith) of Fairfax Co.,
VA as representative of the outer Piedmont
Crystalline Province of Virginia and Maryland
(Thorhbury, 1965). The area has a high drainage den-
sity with most of its perennial streams incised into uh-
weathered bedrock. .
, Given the great age of the rocks, the high degree of
weathering, and absence of quaternary glaciatipn, the
regolith (weathered rock, saprolite, subsoil, and soil)
in< the Piedmont can be quite deep. The maximum
thickness of regolith is beneath flat upland hilltops.
On schist, gneiss, and granite it is typically 15 to 30
m deep. Rocks such as serpentine and quartzite which
weather slowly have thin-regolith (Pavich et al,
1989). Throughout, the outer Piedmont Crystalline
Province, unweathered bedrock crops out in streams
and regolith is generally thin or absent in valleys of
. perennial streams (Pavich et al., 1989). The contact
between weathered and unweathered rock can be es-
timated on the side slopes of valleys by the location of
heads of perennial streams at minor springs.
Groundwater drains along the contact between weath-
ered and unweathered rock and enters surface flow
through springs (Pavich et al., 1989). Most of the
groundwater storage in the Piedmont is within the re-
golith above.the unweathered bedrock (Pavich et al.,
1989). The saprolite acts as a relatively porous reser-
voir for groundwater. To a large extent, the depth of
the regolith controls the hydrology of most Piedmont
areas. . •
Based on hydrograph separations .in the Piedmont
of Chester County, Pennsylvania, Sloto (1994) found
that basefiow ranges from 57 to 66% of watershed
discharge, similar to estimates for the Virginia
Piedmontof 60% (Pavich et al., 1989). The remainder
of streamflow occurs during and following storms,
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32
but the proportion that is surface runoff, as opposed to
enhanced subsurface flows (e.g., through near-stream
rise in groundwater, drainage from soil layers, or
rapid lateral transport through macropores), is diffi-
cult to determine. In forested watersheds, very little
surface runoff occurs except from near-stream zones
of high soil moisture. However, cultivated fields in the
Piedmont generate greater surface runoff than fields
in lower gradient Coastal Plains.
2. Control of Nonpoint Source Pollutants
Direct studies of NPS pollution control by RFBS
have generally begun since 1990 so most results are
preliminary at this point. The complex hydrogeology
of the Piedmont Province will make generalizations
from ongoing studies difficult even when final results
are available. Most of the discussion to follow will
focus on preliminary results from the North Carolina
Piedmont which are most applicable to the southern
portion of the Piedmont in the Bay watershed.
Discussions of the geohydrology of the Piedmont and
recent studies of the sources of water reaching,
streamflow will also be used to make inferences about
the roles of RFBS in this province.
a. Removal of Nitrate from Groundwater
Groundwater in the Outer Piedmont Crystalline
Province drains along the contact between weathered
and unweathered rock and discharges through springs
(Pavich et al., 1989). There are thought to be three
pathways for groundwater discharge. In valleys un-
derlain by weathered saprolite (often near headwa-
ters), flow through the saprolite dominates baseflow.
Water in the flow system is often oxic and may dis-
charge nitrate directly to the stream channel. In val-
leys where streams have cut through the regolith to
bedrock, springs begin in the valley flanks. Where
streams have eroded to bedrock, discharge from frac-
tures in the bedrock also contribute to streamflow.
Stream discharge from the bedrock groundwater sys-
tem is bypassed by the shallower systems if the re-
golith is not entirely eroded away. Even where
bedrock contributes, most of the water in streams
originates in the regolith because the volume of water
in storage is so much greater than in the fractured
bedrock.
Most groundwater recharge in the marble valleys
occurs rather rapidly into fracture zones close to the
land surface. The regolith, although variable in thick-
ness, is usually thin and discharge to streams is prob-
ably from discrete fracture zones (in springs or di-
rectly into stream channels). As a result, there is prob-
ably little interaction of the groundwater with the root
. zone of riparian systems in the marble valleys.
In one study of nitrate transport in the Maryland
Piedmont, McFarland (in press) found that streams
contained nitrate concentrations of 5 to 10 mg NOs-N
I/1. Most of the nitrate was attributed to discharge of
water that was 0 to 5 years old from springs and from
shallow flow systems in the regolith. Water in the
bedrock was 20 to 30 years old with low or zero ni-
trate concentrations. Denitrification was suspected
along older flow paths because of low dissolved oxy-
gen and high iron concentrations in the water. This
study indicates that riparian systems with deeply
' rooted vegetation may reduce nitrate.in streams by re-
moving nitrate from spring flow and the shallow flow
systems through the regolith.
The only experimental study from the Piedmont
that addresses the effectiveness of riparian buffers in
mitigating subsurface flows of nonpoint pollutants is
that of Daniels and Gilliam (in press), in which spa-
tial and temporal patterns of groundwater nitrate at
three sites in North Carolina were examined,.
Cultivated fields were separated from ephemeral or
intermittent stream channels by 3 to 20 m of grass and
naturally forested riparian buffers. Nitrate concentra-
' tions in groundwater under the cultivated fields ex-
ceeded 10 mg I/1, but declined to lower levels in
downslope wells. At one site, concentrations declined
by as much as 30 mg I/1 over a distance of 20 m from
the field edge. The study did not include mass balance
analyses of nitrogen losses, and interpretation is com-
plicated by the fact that streamflow nitrate concentra-
tions exceeded those in near-stream wells. Thus, the
. authors were unable to partition actual nitrogen re-
moval within the'riparian zone from mixing (or dilu-
tion) effects, although they speculated that both were
involved.
The results of Daniels and Gilliam (in press) are
consistent with findings from Coastal Plain studies
showing that high rates of nitrogen removal occur in
areas with high water table conditions and shallow
groundwater movement near the root zone. This sug-
gests that the effectiveness of RFBS at particular sites
throughout the Piedmont will depend strongly on the
flowpaths of subsurface water in and near the riparian
zone. Whatever the outcome of additional site-sper-
cific studies, it seems likely that regional estimates of
RFBS effectiveness will also require data regarding
hydrologic properties of near-stream zones.
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33
The Piedmont is topographically diverse. In areas
of gentle slopes and broad .alluvial floodplains, the
depth to ground-water in near-stream areas is probably
1 to 2 m 'as was the case in the North Carolina
Piedmont study. In such areas there may be ample in-
tersection of the saturated zone with the» root system
of riparian vegetation. This may also allow interaction
between the saturated zone and soil layers containing
sufficient 'organic matter to induce rapid denitrifica-
tion. However in upland areas, the'water table typi-
cally lies 3 to 10m below the ground surface. In areas
of steep terrain, it is common for smaller streams to
be incised in relatively steep valleys. Near the fall
line, larger streams also tend to have steep valley
walls and minimal floodplains. Under these circum-
stances, the area of riparian forest in which the soils
and root zone intersect the water table may be quite
small. ' .
Perhaps equally; as important as water table eleva-
tion is depth or thickness of the aquifer in the near
stream zone. While the bulk of subsurface water stor-
age in the Piedmont occurs.in the regolith, which may
vary in depth from less than one meter to approxi-
mately 3 0 m, substantial storage occurs within a
deeper zone of unweathered but fractured bedrock.
The depth of .the fractured zone^ as indicated by
well-water yields, may range roughly from 60 to 200
m, depending on rock type (Sloto, 1994). The likeli-
hood, that water reaches streams via shallow path-
ways, therefore, would depend both on the depth of
the regolith in the vicinitylof the stream,.and on the
proportionate contributions to streamflow from the re-
golith and from the fractured zone. Olmstead and
Healy (1962) concluded from analyses of temporal
patterns in, baseflow and water table elevations in the
Brandywine Valley of Pennsylvania that most stream-
flow originated from the regolith. Rose (1992, 1993)
reached a similar conclusion from analyses of tritium
variations in. streamwater and groundwater in the
Georgia Piedmont. If the regolith is beneath alluvial
deposits, near, streams, much of the water reaching
streams may: pass through the riparian zone at sub-
stantial depths. '
Another aspect of subsurface water movement that
may prove important to RFBS effectiveness is the po-
tential for lateral flow through near-surface soil lay-
ers. Lateral ;downslope water flow through unsatu-
rated or briefly saturated soils may occur through
macropores (Bevin and Germann, 1982) or where
vertical flowis impeded by a soil horizon of low per-
meability (Gaskin et al., 1989; Schoeneberger and
Amoozagar, 1990). There has been considerable, in-
vestigation of shallow lateral drainage in .other re-
gions (e.g*, Mosley, 1982; Mulholland et al., 1990;
McDonnell, 1990) but only a few studies from the
Piedmont Province.
Hooper et al. (1990) used end-member-
.mixing analysis (EMMA) of water chemistry to dis-
tinguish water sources in a Georgia Piedmont water-
shed. Their model used alkalinity, sulfate, sodium,
magnesium, calcium and dissolved silica to identify
three water sources: an organic soil horizon, hillslope
drainage through subsoil and saprolite, and ground-
water in bedrock. They concluded that hillslope
drainage contributed a large portion of both baseflow
and stormfiow drainage during the wet winter
months. Groundwater dominated the baseflow during
the dry summer months with significant contributions
from the organic horizon during storms. Comparable
results were obtained by Rose (1992,1993) in another
study in the Georgia Piedmont. Rose inferred from
analyses of tritium arid inorganic analyses, that while
baseflow during dry summer months originated from
groundwater with an average residence time of 15 to
30 years, higher winter baseflows included a substan-
tial component of water with a much shorter resi^
dence time (less than 10 yr) and lacking the chemical
signature of groundwater.
In the North Carolina Piedmont," Daniels and
Gilliam (in press) noted that soil water in an alluvium
overlying saprolite was chemically distinct and appar-
ently isolated from thex deeper groundwater in the
saprolite. They attributed the isolation to low perme-
ability of the Bt soil horizon (subsoil compacted by
tillage), and inferred that water in the soil layers trav-
eled laterally above the Bt horizon into the riparian
zone. Kaplan and Newbold (1993) hypothesized ex-
tended periods of soil water drainage following
storms to explain patterns of dissolved.organic carbon
concentrations in a Pennsylvania'Piedmont stream.
The Bt horizon is well developed in Typic Hapludalfs
and Typic Hapludults, soil groups.which are common
throughout the Piedmont, particularly in agricultural
areas. In and near riparian zones, Aquic Fragiudults,
are common. The fragipan associated with the latter
soils probably also promotes lateral flow.
b. Removal of Sediment and Nutrients in
Surface Runoff
The ability of RFBS to reduce nonpoint-source
pollutants in overland flow may be of greater signifi-
cance in the Piedmont than in the Coastal Plain be-
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34
cause the steeper topography promotes greater veloc-
ities of overland flow. Daniels and Gilliam (in press)
studied sediment and chemical reduction by GVFS
and riparian areas for two years at six sites in the
North Carolina Piedmont. They reported that the total
sediment load reduction by the vegetated buffers dur-
ing the study period ranged from 30 to 60 percent.
However much of the sediment (mostly sand) passed
through the vegetated buffers during one storm. When
the results of that one storm were omitted from the
calculations, the buffers removed approximately
80 percent of the sediment. Removals of silt plus clay
averaged approximately 80% for the two-year study
period. Total P removals in the filters ranged from 50
to 70%. Soluble orthophosphate removal was highly
variable and usually was 50% or less. Removal of var-
ious forms of N was also variable and generally
ranged from 40 to 60%. Most of the reductions were
observed within 7 m of the field edge. The authors
noted that the slope of the GVFS was less than that of
the fields, so some of the sediment removal could be
attributed to the change in slope alone. They further
cautioned that the effectiveness of GVFS on steeper
slopes might be limited. Where runoff from fields was
directed as concentrated flow into riparian areas with-
out complete vegetative cover, little or no reductions
in either sediment or nutrients were observed. From
these observations, Daniels and Gilliam (in press) rec- .
ommended upslope dispersal of drainage water di-
rected into forested areas.
Parsons et al. (1994a, b) conducted plot-scale stud-
ies in the North Carolina Piedmont on sediment and
nutrient removals in grass and forest vegetated filters.
They used 4 and 8 m grass buffers and 4 and 8m
forested filters to determine removal efficiencies. To
date, they have monitored 50 storms over a three year
period. They have observed that grass filters were
somewhat more effective for sediment removal than
the forest filters because of greater tendency for chan-
nelization in the forested area. Comparison of the
grass and forest buffers is difficult because slopes
were 4 to 6% in the grass filters as compared to slopes
of 12 to 16% in the natural forested area. There was
generally more ground cover on the grassed plots than
in the forest, especially after grasses were reestab-
lished in grass buffers.
Preliminary data from these studies are available
for a maximum of four storms in 1991 (Parsons et al.
1994b). They found reduction of field edge sediment
loads was consistently over 90% for both 4.3 m and
8.5 m forest buffers. Sediment loads were reduced
94% in the 4.3 m forest buffer (three storms) and 92%
in the 8.5 m buffer (two storms). Reductions of ni-
trate, total Kjeldahl N, ortho-P and total P were more
variable in these riparian forest buffers in the four
1991 storms. Although data are not available for all
storms, it appears that the tendency to have channel-
ized flow through the riparian forest area caused the
high variation. For the available storm data lumped
together, nitrate was reduced 41% compared to edge
of field load for the 4.3 m forest buffer (four storms) ,
and reduced 63 % in the 8.5 m buffer (two storms).
Total kjeldahl N was reduced 67% in the 4.3 m buffer
(three storms) but increased 14% in the one storm
monitored in the 8.5 m buffer. Ortho-P, all of which
was dissolved, decreased 6% in the 4.3 m forest buffer
(three storms) but increased 17% in one storm
through the 8.5 m forest buffer. Total P (sediment-
bound + dissolved) decreased 50% in the 4.3 m buffer
(three storms) but only 17% in the 8.5 m buffer (one
storm). Although these data are preliminary, they
show similar trends as some of the Coastal Plain
runoff data with good control of sediment and sedi-
ment associated P' but variable control of dissolved
nutrients in surface runoff, especially dissolved P.
More complete data from these studies should help
guide design of RFBS for Piedmont landscapes.
3. Conclusions
Limited data from riparian forest studies in the
Piedmont makes, quantitative estimates of the NPS
pollution control functions difficult. Patterns similar
to ICP results seem to be present in studies from the
North Carolina Piedmont with good control of nitrate
in shallow flow paths and good control of sediment
and sediment-borne pollutants in surface runoff.
Knowledge of the hydrology of certain parts of the,
Piedmont, such as the marble valleys of Pennsylvania
and Maryland indicate a minor role for RFBS in con-
trol of groundwater borne pollutants. On smaller
streams and in areas with thinner regolith, it appears
that shallow groundwater movement through the sub-
soil and saprolite may be affected by RFBS. Buffer
systems in the headwaters of streams where springs
enter surface runoff may be effective, especially if the
RFBS promotes the eventual presence of high organic
matter soils in the areas of springs and permanent
groundwater seeps. RFBS probably also intercept
water moving in relatively shallow flow paths above
texture discontinuities which promote lateral move-
ment in the soil and subsoil. If extended periods of
soils drainage above these texture discontinuities does
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35
occur, these waters should be subject to nutrient re-
moval rates in RFBS similar to those in the ICP situ-
ations. ,
Although data are also limited on effects of RFBS
on surface runoff, preliminary results indicate that the
slope of RFBS may limit effectiveness because, of
channelization through forests. Relatively steep
RFBS will certainly benefit from the presence of a
well managed Zone 3 at the field edge and may re-
quire level-lipped spreaders,to control the tendency of
surface runoff to create permanent channels. In rees-
tablished areas on relatively steep slopes, such as the
12 to 16% slopes reported on from North Carolina, a
.high stocking density of trees in Zone 2 is warranted.
This would have the effect of both increasing resis-
• tance to surface flow by increased numbers of stems,
as well as providing a high level of root biomass more
quickly than lower stand densities.
C. VALLEY AND RIDGE
1. General Land Use and Hydrology
.' Valley" and Ridge physiographic province is the
area in which structures due to folding dominate the
topography. The Valley and, Ridge and Appalachian
Plateau make up about 60% of the CBW (Table 1).
Geomorphologically, the Valley and Ridge province is
one of folded mountains in which resistant strata form
ridges and weaker rocks are-worn down to lowlands.
Valleys within this province are underlain by lime-
stone or shale and the ridges are capped by the moire
resistant rocks (well-cemented siliceous sandstone
and conglomerate). . .
The physical characteristics of this province are in-
timately connected with its-streams which are primar-
ily causes of the present topography. Streams develop '
mostly on belts of soft rock crossing hard rock ridges
infrequently and usually at right angles. The .Bay wa-
tershed encompasses the middle section of the Valley
and Ridge. Distinctive features of this section are con-
spicuous trellised drainage patterns and a comparative
absence of ridges on its southeastern one-quarter to
one-third, the Great Valley (Fenneman, 1938).
Heath (1984) placed the Valley and Ridge in the
' Central Nonglaciated groundwater region. The region
. is characterized by thin regolith underlain by frac-
tured sedimentary bedrock. The principal water-bear-
ing openings in the bedrock are fractures which de-
velop both along bedding planes and across them at
steep angles. Openings developed, along the fractures
are usually less than 1 mm wide. The principal ex-
» , i / -
ception to this is in limestone,, where water moving '
through the original fractures has enlarged them to
form, at the extreme, extensive cavernous systems ca-
pable of transmitting large amounts of subsurface
flow. Recharge of groundwater in this region gener-
ally occurs in outcrop areas of the bedrock aquifers in
the uplands between streams. Discharge from the
groundwater system,is by springs, seepage areas, and '
direct inflow to the stream bed, and by evaporation
and transpiration in the near-stream areas where the
water table is near the land surface. ' .'••.'
. The aquifers, in the Valley and Ridge are uncon-
fined with little matrix permeability and low storage
coefficients. Groundwater circulation is limited at '
depths greater than 100 m due to the decrease in frac-
ture size and frequency. Even though the entire Valley
and Ridge falls within the Central Nonglaciated
groundwater region^- there are substantial differences
in flow characteristics between the limestone and
shale valleys, arid among the limestone valleys. Flow
characteristics are most complicated within the lime-
stone aquifers and connections between lower-order
streams and regional groundwater are quite variable
in time and space. •".;-.
2. Control of Nonpoint Source Pollution
Despite the Valley and Ridge and Appalachian
Plateau comprising a large portion of the CBW, only
a small number of research projects have been con-
ducted to evaluate the amelioration of NFS pollution
in riparian buffers within this area. These projects ad-
dressed within-stream water quality (e.g., cold water
fisheries habitat, macroinvertebrates, and sedimenta-
tion) and Bay-scale water quality (export of plant nu-
trients, pesticides and suspended solids).
The entire study of riparian ecosystems relative-to
stream quality was begun within the last 20 years.
That little of it was conducted in the Valley and Ridge
may be explained by the small amount of wetlands in
this province. The conditions which promote im-
provements in the chemical composition of waters
discharging through riparian ecosystems (small land
surface slopes, high water tables and low aeration sta- •
tus) are commonly associated with wetlands. Only
1% of the Valley and Ridge located in the CBW is
classified as wetlands, which constitutes 7% of the •'.
wetlands in the CBW (Table 1). This contrasts with
the 57% of CBW wetlands on the Coastal Plain com-
prising 21% of the Coastal Plain within the CBW.
. The most intensive agricultural NFS pollution oc-
curs in the limestone valleys of the Valley and Ridge.
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36
Study of subsurface hydrology necessary to deter-
mine the extent of groundwater renovation in the ri-
parian zone is difficult and expensive in karstic lime-
stones. A paucity of studies in the Valley and Ridge
may be due to the major source of NFS pollutants
being located over an aquifer with complex hydrology
for the scale of riparian zone studies.
The processes which renovate surface and ground-
water within riparian ecosystems are the same in the
Valley and Ridge as in the other physiographic
provinces and much can be inferred from research
done in the Atlantic Coastal Plain. When watershed
mprphology and aquifer characteristics are compared
(Schnabel et al , 1994), general statements'can be
made about the likelihood of ground and surface
water renovation in riparian zones of the Valley and
Ridge relative to the Coastal Plain. However, research
conducted within the Valley and Ridge must be eval-
uated to quantify the impact of riparian buffers on
stream hydrology, chemistry and biology.
a. Removal of Nitrate from Groundwater
The Mahantango Creek Watershed, a USDA-ARS
research watershed, is located within the
Susquehanna R. Basin approximately 40 km north of
Harrisburg, Pennsylvania. Topography, geology, and
land use of the Mahantango Creek Watershed are typ-
ical of upland watersheds in the unglaciated, intensely
folded and faulted Valley and Ridge Province. These
watersheds generally have relatively steep land-
surface slopes and minimal floodplain development
or alluvium. Most stream reaches expose bedrock
over all or part of their length. Land use within the
watershed is approximately 57% cropland, 35% forest
and woodlots, and 8% permanent pasture. Elevation
ranges from 240 to 480 m msl. The northern ridge is
covered with a mature deciduous forest, while agri-
cultural land use predominates .in the remainder of the
watershed. Climate is humid and temperate, and rain-
fall averages about 1150 mm yr1.
Groundwater provides most of the 60 to 80% of
streamflow estimated to be subsurface return flow
(Gburek et al., 1986). Primary recharge occurs in the
late fall, winter, and early spring months, but minor
recharge can occur during the growing season follow-
ing large single or grouped precipitation events.
Because ridge-top soils are highly permeable, nearly
all rainfall infiltrates. In contrast, the finer-textured
poorly drained soils adjacent to the stream often func-
tion as groundwater discharge zones during the dor-
mant season.
The Mahantango Creek Watershed is underlain by
two geologic formations, Trimmers Rock (Late
Devonian) and Catskill (Late Devonian—Early
Mississippian). Previous analysis of well yields indi-
cated that rock fracture patterns are as important to
formation permeability as rock type, and based on
specific capacity data (Urban, 1977), the two forma-
tions are hydrologically similar. A shallow, approxi-
mately 10 to 15 m layer of weathered fractured
bedrock overlays the entire watershed and has hy-
draulic properties different from those of the deeper,
less-fractured layer (Gburek and Urban, 1990). The
two-layer aquifer, with its upper, highly transmissive
layer, permits rapid horizontal groundwater, through-
flow while also leaking to recharge the deeper layer.
Differing land uses in the area recharging groundwa-
ter, the layered subsurface permeability distribution,
and the general pattern of groundwater flow are ex-
pected to result in a general pattern of higher nitrate
concentration in shallow groundwater and lower .con-
centration in the deep groundwater. In the experimen-
tal area, all aquifer waters, both shallow and deep, dis-
charge to the surface streams.
Although it is a very small portion of the watershed
area, the near-stream zone exerts major controls on
stream flow chemistry and hydrology. Because it is
hydrologically dynamic, particularly as related to
seep zone formation, the near-stream zone can control
the amount and timing of surface runoff and, thus
downstream flooding. The water table response to
storms strongly influences or controls subsurface dis-
charge, the nature and extent of riparian, vegetation,
stream bank stability, "and the nature of the chemical
and biological systems to which chemicals in the dis-
" charge are exposed.
Nitrogen and phosphorus species were measured in
surface runoff and seepage waters in a grassed buffer
between a first-order stream and a cropped field, dur-
ing and immediately after storms to determine how
surface and subsurface waters interact to generate
streamflow during storms (Pionke et al., 1988).
Nitrate concentration in seepage and base flow were
similar and typically exceeded concentrations in sur-
face runoff, rainfall, and peak storm flow by 5 to 20
times. The median nitrate concentration observed in
seepage was similar to mean concentration observed
in stream base flow at the outlet to a 9.9-ha catchment
over a 2-yr period and similar to those computed from
a hydrologic and nitrogen mass balance for agricul-
tural groundwaters of Mahantango Creek Watershed
(Pionke and Urban, 1984). They concluded that hy-
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37
drblogic conditions mainly determined nitrate con-
centration and load delivered 'to the stream over the
short-term, as hydrology affected both the surface
runoff: subsurface discharge ratio and the volume of
subsurface discharge. This indicates that surface-
groundwater interactions are more frequent and
longer lasting on lower portions of the watersheds.
Consequently, in this and similar watersheds the ni-
trate content of groundwater is less likely to be altered
in the riparian zone in the,upper portions of head
water streams. !
Another study on Mahantango Creek focused on ,
the role of existing riparian zones. A strip of woods
(20-60 m) was located between the stream and crop-
land on both sides at the study site. There was a break,
in slope on one side of the stream where the land sur-
face flattens as it approached the stream. The woods
were removed from this flatter area bordering the
stream 15 to 20; years before the study and it was
seeded to grass. Thus the vegetation pattern moving
up-gradient on one side of the stream was a relatively
flat, well-maintained grass strip, a .steeper strip of
woods and then cropland. On the other side a,steep
strip of woods separated the stream from cropland.
Nitrate-N concentrations in shallow groundwater
under the grass strip were reduced by 25 to 50% be-
tween 9 m and 6 m from the stream during the grow-
. ing season. There were generally small differences in .
• nitrate-N concentrations in shallow groundwater 3 m
from the stream and baseflow in the stream. The water
table was frequently deeper than 1 m, particularly on
the wooded side of the stream.' The wooded side was
much steeper and didn't develop seepage zones as fre-
quently as on the less steep grassed side of the stream
(Schnabel, 1986). A pattern similar to the nitrate con-
centrations measured in the grassed riparian zone was
found in deeper groundwater (3 m) beneath the
wooded riparian zone; Gburek et al. (1986) estimated
that nitrate reduction within'the riparian zone of the
Mahantango Creek Watershed was equivalent to only
4% of the mineral N exported from the watershed dur-
ing the year. The limited impact of riparian processes
on .total N export resulted from the small area near the
stream thought to support denitrification at optimum
rates, combined, with the fact that the area generally
expands after soil temperatures begin to decrease,
presumably limiting denitrification and plant uptake
rates. , .
. The chemical composition of the aquifer differs,
with .depth. While recharge for the deeper part of the
aquifer originates at the wooded ridge tops, the shal- .
lower part of the aquifer is recharged in the agricul-i
rural interior of the watershed. From simulation with
a mixing nfbdel which, viewed baseflow as a mixture
of discharge from the shallow fractured part of the
aquifer and deeper, less fractured portion of the
aquifer, Schnabel et al. (1993) concluded that the ri-
parian zone was not the major control on temporal
variation in nitrate concentration at the outlet to
Mahantango Creek Watershed. "
A study designed to examine groundwater nitrate
dynamics was-conducted in the western portion of the
Valley and Ridge Province. Altman and Parizek
(1994) conducted a study of nitrate movement from a
field through the riparian zone of a tributary of Bald
Eagle Creek at the western edge of the Valley and
Ridge in Pennsylvania. They found that nitrate levels
in groundwater decreased from 5 to 8 nig NOs-N L'1
beneath the field to Jess than 0.5 mg NOa-N I/1 in the
-riparian zone. Based on flow-net analysis, they con- •
eluded that water sampled in the riparian zone appar-
ently did not originate from the crop area with ele-
vated nitrate levels. The groundwater flow direction
did not follow the surface topography but instead fol-
lowed the local bedrock topography. Groundwater
was actually flowing toward the larger creek which
the tributary stream was feeding. Their report did not
address the,fate of the nitrate enriched water as it
moved through the riparian,system associated with
Bald Eagle Creek, This study does point out the diffi-
culty of research on groundwater and associated
solute movement in areas of complex hydrogeology
such as the Valley and Ridge:
b. Streamflow Transport of Phosphorus
Sediment and water associated phosphorus ex-
port from the Mahantango Creek Watershed and a
9.9-ha subcatchment was examined for a 4-yr period
to determine the mode of phosphorus transport from a
typical Valley and Ridge upland watershed (Piorike
and Kunishi, 1992). During storms, most of the labile
P (sum of total soluble P and sediment P extracted by
Cl. resin) was exported from Mahantango Creek
Watershed in the dissolved rather than the particulate
phase. The dissolved P dominated because the dilu-
tion of sediment by runoff (~3000:1) more than com-
pensated for the greater concentration of labile P
compared to soluble P..concentration (~1000':1). .In
contrast, storm flow transport of algal-available (sed-
iment P extracted by 0.1 N NaOH) total P was mostly
with-sediment, largely because concentrations of both
on sediment greatly exceeded labile sediment P con-
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38
centrations. When combined with P concentrations in
base flow, which accounts for approximately 80% of
total flow, 50 and 28% of algal-available and total P,
respectively, were exported from Mahantango Creek
Watershed in the dissolved phase. Thus, the most
readily available P components are transported in the
dissolved rather than particulate forms. This has im-
portant implications for use of RFBS to control P
losses from agricultural land. If RFBS are more ef-
fective at controlling particulate P than dissolved P,
higher proportions of P in the dissolved phase would
imply less effective overall control of P transport.
c. Removal of Sediment and Nutrients in
Surface Runoff
Studies by Dillaha et al. (1988, 1989a,b) have
shown the potential efficacy and limitations of
grassed filter strips for controlling NPS pollution.
Near Blacksburg, Virginia, Dillaha et al. (1988) stud-
ied the use of orchardgrass (Dactylis glomeraia)
GVFS for controlling potential sediment and nutrient
losses from feedlots. Plots received 7,500 and 15,000
kg ha"1 of fresh dairy manure and had slopes of 11 and
16%. In plots with shallow, uniform surface flow, 81
and 91% of the sediment and soluble solids were re-
moved by 4.6 and 9.2 m GVFS, respectively. In plots
where concentrated flow was allowed to occur, re-
moval was much less. The GVFS were ineffective for
controlling dissolved nutrients and nutrients associ"
ated with fine sediment. Concentrations of soluble N
and P in effluents from GVFS were found to be high
enough to cause eutrophication in receiving waters.
Concentrations of soluble inorganic N were as high as
8.2 and 5.1 mg N I/1 from the 4.6 and 9.2 m GVFS,
respectively.
In a similar study of orchardgrass filter strips
below fertilized cropland, Dillaha et al. (1989b) ob-
tained comparable results to the feedlot experiment.
The sediment was initially trapped at the top of the
GVFS. However, the GVFS became ineffective as it
gradually became inundated with sediment.
In surveys of farms that employed GVFS along
streams in Virginia, Dillaha et al., 1989a,b) found that
in Valley and Ridge areas, the GVFS tended to be less
effective than in flatter Coastal Plain sites. Except for
localized erosion control along the stream .bank,
GVFS did little to mitigate NPS pollution from.the
upland in the Valley and Ridge because surface runoff
usually became concentrated within the fields in nat-
ural drainageways before entering the GVFS. In gen-
eral, the GVFS were most effective below smaller
fields where water could enter the GVFS before it had
an opportunity to concentrate.
Even where the GVFS had potential for sediment
trapping, in many cases inadequate maintenance had
rendered them ineffective (Dillaha et al., 1989a). Lack
of mowing sometimes allowed taller weeds to shade
put low ground cover, thereby reducing the capability
of the GVFS to trap sediment. Erosion across the
GVFS had caused severe gully problems in some
cases. Heavy traffic had sometimes damaged the sod
and created ruts. Sediment buildup on some sites had
caused the upper margin of the GVFS to be higher
than the adjacent field. Or sometimes, ditches from
moldboard plowing were created parallel to the upper
edge of the GVFS. In either of these two latter situa-
tions, water would run parallel to the edge of the
GVFS until it could get across it in concentrated flow.
3. RFBS in Forested Watersheds
Although the RFBS is designed for use adjacent
to agricultural areas, a number of forestry experi-
ments in the Valley and Ridge Province-provide use-
ful general information on hydrology, sediment trans-
port, and sustainability of riparian forest buffers. A
series of experiments were begun in the late 60's and
early 70's by Forest Service personnel to design
BMPs for logging operations in response to the
Federal Water Pollution Control Act Amendments of
1972 (P.L. 92-500). In many of these experiments, a
strip of trees was left standing along perennial
streams to protect the stream from NPS pollutants.
The experimental sites included locations in
Pennsylvania, West Virginia, and Tennessee.
The Leading Ridge Experimental Watershed
Research .Unit is located in the Ridge and Valley
Province of central Pennsylvania and consists of three
adjacent watersheds. BMPs used on a commercially
clearcut watershed were designed to minimize stream
sedimentation from silvicultural operations. These
practices included maintaining a 30 m buffer strip on
each side of perennial streams and restricting,slash
piles and log-landing sites'from the vicinity of stream
channels (Lynch and Corbett, 1990). A comparison of
suspended sediment concentrations on the Leading
Ridge Experimental Watersheds for the first two years
after clearcutting shows that the BMPs were effective.
Average suspended sediment concentrations of 1.7,
10.4 and 5.9 mg L'1 were reported for an uncut con-
trol watershed, a clearcut and herbicide-treated water-
shed without BMPs, and a commercially clearcut wa-
tershed with the riparian buffer strip, respectively, for
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the first year after treatment (Lynch et al., 1985; '
Lynch and Corbett, 1990). Water, yield increased fol-
lowing all clearcutting treatments. The greatest in-
crease, equivalent to 32 cm over the area cut, occurred
the first year after'clearcutting and herbicide'treat-
ment to control regrowth. Annual yield was sharply.
lower the'second year and was not statistically differ- .
ent from water yield on the control watershed at the
end of the fourth year following, harvesting (Lynch.
and (Corbett, 1990). Average suspended sediment
concentrations in the clearcut/herbicide .treatment.
jumped to 78.7 mg L'1 during the second year after
treatment compared tp 5.1 and 9.3 mg LT' for the con-
trol watershed and clearcut watersheds with riparian
buffers, respectively (Lynch et al., 1985; Lynch and
Corbett, 1990). The extremely high sediment concen-
trations on the clearcut/herbicide treatment were at-
tributed to channel cutting, and bank erosion and .
slumping on the lower portion of the channel.
Increased sediment concentrations on the commer-
cially clearcut watershed the second year (9.3 mg L'1 ,
compared to 5.9 mg L"1) were attributed to tree blow-
down along a 460 m length of intermittent stream
which did not have a buffer along it. The few remain-
ing trees'in the section blew over and loosened the
soil. The increased 'water* yield resulting from the
clearcut caused the intermittent stream to flow, per-
mitting the transport of soil to the stream- channel.
This illustrates the need for maintaining riparian
buffer strips along intermittent streams (Lyrtch et al.,
1985). , . ---/':•,••••'
Average annual' nitrate concentration in the
clearcut/herbicide watershed was substantially in-
creased compared to the control watershed (2.54 vs.
0.11 mg NOs-N L'1). Increased nitrate concentrations
combined with increased water yields from the
treated watershed resulted in significantly greater N
loading .than the control. However, rapid revegetation,
which is almost impossible to prevent in the humid,
East, generally prevented any major stream enrich-
ment problems. Where BMPs were used, nitrate con-
centrations were substantially less than the clearcut/
.herbicide treatment, although significantly higher •
than control (0.37 vs. 0.08 mg NO3-N L'1) for the first
two years after clearcut (Lynch et al., 1985).
Mulholland et al. (1990) studied an area similar to
the Valley and Ridge region of the Chesapeake Bay.
They investigated the hydfogeochemical response of
the West pork of the Walker Branch Watershed in
eastern Tennessee to four large storms. The study area
was a 38.4-ha forested watershed with.deep, highly
weathered soils, a network of ephemeral stream chan-
nels, and a spring-fed perennial stream which flowed
over dolomite bedrock in.the lower portion of the wa-
tershed. The watershed has broad ridges which slope
steeply to harrow valleys. Surface soils have very high
hydraulic conductivities due to macroporosity associ-
ated with forest soil formation processes. Reduced
hydraulic conductivities at depth in the soil are asso-
ciated with increasing clay content. The weathered
zone ranges in depth from a meter near the stream to
about 30 m at the basin divides.
In this watershed, water held above the shallow re-
strictive layer flowed through the rhizosphere and was
virtually depleted of nitrate. However, water passing
through the restrictive layer (apparently the layer was
quite leaky) had higher nitrate concentrations.
Grburidwater transferred between catchments or
leaked to deeper groundwater and discharged near the
watershed outlet bypassed the riparian zone closest to
the source of NFS pollutants. Where these transfers
occurred, groundwater was less likely to be renovated
by riparian zone processes'.
4. Conclusions
Forested riparian buffers have proven effective in
controlling water temperature and sediment delivery
to streams in forest and agricultural settings within
the Valley and Ridge. Our knowledge of groundwater
renovation in riparian ecosystems is less certain.
Where regolith is thin and bedrock controls subsur-
face flow, seepage faces or springs produce relatively
small saturated areas with wetland characteristics.
Attenuation of nitrate concentrations, may occur if
RFBS are restored in these seepage areas. In contrast,
where regolith is deep with flow-restrictive layers
near the land surface, shallow flow systems develop
on the confining layers resulting in more extensive ri-
parian ecosystems where groundwater discharges to
the streams. These conditions, more likely in the
glaciated Appalachian Plateau than in the Valley and
Ridge, are likely to have higher overall nitrate re-
moval rates. .
-------
-------
Applicability of the Three Zone
Riparian Buffer System
The three-zone RFBS specification is based on
studies of naturally. occurring riparian forests along
low order (1st to 4th order) streams and experimental
scale grass filter strips. Under natural conditions, ri-
parian forest ecosystems formed a dynamic yet stable
buffering system .along most shorelines, rivers and-
streams in the Bay watershed. Although few studies
have documented the specific changes in water qual-
ity functions during the establishment period of a ri-
parian forest, established RFBS are expected to sus-
tain water quality functions over the long term in a *
manner similar to the natural system.
The effect of upstream activities which modify hy-
drology or pollutant loads, loading rates, or the
change in functions due to management of the'RFBS,
, such as timber harvest, add uncertainty and risk to
•predicting changes in some water quality functions
over time. However, existing research, knowledge of
riparian ecology, and experience with related hydro-
logic systems can form the basis for recommenda-
tions on the applicability of RFBS. The 12 member
scientific panel that prepared this report utilized these
resoufpes to produce the following set of Best
Professional Judgements (BPJ) of conditions and cri-
teria for assessing the effectiveness of the three-zone
RFBS for use in the CBW.
A. CONTROL OF THE STREAM
ENVIRONMENT
Control of the stream environment will occur in al-
most all cases along smaller streams with Zone 1 veg-
etation. The environments of tidal streams, tidal por-
tions-of the bay, and larger, rivers maybe controlled by
other factors more than the immediate riparian
ecosystem. The consensus BPJ are:
1) Gontrol of the stream-environment..for aquatic
ecosystems is most likely to be achieved with
vegetation approximating the original native
vegetation along streams.
2) Control of the stream environment will be af-
fected less by physiographic regions than by
size of stream. As the size of stream or water
body increases, most effects of the riparian
system, on the stream environment decrease.
However, the habitat functions of large woody
debris are important even on large river banks
and on Bay shorelines.
3) Just as Zone 1 may also play an important role
in NPS pollution removal, Zone 2 may play an
important supporting role in controlling the
stream environment.
In many cases, especially along higher order
streams, quality of the stream environment will reflect
the influence of both Zone 1 and Zone 2. Sus-
tainability of Zone 1 function may depend on proper
management of Zone 2. Where windthrow of trees or
stream bank stability is a problem, Zone 2 vegetation
should be managed with long rotations, thinning cuts,
or other practices which minimize the time and areal
extent of a non-forest Zone 2. The general goal would
be to miniriiize both the amount of time and the
stream length for which Zone 1 would be the only ri-
parian forest. Where Zone 1 will function alone, in-
creased width and/or other adjustments may be re-
quired to enhance sustainability.
B. CONTROL OF NONPOINT SOURCE
POLLUTION
Unlike the processes involved :With control of the
stream environment, the functions of riparian systems
,10 control NFS pollution are dependent on hydrologic
Connection(s) of pollutant source(s) with the riparian
for,est buffer system. Although generalizations will be
made, the extent, timing, and spatial variability of the
hydrologic connections add uncertainty to BPJ, as-
sessment of NPS pollution control. The hydrologic
connection between source areas and riparian ecosys-
41
-------
42
terns probably ranges from nearly 100% of the water
moving across the surface or in shallow groundwater
through the biologically active soil zones (e.g., ICP
Watersheds) to a very low percentage of flow moving
through riparian ecosystems. This lower limit is not
well defined, but a conservative estimate can be made
by hydrograph analysis to separate storm flow from
baseflow. At a minimum, most stormflow should
move in either surface runoff or shallow groundwater
and should be subject to processing in a RFBS.
For either surface runoff or shallow groundwater,
removal of NFS pollutants in RFBS is first deter-
mined by the hydrologic pathways and then modified
by interactions of hydrology, soils, geochemical envi-
ronments, management, loading, and vegetation. As
pointed out above in Sections I and II, some of these
factors are poorly understood and most are poorly
quantified, especially outside the ICP from which
much of the existing information is derived.
As a means of conceptualizing the NFS pollution
control functions of riparian ecosystems in the CBW,
a series of flow diagrams for different physiographic
settings was developed (Fig. 6 through 14). These fig-
ures are generally representative of many of the dif-
ferent hydrologic settings within the regions and pro-
vide reference points for discussions (below)
concerning hydrologic controls on the NFS removal
function. It is important to note that these diagrams
are generalized and that more than one hydrologic
setting may be present in larger watersheds. The con-
sensus BPJ decisions are summarized for nitrate re-
moval, sediment and sediment-borne pollutant re-
moval, and phosphorus (from all sources) removal in
Tables 7 through 9.
1. Coastal Plain
a. Inner Coastal
The best information on RFBS comes from Coastal
Plain systems represented by Figure 6. In these ICP
systems, most of''the excess precipitation moves to
streams via subsurface runoff or shallow groundwater
movement. Most or all of this water moves in or near
the root zone or is subject to capillary transport due to
transpiration from the root zone. The ICP represents
one end of the spectrum of riparian ecosystems func-
tion for removal of NFS pollutants. In these systems,
riparian ecosystems exert substantial control. over
both the hydrologic and nutrient transport response of
agricultural watersheds. ICP areas, represented in
Section II by Rhode R. in Maryland and areas in
Georgia and North Carolina, are typically areas with
a high density of stream channels, well developed
"natural" riparian forests, and extensive connections
between agricultural fields and riparian forest ecosys-
tems. Most of the Western Shore and the upper
Eastern Shore Coastal Plain in the .CBW is considered
ICP Because of the relatively large amount of scien-
tific data collected from ICP type systems, primarily
in MD, NC, and GA (see Section II), the scientific
panel was able to make the most comprehensive con-
sensus BPJ for these areas. Among these conclusions
are the following:
1) Based on mass balances, established RFBS re-
move 20 to 39 kg NO3-N ha'1 yr1 from sub-
surface flow.
2) Based on mass balances, total N retention in
established systems ranges from 26 to 74 kg N
ha'1 yr1.
3) For the RFBS to be applicable in systems with
artificial drainage near streams, the drainage
system will have to be modified to work in
conjunction with the RFBS.
4) Newly established systems are likely to have a
substantial effect on subsurface nitrate loads
in (at most) 5 to 10 years if anoxic sediments
and high organic matter surface soils are al-
ready in place. By 15 to 20 years, reestab-
lished RFBS should control groundwater ni-
trate loads in most (if not all) ICP situations.
Reestablishment of RFBS along all streams in
the ICP is likely to lead to water quality im-
provements.
5)' The nitrate concentration data from ICP sys-
tems indicates that higher nitrate loadings
could be removed in the RFBS if it was ex-
posed to higher loadings than represented in
the mass balance studies. This is most likely to
be true in systems with highest denitrification
rates or potentials.
6) Based on the mass balances, net retention of
phosphorus in established systems is 1.2 to 2.9
kg P ha-1 yr1. Retention of phosphorus in sur-
face runoff appears to be mainly through re-
tention of particulate phosphorus and infiltra-
tion in the RFBS. Retention of dissolved
ortho-P appears to be considerably less effec-
tive for both surface runoff .and subsurface
flow.
-------
'43
ui
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1
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-------
46
FIGURE 6. Inner Coastal Plain flow system.
10-40m
Aquiclude
FIGURE 7. Outer Costal Plain—Well-Drained Upland flow system (from Phillips et al., 1993)
7) For a contributing area to RFBS area ratio of
about 2:1, the range of sediment and sedi-
ment-borne N and P reductions that could be
expected under worst-case conditions is about
96% for sediment, 75% for total N and 77%
for total P. Most other cases— with a 2:1 area
ratio and better upland conservation practices
—would be expected to have lower concentra-
tions leaving the RFBS. These numbers are
based on the assumption of non-channelized
flow through the RFBS.
Because of the lack of quantitative information on
RFBS functions in other hydrologic/physiographic/ •
transpiration settings, the more detailed information
from ICP settings will be used to guide quantitative
estimates for the other settings. The consensus of the
scientific panel was that the ICP data represented an
upper limit on the functions of essentially unmanaged
RFBS. Numerous management options and manage-
ment factors discussed below could lead to increases
in the effectiveness and sustainability of nonpoint pol-
lution control functions. But in general practice, with-
out depending on the management improvements, the
effects of RFBS in the ICP would be representative of
other systems in the CBW where essentially 100% of
excess precipitation moves through an unmanaged
RFBS. Where less than 100% of, excess precipitation
moves through the RFBS, the NPS pollution control
effects would be proportionally less.
b. Outer Coastal Plain
1) Well Drained Upland: Because much of the
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47
groundwater flow reaches the stream channel through
the hyporheic zone, interactions with biologically ac-
.tive soil, layers appear to be limited in the Well-
Drained Upland (Figure 7). The consensus of the BPJ
group was that for Coastal Plain systems, the WDU
represented the other end.of the spectrum,from the
ICP. Processing of groundwater-borne ]STPS pollu-
tants, including nitrate, would be least in the WDU.
Based on, the present knowledge of. these systems,
RFBS in the WDU would remove some nitrate from
groundwater. This removal function might be en-,
hahced by vegetation management, especially in the
Zone 1 area where tree roots could access groundwa-
ter discharge. Consensus decisions for the WDU are:
. 1) Where hydrologic connections between
groundwater and biologically active soil lay-,
ers are made, RFBS in the WDU should have
about the same capacity for nitrate removal as
• in the ICP areas.
2) The Zone 1 vegetation (adjacent to the stream
channel) is very important because of poten-
tial access to water 'and pollutants in the hy-
porheic zone. Zone 1 vegetation should be
managed for N uptake and for formation of
high organic matter surface soils. Provision of
leaf litter and other organic matter to the
stream channels may increase denitrification
in the channel and hyporheic zone.
3) RFBS in the WDU portion of the Coastal
Plain would have about the same capacity to
filter sediment and sediment-borne pollutants
from surface runoff as RFBS, in the ICP
4) RFBS in the WDU may have higher capacity
. for removing dissolved chemicals from sur-
face runoff because of higher available storage
' for infiltrated surface .runoff. This function is
directly related to lower water tables in the,
RFBS.
5) Reestablishment of RFBS in the WDU should
focus on headwater- streams, many of which
have been ditched. Enhancement of existing
forests along both small and large streams
should focus on control of surface runoff and
surface-borne pollutants and on management
•} of Zone 1 to intercept nitrate 'enriched ground-
water. '•
2) Poorly Drained Upland/Surficial Con-
- fined: Functions of RFBS in the Outer Coastal Plain
hydrologic systems designated Surficial Confined and
Poorly Drained Upland are thought to be intermediate,
between the"WDU and the ICP. These flow systems,
are represented in Figure 8. Specifically, the consen- ,
sus BPJ on these1 regions included the following: '
1) Potential for nitrate removal is intermediate
between WDU and ICP. Generally lower re-
gional groundwater concentrations of nitrate
will lead to lower actual removal rates and to
less important role for nitrate removal.
2) Agriculture in these regions is commonly as-
sociated with artificial, drainage which will
need to be integrated with RFBS system.
3) Potential for control of sediment and sedi-
, ment-borne chemicals should be similar to
RFBS in the ICP, but actual removal is proba-
bly less because of lower loads of surface-
borne pollutants.
4) Potential for control of dissolved chemicals in
' surface runoff may be less than in WDU be-
cause of higher water tables and generally less
available storage.
c. Tidally Influenced
Tidally influenced areas of the Coastal Plain pre-
sent unique situations for a number of reasons. First,
water and pollutants moving through the terrestrial/
aquatic interface move directly into the bay or tidal
reaches of streams, providing a direct input of pollu-
tants. Secondly, movement of groundwater through
1 these tidal systems are affected by tidal movements of
bay water which serve to restrict discharge from
freshwater aquifers. Thirdly, two main types of terres-
trial/aquatic interfaces appear to exist, especially for
groundwater fluxes. One case is a tidal stream, em-
bayment, or main stem location where a marsh system
forms a buffer at the terrestrial/aquatic interface. In
areas with marsh, the nitrate removal function of the .
RFBS is less significant due to groundwater discharge
through the marsh being stripped of nitrate in anaero-
bic marsh sediments. The second case is when the in-
terface does not include the marsh system and dis-
charge takes place from a sand aquifer directly into
the bay or tidal stream. The second case is the one that
is shown diagrammatically in Figure 9.
The nonpoint pollution control functions of RFBS
in tidally influenced areas are dependent on two fac-
tors: depth to Water table and bank stability. The in-
teraction of water table depth and nitrate removal via
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48
denitrification has been discussed extensively in pre-
vious sections. Bank stability is a major factor in
tidally influenced areas because of wave action, boat
wakes, storms, and rising sea level undermining trees
at the waters edge. It is likely that in tidal areas with
eroding shorelines, trees in a Z one 1 position will
contribute to erosion and de-stabilization.
The consensus BPJ on tidal areas of the Coastal
Plain include the following:
1) In areas without a tidal marsh at the terres-
trial/estuarine interface, nitrate removal
should be significant if the water table is
within or near the root zone of trees in Zone 2.
This removal would be both through direct
vegetation uptake and through coupling of
vegetation uptake/denitrification in surface
soil. Where the water table is consistently
below the root zone significant nitrate reduc-
tion is unlikely to occur.
2) In areas where shoreline erosion is a problem
or potential is high, Zone 1 trees at the water's
edge are likely to contribute to shoreline ero-
sion due to undermining of trees and tree fall
into tidal waters. If established in these situa-
tions, Zone 1 trees need to be put in a position
that is not likely to contribute to,active ero-
sion, cliff destabilization, or shading of
marshes.
3) Functions of Zone 3 for sediment and some
nutrient removal should be similar to function
in ICP systems.
4) Restoration of RFBS in tidal areas should con-
centrate on areas with shallow water tables, an
absence of tidal wetlands, limited shoreline
Aquiclude
FIGURE 8. Outer Coastal Plain—Poorly Drained Upland/Surficial Confined flow system.
2-1 Om
Aquiclude
FIGURE 9. Coastal Plain-Tidal Influenced flow system (based on Staver and Brinsfield, 1994).
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49
. erosion problems and in areas with substantial
surface runoff into tidal waters from adjacent
land uses.
2. Piedmont
Although substantial work on RFBS has been done
in the North Carolina Piedmont and is underway in
Piedmont areas of the CBW (See Section II), less in-
formation is available in the Piedmont than for the
Coastal Plain. The consensus BPJ of the scientific
panel was that RFBS in the Piedmont represented a
range of conditions for NFS pollution control, de-
pending on both the localized and watershed hydrol-
ogy and the proportion of excess precipitation which
moves through the. RFBS. When hydrblogic condi-
tions lead to surface runoff to streams and movement
of groundwater in or near the root zone of the RFBS,
the degree of NFS pollution control-should be similar -
to conditions measured in the North Carolina
Piedmont and potentially as effective as the ICP con-
dition. When excess precipitation moves into deeper
groundwater and into larger streams through the hy-
porheic. zone, control of groundwater pollutants such
as nitrate may be minimal. As described above, base-
flow/stormflow separations for Piedmont watersheds
should provide a conservative estimate of the quantity.
of water moving through RFBS.
The first hydrologic condition represented in" the
Piedmont is in areas with thin soils, direct flow paths
to streams^ and a large amount of water movement
through surface runoff and seepage faces (Figure 10).
These conditions are most likely in the Virginia
rPiedmont in the southern portions of the CBW. Under
these conditions, the consensus BPJ are:
1) Nitrate removal would be approximately as ef-
fective as in ICP systems. Nitrate removaj
may be more dependent on vegetation
processes because of potential for deeper root-
ing depth in more aerated soils and the poten-
tial for longer residence time- for water in
Piedmont RFBS.
2). Control of sediment and sediment-borne pol-
lutants in surface runoff should be as effective
as ICP and North Carolina Piedmont systems.
Control of sediment in surface runoff is likely
to be limited by development of concentrated'
flow channels, especially in steeper RFBS
areas of the Piedmont. These areas may re-
quire an expanded Zone 2. .
3) Control of all sources of P should be repre-
• Dented well by ICP conditions and conditions
from North Carolina studies. As in these sys-
tems, control should be more effective for sed-
iment-borne P than for dissolved P in either
surface runoff or groundwater.
Piedmont areas with deeper soils and saprolite are
likely to have longer flow paths and more water en-
tering the stream channel directly from these longer
flow paths and the hyporheic zone. These types of
Piedmont systems are represented by areas with"pri-
marily, gneiss/schist bedrock and primarily marble
bedrock (Figures 11 and 12). Areas with primarily
. schist bedrock should have substantial seepage which
should be subject.to treatment in RFBS.
For Piedmont areas represented in Figures 11 and
12, the Consensus BPJ include:
. 1) Nitrate removal would be medium in the
Piedmont areas with Schist/Gneiss bedrock
and should be used to control movement of
water in both shallow water table conditions
• and in seepage areas near streams. Nitrate re-
moval should be least important in Piedmont
areas underlain by marble because of move-
ment of groundwater and associated nitrate
, into regional aquifer systems which will
recharge larger streams. This component of
groundwater flow is likely to by-pass riparian
systems. In both systems, nitrate removal will
likely be enhanced by deeply rooted vegeta-
tion.
2) Control of sediment and sediment-borne
.chemicals will depend on management of
Zone 3 to reduce the effects of concentrated
flow and to protect reestablished forests.
Steeper slopes in riparian areas may limit both
; the sediment filtering capacity and. the reten-
tion time of water. These conditions may re-
quire expanded Zone 3 and/or .Zone 2.
3) Control of all sources of phosphorus will be
limited by ability to remove dissolved P in
surface runoff. Areas with high sediment
borne surface runoff P loads .should be re-
stored on a priority basis because of potential
for controlling these P types.
3. Valley and Ridge/Appalachian Plateau
The Valley and Ridge is represented by larger order
-------
FIGURE 10. Piedmont (thin soils) flow system.
Bedrock
FIGURE 11. Piedmont (schist/gneiss, bedrock) flow system.
T
10-30m
FIGURE 12. Piedmont (marble bedrock)/Valley and Ridge (limestone bedrock) flow system.
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51
streams draining the main valleys with either lime-
stone bedrock (Figure 12) or shale/sandstone bedrock
(Figure 13) and by smaller order streams draining the
ridges (Figure 14). The situation for^sediment and P
sources is thought to be similar to the Piedmont hy-
drologic settings. Nitrate removal will probably show
the most variability among the different Valley types
and with different valley configurations and flood-
plain extent. Consensus BPJ for larger order streams
in the Valley and Ridge for nitrate removal functions ,
are:
!)• Valley and Ridge'areas with limestone
bedrock (Figure 12) will have the least poten-
tial for nitrate removal due to most flow going
into regional aquifers which are intercepted
primarily by major rivers. Seepage areas,
springs, and floodplains will have'the most
potential for nitrate removal1! Deep rooted veg-
etation should be used io control nitrate in
v these areas. " • -
'.^-, .y • • '•''.' ~ • • •
2) Valley and Ridge with sandstone/shale <
bedrock (Figure 13) will have more potential
; for nitrate removal due to less movement of
groundwater and nitrate into regional aquifers
; and the importance and prevalence of seepage
areas, moving nitrate into biologically active
soil horizons.-The processing of nitrate is con-
trolled by the presence and size of the flood-
plain arid by the presence of seepage areas and
springs. As in other Piedmont and Valley and
Ridge settings, deep-rooted vegetation should
be used to maximize the potential for N up-
, take. ' •'.;'.
3) Nitrate removal from low order streams in
both Valley and Ridge and Appalachian
Plateau (Figure 14) settings will depend on
residence time of water arid the presence of
Bedrock
FIGURE 13. Valley and Ridge (sandstone/shale bedrock) flow system.
10-30m
FIGURE 14. Valley and Ridge/Appalaehian (low order streams) flow system (based on Mulhollandet al, 1990).
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52
seeps and floodplains. In these cases, as in
other situations without extensive wetlands,
the use of deeply rooted vegetation should en-
hance nitrate uptake. Because of the limited
extent of riparian systems in areas of high re-
lief, Zone 1 will important for nitrate removal
in these smaller streams.
C. LOADING RATES AND NONPOINT
SOURCE POLLUTION CONTROL
As a nonpoint pollution control practice, Riparian
Forest Buffer Systems represent a long-term invest-
ment which can change the structure of the agricul-
tural landscape. As a long-term management option, it
is quite likely that RFBS will be exposed to a wide
range long-term management option, it is quite likely
that RFBS will be exposed to a wide range of pollu-
tant loadings due to both interannual variation, and
changes in management practices in source areas.
Information on how mature RFBS respond to chang-
ing pollutant loads is essential to understanding long
term sustainability of RFBS.
As discussed above and in Section II, research on
some ICP systems indicates that higher rates of nitrate
removal would be possible under higher loadings of
nitrate. Published studies indicate that this is most
likely to be true in areas where denitrification is the
primary means of nitrate removal. Given the range in
nutrient uptake observed both among different plant
species and within the same plant species, it is likely
that vegetation uptake will increase with increasing
loads, if there is significant hydrologic interaction
with vegetation.
Increasing loads of P are likely to be less effec-
tively controlled than increasing loads of N, because
of the lack of biological processes to remove or se-
quester P in the RFBS. If increasing P loads are to be
controlled, it will require both effective.management
of Zones 3 and 2 for sediment removal and manage-
ment of Zone 2 for infiltration. If dissolved or partic-
ulate P can be retained in the root zone, it will be
available for both biological and chemical removal
processes. If RFBS have some absolute removal po-
tential for P, reducing input loads should increase the
efficiency of removal.
Management to control increasing loads of sedi-
ment and sediment-borne chemicals will require spe-
cific management of Zones 3 and 2 for sediment re-
tention. As described above in Sections I and II, most
of the mass of sediment will be deposited in Zone 3
and most of the sediment-borne nutrients will be de-
posited in Zone 2. Increased sediment loadings to
Zone 3 will require increased management to elimi-
nate concentrated flows, remove accumulated sedi-
ment especially in berms, and restore the herbaceous
vegetation. Increased sediment and sediment-borne
chemicals to Zone 2 should lead to higher amounts of
chemical deposition in surface litter. As with other
dissolved P in surface runoff, the ability of Zone 2 to
retain P may be limited, especially under high load-
ings of dissolved P.
Loading rate/buffer width relationships are only
poorly defined, especially for dissolved pollutants. In
published studies with water clearly in contact with
surface litter or the biologically active root zone,
buffers of about 100 feet have been effective for at
least sediment and nitrate removal. One of the diffi-
culties in describing these relationships is that in-
creasing pollutant loads may also be accompanied by
increasing water volumes in either'surface runoff,
groundwater, or both. In the presence of increased
water movement^ denitrification for nitrate removal
should be enhanced and sedimentation and infiltration
may be decreased. Increased surface runoff and load-
ing-of sediment and sediment-borne chemicals can be
accommodated by management of Zones 3 and 2 to
increase roughness and control channelized flow.
Although mass balance approaches, used in Section II
may be extrapolated to higher loading rates, they pro-
vide only an estimate and may not predict real-world
responses.
D. STREAM ORDER/SIZE
Regardless of the size of stream or the hydrologic
setting, water moving across the surface or through
the root zone of a RFBS should show reduction in ei-
ther nitrate (groundwater) or sediment and sediment-
borne chemical loads reaching the stream. As streams
increase in size, the integrated effects of adjacent ri-
parian ecosystems should decrease relative to the
overall water quality of the stream. On lower order
streams there is greatest potential for interactions be-
tween water and riparian areas. For NPS pollution
control, the change in impact of RFBS as stream order
increases can be estimated based on hydrologic con-
tributions from upstream and from' the riparian
ecosystem. For first-order streams, the potential im-
pact of the RFBS on chemical load or flow-weighted
concentration is directly proportional to the propor-
tion of the excess precipitation from the contributing
area which moves through or near the root zone or
surface of the RFBS. For all streams above first order,
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53
the contributing area is only one source of pollutants,
with upstream reaches providing the other source. For
second-order and above, the NFS pollution control
function of a given RFBS is based on both the pro-
portion of water from the contributing/area which
moves through the riparian system and the relative
sizes of the two potential pollutant loads - upstream
sources or adjacent land uses.. Clearly, the larger the
stream, the less" impact a RFBS along a particular.
stream reach can have on reduction in overall load
within that reach. If there are no RFBS upstream from
a particular stream reach, the water entering the
stream reach is likely to be already contaminated.
On a watershed basis, the higher the proportion of
total streamfiow originating from relatively short
flow-paths to small streams, the larger the potential
impact of RFBS. In comparing the potential effective-
ness of RFBS among watersheds, drainage density
(length of channel per unit area of watershed) should
provide a useful starting point. Higher drainage den-
sity implies greater potential importance for RFBS in
NFS pollution control.
Control of the stream environment is most effective
when native vegetation forms a complete canopy over
the,stream. This is obviously only possible on rela-
tively ;Small streams. The effect of the RFBS on the
stream environment is not simply proportional to the
amount -of the channel which is shaded As noted
above in Section I, besides direct shading of the
stream channel, cooling of groundwater recharging
streams and provision of bank habitat will occur even
on larger streams. Bank habitat, provision of coarse
woody debris and provision of leaf detritus remain
important functions, regardless of stream size.
E. ESTABLISHMENT AND
SUSTAINABILITY
Some aspects of establishment are discussed
above. RFBS should be used as part of an integrated
land management or conservation system which con-
sists of 1) watershed scale management, 2) NFS pol-
lution management, and 3) active management of the
RFBS. In this way, RFBS become part of conserva-
tion, stormwater, nutrient and farm management, tim-
ber harvest, and other land management planning ef-
forts.
Watershed.management is essential to reduce over-
all pollutant loadings and integrate the riparian area as
part of a landscape influenced by upstream hydrology.
In a landscape context, RFBS which mimic the nat-
ural ecosystems of the area will increase the likeli-
hood of long-term sustainability. Consideration, of ex-
isting riparian forests and linkage of RFBS as contin-
uous stream 'corridors is desirable. Source manage-
ment and land conservation measures are important in
conserving natural resources, reducing overall pollu-
tion, and limiting stress on the RFBS. These mea-
sures, along with maintenance of buffer plantings, are
especially important during the establishment phase
and in preventing excessive runoff or sediment and
nutrient loading beyond the capacity of the buffer.
RFBS management such as periodic harvesting,
runoff control maintenance, control of invasive plants,
etc., is desirable to maximize performance and ensure
long term effectiveness..Continued runoff control and
protection of Zone 1 functions are essential to main-
taining optimum performance in RFBS.
• Integration of,RFBS within land management
helps to prevent some of the primary reasons for
"acute" failure such as runoff inputs which exceed the
design of.the RFBS and cut gullies or channels, or
failure to address "chronic" problems such as a grad-
ual decrease in phosphorus retention. Where gullies
have formed into or through riparian forests, mea-
sures other than flow-spreading in Zone 3, will be"
necessary to control channelized flow. Because of the
commitment of land required for RFBS establish-
ment, the approaches used for establishment and sub-
sequent management should contribute to a RFBS
which is sustainable for decades.
At least one sustainability .question has been raised
relative to each zone of the RFBS. The major sustain-
ability question for Zone 3, discussed in .Section I,
above is the need to remove accumulated sediment
and reestablish herbaceous vegetation periodically.
Functions of Zone 3 should be sustainable given
proper management of the sediment and vegetation.
The other two sustainability questions are closely re-
lated to Zones 2 and 1. In most cases, the sustainabil-
, ity of Zone 1 functions will depend on having a Zone
. 2 which is harvested infrequently. Biomass planta-
tions which require frequent coppicing of trees or
grassed Zone 2 areas are likely to expose Zone 1 veg-
etation to catastrophic failure due to blow down of
.trees. Zone 2 functions,-if dependent on particular
types of vegetation, such as deep-rooted species or
vegetation specific for high levels of nutrient uptake,
will require some management to control invasive
species.
-------
-------
Research Needs
Research needs are grouped into four general ob-
jectives: 1) assessment of existing riparian forest
ecosystems relative to the minimum RFBS standards;
2) assessment of the potential for RFBS restoration
areas to control NFS pollution; 3) assessment of ef-
fectiveness of NFS pollution removal in pilot restora-
tion and enhancement areas; and 4) determination of
the effects of management factors on the NFS pollu-
tion control functions of restored and enhanced
RFBS, Ideally, objectives 1 and 2 would be completed
as guidance for pilot restoration and enhancement
studies or large scale research projects which would .
be u'sed as the basis to achieve' objectives 3 and 4. If
ongoing assessment work related to these first two ob-
jectives is done in a timely manner, it will provide
substantial guidance to achieve, objectives 3 and 4.
The assessment and evaluation of existing riparian
forest ecosystems will require the use of remotely
sensed data for delineation and classification of ripar-
ian forests. Significant progress has been made in as-
sessment of the forest resources of Maryland in a
"Comprehensive Forest Resources Inventory for the
State of Maryland" undertaken by the Maryland Dept.
of Natural Resources (Lade, 1994). The objective of
this project was to use Thematic Mapper data 'to cre-
ate maps, statistical summaries and digital data sets to
describe the location and extent of forest (especially
-streamside forests) in the state of Maryland. The
study was designed for delineation of all forest re-
sources with a minimum mapping unit of 1 acre and a
minimum mapping unit of 100 feet for linear forest
areas associated with streams. The data are then used
in a Geographic Information System to characterize
the extent and types or absence of forest in 30 feet arid
100 m riparian buffers. This was done to explicitly as-
sist in the identification of potential riparian forest
buffer restoration sites for the entire state. The char-
acterization of linear forest should be done at a finer
resolution (10 to 20 m) in order to delineate riparian
forest buffers of the width recommended in the RFBS
specification. Data for these narrower linear forests
are needed for the entire CBWin order to character-
ize the riparian forest resources and forest cover in ri-
parian'areas. ' , '
One use of the forest inventory will be to overlay
other digital layers for further analysis of the relation-
ship of riparian forest buffers to other landscape char-
acteristics (Lade, 1994). The classification scheme
developed here could be used as the starting point for
an assessment of the potential for existing, enhanced,
or restored RFBS to intercept surface runoff or sub-
surface borne NFS pollutants. Refinement of the clas-
sification scheme based on existing and new geohy-
drology data could be used to.produce basin .wide
maps of the relative potential for control of surface
and subsurface borne pollutants. These maps, overlain
with maps of riparian buffer vegetation and other data
layers such as wetland soils would make it possible to
quantify the riparian areas with different potential for
NFS pollution control which-Were available for
restoration on a subwatershed basis. Research sum-
marized in this report, as well as forthcoming research
results, could be used in conjunction with the mapped
information to make quantitative or comparative esti"
mates of the amount of NFS pollution reduction rela-
tive to the load reduction goals set for the Chesapeake
Bay.
.Concurrent with development of a mapping ap-
proach is the need to make field assessments of the
potential for hydrologic interaction between nonpoint
pollutant sources and the RFBS. Reliable indicators
of the degree of interaction between groundwater/sur-
face water and the RFBS will be necessary when
making .field/farm/subdivision/ or watershed scale as-
sessments of nonpoint pollution control potential.
Streamflow data from USGS and other sources could
be used to assess stormflow/baseflow proportions as a
screening technique. Watersheds with higher propor-
tions of stormflow could be targeted for more inten-
sive reconnaissance investigations to .determine the
potential applicability of RFBS.
The outputs of objectives 1 and 2 should be used to
55
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56
guide the establishment of pilot restoration and en-
hancement projects and large scale research projects.
Only limited objectives can be accomplished in RFBS
restoration and management research conducted at
the small scales normally associated with agricultural
plot research. A number of hypothetical examples can
be used to show the potential and the limitations of
small scale research in RFBS. For instance, small
plots are being used to examine the effects of vegeta-
tion management on surface runoff spread evenly
through a restored RFBS. These same small plots can-
not be used for the study of effects of concentrated
flow on the filtering capacity of the RFBS. Similarly,
the effects of vegetation management on subsurface
flow cannot be studied on small plots. The minimum
size for plots to look at long term effects of clear-cut-
ting Zone 2 vegetation is constrained by the ability of
trees in adjacent reference areas to put roots into the
clear-cut areas which will affect the hydrology of both
reference and clear-cut areas.
The ideal scale to .accomplish RFBS research
should be based on the land uses contributing non-
point pollution and the hydrology of the system. It
may be necessary to conduct work at the watershed
scale where accurate streamflow gaging data can be
used to assess the effects of RFBS on watershed re-
sponses over time. At a minimum, the scale of re-
search is probably that of the representative hillslope.
Ideally, integrated research programs at a number of
spatial scales would be pursued simultaneously. For
instance, a number of hillslope studies with different
nonpoint pollution sources might be conducted in one
watershed. The hillslopes studies could be used to: 1)
examine the effects of, differing pollutant
loads/sources on similar RFBS; 2) examine the effects
of differing RFBS management on similar pollutant
loads, or a combination of the two approaches; or 3)
examine the effects of differing hydrologic conditions
on RFBS functions. These research projects should
include sub-objectives to understand the processes re-
sponsible for removal of NFS pollution. At the same
time, restoration or enhancement of RFBS for signif-
icant portions of entire subwatersheds could provide
for a comparison with,other watersheds without the
RFBS restoration/enhancement.
The above discussion of objective 3 amounts to an
argument for the integration of research and demon-
stration projects on RFBS in the CBW and elsewhere.
The advantages to the research programs are in both
the ability to conduct research at the appropriate
scales and the ability to relate the research to "real-
world" restorations. The advantage to the demonstra-
tion or operational restoration and/or enhancement
project is the potential for direct quantification of the
water quality benefits of RFBS in different land
use/hydrologic/buffer management settings.
A long list of sub-objectives is possible for objec-
tives 3 and 4 of a general research program. Among
the potential research topics are: 1) effects of vegeta-
tion type and management on sustainabiliry of RFBS;
2) effects of vegetation type and management on NFS
pollution control by RFBS; 3) effects of chronic stres-
sors such as long-term N loading and N-saturation on
NFS pollution control; 4) effects of acute stressors
such as large storms or extremes in temperature or
growing season rainfall. For any given size and loca-
tion of RFBS, the actual degree of NFS pollution con-
trol may be dependent on management factors.
Although the existing research provides little guid-
ance in this area, management factors are likely to
help control the effects of both chronic and-acute
stressors. ,
A viable approach to these research needs "would
be to continue funding for the assessment and map-
ping work under objectives 1 and 2 while develop-
ing/enhancing the coordination between institutions
and individuals involved with pilot programs and
demonstration projects and institutions and individu-
als interested in pursuing research associated with
these projects.
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